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Limnetica 25(1-2)01
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Volumen 25. Número 1-2. 2006
LIMNETICA
Revista de la
Asociación Española de Limnología
The ecology of the Iberian inland waters:
Homage to Ramon Margalef
Editores
Joan Armengol (Universitat de Barcelona)
Ramon Moreno-Amich (Universitat de Girona)
Antoni Palau (Universitat de Lleida)
Con la colaboración de:
UNIVERSITAT DE BARCELONA
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LIMNETICA
Vol. 25(1-2), 2006
GRAÇA, MANUEL A. S., AND CRISTINA CANHOTO Leaf litter processing in low order streams ..............................................
SERRANO, L., M. REINA, G. MARTÍN, I. REYES, A. ARECHEDERRA, D. LEÓN, AND J. TOJA The aquatic systems of Doñana
(SW Spain): watersheds and frontiers ............................................................................................................................
RUEDA VALDIVIA, FRANCISCO Basin scale transport in stratified lakes and reservoirs: towards the knowledge of freshwater
ecosystems .....................................................................................................................................................................
SOUSA, ARTURO, LEONCIO GARCÍA-BARRÓN, JULIA MORALES, AND PABLO GARCÍA-MURILLO Post-Little Ice Age warming
and desiccation of the continental wetlands of the aeolian sheet in the Huelva region (SW Spain) ...............................
GARCÍA MURILLO, P., R. FERNÁNDEZ ZAMUDIO, S. CIRUJANO, AND A. SOUSA Aquatic macrophytes in Doñana protected
area (SW Spain): An overview .......................................................................................................................................
MARTÍNEZ -ABAIGAR, JAVIER , ENCARNACIÓN NÚÑEZ -OLIVERA , M ARÍA ARRÓNIZ-C RESPO, R AFAEL TOMÁS, NATHALIE
BEAUCOURT, AND SAÚL OTERO Effects of ultraviolet radiation on aquatic bryophytes ...................................................
GUERRERO, FRANCISCO, GEMA PARRA, FRANCISCO JIMÉNEZ-GÓMEZ, CARLOS SALAZAR, RAQUEL JIMÉNEZ-MELERO, ANDREA
GALOTTI, ENRIQUE GARCÍA-MUÑOZ, Mª LUCÍA LENDÍNEZ, AND FERNANDO ORTEGA Ecological studies in Alto Guadalquivir
wetlands: a first step towards the application of conservation plans .................................................................................
ÁLVAREZ-COBELAS, MIGUEL Groundwater-mediated limnology in Spain ............................................................................
ELOSEGI, ARTURO, ANA BASAGUREN, AND JESÚS POZO A functional approach to the ecology of Atlantic Basque streams ...
SORIA GARCÍA, J. M. Past, present and future of la Albufera of Valencia Natural Park .......................................................
BÉCARES, ELOY Limnology of natural systems for wastewater treatment. Ten years of experiences at the Experimental Field
for Low-Cost Sanitation in Mansilla de las Mulas (León, Spain) ..................................................................................
CASAS, J. JESÚS, MARK O. GESSNER, PETER H. LANGTON, DEMETRIO CALLE, ENRIQUE DESCALS, AND MARÍA J. SALINAS Diversity
of patterns and processes in rivers of eastern Andalusia.......................................................................................................
MORALES-BAQUERO, RAFAEL, ELVIRA PULIDO-VILLENA, OTILIA ROMERA, EVA ORTEGA-RETUERTA, JOSE Mª CONDE-PORCUNA,
CARMEN PÉREZ-MARTÍNEZ, AND ISABEL RECHE Significance of atmospheric deposition to freshwater ecosystems in
the southern Iberian Peninsula .......................................................................................................................................
CASAMITJANA, XAVIER, JORDI COLOMER, ELENA ROGET, AND TERESA SERRA Physical Limnology in Lake Banyoles...........
CARRILLO, PRESENTACIÓN, JUAN MANUEL MEDINA-SÁNCHEZ, MANUEL VILLAR-ARGAIZ, JOSÉ ANTONIO DELGADO-MOLINA,
AND FRANCISCO JOSÉ BULLEJOS Complex interactions in microbial food webs: Stoichiometric and functional approaches...
MORENO-OSTOS, E., L. CRUZ-PIZARRO, A. BASANTA-ALVÉS, C. ESCOT, AND D. G. GEORGE Algae in the motion: Spatial
distribution of phytoplankton in thermally stratified reservoirs .....................................................................................
TORO, MANUEL, IGNACIO GRANADOS, SANTIAGO ROBLES, AND CARLOS MONTES High mountain lakes of the Central Range
(Iberian Peninsula): Regional limnology & environmental changes ..............................................................................
DE VICENTE, I., V. AMORES, AND L. CRUZ-PIZARRO Instability of shallow lakes: A matter of the complexity of factors involved
in sediment and water interaction? .................................................................................................................................
PARDO, ISABEL, AND MARUXA ÁLVAREZ Comparison of resource and consumer dynamics in Atlantic and Mediterranean
streams ...........................................................................................................................................................................
PALAU, ANTONI Integrated environmental management of current reservoirs and regulated rivers......................................
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DE LIMNOLOGÍA
The ecology of the Iberian inland waters:
Homage to Ramon Margalef
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ISSN 0213-8409
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ASOCIACIÓN ESPAÑOLA
DE
LIMNOLOGÍA
Presidencia:
Vicepresidencia:
Secretaría:
Tesorería:
SERGI SABATER. Girona
JULIA TOJA. Sevilla
JUAN MIGUEL SORIA. Valencia
EUGENIO RICO. Madrid
LIMNETICA
LIMNETICA es una revista internacional publicada por la Asociación Española de Limnología.
Editor
JOAN ARMENGOL
Editores adjuntos
ISABEL MUÑOZ
Comité editorial
J. Alba Tercedor. Granada
M. J. Boavida. Lisboa, Portugal
X. Casamitjana. Girona
J. Catalán. Barcelona
G. George. Lancaster, Reino Unido
H. L. Golterman. Francia
M. A. S. Graça. Coimbra, Portugal
C. Granado. Sevilla
B. Malmqvist. Umea, Suecia
L. Naselli-Flores. Palermo, Italia
A. Palau. Lleida
C. Pedrós-Alió. Barcelona
D. Planas. Montreal, Canadá
J. Pozo. Bilbao
N. Prat. Barcelona
A. Quesada. Madrid
A. Rodríguez Capítulo. La Plata, Argentina
K. Simek. Ceske Budjovice, Rep. Checa
J. G. Tundisi. Sao Carlos, Brasil
E. Vicente. Valencia
H. Zagarese. Chascomús, Argentina
Secretaria de Redacción
JAIME ORDÓÑEZ
CARLOS COVIELA
Toda correspondencia relativa a la ASOCIACIÓN ESPAÑOLA DE LIMNOLOGÍA y a la revista LIMNETICA, incluida la
petición de altas, bajas, intercambios, suscripciones y ejemplares atrasados debe dirigirse a la Secretaria de la Asociación
Española de Limnología, C/ Los Ángeles, 33, 46920-Mislata (Valencia). Web: http://www.aelimno.org/Limnetica.htm.
Los manuscritos de trabajos científicos para su publicación en LIMNETICA deben ser enviados a Joan Armengol,
Departament d'Ecologia. Facultat de Biologia. Universitat de Barcelona. Av. Diagonal, 645. 08028-Barcelona.
Limnetica está indexada en las siguientes bases de datos:
Aquatic Sciences and Fisheries Abstracts (ASFA); Zoological Record of BIOSIS® database; Freshwater Biological
Association (FBA); NISSC’s FISHLIT database; Sistema de Información en Línea para Revistas Científicas de América Latina,
Caribe, España y Portugal (LATINDEX); Library of Natural Sciences of Russian Academy of Science (LNS); Indice Español
de Ciencia y Tecnología (ICYT).
QUINTANA, X. D., D. BOIX, A. BADOSA, S. BRUCET, J. COMPTE, S. GASCÓN, R. LÓPEZ-FLORES, J. SALA, AND R. MORENO-AMICH
Community structure in mediterranean shallow lentic ecosystems: size-based vs. taxon-based
approaches......................................................................................................................................................................
MORENO-AMICH, RAMON, QUIM POU-ROVIRA, ANNA VILA-GISPERT, LLUÍS ZAMORA, AND EMILI GARCÍA-BERTHOU Fish ecology
in Lake Banyoles (NE Spain): a tribute to Ramon Margalef ..........................................................................................
SABATER, SERGI, HELENA GUASCH, ISABEL MUÑOZ, AND ANNA ROMANÍ Hydrology, light and the use of organic and inorganic
materials as structuring factors of biological communities in Mediterranean streams ...................................................
ENCINA, L., A. RODRÍGUEZ, AND C. GRANADO-LORENCIO The Iberian ichthyofauna: Ecological contributions....................
GUISANDE, CÁSTOR Biochemical fingerprints in zooplankton .............................................................................................
PRENDA, J., M. CLAVERO, F. BLANCO-GARRIDO, A. MENOR, AND V. HERMOSO Threats to the conservation of biotic integrity
in Iberian fluvial ecosystems..........................................................................................................................................
GONZÁLEZ DEL TÁNAGO, MARTA, AND DIEGO GARCÍA DE JALÓN Attributes for assessing the environmental quality of
riparian zones .................................................................................................................................................................
LÓPEZ-RODAS, V., E. MANEIRO, AND E.COSTAS Adaptation of cyanobacteria and microalgae to extreme environmental
changes derived from anthropogenic pollution...............................................................................................................
FERREIRA, M. TERESA, AND FRANCISCA C. AGUIAR Riparian and aquatic vegetation in Mediterranean-type streams
(western Iberia) ..............................................................................................................................................................
VASCONCELOS, VITOR Eutrophication, toxic cyanobacteria and cyanotoxins: when ecosystems cry for help ......................
FERNÁNDEZ ALÁEZ, CAMINO, MARGARITA FERNÁNDEZ ALÁEZ, CRISTINA TRIGAL DOMÍNGUEZ, AND BEATRIZ LUIS SANTOS
Hydrochemistry of northwest Spain ponds and its relationships to groundwaters..........................................................
CAMACHO, ANTONIO On the occurrence and ecological features of deep chlorophyll maxima (DCM) in Spanish stratified
lakes ...............................................................................................................................................................................
OLIVEIRA, S. V., AND R. M. V. CORTES Environmental indicators of ecological integrity and their development for running
waters in northern Portugal ............................................................................................................................................
LOPEZ, PILAR, ENRIQUE NAVARRO, RAFEL MARCE, JAIME ORDOÑEZ, LUCIANO CAPUTO, AND JOAN ARMENGOL Elemental
ratios in sediments as indicators of ecological processes in spanish reservoirs ..............................................................
MARTI, E., F. SABATER, J.L. RIERA, G.C. MERSEBURGER, D. VON SCHILLER, A. ARGERICH, F. CAILLE, AND P. FONOLLÀ Fluvial
nutrient dynamics in a humanized landscape. Insights from a hierarchical perspective .................................................
MARCÉ, RAFEL, ENRIQUE MORENO-OSTOS, JAIME ORDÓÑEZ, CLAUDIA FEIJOÓ, ENRIQUE NAVARRO, LUCIANO CAPUTO, AND
JOAN ARMENGOL Nutrient fluxes through boundaries in the hypolimnion of Sau reservoir: expected patterns
and unanticipated processes ...........................................................................................................................................
PRAT, NARCÍS, AND MARIA RIERADEVALL 25-years of biomonitoring in two mediterranean streams (Llobregat and Besòs
basins, NE Spain) ...........................................................................................................................................................
CATALAN, JORDI, LLUÍS CAMARERO, MARISOL FELIP, SERGI PLA, MARC VENTURA, TERESA BUCHACA, FREDERIC BARTUMEUS,
GUILLERMO DE MENDOZA, ALEXANDRE MIRÓ, EMILIO O. CASAMAYOR, JUAN MANUEL MEDINA-SÁNCHEZ, MONTSERRAT
BACARDIT, MADDI ALTUNA, MIREIA BARTRONS, DANIEL DÍAZ DE QUIJANO High mountain lakes: extreme habitats and
witnesses of environmental changes...............................................................................................................................
MIRACLE, M. R., B. MOSS, E. VICENTE, S. ROMO, J. RUEDA, E. BÉCARES, C. FERNÁNDEZ-ALÁEZ, M. FERNÁNDEZ-ALÁEZ,
J. HIETALA, T. K AIRESALO, K. VAKKILAINEN, D. STEPHEN, L. A. H ANSSON & M. G YLLSTRÖM Response of
macroinvertebrates to experimental nutrient and fish additions in European localities of different latitudes.................
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© Asociación Española de Limnología
Depósito Legal:V-2404-1986
ISSN: 0213-8409
Autoedición: Servei Gràfic NJR, SL
Impresión: Gráficas Rey, S.L.
Impreso en España / Printed in Spain
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Ramon Margalef (1919-2004): teacher and researcher
Joan Armengol
Dept. Ecology.
Univ. Barcelona
The May 23th, 2004 Professor Ramon Margalef died in Barcelona at the age of 85. While not unexpected, his death equalled his life in simplicity and dignity. The professor had refused to be subjected
to a treatment that could artificially prolong his life, wholly in keeping with the tenor of his character. Already in 1979, Margalef presented some very interesting thoughts, still valid today, on r- and
K- strategy behaviours amongst human populations, the generational problem and the lengthening of
life span in some human populations in the article “El precio de la supervivencia. Consideraciones
ecológicas sobre las poblaciones humanas” (Margalef, 1979). In it, there is a sentence which has kept
its full force over time, considering the circumstances that led to his death. I remember it quite clearly as, even back then, I found it profoundly disturbing and, quoting from memory, it goes something like:” I would not like to enjoy the privileges medicine granted to Franco and Tito”. It looks to
me as if this sentence were what we call a living will “avant la lettre” and, in it as in so many respects, professor Margalef was way ahead of his time. However, I would not like to dwell on this subject which leads me to very painful recent memories, but to write about his life as a teacher and researcher at the University of Barcelona from the perspective of one of his pupils who got introduced the
world of ecology by the hand of professor Margalef and lived side by side with him during part of his
“golden years” of scientific research.
Margalef was not especially didactic as a teacher, at least not for those who preferred well organised lectures that allowed the taking of clear and methodical notes, with outlines to complement the
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explanations in class. We need not forget that until 1974 we cannot talk of a “book on ecology that
could complement his lectures”. Until the publishing of Ecología (Margalef, 1974), we felt fortunate
enough if we had a barely readable cyclostyled copy in thick paper of “Comunidades naturales”
(1962), a compilation of some of the lectures he had given at a course on ecology in Puerto Rico
which had been published in an unconventional way. His lectures would be a continuous improvisation and, even though there was a well-outlined thread in the syllabus, you could hardly follow it
during the lecture on any given day. He followed, or better, pretended to follow some notes he had
scribbled on one of those index cards we used to write down bibliographical notes. During the class,
however, he would keep bringing up new topics to end up talking about the ones that were of his
interest at that moment. More than once, you would write down that there were different ways of
facing a specific aspect of an issue only to realise that he expanded on one of them and forgot about
the rest. They could of course have been dealt with, or not, but that didn’t seem to worry him at all.
But please do not reach false conclusions from my previous words: his lectures were speeches in
the creative sense of the word as he was giving us a state-of-the-art account of key aspects of contemporary ecology, continually updated, as he was leading it himself together with E. Hutchinson,
R. H. McArthur, the Odum brothers (Tom and Eugene) and R. Lewontin, amongst others. In his
classes, he would bring up the latest books and articles from the latest issues from the most prestigious magazines and he would use them as the backbone of the lesson. I clearly remember as one
time, during a lesson on marine plankton, he got sidetracked into talking about a most interesting
book he was reading at the moment and about biology of leaves and he started to argue on how
many times the surface of the earth could be covered if all the leaves were put one right next to the
other one. A kind of biospheric foliar index which led to his reflecting on the idea of why life had
not evolved towards one unique species that would cover the whole surface of the Earth, with an
autotrophic top layer and an heterotrophic bottom layer, and he even predicted that its thickness
would have to be no more than a few millimetres at the most, enough so that there would be a
redox potential difference between layers, enough to balance production with respiration. This idea
of a planet covered by just one species was the complete antithesis of the concept of biosphere but
he used it to stimulate our thinking about what the biogeochemical cycles would be like within a
system with no diversity, little biomass, but possibly a lot more efficient in the capture of energy
from sunlight through photosynthesis. Margalef underwent cataract surgery in the days before
laser surgery and with techniques that were a lot more invasive and required several days in hospital, and therefore we can imagine what it meant for him to spend those days with the eyes bandaged and with nothing else to do but to meditate on some of his favourite subjects. He asked for a
cassette player to be brought to him and he recorded a story about a human expedition to a planet
that fulfilled the requirements mentioned above, too long to relate now. Unfortunately, the recording is lost, even though it would nowadays be more relevant as a testimony of Margalef ’s personality rather than for the subject itself. What we can infer from these anecdotes is that Margalef enjoyed these kind of theoretical approaches similar to Einstein’s Gedankenexperimente and the ones
by other physicists of his time, although they were not quite the same. I am referring to experiments whose realization is frequently impracticable but which nevertheless lead to reliable results.
In Margalef ’s case, these mental experiments were not merely theoretical, but were based on a
deep and perceptive observation of nature, on simple experiments and the application of regularities he had observed in nature that were based on ecological successions. For Margalef, perfect
crime didn’t exist even in nature and the observation of natural phenomena allowed him to detect
casual linkages that led him to discover principles that had gone unnoticed until then. That’s why
Margalef had always regarded himself as a naturalist. “He dignified the meaning of naturalist”,
wrote Joandomènec Ros (Ros, 2004) not too long ago to recall Margalef ’s passion for nature, and
Margalef himself preferred this term to all others to describe his scientific activities. For this rea-
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son, some authors have adopted Josefina Castellví’s views that “talking about ecology is talking
about Margalef, but talking about Margalef certainly implies a lot more than talking about ecology”. With these words, “more than ecology”, we mean the observation and study of nature along
with deep intellectual interests.
I do not want to expand on the emerging principles of ecology that Margalef developed together
with the most prestigious ecologists of his time, Hutchinson, McArthur and the Odums amongst
others, or on his lifelong contributions to theoretical ecology as they have been described in detail by
other authors (Bascompte and Solé, 2005; Flos, 2005; Walter, 2005). However, I would like to
emphasize that, in my opinion, the most relevant article published by Margalef is “On certain unifying principles in ecology” (Margalef, 1963). Very few times more has been said with fewer words.
In this paper, Margalef presented a series of emerging principles based on the ecological succession
and with them he started dissecting nature. In other words, he started to study and measure all ecosystems, from the least productive seas, such as the Mediterranean, to fertile ones like the Sahara
upwelling. Likewise, the Mediterranean forest, the rainforest, the small pond, the biggest lakes or
dams, the coral reefs or caves, they all became the subject of his studies. Nothing escaped his ability
to discern patterns and the results were spectacular. The best of his comparisons can be found in
“Perspectives in ecological theory” (Margalef, 1968), where we are able to realize how powerful the
tool he had created was. No wonder this book is one of the top 10 most cited works in ecology and is
fully up-to-date. Just to mention a few examples that are far from exhaustive of the application of
these emerging principles: Margalef deduced that the natural evolution of lakes was from eutrophic
to oligotrophic aquatic systems if the influx of nutrients or organic matter was cut off (Margalef,
1968). He also explained the dynamics of a river population as an equivalent to space succession
(Margalef, 1960) and the seasonal dynamics of phytoplankton as a microsuccession (Margalef,
1978). The direct consequence of this last idea led him to develop the concept of biological types of
phytoplankton as an adaptation of the species to a double gradient of concentration of nutrients and
of turbulent kinetic energy, with his famous mandala model (Margalef, 1980). From those research
topics he developed the concept of external or exosomatic energy and its relevance in the organisation of communities. Societies or systems that use more exosomatic energy are the ones that exploit
or dominate the other ones. I would suggest a “Gedankenexperimente” to you and to apply this thesis
to the present geopolitical situation for the control of the non-renewable natural resources and reach
your own conclusions. Margalef used to do it as well, whether to study a coral reef or to analyze any
level of organisation of human populations (Margalef, 1992).
And, going back to the topic of Margalef as a teacher, I have to stress that all the advantages and
disadvantages I mentioned before helped split his students in two groups: the ones that liked his classes and the ones that didn’t, with no intended disrespect towards the latter. Margalef was passionate of
natural selection and he considered it could be applied to all aspects of life and at all levels of human
organisation. He was, therefore, capable of giving a pass to some students who didn’t deserve it while
telling them “life will fail you” or “look, I give you a pass but promise me you will never teach the
subject or work in anything related to ecology”. It is true that he didn’t like being too hard on students
during exams. He was, however, strict in his selection of the students that deserved the best marks.
Exams are always a source of stress no matter the subject or the professor, but with respect to the
exams on ecology, they had the disadvantage they were also atypical as far as the questions were concerned. Many times the problem lay in the way he formulated the questions and not in the subject itself.
Margalef was always on the lookout for the bright student who could become a disciple and would
show some degree of originality and he would pick the best by asking questions in his particular way.
Some questions were handed down from year to year by senior students to the freshmen so that they
knew what to expect. The questions might be of the sort: “Why are the taxis in Barcelona black and
yellow? They may seem a bit esoteric to the students that are being introduced to the subject for the
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first time but it would not be an insurmountable obstacle if you knew anything about aposematic coloration. Other questions such as “effect of the Coriolis force in the curvature of the antlers of antelopes,
in the growth of branches in the tree trunks, and in the distribution of the genus Velella” were meant to
sort out outstanding students who could have otherwise remained unnoticed. No matter how hard the
exams were, the percentage of passes and failures never changed, with passes to failures at about 2-to1 ratio. However, many of the students that got a pass were aware of Margalef ’s opinion of them when
they got back the exam together with a mark which was obviously a fail. The exams of the ones that
didn’t pass do not even deserve to be mentioned. Regarding the exams we had to take during our own
1970-1971 ecology course, Margalef suffered from an extra dose of originality as he decided to abolish
the traditional Napoleonic exams, with the students locked up in a classroom while they were answering questions. The novelty consisted in a short meeting with all the students early in the morning in the
Department library where he hand us two topics to expand upon: we had from 9 a.m. until 4 p.m. when
we had to stop by his office and hand him the paper we had written on one of the two topics we had
chosen. I have to admit that I had a very bad time over it and many of my classmates shared my feelings
due to the difficulty of trying to write something original while having all the notes, books and other
means at hand. A few days later he told us we didn’t deserve this kind of exams as we had done so
badly in general. To a chosen group of us, who had done well, he let us take a second non-Napoleonic
term exam but we had no chance of a third for the final exam and we all went back to the traditional
system. I remember that during this first exam four of my classmates handed in an essay which was the
result of a joint effort, probably very well thought out as they got an A. They had, however, to share the
mark democratically amongst the four of them, with the result of an obvious fail.
I have so far commented on professor Margalef ’s teaching career, but he pursued a career in research beyond this aspect of teaching which I would downright call frantic. In the first years of existence of the Ecology Department, Margalef combined his work between the University of Barcelona and
the Fisheries Research Institute (IIP) of CSIC. He would go to IIP on Tuesdays and Thursdays and
spend the rest of the week at the university. He had his own research team at each one of the centers:
the marine biologists Marta Estrada in Barcelona and Miguel Alcaraz and Xavier Niell in Vigo, while
at the university, the limnologists Dolors Planas and Rosa Miracle, who were at the time, early 70’s)
beginning their research work at the lake of Banyoles plus a group of students who would go during
their free time and amongst which I counted myself. Tecla Riera was Margalef ’s assistant and was
soon joined by Joandomènec Ros and the department became divided into two kind of doctorate students, the marine ecologists and the fresh water ecologists.
The writing of Ecología (Margalef, 1974), with its 951 pages, dates from that period. I suppose
that, as with anything else, some people are better at writing than others but the way Margalef would
write can only be described as extraordinary. His Olivetti typewriter sounded like a machine gun that
only stopped when the letters hit the rubber cylinder with a different sound as when there was paper.
It was time to stop, pick up the paper from the floor if it was handy or at least the carbon paper, as he
used two sheets and some carbon paper to keep a copy. The writing began early in the morning, right
after the ecology class, which started at 8 a.m. to allow him more time for his writing. He would seldom have a break, just enough for a coffee and he dealt quickly with any visits. He stopped writing at
around 2 p.m., picked up the sheets that might have fallen to the floor, sorted them out, numbered
them and piled them up at one end of the table and would call it a day just to continue two days later
as if nothing had happened in between. We have to remember that on alternate days he went to IIP
and he used the afternoons to attend to other matters. He kept the typed sheets inside a metallic cabinet in brown folders bound with a rubber band. On the cover of the folder he would leave handwritten
notes and some of the sheets inside would also be full of them. The 951 pages could easily consist of
3000 or more sheets which made quite a considerable stack. While writing, he would include all the
bibliography he remembered and then he would go over the text and insert the missing references by
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hand. The draft copy was finished in one year. The final writing of the book was not a mere copy of
the first one but a full rewriting that took almost as long. If we take a look at Margalef ’s bibliography
during those years, 1971-73 (Ros, 1991), we realise that he had time to write articles on the side that
can match the amount of articles published in the previous and later years. The writing of Limnología
(Margalef, 1983), with its 1010 pages, followed a similar pattern to the one described above and I
will obviously not go over it again.
Peter Wangersky, from the University of Halifax, who spent some sabbatical stays in Barcelona,
used to say that Margalef could work right through a three-ring-circus show without losing track and
being at his most efficient.
Margalef was a person who didn’t get out of the office much but his door was always open and students and graduate students alike could visit him there any time we wanted, although we could always
tell if he was eager to get on with something else or deeply involved in his thoughts. Tecla Riera was
in a way a kind of transmission belt that would keep him connected to the department despite his
many other information sources based on his observation skills. He knew what was going on, even
though he didn’t interfere much. Whenever he proposed a research topic, he felt enthusiastic about it
and even anticipated the results he expected if everything turned out well. On many occasions, he
would use the pages of his desk calendar to scribble and sketch data to supplement his initial exposition. When he was done, he would tear the page and somehow you would find yourself standing in
the corridor, or in the office or the library staring at it, trying to figure out what it said while trying to
remember what Margalef had said in relation to what while he was going on about his ideas. We all
had to work in a specific taxonomic group and from there we could fit in all the ecology we were able
to develop. In those days, the zoological and botanical taxonomists that worked in Margalef ’s department were equivalent in numbers to the ones that made up the respective departments. Quoting
Xavier Ferrer, “he would send us on a single-handed voyage along the seas of research and, as a rule,
he wouldn’t warn you of any possible dangers” (Ferrer, 2004), always consistent with his belief in
natural selection. The results would be uneven and, the same as with his students, some would just
disappear discreetly without him losing any sleep over it.
As I have mentioned earlier, he had this incredible capacity for transmitting enthusiasm for the
ideas that interested him. You would come out of his office holding the calendar sheets feeling you
were going to start a research project that would achieve a major breakthrough in ecology. Other times,
he would ask you offhand about your progress and he liked to be shown the results and would get all
excited if he considered them relevant and had no qualms about mentioning these results in his papers.
Margalef founded three scientific magazines and he was a regular contributor with his papers
Publicaciones del Instituto de Biología Aplicada (PIBA), Investigación Pesquera (IP) and Oecologia
aquatica. The issues of PIBA or IP are hard to find and the articles published in them, quite often
written by Margalef himself, are very rarely read. Big mistake, as you can find some gems amongst
them, as not only would he present and interpret data, but he would also anticipate some of the results
and conclusions and formulate hypotheses that he would develop later on. Nowadays this type of
approach or projection of the results is called speculative science. “Too much speculative” is the fatal
sentence that you can usually find in the letter editors send to reject a paper for publication when you
spend too much time on the data assessment or on the conclusion. Margalef was not afraid to
expound his ideas even though many times he himself admitted he was not able to prove them at the
present stage of information available. Many of the criticisms he received from later ecologists were
of the kind that he had this habit of jumping ahead while leaving many gaps to be filled, some of
which have already been filled and some are still pending. The wealth of ideas we find in his writings
in PIBA or IP can already be found in his earlier works, many of them geared towards the general
public. In that sense I can recommend some booklets from the end of the 40’s published by Seix y
Barral that took up less than a hundred pages and that he wrote as a complement to a meagre salary
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J. Armengol
to help support his family of six. That’s why, with a mischievous smile, he used to call these papers
“nutritional ecology”. The topics, of course, were varied but all juicy nevertheless: La vida en el mar,
Los insectos sociales, Las plantas carnivoras are some of the titles I have been lucky enough to read
many years after they were written. One of my favourites has always been the latter as in it he predicted the new food adaptation of carnivorous plants, an example of allotrophy. According to Margalef,
this adaptation came about because they had no other chances of obtaining nutrients through more
orthodox means. A few years later we had the chance to prove his theory right at the old department
in University old building with a specimen of Sarracenia he had brought from Canada and which we
had kept for a long time in a crystallizer, watering it with distilled water and with regular visits to the
genetics department to get a pot full of Drosophyla to feed to it.
From the many activities going on at the department, the so-called magic soirees on Thursday
afternoon were of special interest. We euphemistically called that to the seminars held by Margalef.
They were open activities and they were not based on a previously announced topic; we would just
attend and if it was time and nobody came up with a topic, Margalef would stand up and start talking
about something that could lead to a discussion, without necessarily having to reach any conclusions.
Many of the graduate students at IIP used to take part in those seminars and also many physicists,
Jorge Wagensberg amongst others, and many of the physicists involved in the group of complex
systems. Jordi Flos was the one that started calling these seminars magic soirees not because of the
topics being discussed but for the way the ideas would flow, just like rabbits coming out of a magician’s hat. Flos gives a short but interesting account of those seminars in his book Ecología, entre la
magia y el tópico (Flos, 1984).
Ramon Margalef kept up his activities until his illness prevented him from leaving his house, and
that was for a very short time. He kept coming to his office at the department, mostly as an incentive
to walk around the libraries of the faculties of Biology, Geology and Physics and Chemistry. He
remembered what day the issues from Science, Nature or many other magazines were expected and
there he was, ready to be the first one to read them. His personal evolution during his last years was
clearly the one of a K strategist, with a mental lucidity and incredible observation skills which he
now used on himself. He didn’t mind talking about his illness and how his life had been altered
because of it. He used to say he found interesting the way we lose memory, “just like the hard disk of
a computer; clusters get deleted without having any links with one another”.
He used to come and see us and he liked to stop by for a chat and tell us about his ideas and projects he thought interesting and could no longer embark on. He was concerned about the big manmade changes to the landscape, and he used to call them “the inversion in the landscape topology”.
At the same time, he was interested in the number of cells of many species from a same taxonomic
group that, according to him, was discontinuous at the species level. He used to compare those discontinuities to shoe size, “sort of a quantic cytometry”, and was as always worried about nutrients,
with a special emphasis on phosphorus. During the opening speech of the Second Iberian Congress
of Limnology in Valencia (June 2000), he insisted on his concerns over the pending issues and the
relevance of their study in the future. He wrote these words in a short but delightful article, “Cabos
sueltos” (2001), published one of the previous volumes of Limnetica, and it can be considered as a
sort of future projection of his ideas.
He used to enjoy our visits to him at his home. Delivering his mail was always a good excuse; just
that many times there were several of us just for a few letters. Even though his memory was failing
him, you could immediately tell if the subject caught his attention as he would awaken, his eyes
would sparkle and would start up a typical Margalefian discussion. He admitted that our visits helped
him while away the “black hours”, as he called the hours he spent by himself or in the company of his
dear wife Maria. He died as he would have liked, on a Sunday, surrounded by his whole family and
able to say his last goodbye to them.
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vii
It was then when many of us found out he had been a religious person and were finally able to
understand some moments in his life when he had shown extreme fortitude. Pere Ynajara, parish
priest from Sta. Eugenia del Congost and a good friend for many years, presided over the funeral service and during the homily he spoke about many aspects of his personality amongst which I would
like to single out the sense of irony Margalef would display on many occasions. “He was worried
about what would happen to his nutrients”, the priest told us. Which is logical as, being a religious
person, he couldn’t have had many doubts regarding more spiritual matters. I can assure you I have
no doubts he said it, nutrients being an issue that interested him and one that, once again, he applied
on himself. Well then, I can only say that I truly hope his nutrients soon get to an oligotrophic ecosystem, such as the Mediterranean Sea, the waters around Mallorca or the Gulf of Lyons or along the
coast of Castellón, the places he studied, described and became the basis of many of his scientific
hypotheses. In those waters of great diversity and biodiversity, with low P/B values, with an internalization of the nutrient cycle, great pigment diversity and big sized K- strategist species, there is
where I hope he can continue to enjoy the wonderful world he helped us understand.
May he rest in peace.
REFERENCES
Bascompte, J. y R. Solé. Margalef y el espacio o porqué los ecosistemas no bailan sobre la punta de una aguja.
Ecosistemas, Año XIV, nº 1. http://www.revistaecosistemas.net/index.asp?id_numero=8
Ferrer, X. 2004. Margalef, el naturalista que yo conocí. Quercus, 221: 8-11.
Flos, J. 1984. Ecología entre la magia y el tópico. Ed. Omega. Barcelona. 129 pp.
Flos, J. 2005. El concepto de información en la ecología margalefiana. Ecosistemas, Año XIV, nº 1.
http://www.revistaecosistemas.net/index.asp?id_numero=8
Margalef, R. 1960. Ideas for a synthetic approach to the ecology of running waters. Int. Rev. ges. Hydrobiol.,
45: 133-153.
Margalef, R. 1962. Comunidades naturales. Instituto de Biología Marina de la Universidad de Puerto Rico.
Mayagüez. 469 pp.
Margalef, R. 1963. On certain unifying principles in ecology. Am. Nat., 97: 357-374.
Margalef, R. 1968. Perspectives in ecological theory. University of Chicago Press. 111 pp.
Margalef, R. 1974. Ecología. Ed. Omega. Barcelona. 951 pp.
Margalef, R. 1978. Life-forms of phytoplankton as survival alternatives in an unstable environment. Oceanol.
Acta, 1: 493-509.
Margalef, R. 1979. El precio de la supervivencia. Consideraciones ecológicas sobre las poblaciones humanas.
Étnica, 15: 103-115.
Margalef, R. 1980. La biosfera: entre la termodinámica y el juego. Ed. Omega, Barcelona. 236 pp.
Margalef, R. 1981. Limnología. Ed. Omega, Barcelona. 1010 pp.
Margalef, R. 1992. Planeta azul, planeta verde. Prensa Científica SA. Barcelona. 265 pp.
Margalef, R. 2001. Cabos sueltos. Limnetica, 20: 1-2
Ros, J. D. 1991. Ramon Margalef, limnologist, marine biologist, ecologist, naturalist”. En: Homage to Ramón
Margalef, or Why there is such pleasure studying nature. J. D. Ros y N. Prat (eds.). Oecologia aquatica, 10:
413-423.
Ros, J. D. 2004. Dignificà l’apelatiu “naturalista”. Notícies de la Institució. Circular de la Institució Catalana
d’Història Natural, 54: 1-3.
Walker, L. R. 2005. Margalef y la sucesión ecológica. Ecosistemas, Año XIV, nº 1. http://www.revistaecosistemas.net/index.asp?id_numero=8
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Limnetica, 25(1-2): 1-10 (2006)
The ecology of the Iberian inland waters: Homage to Ramon Margalef
© Asociación Española de Limnología, Madrid. Spain. ISSN: 0213-8409
Leaf litter processing in low order streams
Manuel A. S. Graça & Cristina Canhoto
Departamento de Zoologia, Universidade de Coimbra, 3004-517 Coimbra, Portugal
[email protected]; [email protected]
ABSTRACT
Forests produce a large amount of detritus, that inevitably end up in streams, subsidizing aquatic systems with organic matter
and nutrients. Here we review some of the research carried out at the University of Coimbra with the objective of getting a better understanding of the breakdown process of these materials and its incorporation to secondary production. Litter-fall in
deciduous forests in Central Portugal can reach up to 750 g AFDM of leaves m-2 yr-1, with 73% of the litter produced between
October and December. In several retention experiments, we measured a 90% leaf retention in low order (1st- 4th) streams
within 15 – 70 m, and a standing stock of up to 450 g AFDM m-2. The amount of nutrients in the water and the plant physical
and chemical defenses can be an indicator of the rate at which plant material is incorporated into secondary production or
exported as dissolved and fine particles of organic matter. Respiration rates of decomposing leaves incubated with fungicides
were severely reduced, supporting the idea that fungi are very important agents in litter breakdown. The fungi group known as
aquatic hyphomycetes are capable of producing enzymes able to cause leaf maceration, and by 2 to 3 weeks, up to 15 % of the
decomposing leaf biomass corresponds to fungi. Shredder invertebrates are also biological agents involved in litter breakdown.
Given their densities and feeding rates, we measured consumption rates of 12 – 54 g of leaves m-2 yr-1 in a stream in Central
Portugal, corresponding to 2 to 9 times the litter standing stock. Feeding rates were high in nutrient rich leaves and low in chemical and physically protected leaves with low nutrient content. According to several experiments, fungal colonization facilitates the access of invertebrates to the energy trapped in deciduous leaves in streams. Some invertebrates have strategies to cope
with low quality food (leaves with low microbial biomass or high chemical defenses). Those include high mobility, small size,
compartmentalized digestion in the gut, presence of endosymbionts, and the capability to decrease respiration rates. The relative importance of fungi and invertebrates in the incorporation of plant litter material into secondary production varies across
rivers and biomes. Shredder invertebrates seem to play a key role in litter breakdown in headwaters, but their importance appears to decrease downstream. In the same way, some systems where leaves are hard or protected, shredder invertebrates may be
less abundant and the energy may be mainly recovered from litter by fungi. Eucalyptus plantations are systems with low diversity of invertebrates and aquatic hyphomycetes. Streams running through eucalyptus plantations seem therefore ideal to experimentally investigate relationships between structural parameters (biodiversity) and function. Finally, our research has been
extended to other climatic areas including the Mediterranean and tropical streams. We reported a wide variety of situation in
those systems. A general rule applying to all of them is that if leaf litter is abundant and high quality, the incorporation of
energy into detrital food webs can be processed very quickly. However, if leaves are well protected and nutrients in the water
are low, processing rates are equally very low, independently of the ambient temperatures.
Key words: litter balance, decomposition, fungi, detritivores, Mediterranean and tropical streams.
RESUMEN
Los bosques producen una gran cantidad de detritus orgánicos, que inevitablemente llegan a los ríos, subsidiando los sistemas acuáticos con materiales y nutrientes. Aquí se revisan algunos de los trabajos que se han hecho en la Universidad de
Coimbra con el objetivo de entender mejor el proceso de descomposición de este material y su incorporación en producción
secundaria. La entrada de hojarasca en bosques caducifolios del Centro de Portugal puede alcanzar hasta 750 g PSLC (peso
seco libre de cenizas) m-2 año-1, con 73 % de este valor ocurriendo entre Octubre y Diciembre. En varios experimentos de
retención medimos que cerca de 90 % hojas que entran en ríos de baja orden (1ª- 4ª) eran retenidas entre los 15 y 75 m, y que
la biomasa de hojarasca acumulada era de hasta 450 g PSLC m-2. La cantidad de nutrientes en el agua y las defensas físicas
y químicas de las plantas pueden ser un indicador de la tasa a que el material orgánico es incorporado en producción secundaria o exportado como material disuelto o finamente particulado. Las tasas de respiración de hojas incubadas con fungicidas disminuyeron severamente apoyando la idea de que los hongos son agentes muy importantes en la descomposición de
hojarasca el los ríos. El grupo de hongos conocido como hifomicetos acuáticos producen enzimas que causan la maceración
de hojas, y en 2 o 3 semanas, hasta 15 % de la biomasa de una hoja en descomposición puede corresponder a hongos. Los
invertebrados desmenuzadores son también agentes biológicos en la descomposición. Dadas las densidades de desmenuzado-
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M. A. S. Graça & C. Canhoto
res y sus tasas de ingestión de alimento, hemos calculado tasas de consumo de hojas en ríos de12 – 54 g m-2 año-1, lo que
corresponde a 2 a 9 veces la cantidad de hojarasca presente. Las tasas de consumo son generalmente altas en substratos
ricos en nutrientes y bajas en hojas pobres en nutrientes o protegidas del punto de vista químico y físico. De acuerdo varios
experimentos, la colonización por hongos facilita el acceso de los invertebrados a la energía de las hojas. Algunos invertebrados han desarrollado estrategias para poder vencer la baja calidad de las hojas, incluyendo un alta movilidad, tamaño
pequeño, compartimentalización de la digestión en el intestino, presencia de endosimbiontes y la capacidad para disminuir
las tasas respiratorias. La importancia relativa de los hongos e invertebrados en la incorporación de la hojarasca en producción secundaria es variable entre ríos y biomas. Los invertebrados desmenuzadores parecen jugar un papel importante en la
descomposición de hojarasca en los ríos de bajo orden, pero su importancia parece disminuir rió abajo. Del mismo modo, en
algunos sistemas en que las hojas son duras o protegidas, los invertebrados pueden ser menos abundantes y la energía canalizada en producción secundaria principalmente por los hongos. Las plantaciones de eucaliptos son sistemas con una baja
diversidad de invertebrados e hifomicetos acuáticos. Los ríos que corren por plantaciones de eucaliptos parecen ser por este
motivo sistemas ideales para investigar las relaciones entre parámetros estructurales (biodiversidad) y función. Finalmente,
nuestra investigación ha sido extendida para otras zonas climáticas, incluyendo el Mediterráneo y las zonas tropicales.
Hemos reportado una gran variedad de situaciones en esos sistemas. Una regla general a todos ellos es que si la hojarasca es
abundante y de alta calidad, la incorporación de la energía de las hojas en las cadenas alimentares se procesa de forma muy
rápida. Sin embargo, si las hojas están bien protegidas y los nutrientes el agua son bajos, estas tasas son igualmente muy
bajas, independientemente de las temperaturas ambientales.
Palabras clave: Balance de la hojarasca, descomposición, hongos, detritívoros, arroyos mediterráneos y tropicales.
ALHOCHTHONOUS ORGANIC MATTER
IS AN IMPORTANT ENERGY SOURCE
FOR FORESTED LOW ORDER STREAMS
Forests are among the most productive systems
on Earth with primary production reaching
1800 g dry mass m-2 year-1 in the tropics. Even
boreal forests are more productive than cultivated lands (850 vs. 750 g dry mass m-2 year-1,
respectively; Ricklefs, 2000). In forested systems less than 5 % of the primary production
will be lost to herbivores (Ricklefs, 2000); this
implies that a very large proportion of the
energy fixed in forests will end in the detrital
pathways (Fig. 1). This is particularly evident in
deciduous forests with litter-fall ranging from
300 to 800g dry mass m-2 year-1, or with
> 1000g dry mass m-2 year-1 in tropical forests
(reviewed Abelho, 2001).
With such an amount of litter production, it is
virtually impossible that leaves, fruits, seeds,
twigs, and other plant remains, will not end in
streams. Moreover, trees in the riparian zones
shade the small streams, decreasing in this way
the amount of solar energy which could be used
by primary producers. Therefore, litter shed by
trees is likely to be a key energy source for low
order streams running through forests. It is therefore ecologically relevant to understand the
fate of energy and nutrients in those systems. At
the University of Coimbra, Portugal, we have
been addressing several aspects of litter decay in
small streams for the last 15 years. Here we
review the main findings of our research.
LITTERFALL AND THE DYNAMICS
OF ORGANIC MATTER
How are leaves retained in streams? Can we predict decomposition rates of leaves based on
their intrinsic characteristics? What is the relative role of the environment in litter decomposition? What are the main agents affecting litter
decomposition? To address some of those questions we began measuring litter dynamics in
deciduous forests in Central Portugal. In a forest
dominated by Castanea sativa Mill., annual litter-fall reached 750 g m-2 yr-1, with 73 % of litter produced between October and December,
which is consistent with other results reported
for deciduous forests. Nearly 90 % of the leaves
falling into low order streams in Central
Portugal were retained in within 10 – 70 meters,
with retention decreasing downstream (Canhoto
& Graça, 1998). Retained litter accumulates in
the stream-bed before being processed or washed away during floods; we measured standing
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Leaf decomposition in streams
stocks of organic matter of 50 – 450 g AFDM
m-2 in streams of central Portugal. These values
were much higher than the standing stock of
periphyton (6 g m-2; Abelho & Graça, 1998) in
the same river. Moreover the amounts of coarse
particulate organic matter in rivers tend to
decrease downstream, whereas the standing
stocks of benthic algae tend to increase in the
same direction (Cortes et al., 1995).
Decomposition is therefore a critical ecosystem process, determining the availability of
nutrients for primary producers. Can we predict the rate at which leaves decompose? The
answer, to some extent, is yes. We found that
decomposition rates increase with nitrogen
content of leaves and decrease with the amount
of plant chemical and physical defenses
(Cortes et al., 1994; Canhoto & Graça, 1996).
Decomposition rates also tend to increase with
nutrient content in the water. This information
is important for conservation, restoration and
management of riparian zones. “Cleaning”
3
streams by removing wood and other retentive
features and removing stream-shading vegetation is a bad environmental practice. Although
litter decomposition proceeds until all material
is mineralized, this paper will refer to the
breakdown of large particles of organic matter
and not to the processing of fine particles or
dissolved organic matter.
DECOMPOSERS
When leaves enter the streams, their nitrogen content generally increases. This is evidence of
microbial colonization, which can be corroborated by the increase of ATP and oxygen consumption of leaves (Abelho et al., 2005). Moreover,
leaves start loosing mass at a rate proportional to
microbial colonization (Suberkropp & Chauvet,
1995); decomposition is therefore a biological
process and a measurement of the rate of incorporation of leaf material into secondary production.
Figure 1. Leaf litter accumulated on soil in a Eucalyptus plantation. Hojarasca acumulada en el suelo de una plantación de Eucaliptus.
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Which are the microorganisms involved in litter
decomposition? There is evidence from the literature, that fungi are more important than bacteria in this process in terms of biomass and production (e.g. Pascoal & Cassio, 2004; Abelho et
al. 2005). In a tropical stream, we found that leaves exposed to fungicides had lower respiration
rates and lower microbial biomass than leaves
exposed to bactericides. Other authors concluded that even under organic pollution conditions,
production of bacteria in leaves is lower than
fungal production (Pascoal & Cassio, 2004).
It is also generally accepted that fungal
decomposers of leaves are aquatic hyphomycetes
(Fig. 2; Gessner & Chauvet, 1994; Gulis &
Suberkropp, 2003), since just after submersion a
large amount of conidia start detaching from leaves (e.g. Bärlocher, 2000). However, many geofungi have been also isolated from decomposing
submersed leaves plated over agar. What is the
relative role of both types of decomposers in the
decomposition of organic matter in streams? To
answer this question we measured the capability
of several species of geofungi and aquatic
hyphomycetes to cause leaf maceration in water
and under laboratory conditions. Only aquatic
hyphomycetes caused significant leaf maceration
(measured as mass loss and decrease in tensile
strength) and had higher enzymatic (xylanase,
pectinlyase and polygalacturonase cellulose C1
and Cx) activity in submerged substrates than
terrestrial fungi isolated from leaves (Graça &
Ferreira, 1995; Rodrigues & Graça, 1997). Softening was correlated with the activity of all enzymes, especially xylanase (rs = 0.94; P< 0.001).
Our conclusion is that when falling in the
water, leaves are already colonized by terrestrial fungi, but their activity is severely depressed. In the water, leaves are rapidly exposed
to thousands of spores of aquatic hyphomycetes (e.g. Bärlocher & Graça, 2002) that germinate and grow into the leaf substrates (Canhoto
& Graça, 1999) and produce degrading enzymes (Canhoto et al., 2002).
Many of the chemical and physical plant
defenses against pathogens and herbivores may
remain active after senescence. Thick cuticles
may have a two-fold role in plants, by decreasing water losses and retarding fungal attack.
One of the explanations for the lower decomposition rates of some eucalyptus leaves in nutrient
poor streams is the presence of a thick cuticle.
Electronic microscopy observations showed that
fungi can only penetrate into the leaf mesophyll
of eucalyptus leaves through stomata and cracks
at the waxy cuticle (Canhoto & Graça, 1999).
Figure 2. Spores of aquatic hyphomicetes: left and right: Tricladium splendens; center: Clavariopsis aquatica, Articulospora tetracladia and a sigmoid. (Photos by Felix Bärlocher). Esporas de hifomicetes acuáticos: izquierda y derecha: Tricladium splendens;
centro: Clavariopsis aquatica, Articulospora tetracladia y un sigmoide (Fotos de Felix Bärlocher).
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Another defense of eucalyptus leaves is the
presence of oils, allocated in glands. In
eucalyptus leaves, oils may account for up to
5 % of leaf mass (Canhoto & Graça, 1999) and
they are known to be have antibiotic properties.
They were found to also reduce or suppress
growth of aquatic hyphomycetes in vitro
(Canhoto & Graça, 1999) and interfere with
microbial enzymes (Canhoto et al., 2002).
Fungal sporulation from eucalyptus leaves was
retarded when compared with other leaves, but
the removal of cuticle and oils resulted in accelerated sporulation (Canhoto & Graça, 1996;
Bärlocher et al., 1995). The extraction of oils
from eucalyptus leaves also resulted in an
increase of consumption by the shredder Tipula
lateralis, whereas the transference of eucalyptus oils to alder leaves resulted in a decrease in
food consumption by the same shredder.
DETRITIVORES
Many stream invertebrates use leaf litter as a
food resource. Besides incorporating leaf material into secondary production, shredder invertebrates fragment leaves and produce a large
quantity of fecal pellets. The result is the transformation of coarse particulate organic matter
(C.P.O.M.) into fine particulate organic matter
(F.P.O.M.), which may constitute an important
food source for other organisms we call “deposit
feeders” and “filter feeders”.
Feio & Graça (2000), González & Graça
(2003), and Azevedo-Pereira et al. (2006) calculated for a mountain stream in Central
Portugal that the mean annual consumption of
leaves by the caddisflies (Sericostoma vittatum
Rambur and Lepidostoma hirtum (Fabricius))
was, respectively, 12 – 22 g m-2 year-1 and 54
g m-2 year-1. These values correspond to 2 – 9
times the leaf standing stock of the stream.
Shredder invertebrates have therefore a key role
in the trophic ecology of low order streams
(reviewed Graça 1993, 2001).
Several factors can constrain the access of
invertebrates to the energy trapped in leaves. To
start with, a reduced number of animals have the
5
enzymatic capability to use the structural compounds of leaves. How do they manage to access
the plant energy? We have been studying energy
transference from litter pool to invertebrate
shredders, using the caddisflies Sericostoma vittatum Rambur, and Lepidostoma hirtum
(Fabricius), as well as the crane fly Tipula lateralis Meigen (Fig. 3) as test organisms. Leaves differ in their quality for shredders as asserted from
measurements of feeding rates, food choice experiments and growth rates (e.g. González & Graça,
2003). The incorporation of leaf material into
invertebrate secondary production proceeds at a
faster rate in nitrogen rich and soft leaves, when
compared with nitrogen poor, chemically protected, hard leaves (Canhoto & Graça, 1995;
González & Graça, 2003). The implication is that
changes in the frequency of leaf types and therefore forest practices may affect the dynamics of
invertebrates in streams. Moreover, litter-fall in
temperate areas occurs mainly during autumn,
and litter is composed by a mixture of leaves differing in their quality. Leaves of high quality
such as alder are quickly consumed, whereas leaves of more recalcitrant species, such as oak, take
longer time to be fully colonized and degraded by
microorganisms, but they can be a good resource
for later in the season. If the mixture of leaves is
replaced exclusively by leaves of high quality, it
may supply shredders with a large input of high
quality food for a short period of time, but energy
may lack in later stages. On the other hand, if
streams are provided only with low quality
resources, food may be scarce early in the season.
Figure 3. Larvae of Tipula lateralis, a stream shredder. Larva
de Tipula lateralis, un triturador fluvial.
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6
INTERACTIONS BETWEEN
DETRITIVORES AND FUNGI
One common observation on the ecology of
shredder detritivores is that they preferentially
feed on fungal colonized leaves in laboratory
(Suberkropp, 1992; Graça et al., 1993b) and
field conditions (Graça, 1992). They also feed
and grow faster, survive better and have a larger reproductive output when leaves are
fully colonized by fongi (Graça et al., 1993b).
The reason seems clear: fungal colonization
cause leaves to increase nitrogen content
(because of fungal biomass) and leaf maceration, benefiting in this way from microbial
enzymes (Suberkropp, 1992; Graça et al.,
1993b; Graça, 1993). Some shredders do selectively consume the leaf patches with high fungal mass or selectively feed on fungal biomass
growing on the surface of the leaves (e.g.
Graça et al., 1993b; Graça et al., 2000).
SOME NOTES ON THE ECOLOGY OF
SHREDDERS
Fast moving invertebrates are very active visiting
patches were litter accumulates and probably
remaining for short periods of time in the patches
if the food quality is low. For invertebrates with
low mobility, high selectivity may not be an
option because if they reject less-profitable food
they may spend a long time searching before they
encounter food again. Invertebrates with low
mobility may be more efficient in taking their
energetic requirements from low quality food.
Tipulidae larvae are slow moving invertebrates that inhabit streams. Unlike carnivore tipulids, shredder tipulids have an alkaline anterior
gut with a pH 10.5 – 11 (e.g. Bärlocher & Porter,
1986; Graça & Bärlocher, 1998; Canhoto, 2001).
At such a high pH, the gut proteolytic activity of
these tipulids remain active and is not affected
by polyphenolics from leaf extracts (Graça &
Bärlocher, 1998). This strategy therefore, allows
for a maximum protein extraction and, at the
same time, the plant defenses are overcome. In
the posterior section of these tipulids gut, pH
values are neutral/alcaline and a high number of
endosymbionts seem to have a key role in the
digestion of the plant polysaccharides.
In a series of laboratory experiments, we
found that Gammarus pulex (L.) was able to
maintain growth even when low quality food
was supplied whereas that did not happen with
the less active Asellus aquaticus L. (Graça et
al., 1993a). G. pulex compensated for low quality food by reductions in respiration rates.
Although another form of compensation may be
the increase of food intake to maintain a constant energy / nutrient income (e.g. Calow, 1975;
Rollo & Hawryluk, 1988), in most cases, shredding invertebrates decrease their energy intake
when fed low quality food.
WHAT IS THE RELATIVE
IMPORTANCE OF INVERTEBRATES
AND FUNGI IN THE INCORPORATION
OF LEAF ENERGY INTO FOOD WEBS?
The relative importance of invertebrates and
fungi in litter breakdown, and therefore in the
incorporation of energy trapped in leaf tissues
into food webs has been a matter of debate (see
references in Graça, 2001). Apparently, whereas
fungi are omnipresent in all flowing waters, the
densities of shredder invertebrates can be controlled by other factors, including the quality
and quantity of the litter. Therefore, in some
systems invertebrates can be considered as
unimportant in energy transference in detritus
based systems, while in other cases they may be
the key elements. For example, Hiebber &
Gessner (2002) calculated that, in a stream,
fungi were responsible for removing 15 – 18 %
of leaf mass, whereas the values for shredder
invertebrates were 51 – 64 %. On the other hand,
Gonçalves et al. (2006) calculated that almost
no litter energy in the form of coarse particulate
organic matter was taken into secondary production by invertebrates due to the high recalcitrant
properties of Savannah Cerrado streams.
As the availability of coarse particulate organic matter tends to decrease downstream and
nutrients in the water to increase in the same
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Figure 4. Eucalyptus leaf with oil glands in white. Hoja de
eucalipto con las vesículas de aceite en blanco.
direction, it is plausible that the role of both
types of organisms change along the longitudinal gradient. We tested this hypothesis in a
series of streams, ranging from 2nd to 6th order
in Central Portugal. Decomposition rates did not
differ along the longitudinal gradient (see also
Cortes et al., 1995). However, microbial role on
litter decomposition increased downstream as
judged by the difference in mass loss in leaves
incubated in coarse and fine mesh bags.
Consistently, the density in spores in the water
column increased downstream, whereas the density and percentage of shredder invertebrates
increased upstream. This relationship was
observed only in spring / summer. It is conceivable that during autumn / winter there might be
a surplus of energy in the form of leaves and the
impact of invertebrate feeding on litter breakdown may then be small (Graça et al., 2001b).
STRUCTURAL AND FUNCTIONAL
PARAMETERS IN DETRITUS BASED
SYSTEMS
Detritus based systems are a ground for testing
some ecological theories. For instance, species
replacement has been analyzed from the structural point of view but we can learn a lot on the
functional changes related to species replacement by invasions. Species invasions have
shown to affect community structure, sometimes with the reduction of biodiversity due to
7
local extinctions and the dominance of introduced species (Towsend et al., 2000). Given that
decomposition is controlled by nutrient related
factors and plant defenses, can we predict the
ecological effects of species introductions?
If the plant invader is a nitrogen fixing species, then we may expect that the turn over of
organic matter to be accelerated. However, if
the invaders are chemically or physically protected, decomposition and therefore the rate at
which energy re-enters the biota component of
ecosystems to be retarded. Invaders are very
common in riparian areas (e.g. Vitousek, 1996)
and we have been testing these assumptions by
looking at soil and aquatic systems.
In a series of litter breakdown experiments in
which introduced vs. native and high quality
(N content) vs. low quality (high protection)
leaves in soils and water were compared, it was
found that decomposition rates and associated
processes such as microbial and invertebrate
colonization were independent of plant origin,
but could be explained by intrinsic leaf proprieties (Pinto et al., 1997; Pereira et al., 1998).
In aquatic systems we compared streams bordered by native deciduous and eucalyptus plantations. Eucalyptus are originally from
Australia, but they are nowadays ubiquitous in
several parts of the world. Vast areas in the
Iberian Peninsula are planted with eucalyptus.
This subject was reviewed by Graça et al.
(2002) and will not be treated in detail here, but
we can summarize the changes associated to
eucalyptus plantations in the following way:
In eucalyptus plantations the seasonality of
litter-fall is altered from an autumn peak to an
even litter-fall along the year or a summer peak
if the hydrological stress is high. The average
standing stock of organic matter was not different between native deciduous and eucalyptus
plantations; streams or tended to be higher in
eucalyptus plantations, probably because of spates and bark accumulation, which increases litter
retention. Fungi accumulate in decomposing
leaves at similar rates in both stream types.
Eucalyptus leaves are a low quality substrate for
shredder invertebrates and fungi, as judged
from: (a) their oil content with antibiotic proper-
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Página 8
ties (Fig. 4), affecting fungal growth, fungal
digestive enzyme activity (Canhoto et al., 2002),
and (b) feeding experiments with invertebrates
in which there was a decrease in surviving,
growth and feeding rates when fed with eucalyptus leaves. Oils inhibit fungal growth and invertebrate consumption in vitro (Canhoto & Graça,
1999). Elimination of leaf lipids resulted in faster decomposition and high sporulation by aquatic hyphomycetes (Bärlocher et al., 1995).
Maybe for those reasons, invertebrate and
fungal richness was low in Portuguese streams
running through eucalyptus plantations. Because
assemblages of decomposer and detritivore species are poor in eucalyptus streams, we have an
ideal model system to investigate relationships
between community structure and ecosystem
functioning. For instance, Bärlocher & Graça
(2002) reported that although streams running
through eucalyptus forests had lower number of
aquatic hyphomycete species, decomposition
rates of chestnut (Castanea sativa) were similar
(but see Abelho & Graça, 1998).
TROPICAL SYSTEMS
The ecology of low order streams is well established for temperate areas, but scarce in other
zones. Most of the literature on the dynamics of
litter-fall and the fate of organic matter entering
streams is based on research carried out in
North America and Europe. A quick survey in
the “Web of Science” was run for citations on
papers dealing with litter breakdown in streams
from 2000 to 2004 and 110 references were
found, 44 % from North America, 30 % from
Europe, 8 % from the Mediterranean, 8 % from
Australia and New Zealand and 2 % for the rest
of the world, revealing that patterns of litter
dynamics in forested stream systems are based
upon research carried out in a restricted geographic area. Do the reported patterns apply to
areas with different productivity, seasonality
and hydrology? Do invertebrates and microbes
play a similar role in other climates?
In a series of feeding trials we found that, as
reported for tempered shredder species, tropical
shredders also selectively feed on microbial
colonized leaves, and there was a tendency for
growth rates to be reduced in the absence of
microbial assemblages in the leaves. The rate at
which leaves are incorporated into secondary
production was more variable in the tropical
areas than in temperate ones. In experiments
carried out in tropical cloudy forests in
Venezuela, decomposition rates were fast, with
50 % of leaf mass loss in less than 10 days in leaves of Hura crepitas L. The leaves of this species
were found to be equivalent to those of Alnus
glutinosa (L.) in terms of food resources and
decomposition rates (Graça et al., 2001a and
unpublished data). However, in Savannah streams, in Brazil (Cerrado), it took 90 days for alder
leaves to loose 50 % of their mass (Gonçalves et
al., 2006). Apparently, the availability of leaves,
their quality, and water chemistry are important
factors explaining the differences.
CONCLUSION
Detritus based systems are ideal to test many
current ecological theories. They can be studied
at community, population, and auto-ecology
levels. Litter decomposition is also a research
field in which the knowledge of several areas of
science (plant ecology, biochemistry, mycology,
population ecology, and others) is needed. If
organic matter breakdown is an important process in streams, factors interfering with the activities of fungi and invertebrates are likely to
affect the functional process of decomposition.
Therefore, decomposition rates may be used as
indicators of functional status of streams, as
proposed by Gessner & Chauvet (2002).
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Limnetica, 25(1-2): 11-32 (2006)
The aquatic systems of Doñana (SW Spain): watersheds and frontiers
L. Serrano1, M. Reina, G. Martín, I. Reyes, A. Arechederra, D. León & J. Toja
Dep. Plant Biology and Ecology. University of Sevilla. P.O. Box. 1095, Sevilla 41080.
1corresponding author: [email protected]
ABSTRACT
Doñana includes an extraordinary variety of aquatic systems. They are broadly classified according to their location (on
either aeolian sands or marshland) as their hydrology largely depends on the geomorphology of their basins. Their chemical
composition is mainly influenced by rainfall, evaporative concentration, groundwater discharge, biogeochemical interactions at the sediment-water interface, and the quality composition of their watersheds. The influence of the watershed can
be studied at different scales of observation. Rainfall infiltration in sandy soils is usually high so surface runoff becomes a
rare event of very short duration during floods which, nonetheless, exerts a huge influence on the limnology of temporary
ponds on aeolian sands. The water quality of the Doñana marshland, on the contrary, is influenced by long-term processes
taking place on large-scale areas: sediment deposition, eutrophication and heavy metal pollution. The review of the main
literature on the limnology of the Doñana aquatic systems during the past two decades, enable us to make a comparison in
time focusing on the interactions at the frontier between terrestrial and aquatic systems within watersheds. Presently, the
eastern area of the Doñana marshland is particularly affected by the low quality of the incoming flowing water compared
with the more isolated southern marshes within the National Park. Water from the lower strech of the Guadiamar River
(“Entremuros”), that floods the marshes of “Lucio El Cangrejo Grande”, showed a significant correlation between inorganic suspended solids and total P (r=0.807, p<0.05) during 2003-04, indicating an important contribution of inorganic particulates to the eutrophication of this area. The northern streams of the “Arroyo del Partido” watershed have not significantly
improved their water quality in the last two decades despite the construction of two waste-water treatment plants, being total
P correlated to dissolved phosphate concentration (r=0.995, p<0.01) during 2003-05. A general increase in NO3- concentrations have been detected in all studied aquatic systems of the Doñana marshland, including those with the highest water
quality (“Arroyo de la Rocina”) during the last two decades. Despite wetland management requires a watershed approach,
successive hydrologic projects in Doñana have failed to address the great spatio-temporal variability of processes affecting
water quality in this area.
Keywords: temporary ponds, marshland, streams, water quality, long-term study, eutrophication.
RESUMEN
Doñana alberga una extraordinaria variedad de sistemas acuáticos que se clasifican de forma general según su localización,
bien en las arenas o en la marisma, ya que su hidrología depende, fundamentalmente, de la geomorfología de sus cuencas.
La composición química de sus aguas varía en función de la lluvia, la evaporación, la descarga freática, las interacciones
biogeoquímicas en la interfase agua-sedimento y el estado ecológico de sus cuencas. La influencia de la cuenca se puede
estudiar a escalas distintas. La lluvia se infiltra fácilmente en la arena por lo que la escorrentía se convierte en un episodio
raro y breve que, sin embargo, afecta considerablemente al funcionamiento limnológico de las lagunas temporales sobre arenas. Por el contrario, la calidad del agua en la marisma de Doñana está afectada por procesos extensos y largos, como la
sedimentación, la eutrofización y la contaminación por metales pesados. Una revisión de la bibliografía limnológica permite realizar un estudio comparativo de las últimas dos décadas, centrado en las interacciones que tienen lugar en las fronteras
entre los ecosistemas terrestres y acuáticos que comparten las cuencas de estos cuerpos de agua. Actualmente, la zona Este
de la marisma está especialmente afectada por la baja calidad de las aguas de entrada en comparación con la zona Sur del
Parque Nacional que se encuentra más alejada de estos aportes. El agua que discurre por el último tramo del encauzamiento del río Guadiamar (“Entremuros”) inunda las marismas cercanas (“Lucio El Cangrejo Grande”) y mostró una correlación significativa entre la carga de materia inorgánica en suspensión y la concentración de P total (r=0.807, p<0.05) durante el periodo 2003-04, indicando la importante contribución del material particulado inorgánico en la eutrofización de esta
zona. En la zona Norte, la calidad del agua en los arroyos de la cuenca del Partido no ha mejorado significativamente en la
últimas dos décadas, a pesar de la construcción y funcionamiento de dos estaciones depuradoras de aguas residuales. Las
concentraciones de P total y fosfato disuelto en el agua se encontraron altamente correlacionadas (r=0.995, p<0.01) durante el periodo 2003-05. En las dos últimas décadas, se ha detectado un incremento de la concentración de NO3- en los siste-
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Serrano et al.
mas acuáticos estudiados en la marisma, incluído áquel con la mejor calidad del agua (“Arroyo de la Rocina”). A pesar de
que nadie duda que la gestión de los humedales require una estrategia a nivel de cuenca hidrográfica, los sucesivos proyectos hidrológicos que se desarrollan en Doñana no llegan a abarcar la gran escala espacio-temporal de los procesos que
afectan a la calidad de su aguas.
Palabras clave: lagunas temporales, marisma, arroyos, calidad del agua, eutrofización.
INTRODUCTION
Doñana is considered the most relevant wetland
area in Spain. Despite its significance for wildfowl, research into the limnological processes of
this vast wetland area started merely two decades
ago. Prior to this, there were some scattered
information published by re-known specialists in
aquatic invertebrates that visited the area following the tradition of naturalists “exploring”
Doñana in the previous 19th century and providing new zoological and botanical cites to the region. This was the case of surveys for the collection of rotifers (De Ridder, 1962), crustaceans
(Dussart, 1962, 1967, Estrada, 1973, Armengol,
1976) aquatic insects (Bigot & Marazanof, 1965,
Marazanof, 1967) and phytoplankton (Margalef,
1976). Later, microinvertebrates continued to be
studied in the ponds, particularly ciliates (PérezCabrera & Toja, 1989), rotifers (Mazuelos et al.,
1993) and crustaceans (Galindo et al., 1994 a,b,
Ruiz et al., 1996, Serrano & Toja 1998, Fahd et
al., 2000, Serrano & Fahd, 2005). The study of
macroinvertebrates was mainly focused on
Odonata, Heteroptera and Coleoptera from the
marshes (Montes, 1980, Montes & RamírezDíaz, 1982), and later resumed with the impact of
the red swamp crayfish (Gutierrez-Yurrita et al.,
1998, Alcorlo et al., 2004). An extensive survey
of aquatic and semiaquatic Coleoptera has been
recently performed (Millán et al. 2005). Aquatic
vertebrates such as amphibians have been extensively studied by Díaz-Paniagua (1979, 1988,
1990, Díaz-Paniagua et al., in press) while fish
have received some attention much later
(Fernández-Delgado et al., 2000). The study of
aquatic vegetation in Doñana started also with
early “explorations” to be later focused on particular aspects (García-Murillo et al., this issue). A
floristic revision of aquatic macrophytes is provided by García-Murillo et al. (1993) and Espinar
(2000). Aquatic vegetation has also been studied
with a limnological perspective (Bernués, 1990,
Duarte et al., 1990, Sousa & García-Murillo,
1999; Espinar et al., 2002), being the work by
Espinar (2004) the most extensive study on the
ecology and distribution of aquatic macrophytes
in the Doñana marshland.
The water composition of the Doñana marshland and the quality of the surface waters entering the marshes were thoroughly studied
during the 80’s and reviewed by Arambarri et al.
(1996). The first ecological studies (aimed at
relating biological populations to environmental
variables) were performed by Furest & Toja
(1981) and Montes et al. (1982). Later, the
Doñana ponds were typified according to their
hydrology and chemical composition (GarcíaNovo et al., 1991; Manzano, 2001), their hydrology and substrate (Bravo & Montes, 1993) or
their hydro-chemistry and littoral vegetation
(Muñoz-Reinoso, 1996). The larger size of
Santa Olalla and Dulce ponds made them suitable for many limnological studies (López et al.,
1991, Toja et al., 1991, Sacks et al., 1992,
Serrano et al., 1994, 1999, Toja et al., 1997,
López-Archilla et al., 2004) compared to the
rest of ponds (Bernués 1990, López et al., 1994,
Serrano & Toja, 1995). Additionally, the relevance of sediment in the functioning of these
shallow aquatic systems has been brought forward in numerous publications (Grimalt et al.,
1991, Jaúregui & Toja, 1993, López et al.,
1997, Díaz-Espejo et al., 1999, Álvarez et al.,
2001, Serrano et al., 2003). The interaction with
their surrounding terrestrial ecosystems has
also been studied under different perspectives,
such as vegetation-groundwater interactions
(García-Novo et al., 1996, Zunzunegui et al.,
1998), climate change (Sousa & GarcíaMurillo, 2003) and landscape management
(Muñoz-Reinoso & García-Novo, 2005).
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The aquatic systems of Doñana
13
Figure 1. Location of Doñana featuring different landscapes: marshland (1), stabilized sands (2), moving dunes (3), and the ecotone between aeolian sands and marshes. Antropic activities have deeply modified the landscape into: pine/gum-trees plantations (4),
irrigation fields, rice fields, dry-land plant cultivars, artificial wetlands for fish cultives, salt pans, and villages. Localización del
área de Doñana y sus diferentes tipos de paisaje: marisma (1), arenas estabilizadas (2), arenas móviles (3) y ecotono entre el
manto arrasado y la marisma. Se incluyen paisajes muy antropizados como plantaciones de pino/eucalipto (4), regadíos, arrozales, cultivos de secano, piscicultura extensiva, salinas y núcleos urbanos.
This wealth of limnological information enable
us to make a comparison of the main aquatic
systems of Doñana. We will focus on processes
affecting water quality that take place at the
frontier between terrestrial and aquatic ecosystems at different spatio-temporal scales.
STUDY AREA
The Doñana region (37°N, 6°W), extends along
the coastal plain of the Gulf of Cádiz from the
left bank of the estuary of the Guadalquivir river
to the estuary of the Tinto river, and inland from
the lower Guadalquivir River valley to the
uplands of “El Aljarafe” (Sevilla) and “Condado
de Niebla” (Huelva). It includes several territories with a different degree of environmental
protection covering over 100 000 ha: a Biological Reserve created in 1964, a National Park
(designated as a Ramsar site in 1982 and a
World Heritage Site by UNESCO in 1995) and
a Natural Park created as a surrounding protective area in 1989 (Fig. 1). At the same time, the
Doñana region constitutes a space featuring
the widest variety of pressures regarding the use
and assignment of water resources. In 1990,
Doñana entered onto the Montreaux Record of
Ramsar sites under threat because a number
of disturbances related to the conservation of
the marshes had the potential to change its ecological character. The Doñana region hosts a
population of nearly 180 000 inhabitants whose
activities are devoted to agriculture and tourism.
Rice fields occupy a vast extension on the east
margin (about 35 000 ha). Water for the growth
of rice is mainly provided by the Guadalquivir River, while 15 000 ha of irrigation
fields, scattered over the sandy soils on the west
and northern areas, are watered by the aquifer
which produces a groundwater withdrawal of
55-60 hm 3 per year (Cruz Villalón, 2005).
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Serrano et al.
Additionally, two large tourist resorts lie bordering the coast (“Matalascañas”) and the marshes
(“El Rocío”). The former concentrates over
63 000 people in summer, while the latter
attracts over half a million people during a traditional pilgrimage held in spring.
Doñana has a Mediterranean climate with
Atlantic influence, generally classified as dry
subhumid. Rainfall is quite variable, both within
a year and over the years, with a 580 mm yearly
average, about 80 % of which is distributed
throughout a wet period from the end of
September to the beginning of April. Summers
are very dry and hot, while winters are short and
mild. Water balance is generally deficient as
rainfall exceeds evapotranspiration only during
3-4 months of the year (Siljeström & Clemente,
1990). Potential evapotranspiration is very high
with a yearly average of about 900 mm
(Ménanteau, 1982). The deviation of rainfall to
the yearly average (CDYP) showed an irregular
sequence of hydrologic cycles (1st October-30th
September) during the past 16 years: 6 wet, 5
dry and 5 moderate cycles (Fig. 2).
Doñana started to be formed in the Quaternary
age when the estuary of the Guadalquivir River
was enlarged and reshaped by the formation of
sandy spits after the last postglacial transgression. The alluvial deposition of fine materials
brought about the filling of the former estuary
and progressively isolated it from the sea. In
1984, the construction of a levee on the right
bank of the Guadalquivir River (“Montaña del
Río”, Fig. 1) minimized the tidal influence on
the marshland which eventually became a continental formation (Clemente et al., 2004).
Consequently, the deposits on this ancient plain
present a rather heterogeneous lithology as it is
partially covered by aeolian sands, while the
central plain presents a saline silty-clay layer of
up to 100 m thickness with deltaic deposits of
sand and gravel increasing towards the north.
The permeability of the main geomorphological
units is very different: the aeolian sands correspond to an unconfined aquifer (with a shallow
watertable and several flow systems) while
groundwater is confined below the silty-clay
deposits of the floodplain. Both units composed
Figure 2. Deviation coefficient of rainfall to a yearly average (CDYP) of 563.2 mm recorded in the past 16 hydrologic cycles
(1989/90-2004/05). Confidence limits at 95 % of significance (±128 mm) are indicated by dotted lines. Each hydrologic cycle was
classified as wet (CDYP> 128 mm), dry (CDYP> -128 mm) or moderate (128 mm >CDYP< -128 mm). Coeficiente de desviación
a la media anual de lluvia (CDYP) de 563.2 mm durante los últimos 16 ciclos hidrológicos (1989/90-2004/05). Los límites del
intervalo de confianza al 95 % de significación (±128 mm) se indican con líneas punteadas. Cada ciclo hidrológico se ha clasificado como húmedo (CDYP> 128 mm), seco (CDYP> -128 mm) o moderado (128 mm >CDYP< -128 mm).
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an aquifer system of about 3400 km2 bottomed
by impermeable marine marls known as the
“Almonte-Marismas” aquifer (Llamas, 1990).
The depth of the aeolian sand deposits varies
from over 100 m on the coast to barely 10 m at
the northern edge of the Park. Groundwater
recharge is produced by rainfall infiltration in
the unconfined aquifer at an estimated rate of
200 mm y-1 (Vela, 1984). Groundwater in the
confined aquifer, however, flows at a rate of 0.11 cm y-1 (Konikow & Rodríguez-Arévalo,
1993). The silty floodplain and the sand mantle
also create two contrasting landscapes which
make contact with each other along an ecotone
locally known as “La Vera-Retuerta” (Fig. 1).
The aquatic ecosystems of Doñana are broadly
classified according to their location (on either
aeolian sands or marshland) as their hydrology
largely depends on the geomorphology on their
basins. Outside the protected areas, watersheds
are severely altered by human activities (Fig. 1).
The lower valley of the Guadalquivir River is
devoted to agriculture (traditional cultivars of
olive trees, irrigated crops, and rice fields) and
farming of cattle and horses. The watershed of
the Guadiamar River extends to the highlands on
the north (Sierra de Aracena) where soils are partially covered by scrub vegetation and plantations of gum-trees. Downstream, the river receives the drainage and seepage of the spoil heaps
from an open-cast polymetallic sulphide mine
(Aznalcóllar) through one of its tributaries
(Arambarri et al., 1996). The dumping of 5 hm3
of mud and acid water with high concentrations
of heavy metals in the mining accident of April
1998 flooded an extension of 2600 ha downstream. An extensive cleaning activity took place
in the river floodplain and the riverbanks were
later protected as a buffer area (“Corredor Verde
del Guadiamar”). The lower stretch of the
Guadiamar River runs between two levees
(“Entremuros”) built in 1956. Its final stretch is
canalised and connected to a dead arm-river
(“Brazo de la Torre”) that drains to the estuary of
the Guadalquivir River (Fig. 1).
The flatness of the vast floodplain occupied
by the marshland (about 23 000 ha) is altered at
a topographic scale of a few cm that creates
15
depressions (locally known as “lucios”) and
upper areas (“paciles”) which have the appearance of emerged islands (“vetas”) during heavy
floods. On the north-western area, surface water
to this plain is supplied by rainfall and the overflood of small water flows (“Arroyos de La
Rocina, del Partido, Cañada Marín and Cañada
Mayor”) which drain southwards into the Park
through a channel called “Caño de la Madre de
las Marismas del Rocío”. On the north-eastern
marshland, the Guadiamar River used to drain
southwards through numerous small creeks (or
“caños”), but most of its water-flow is presently
deviated to the estuary of the Guadalquivir river
so only a minor part of it reaches the marshes
through both pipes (“Caño del Guadiamar”) and
a complex channel network (“Entremuros-Brazo
de la Torre”) which also carries the drainage of
the nearby rice-fields. Quantitavely less important, but ecologically relevant, is the ground
water seepage along the ecotone (“La Vera”)
that provides humidity to grass meadows and
hygrophitic vegetation (“algaidas”), and feed
small creeks (“caños” and “sotos”) especially
during heavy rainy periods. In some spots of the
marshland, groundwater seepage maintains permanent sub-surface springs (“ojos”).
The Doñana marshland is flooded seasonally
by freshwater, depending on hydro-meteorological
conditions, as the Guadalquivir River is the only
permanent river in the area and its tidal influence
is currently minimal. This marked seasonality of
flooding periods followed by summer drought has
accentuated the endorreic character of the marshland (Clemente et al., 2004). The resulting ionic
composition of the “lucios” is dominated by Cland Na+ as a result of the solubilization of salts
from the sediment, the concentration of salts
being dependent on the frequency and duration of
flooding in each area (Clemente et al., 1998).
Currently operative salt pans are located on the
left bank of the Guadalquivir river. On the southeastern boundary of the Natural Park, 37 artificial
wetlands (total surface about 3000 ha) are devoted
to extensive fish cultures (“Veta La Palma”). Tidal
marshes, in contrast, have been reduced to a
narrow fringe along the banks of the Guadalquivir
River (Gallego & García-Novo 2003).
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Serrano et al.
Figure 3. Location of a variety of aquatic systems on aeolian sands within the Doñana Biological Reserve (wet meadows, temporary ponds, wet dune slacks and temporary streams). The names of the main “peridune ponds” fringing the moving dunes are indicated. Localización de diversos sistemas acuáticos sobre el manto arrasado en la Reserva Biológica de Doñana (pastizales de la
Vera, lagunas temporales, corrales encharcados y caños). Se indican los nombres de las principales lagunas peridunares situadas
a lo largo del frente de dunas móviles.
The Holocene aeolian sand mantle is composed
of several dune generations originally deposited
by marine drift (Vanney & Menanteau, 1985). A
system of moving dunes with several dune fronts
runs parallel to the coast-line with a NW-SE
direction (max. altitude: 30 m). Within the
Doñana Biological Reserve, the sand mantle is
mostly covered by Mediterranean scrub (stabilized dunes) with a species composition closely
following water availability which, in turn,
depends on groundwater flow systems of different
spatial scales (Muñoz-Reinoso & García-Novo,
2005). In this undulating landscape, hundreds of
small ponds appear when the water table rises
above the topographical surface during heavy
rains (Fig. 3). These ponds (locally known as
“lagunas”) are fed by freshwater (rainfall, runoff
and groundwater discharge) and have no surface
or groundwater connection to the sea so they cannot be properly called lagoons though they receive salts of marine origin through airborne deposition. Their groundwater feeding is relatively
complex due to changes in recharge and topographic boundaries that modify their connection to
different aquifer flow systems over time (Sacks et
al., 1992, Muñoz-Reinoso, 2001). They range
widely in size (from rain puddles to shallow
lakes) and in flooding duration (from days to
decades), but they all have been reported to dry
out eventually. Hence, they all are temporary
water bodies exhibiting wide fluctuations of
water level. Many attempts have been made to
classify them into different categories (permanent, semipermanent, seasonal, ephemeral), but a
short number of observations have produced contradictory results (García-Novo et al., 1991,
Bravo & Montes, 1993, Manzano, 2001). As a
whole, the Doñana ponds form a system of temporary water bodies of remarkable singularity in
Europe with a high protection status (Serrano &
Toja, 1995, Williams et al., 2001). The area protected within the Biological Reserve (Fig. 3)
covers a groundwater discharge surface of about
200 km2 (Allier et al., 1974). The density of
ponds in this area during winter floods is 6 ponds
per 100 ha (holding water for 1-6 months) and 1
pond in 100 ha (holding water for more than 6
months, García-Novo et al., 1996). A few artificial wetlands are maintained by groundwater
pumping, while digging water-holes in the ground
near ponds (“zacallones”) for cattle drinking
during dry periods is a very common practice.
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17
Aquatic systems on aeolian sands within the
Doñana Natural Park are also interesting and heterogenous (Fig. 1). A cliff formed by fossils dunes
(“Médano del Asperillo”) runs along the coast to
the west of the National Park and reaches a maximum altitude of 100 m. Rainfall and groundwater
seepage drain to the sea through small ravines
crossing the cliffs, providing shelter to some
hygrophytic species (Díaz-Barradas & MuñozReinoso, 1992). Ponds also appeared inland to
this formation, scattered over an area covering
about 25 000 ha called “El Abalario” (Sousa &
García-Murilllo, 1999). Vestiges of Atlantic peatbogs can be found there and along the margins of
La Rocina brook. The extensive plantations of
pines and gum-trees (“Coto del Rey”) shelter
numerous small ponds during floods. A total of
568 temporary wetlands on sandy soils have been
recorded by the local administration in the
Doñana region (Junta de Andalucía, 2002). The
only permanent aquatic system over sands (the
lagoon of Tarelo) fringes the pine plantation of
“La Algaida” on the left bank of the Guadalquivir
River. Its basin was artificially excavated for sand
and gravel extraction, and it is fed by rainfall and
groundwater seepage from the estuary of the
Guadalquivir River (Serrano et al., 2004).
CSIC). New physico-chemical data presented
here corresponded to four different locations.
The experiment with limnocorrals was carried
out in several temporary ponds (Doñana Biological Reserve): at November-December
1991 (Jabata pond) and November 1995 (Las
Verdes and Dulce ponds). Groundwater samples were collected from shallow piezometers
following the methodology used by López et
al. (1994). Limnocorrals (1 m diameter, 1 m
height) and runoff samplers (5 l volume) were
made of translucient impervious plastic as
described in Serrano et al. (1999).
Samplings of the floodplain and water-flows
entering the marsh area of “Lucio El Cangrejo
Grande” (Doñana Natural Park) were carried out
bimonthly from February 2003 to September
2004. Three sampling stations were located in
the central floodplain of this area; one sampling
station in the Guadiamar River at “Vuelta de la
Arena” (“Entremuros”), 3 sampling stations
along its canal (“Canal de Aguas Mínimas”) and
one sampling station at the rice field main outlet
(“Canal Principal de Desagüe”). In May 2004,
five sites within the lower Guadiamar River
watershed were sampled, both in the upper
stretch of “Entremuros” and in several tributaries: “Arroyo de la Cigüeña”, “Arroyo de Gato”,
“Arroyo Chucena” and “Arroyo Algarbe”.
Samplings of the Rocina-Partido watershed
were performed in November 2003, December
2004, March and June 2005. The Rocina brook
was sampled in its main stream just before
MATERIAL AND METHODS
Rainfall data was obtained from the meteorological station of “Palacio de Doñana” (RBD-
Table 1. Chemical composition of water inside limnocorrals (rainfall + groundwater discharge), outside them (rainfall + groundwater discharge + runoff), in surface runoff and phreatic water below several ponds at the onset of their filling period. Composición química del agua dentro de los limnocrrales (lluvia + descarga subterránea) y fuera de ellos (lluvia + descarga subterránea + escorrentía), de la escorrentía
superficial y el agua freática en algunas lagunas al comienzo de su llenado.
La Jabata pond
(7/11/91)
(5/12/91)
phreatic inside outside
E.C. (mS cm-1)
pH
Alkalinity (meq l-1)
i-P (µg l-1)
N-NO3- (µg l-1)
N-NO2- (µg l-1)
NH4+ (mg l-1)
0.81
7.0
0.7
111
1.4
1.9
0.96
0.87
6.4
0.6
9
15.4
2.4
0.28
0.41
7.3
0.2
9
14.0
1.5
0.45
Las Verdes pond
(12/11/95)
phreatic outside runoff
0.74
8.2
1.3
37
0.5
7.3
0.20
2.70
7.2
0.6
251
12.1
642.6
0.45
1.86
6.3
0.2
115
3.9
24.9
0.41
La Dulce pond
inside
12.5
7.3
2.5
208
150.6
0.08
(12/11/95)
outside
5.41
7.5
0.8
242
156.4
0.80
runoff
0.88
7.1
0.6
533
13.6
1.13
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Serrano et al.
draining to the marsh. The Partido stream was
sampled across the “Raya Real”, at its delta, and
in a tributary (“Caño Marín”).
Conductivity (compensated with temperature
at 20 ºC) and pH were recorded in situ. Water
samples (1-2 l) were collected, stored at 4 ºC and
filtered in the laboratory through Whatman GF/C
filters within the next 24 h. Suspended solids
were analyzed gravimetrically, in 3-4 replicates,
using previous dry filters (100 ºC). The concentration of inorganic suspended solids was estimated after ignition (450 ºC, 4 h). The rest of analyses were carried out in duplicates. COD
determinations were performed using KMnO4 as
oxidative agent. Total alkalinity was determined
by titration (Rodier, 1981). The concentration
of i-P was determined following the method of
Murphy & Riley (1962). Tot-P was analyzed as
i-P after acid digestion of the unfiltered water
sample with 0.5 M H2SO4 and K2S2O8 (0.5-1 g)
at 120 °C for 4 h (De Groot & Golterman, 1990).
The concentration of NO2- and NH4+ were determined by colorimetry (Rodier, 1981). The concentration of NO3- was measured as NH4+ after
complete oxidation with TiCl3 (Golterman, 1991).
RESULTS AND DISCUSSION
The chemical composition of the Doñana aquatic systems is mainly influenced by rainfall, evaporative concentration, groundwater discharge,
biogeochemical interactions at the sedimentwater interface, and the quality composition of
their watersheds. The influence of the watershed
can be studied at different scales of observation.
Rainfall usually infiltrates easily through sandy
soils so surface runoff rarely reaches the ponds
(only for very short time-lapses during heavy
rainfall). The water quality of the Doñana
marshland, on the contrary, is influenced by
long-term processes taking place on large-scale
areas (such as sediment deposition, eutrophication and heavy metal pollution).
The chemical composition of the unconfined
aquifer is considered rather uniform, dominated
by Ca(HCO3)2 and with a salinity usually lower
than 500 µg l-1, except at discharge areas becau-
se of the influence of biochemical processes
(Llamas, 1990). As a whole, the aquifer shows a
vertical gradient in salinity, from brine-water
near the land surface to freshwater at 80 m of
depth (Konikow & Rodríguez-Arévalo, 1993).
Below the confined aquifer, salinity increases in
a NE-SW direction pushing the interphase between the aquifer recharge and the fossil marine
groundwater to the NE boundary of the Park
where extensive irrigation fields happen to be
developing (Plata & Ruiz, 2003).
The high proportion of NaCl in rainwater
due to marine influence affects the composition
of shallow groundwater, but the high ratio of
Mg 2+ over Na + suggests that wet and dry
atmospheric deposition has not yet been properly addressed (Lozano, 2004). The successive
cycles of flooding and evaporation in the discharge areas have enriched in NaCl the shallow
free groundwater (or phreatic) below the pond
basins as it was shown by López et al. (1994) in
water samples collected in piezometers (<2 m
depth) during a dry hydrologic year. Ca2+ dominated over Na + only in a small depression
within the dune tail where recharge dominated
over discharge flow and water was temporary
deposited. Conductivity and total alkalinity of
groundwater below the ponds showed little
relationship to morphometry or trophic state,
suggesting the existence of local flow systems.
Limnocorral experiments were used in the
temporary ponds to isolated both surface and
groundwater sources at the onset of their filling
period. The chemical composition of water inside the limnocorral (rainfall+groundwater discharge) and outside it (rainfall+groundwater discharge+runoff) was compared to the phreatic
water below the ponds and to surface runoff
collected in the watershed (Table 1). The chemical composition of the water filling the ponds
resulted from an interaction of both surface and
groundwater sources during discharge, but general trends were difficult to predict. Firstly because water budgets have not yet been elucidated
with sufficient resolution and, secondly because
of the impact of sparse local events on these shallow systems. For example, heavy rainstorms
(>80 mm) that only represented 1.4 % of total
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19
observations in a 7-year record (Serrano et al.,
1999) can disrupt the pond development and
revert it to earlier successional stages (Toja et
al., 1991). De Castro-Ochoa & Muñoz-Reinoso
(1997) elaborated a multiple regression model
for water-table fluctuations of the dune wetslacks that depended on rainfall, time lapse between two consecutive measurements and present
depth. They found, however, that groundwater
discharge at the ponds did not fit into their dune
aquifer model because pond feeding did not only
depend on rainfall. A hydrologic budget for the
largest pond during the rainy season (OctoberMarch) estimated that surface sources (rainfall
and runoff) accounted for 48 % of the water
input during very dry years (<250mm of annual
rainfall), but reached 100 % during extremely
wet cycles (>1000 mm of annual rainfall). The
rest of the water input was due to groundwater
discharge, from both a shallow phreatic and a
deep water table (Sacks 1989).
A review of studies on the ionic composition
in the Doñana aquatic systems reveals the
influence of the different scales of observation
in each study. The hydroperiod (or duration of
water on the surface) and the water origin are
relevant hidrologic features that influence the
ionic composition of water, but their assesment
is greatly affected by the duration of the study
period. The first extensive survey (47 ponds) of
water chemical composition was performed by
García-Novo et al. (1991) during the heavy floods of winter 1990. Again, Na+ and Cl- were the
dominant ions in all pond waters. The ratio of
Ca2+ over Na+ was not a good predictor compared to the Mg2+/Na+ ratio, probably due to the
local effect of microtopography and vegetation
on CaCO3 precipitation. A ratio of Mg2+ over
Na+ higher than 0.25 (in meq/l) corresponded to
discharge areas, and a lower ratio indicated
recharge areas where water runoff could be temporary deposited due to the presence of a high
concentration of organic matter and/or clay in
the sediment. They segregated the ponds in
three discharge groups according to their location, mineralization and trophic state (Fig. 3): 1)
ponds in “Las Naves”, 2) those along the ecotone between the moving dunes and the stabilized
Table 2. Maximum and minimum values of water conductivity
(20 ºC) recorded in the water surface of the most visited ponds from
1989/90 onwards, indicating wet (w), dry (d) and moderate cycles (m).
Valores máximos y mínimos de la conductividad (20 ºC) registrados
en la superficie del agua de las lagunas más visitadas desde 1989/90,
indicando los ciclos húmedos (w), secos (d) y moderados (m).
Hydrologic
cycle
Santa Olalla
Max
1989/90w
1990/91 m
1991/92 m
1992/93 d
1993/94 d
1994/95 d
1995/96 w
1996/97 w
1997/98 w
1998/99 d
1999/00 m
2000/01 w
2001/02 m
2002/03 m
2003/04 w
2004/05 d
4.1
4.3
16.5
28.4
1.1
0.7
0.6
11.6*
9.2*
5.7*
8.3
7.7
19.3
Min
Dulce
Taraje
Max Min
Max Min
0.4
5.8
1.3
7.1
2.3 12.5
4.4 10.6
6.0
2.0
0.7
5.9
0.3
0.6
0.4
0.5
1.7* 16.6*
1.9* 3.5*
1.7* 2.5*
2.3
3.7
1.1
3.3
1.9
1.9
0.2
0.8
1.0
1.2
1.7
0.5
0.3
0.4
1.6*
1.0*
0.6*
0.8
0.4
1.0
7.4
10.6
22.0
15.4
13.3
1.8
2.3
8.5
11.1
5.6
3.4
2.7
4.4
14.9
0.1
1.6
4.1
6.9
1.2
0.3
0.4
0.5
6.1
1.5
0.9
1.7
0.4
8.6
* LÓPEZ-ARCHILA et al. (2004)
sands (“peridune pond” system), and 3) those
between the stabilized sands and the marshland
(“La Vera”). Later, Muñoz-Reinoso (1996)
enlarged this classification to 5 groups by including the wet-slack formation of the moving
dunes, and dividing the “peridune ponds” in two
other groups according to pond size.
Water pH is usually alkaline in the discharge
areas, while acidic water (pH 4-5) has been occasionally reported when rainfall is temporary deposited on rich organic soils. The combination of
alkaline waters over siliceous sand basins makes
these water bodies rather singular compared to
other European wetlands (Serrano & Toja, 1995).
Hydrochemical classifications of ponds based
on absolute limits have, so far, proved evasive as
the ionic composition of the Doñana ponds
change widely in time (Serrano & Toja, 1995).
Table 2 shows the change in water conductivity
recorded in some of the most visited ponds
during the last two decades. In Taraje pond, conductivity ranged from 0.1 to 8.5 mS cm-1 during
wet years, and from 1.2 to 22.0 mS cm-1
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Serrano et al.
during dry years. Low maximum conductivity
values corresponded to wet cycles while high
minimum values to dry years in all ponds, but
the relationship between conductivity and rainfall was not linear because conductivity also
reflected the past conditions of previous dry
cycles: the minimum conductivity values were
not attained in 1995/96 despite it was the wettest
cycle in the record (1093 mm) as it had been preceeded by several dry years.
The conductivity range was widest in the
larger pond (Santa Olalla: from 0.3 to 28.4 mS
cm-1). Its larger watershed ensured a higher contribution of rainfall and the flooding of a considerable extension of soil and vegetation which
accounted for the solubilization and leaching of
salts which progresively concentrated in the
water as evaporation proceeded. Sacks (1989)
proved that the higher mineralization of Santa
Olalla pond was also achieved by the downgradient movement of solutes through seepage from
its neighbouring ponds (Dulce and Las Pajas)
due to its larger evaporative discharge and lower
altitude. She estimated that the residence time
for Cl- was 6 years in Dulce pond and 34 years in
Santa Olalla pond. During extremely wet
periods, this pond complex (Santa Olalla-DulceLas Pajas) behaves as a flow-through floodplain
of about 100 ha (García-Novo et al., 1991)
that evacuates water and solutes to the nearby
marshes through intermittent creeks.
Flooding and runoff also contribute to the loading of nutrients and dissolved organic compounds to the ponds. The concentration of i-P in
surface runoff water was relatively high compared to groundwater discharge (Table 1), but these
differences were brief: ten days were enough to
equilibrate a six-fold difference between the i-P
concentrations inside and outside a limnocorral
in Las Verdes pond during the filling period of
1991 (Serrano & Toja, 1995). A careful study
of the P-fractional sediment composition showed
that the incoming i-P was partially adsorbed by
the sediment during the first weeks of the filling
period. A pond with a sustrate rich in Fe
(>10 mg g-1 dw) significantly increased its pool
of inorganic P-bound to Fe, while another rich in
organic matter (>17 %) increased its fraction of
org-P solubilised by EDTA (Díaz-Espejo et al.,
1999). Serrano et al. (1999) showed that the
Doñana ponds received i-P from their watershed
during heavy rainstorms after drought. In Dulce
pond, i-P concentration was 100 times higher in
the littoral than in the open-water area. Soil samples from the sandy uplands and the floodplain
meadow, fresh scrub (Halimium halimifolium),
and cattle manure leached i-P concentrations
higher than 0.9 mg g-1 dw in distilled-water under
laboratory conditions, suggesting that this material was a source of P to runoff water draining to
the pond shore. The slow decomposition rate of
litter in arid sandy soils of Doñana can explain
the accumulation of nutrients in the upland areas
of the pond watersheds where leaching of soluble
compounds from litter can last up to 4 moths
after deposition (Gallardo & Merino, 1993). The
accumulation of organic matter, in turn, accounts
for the dominance of organic P-fractions in the
pond sediment (Serrano et al., 2003). The contribution of terrigenous lipids to the sedimentary
composition (Grimalt et al., 1991) and the detection of organic P-compounds derived from vegetation in the sediment such as phytate (Serrano et
al., 2000a) proved the strong influence of the
watershed on these aquatic systems.
Rainfall itself accounted for the leaching from
fresh vegetation of soluble polyphenols (Serrano, 1992). During floods, dissolved organic matter is washed from the fringing vegetation and
carried by runoff water to the ponds where the
concentration of DOC can reach up to 120 mg l-1
during heavy rainfall (Serrano, 1994). Although
the input of DOC by rainfall and runoff water is
a common process in all aquatic systems: e.g.
10-25 % of total yearly input in Canadian lakes
(Schindler 1992), the singularity of this process
here relies on the unpredictability of the flooding events in the Mediterranean climate.
Consequently, the Doñana ponds exhibit very
contrasting conditions in different years that led
Allier et al. (1974) to state that they had distrophic phases during floods. During distrophic
phases, pond water shows a very dark colour due
to high concentrations of DOM. Flooding itself
produces the dilution of solutes and particulates,
bringing about a drastic reduction of the phyto-
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plankton populations and disrupting their previous succesional development (Toja et al.,
1991). No direct effect of polyphenols on primary production has been proved in the ponds
(Serrano et al., 1993) despite these compounds
are able to chelate micronutrients (Serrano &
Guisande, 1990). Vertical attenuation coefficients of PAR in the water of a pond rich in dissolved polypehnols were similar to those of the
hypereutrophic Santa Olalla pond where chlorophyll a concentrations reached upt to 2 mg l-1
(Serrano et al., 1994). Flooding is, therefore, the
cause of both the phytoplankton dilution and the
polypehnolic enrichment. The perturbation caused by flooding in the ponds is so extreme that it
resembles the flood pulse of tropical floodplains,
in which heavy floods can clean water bodies
and rearrange the communities to earlier successional stages (Junk & Weber, 1996). The drainage and vegetation pattern in the watershed determines the extent and variability of the runoff
input to each pond during floods (García-Novo
et al., 1996). Xerophytic scrub (Halimium halimifolium) growing in the upland areas of the
pond watershed leached a higher amount of
polypehnols than bulrushes from the floodplain
under artificial rainfall experiments (Serrano,
1992). The fate of the dissolved polyphenols was
also different in each pond till they dissapeared
on the dry sediment (Serrano, 1994). Dissolved
polyphenols are easily degraded by sunlight
so photo-oxidation can account for their disappearance in the water. The depth of the water
column limits the amount of sunlight that can
penetrate in each pond and thus the extent of
the photo-degradation of polyphenols in each
pond (Serrano et al., 2000b).
The expected development of the Doñana
ponds during a moderate hydrologic cycle would
be a low concentration of polypehnols in the
water which, in combination with a moderate
water depth, would allow the growth of extensive
macrophyte beds. Biomass of submersed vegetation have been reported to reach up to 450 g dw
m-2 in Dulce pond (López et al., 1991). As the
ponds dry out, organic matter is partially mineralised on the dry sediment (Serrano, 1992).
Hence, the concentration of sediment organic
21
Figure 4. Map of the surface watersheds in the Doñana region.
The Guadalquivir River is the only permanent river. Other
water-flows are intermittent under the tidal influence of the
former or by the effect of urban sewage effluent. The rest of
water-flows are considered seasonal. Some artificial canals
are distinguished. Location of presently working waste-water
treatment plants (WWTP) is indicated. Sampling sites in the
watershed: “Lucio de los Ánsares” (1), “Lucio Cangrejo
Grande” (2), “Guadiamar-Entremuros” (3), “La Cigüeña” (4),
“Gato” (5), “Chucena” (6), “Algarbe” (7), El Partido stream
(8), La Rocina brook (9). Mapa de la cuenca superficial de
drenaje en el área de Doñana. El río Guadalquivir es el único
curso de agua permanente. Otros ríos y arroyos son intermitentes bajo la influencia mareal de aquel o debido al efluente
de las depuradoras. El resto de los cursos de agua se consideran estacionales. Se distinguen algunos canales artificiales.
Se indica la localización de las plantas depuradoras de residuos urbanos (WWTP) operativas. Los puntos muestreados en
la cuenca fueron: Lucio de los Ánsares (1), Lucio Cangrejo
Grande (2), Guadiamar-Entremuros (3), arroyo de la Cigüeña
(4), arroyo del Gato (5), arroyo de Chucena (6), arroyo
Algarbe (7), arroyo del Partido (8), arroyo de La Rocina (9).
matter is inversely correlated to the duration of
flooding (Jaúregui & Toja, 1993). During dry
periods, the vegetation pattern surrounding the
ponds changes rapidly: hygrophytic species showed regression while xerophytic species advanced to lower areas (Zunzunegui et al., 1998).
A flooding period following a long drought, therefore, will produce a larger impact of the
watershed on the pond water composition regarding nutrient and dissolved organic matter concentrations. The variability of the hydro-me-
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22
Serrano et al.
teorological conditions in the area ensures the
unpredictability of this terrestrial-aquatic frontier. Furthermore, vegetation changes induced by
antropic disturbance in relation to groundwater
abstraction add more uncertainty to this interaction (Muñoz-Reinoso, 2001).
Long-term and large-scale processes have also
been reported to alter the water composition of
water bodies on aeolian sands. The alluvial aquifer of the Guadalquivir River has long been
exposed to a severe NO3- pollution due to the
excessive use of fertilizers in its cacthment.
About 8 mM of NO3- were already reported in
the estuary in 1987 (I.T.G.E., 1993). Groundwater pollution was most likely the cause of the
high concentrations of NO3- (up to 1.5 mM)
found in the water of Tarelo lagoon fed by
groundwater seepage from the estuary of the
Guadalquivir River (Serrano et al., 2004). The
study of both ecological and historical records
showed that the vegetation of Doñana has been
deeply affected by management practices since
the first quarter of the 17th century (GranadosCorona et al., 1998). Deep sediment cores from
Dulce and Santa Olalla ponds showed that Tot-P
concentrations had increased exponentially since
1960’s suggesting that recent changes in land
management have contributed to their eutrophication (López et al., 1993). A decrease in the
availabity of water may also have contributed to
a decrease of their water quality. In a climatic
context, Sousa & García-Murillo (2003) sugges-
ted that Doñana is under an overall tendency of
dessication due to an increase in aridity initiated
at the end of the 19th century. More recently, the
groundwater pumping for urban water supply to
the tourist resort of Matalascañas has been
reported to damage nearby ponds (Brezo and
Charco del Toro, Fig. 3) located at less than 1 km
to the pumping area during the drought period of
1992-94 (Serrano & Serrano, 1996).
The Doñana marshland is basically fed by
direct rainfall on its floodplain and by several
watersheds: Guadiamar River (1180 km2), the
Partido stream (300 km2) and La Rocina brook
(about 1000 km2). Secondarily, it is fed by
groundwater discharge along the ecotone and
through seepage streams (“sotos”). Lastly, the
tidal influence from the estuary of the
Guadalquivir River is minimal nowadays. The
relative contribution of each water source is
expected to vary according to dry, moderate and
wet cycles. On average, the water flow from the
Guadiamar River is 3-7 times larger than the rest
of sources, but since its drainage was modified
by channels and levees in 1956, most of its water
drains directly to the estuary of the Guadalquivir
River (Espinar, 2004). At present, only several
depressions located on the eastern and southern
areas of the National Park are flooded by the
Guadiamar River through a network of pipes
and pumping stations. Nine outlets along the
levee (“Montaña del Río”), eight of them provided with floodgates, maintain the confinement
Table 3. Range values and means (or mean and standard deviation) of water conductivity (20 ºC), pH, nutrient concentrations and suspended
solids (s.s.) in two marsh sites at different sampling periods. Intervalo máx-min y valores medios (o media y desviación estándar) de la conductividad (20 ºC), pH, concentración de nutrientes y sólidos en suspensión (s.s.) en dos zonas de muestreo de la marisma durante periodos
distintos de estudio.
1981-82*
Lucio Ánsares
E.C. (mS cm-1)
pH
N-NO3- (mg l-1)
i-P (µg l-1)
Tot-P (µg l-1)
s.s. (mg l-1)
1997**
Lucio Cangrejo
Lucio Ánsares
2002-04***
Lucio Cangrejo
max
min
mean
max
min
mean
mean
SD
max
min
mean
35.7
10.2
6.2
620
628
7.9
6.9
1.3
33
19
21.4
8.6
3.8
195
186
36.7
9.2
10.8
587.4
124
20.2
6.9
3.2
78.3
18
31.2
7.9
5
228.4
50
4.25
9.3
0.2
8
-
0.37
0.42
0.03
3
-
8.0
8.5
4.7
39
974.7
433
1.6
7.8
0.1
0
45.5
33
4.4
8.1
1.2
13
179.6
102
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Table 4. Concentrations of dissolved nutrients and COD at several
sites within the lower Guadiamar river watershed. Concentración
de nutrientes y DQO en varios puntos de la cuenca baja del río
Guadiamar.
Entremuros Cigüeña Gato Chucena Algarbe
N-NO3- (mg l-1)
N-NO2- (mg l-1)
N-NH4+ (mg l-1)
i-P (mg l-1)
COD (mg l-1)
0.33
0.10
0.02
0.26
11.36
0.46
0.21
0.02
0.25
12.48
0.57
0.22
0.02
0.22
15.60
0.46
0.47
0.03
0.11
14.88
0.46
0.04
0.02
0.02
14.56
of the water outflow during winter and spring
though, occasionally, the levee is overflowed at
heavy floods. During moderate and dry cycles,
freshwater is accumulated in the depressions
(“lucios”) causing the solubilization of salts
from the top sediment. As the confined water
evaporates, their salts and particulates become
increasingly concentrated. The resulting ionic
composition of the “lucios” is dominated by Cland Na+, the concentration of salts being dependent on the frequency and duration of flooding
in each area (Clemente et al., 1998). One of the
eight projects planned for the hydrologic regeneration of the marshland (“Doñana 2005”) plans
to eliminate part of this levee in order to increase the connection between the southern marsh
and the estuary of the Guadalquivir River.
The water quality of the waters entering the
Doñana marshland has been extensively studied
in the past decades, while the water composition
within the “lucios” is scarcer. Table 3 shows a
comparison in time of the water composition at
two sites (Fig. 4): “Lucio de los Ánsares” between 1981-82 (I.N.I.A., 1984) and 1995-97
(Espinar, 2004), and “Lucio Cangrejo Grande”
between 1981-82 (I.N.I.A., 1984) and 2002-2004
(present data). Despite the recent data was recorded during a much wetter period, the physicochemical changes recorded in these “lucios”
were mainly due to major hydrologic changes
that took place in the past. The levee that isolated
the marshland from the estuary of the Guadalquivir River (“Montaña del Río”) was built in
1984 and enlarged in 1998. Before its construction, the southern marshland received a large
tidal influence that was reflected in both the large
23
conductivity of the water and the high concentration of nutrients in the sampling of 1981-82.
The hydrology of the “Lucio Cangrejo
Grande” had already been altered in 1956 with
the construction of two levees that directed the
water flow of the Guadiamar River to the
estuary through a channel connected to a dead
arm-river (“Brazo de la Torre”). The 8-fold difference in the average water conductivity after
two decades was most likely due to a dilution of
the tide water. At the end of September, rice
pads located on the east, are drained through an
outlet which is connected to the canal feeding
the study area. The salinity of the water drained
by this outlet is relatively low because rice
require less than 1 g l-1 for growth. As this
freshwater outlet was not operative till 1988, the
study area received the direct influence of
the estuarine water which presented an average
water conductivity of 33.3 mS/cm at the mouth
of the estuary during the sampling of 1981-82.
The evaporation of water in the rice pads during
the growth season increased the water conductivity of the outlet only slightly (2 mS cm-1) due
to the recirculation of water through the rice
fields. Therefore, the freshwater output of
nearby rice pads contributed to reduce the salinity of the estuarine water during high tide.
Although the concentration of i-P was lower in
the recent sampling, the high concentration of
suspended matter produced a high concentration
of Tot-P as both variables were strongly correlated (r=0.978, p<0.01). The concentration of
NO3- did not change considerably after two
decades, suggesting that this area has received a
high nutrient load for a long time. Furthermore,
the nutrient concentrations of several tributaries
to the Guadiamar River during a wet period in
the spring of 2004 (Table 4) was similar to that
found by previous authors more than two decades ago (Cabrera et al., 1984). Additionally, the
Guadiamar River has created a chronic pollution
of heavy metals in the area due to both resuspension of sediments from the river bed during
floods (Cabrera et al., 1984, Arambarri et al.,
1996) and direct overflow of the mine dam.
Such overflow was recorded during the winter
floods of 1989 (Dolz & Velasco, 1990). This
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Serrano et al.
Table 5. List of main studies on water quality of Doñana marshland (before the mining spill of Aznalcóllar in 1998), indicating type of pollutant, location and reference. Lista de los principales estudios sobre la calidad del agua en la marisma de Doñana (antes del desastre minero de
Aznalcóllar in 1998), indicando tipo de contaminante, localización y referencia.
pollutant
location
Reference
Guadiamar-Entremuros
Olive-mill effluents
Petroleum hydrocarbons
Organochlorine plaguicides
Guadalquivir estuary
marshes
marshes
Heavy metals
Guadalquivir estuary
marshes
Fertilizers
El Partido stream
Particulate organic matter
El Partido stream
chronic pollution probably accounted for the
finding of a 3-times higher heavy metal concentration in seston of reference sites from Doñana
compared to unpolluted sites reported in the
literature during the monitoring of the
Aznalcóllar mining spill (Prat et al., 1999).
Unfortunately, the quality of the waters entering the Doñana marshland has imporved little
over the past decades regarding the concentration of dissolved nutrients despite olive mill
effluents have been significantly reduced and
several waste-water treatment plants have been
developed (in both cities and rural areas). The
variety of pollutants detected in the surface
water-flows of the Doñana watershed reflects a
large concern on this issue (Table 5). The evaluation made by Arambarri et al. (1996) over the
1980’s decade concluded that the waters of the
Partido stream ought to be treated as they contained high concentrations of organic matter and
nutrients that were hazardous to other aquatic
ecosystems (other streams and the freshwater
Cabrera et al., 1984; 1986
Albaigés et al. 1987
Cabrera et al., 1986
Arambarri et al., 1984; 1996
Albaigés et al., 1987
Cabrera et al., 1984; 1987
González et al., 1987
Ramos et al., 1994
Arambarri et al., 1996
Zurera et al., 1987
Cabrera et al., 1987
Arambarri et al., 1984;1996
Ramos et al., 1994
marsh of “La Madre de las Marismas del
Rocío”). The Rocina brook, in contrast, rated
the highest water quality, while the estuary of
the Guadalquvir showed an aceptable water quality except for its high salinity.
Another report on the quality of water-flows
entering the Doñana National Park estimated
that the Partido stream carried a yearly nutrient
load of 62.4 Tm of organic matter (COD), 2.7
Tm of P and 6.4 of N (Toja et al., 1992). More
than a decade later and two waste-water treatment plants later (located in Almonte and El
Rocío villages), the Partido stream is still
highly polluted compared to the Rocina brook
which does not receive urban waste-water
(Table 6). Nevertheless, there has been a considerable increase of NO3- concentration in both
water-flows during the past decade probably
due to an increase in cultivated land and fertilizer applications. The influence of this pollution
on the eutrophication of the nearby marshes
should not be overlooked. During dry periods,
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25
such as the spring of 2005, the effluents from
the waste-water treatment plants of El Rocío
and Almonte villages run downstream into the
National Park through the “Madre de las
Marismas del Rocío” creek.
Besides the impact of waste-water treatment
plants on river water quality, which is by itself
an issue of great concern causing a severe disturbance at a regional scale (Martí et al., 2004),
the Doñana marshland also features a large sedimentation rate which is contributing to its eutrophication. The concentration of suspended solids
has increased in many areas during the past
decades (Tables 3 and 6). In the lower strecht of
the Guadiamar River (“Entremuros”), the waterflows feeding the nearby marshes (“Lucio de El
Cangrejo Grande”) showed a significant correlation between inorganic suspended solids and
Tot-P (r=0.807, p<0.05) during 2003-04. Most
of the suspended matter was inorganic (6093 %) containing CaCO3 particles adsorbed to P
in the form of hydroxi-apatite. Previous works
have reported that soil particles are eroded and
resuspended from the Guadiamar River watershed during floods, bringing about an increase
of nutrient concentrations in the downstream
waters (González Quesada et al., 1987,
Arambarri et al., 1996). Even in the absence of
floods, the higher speed of the water current
through artificial canals promote the erosion of
its bed and margins increasing the concentration
of suspended solids in the water (Mintegui,
1999). In constrast, Tot-P concentration in the
Partido watershed was not significantly correlated to suspended solids but to i-P concentration
(r=0.995, p<0.01), indicating the sewage origin
of its water. This area has received great attention because of the huge scale of its riverbank
erosion during floods. Since 1995, a sediment
load of nearly 3 x 106 m3 has been deposited in
the marshes (Mintegui 2005). Both processes of
eutrophication and sediment deposition have run
parallel to the expansion of emergent macrophytes in the Doñana marshland, bringing about a
drastic reduction of its open-water areas since
1956 (Espinar, 2004). At the same time, the
recent expansion of introduced species, such as
Azolla filiculoides, in the Doñana marshes could
be a consequence of eutrophication as this N2
fixing symbiont thrives when the N/P ratio is
unbalanced in the ecosystem.
A proper wetland management requires a
watershed approach. At the same time, the quality of the waters draining through a watershed
will reflect the quality of the corresponding
terrestrial ecosystems (Margalef 1983). In
the Doñana region, groundwater recharge
takes place by rain infiltration on the aeolian
Table 6. Mean and range values of water conductivity (20 ºC), concentration of nutrients, suspended solids (s.s.), and COD within the watersheds of El Partido stream and La Rocina brook during two sampling periods. Valores medios e ntervalo máx-min de la conductividad (20 ºC),
concentración de nutrientes, sólidos en suspensión (s.s.) y DQO en la cuenca de los arroyos del Partido y La Rocina durante dos períodos distintos de muestreo.
1991-92*
El Partido
E.C. (mS cm-1)
N-NH4+ (mg l-1)
N-NO2- (mg l-1)
N-NO3- (mg l-1)
i-P (mg l-1)
Tot-P (mg l-1)
s.s. (mg l-1)
COD (mg l-1)
*TOJA et al. (1992)
2003-04
La Rocina
max
min
mean
max
min
1.42
27.4
0.04
0.14
3.2
5.4
103
90.4
0.53
3.6
<0.01
0.02
0.2
0.3
33
19.3
1.13
9.9
0.02
0.08
1.3
2.8
67
61.4
0.63 0.21
1.4
0.2
0.03 <0.01
0.89 0.10
0.13
0
0.8
0.1
175
22
50.3
21
El Partido
mean
max
0.39
0.5
0.01
0.38
0.04
0.3
97
31.8
1.29
15.0
0.28
11.45
3.4
4.2
311
-
min
mean
0.44 0.97
0.3
8.6
0.14 0.19
<0.01 7.57
0.4
1.7
1.1
2.7
12
161
-
La Rocina
max
min
mean
0.57 0.41 0.49
7.5
0.5
2.9
0.02 <0.01 0.015
1.26 0.01 0.67
0.08 0.03 0.005
2.0
0.4
1.2
92
32
62
-
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Serrano et al.
sands covering about 3400 km2. The Doñana
marshland covers a vast extension of both
deeply transformed areas (dessicated marsh,
rice fields, irrigation fields) and less transformed marshes (National and Natural Parks)
which are hydrologicaly connected to each
other. The surface watershed draining to the
marshland extends by narrow corridors to
the pyritic deposits of the northern uplands,
while the lower valley of the Guadalquivir River
is the recipient of waters draining through a
huge watershed of more than 57 500 km2. Both
surface and groundwater resources ought to be
extremely difficult to manage at this massive
scale and so, successive hydrologic projects
carried out in Doñana have failed in many
aspects. The recently developed hydrologic
regeneration plan for Doñana (Doñana 2005)
will bring a better connectivity to the Guadalquivir River, but rice fields continue to be
segregated in this estrategy despite they are not
isolated from the Doñana aquatic systems, neither by water surface nor by aerial depositions.
The outlet water from rice-pads is flooding the
nearby marshes (on the eastern margin of both
the Natural and the National Park) every year at
the end of the rice growing-season. Instead of
lamenting on the issue, rice fields could become
a source of freshwater in the hydrologic budget
of the nearby marshes which, in turn, will create
more concern on the control and monitoring of
fertilizers and pesticides in the area.
ACKNOWLEDGEMENTS
We are grateful to all the people that have participated in our limnology group during the past
decades: A. Furest, T. López, C. Guisande, N.
Gabellone, M. A. Casco, J. Prenda, J. C. Muñoz,
J. Jaúregui, N. Mazuelos, J. A. García-Sánchez,
M. D. Galindo, F. García-Bouzas, A. DíazEspejo, J. L. Moreiras, I. Calzada-Bujak, K.
Fahd, M. D. Burgos, A. Plazuelos and E. Alcalá.
Recent data corresponded to research projects
supported by “Consejería de Medio Ambiente,
Junta de Andalucía” and the Ministry of
Envrionment (Doñana 2005).
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Doñana. Master Thesis. Universidad Complutense
de Madrid. 221 pp.
WILLIAMS, P., J. BIGGS, G. FOX, N. PASCALE, & M.
WHITFIELD. 2001. History, origins and importance
of temporary ponds. Freshwater Forum, 17: 7-15.
ZUNZUNEGUI, M., M. C. DÍAZ-BARRADAS & F.
GARCÍA-NOVO. 1998. Vegetation fluctuation in
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Mediterranean dune ponds in relation to rainfall
variation and water extraction. Appl. Veg. Sci., 1:
151-160.
ZURERA COSANO, G., F. RINCÓN LEÓN, L. M.
POLO VILLAR, M. JODRAL VILLAREJO, R.
JORDANO SALINAS & R. POZO LORA. 1987.
Contaminación por plomo, cadmio y mercurio en
aguas y sedimentos del río Guadalquivir. In: Actas
del IV Congreso Español de Limnología. Sevilla
(Spain): 307-314.
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Limnetica, 25(1-2): 33-56 (2006)
Basin scale transport in stratified lakes and reservoirs: towards
the knowledge of freshwater ecosystems
Francisco Rueda Valdivia
Instituto del Agua y Departamento de Ingeniería Civil – Universidad de Granada. C/Ramón y Cajal, 4 – 18071 –
Granada. Spain. [email protected]
ABSTRACT
The physical processes of transport and mixing exerts a profound influence on the biogeochemistry of aquatic systems: not
only they determine the position of particulate and dissolved substances contained in the water at any given time, but they also
contribute to determine the environment in which biogeochemical reactions occur. The physical behaviour of four aquatic
systems are examined, trying to illustrate the mechanisms of transport at play in any given environment as determined by both
the properties of the external forcing and the characteristics of the system itself (morphology, in particular). Scaling arguments, based on Lake and Wedderburn numbers, the Rossby radius of deformation and the nominal residence time, provide a
simple and first order four-dimensional framework in which one can sketch and classify the physical behaviour of the system.
The four case examples analyzed occupy, in that four dimensional framework, positions which are far apart. Transport in natural systems is shown, by means of examples, to be a highly dynamic feature determined by the exact details of how mass and
energy enters (their magnitude, frequency, exact location in time, their spatial variability). Some of those ‘second order’
effects, not captured in the first order analysis based on simple scales or non-dimensional parameters, are analyzed.
Keywords: transport, stratified lakes, basin-scale.
RESUMEN
Los procesos físicos de transporte y mezcla ejercen una influencia muy marcada sobre la biogeoquímica de sistemas acuáticos: no solo controlan la posición en el espacio y en un determinado momento de partículas en suspensión o sustancias disueltas en agua, sino que además contribuyen a fijar el ambiente en el que las reacciones biogeoquímicas tienen lugar En este
documento se describen los mecanismos responsables del transporte en una serie de sistemas acuáticos naturales, como el
resultado de la interacción entre fenómenos exógenos de forzamiento y las características de los propios sistemas (la morfología en particular). Argumentos de escala, basados en los números de Lago y Wedderburn, el radio de deformación de Rossby y
el tiempo nominal de residencia, proporcionan un marco conceptual y de primer orden en el cual se puede definir y clasificar
el comportamiento físico de los sistemas acuáticos. Los casos analizados corresponden a sistemas, que en este espacio de cuatro dimensiones, ocupan posiciones distantes entre sí. En este trabajo se muestra que la circulación y el transporte en sistemas
acuáticos son fenómenos muy dinámicos, que en gran medida, están determinados por la magnitud, la frecuencia, la variabilidad espacial y su localización exacta en el tiempo de los flujos de masa y energía. Estos efectos, de ‘segundo orden’, son los
que se analizan en este trabajo.
Palabras clave: transporte, lagos estratificados, escala cubeta.
INTRODUCTION
Ramon Margalef ’s Mandala (e.g. Margalef
1997) describes the composition of phytoplankton communities, the relative abundance of
component species, and their evolution (succession) as the result of a nutrient-turbulence
balance (see also Reynolds 1997). If we accept
Margalef ’s Mandala as a valid interpretation of
succession in freshwater ecosystems, one necessarily concludes that the analysis and understanding of the functional structure of phytoplankton communities and its evolution needs to be
grounded on the knowledge of the physical processes of transport and mixing determining turbulence levels and nutrient distribution in the
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F. Rueda
water column. The physical processes of transport and mixing in lakes and reservoirs occur at
a wide range of spatial and temporal scales, that
go from the finest scales characteristic of turbulent motions – O(mm) and O(s) – to those characteristic of basin scale circulation – O(km)
and O(d) – (e.g. Imboden and Wüest, 1995).
Unfortunately and due to the small physical
dimensions of the containing basin (compared
to the ocean), transport processes in lakes and
reservoirs are complex and, therefore, difficult
to characterize and describe (Imberger 1998).
Only in the last few years, the development of
three-dimensional numerical models that solve
the equations of motion with a high temporal
and spatial resolution and low computational
cost, together with the availability of technology
capable of providing velocity and temperature
field observations at a wide range of spatial and
temporal scales, has allowed aquatic scientists
to initiate the exploration of transport processes
in natural lakes and reservoirs. My research has
been devoted to the analysis of basin-scale
transport patterns in stratified lakes and reservoirs and to the exploration of their consequences on the distribution of substances in these
environments, using three-dimensional hydrodynamic models, scaling analysis and field
observations. My approach to the knowledge of
freshwater aquatic ecosystems (lakes and reservoirs in particular) is, therefore and fundamentally, a physical one. However, it should be
understood within the framework of Ramon
Margalef´s vision of the structure and evolution
of phytoplankton communities (Mandala),
which suggests that physics has a role in determining the biology in aquatic ecosystems. My
latest lines of research, recently initiated at the
Water Institute in the University of Granada, in
collaboration with several research groups in
Aquatic Ecology from Southern Spain, attempt
to examine the interactions between the structure of phytoplankton communities (functional
and in terms of size) and the physical environment in reservoirs. This manuscript is a summary of some of the most relevant aspects of my
previous work, which focuses on transport patterns in stratified lakes and reservoirs.
TRANSPORT PROCESSES IN AQUATIC
SYSTEMS: A BRIEF OVERVIEW
Transport and mixing processes in aquatic
systems, with the exception of molecular diffusion, are driven by fluxes of mass and energy
(thermal and mechanical) that occur either
through the free surface or through inflow or
outflow sections (Fischer et al., 1979; Imboden
and Wuest 1995). These fluxes are subject to
spatial and temporal variations at a wide range
of scales. For example, the heat fluxes through
the free surface of a lake change at temporal
scales ranging from sub-daily (characteristic of
diel cycles of heating and cooling) to interannual scales (a feature of climate cycles).
Mechanical energy fluxes on the other hand,
either through the free surface (as a consequence of wind) or through the inflow or outflow
sections occurs in pulses or events: wind is not
continuously blowing over the surface of a lake,
but it occurs as events with varying frequency,
intensity and duration; inflow in reservoirs is
maximal during rainfall events; outflows in
reservoirs are subject to changes dictated by
water demands, which in the case of hydroelectric power, for example, has peaks followed by
periods of low demands. The specific transport
patterns, either at the basin-scale or at the finest
scale of turbulence, that develop in response to
mass or energy fluxes depend on a balance of
mass and energy fluxes and the specific spatial
and temporal patterns characterizing the fluxes
(the frequency, intensity, duration and location
in time of the events), together with characteristics of the aquatic systems, as their morphometry (Imboden and Wuest 1995). It is a balance of energy fluxes what determines, for
example, the stratification in aquatic systems
and its evolution in time: stratification develops
if and when the magnitude of fluxes inducing
density gradients in the water column (e.g. heating through the free surface) is larger than the
fluxes that introduce kinetic energy available to
mix it (i.e. wind or cooling at the free surface).
Atmospheric (wind) or hydrologic events
acting on a stratified water body tend to perturb
its equilibrium status by tilting the isotherms.
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Basin-scale transport in stratified systems
The behaviour of stratified systems of small and
medium-size in response to those perturbation
events can be parameterized using the adimensional Lake LN and Wedderburn W numbers.
These numbers relate the magnitude of the
external forces acting upon the system and displacing it from its equilibrium status to the forces that tend to re-establish the original equilibrium status as the system is displaced from it
(gravity). The Lake number LN is defined as the
ratio between the stabilizing and the destabilizing moments referred to the centre of volume
(Fig. 1). In the particular case of wind forcing
acting upon a stratified system LN can be estimated as (e.g. Imberger and Patterson 1990)
LN =
Figure 1. Response of a stratified lake to wind forcing: (a) A
system in equilibrium, with horizontal isotherms (thin lines).
The centre of volume is represented with a solid dark circle,
while the empty circle shows the location of centre of mass (displaced towards the bottom from the centre of volume, due to
stratification). (b) The destabilizing moments are associated
to wind force applied to the free surface, separated zm-zg from
the centre of volume. The stabilizing moments are associated to
gravity forces acting on the centre of mass. Diagram b shows a
mode 1 response to wind: the metalimnion moves towards the
surface at the upwind end of the lake. (c) Mode 2 response: isotherms open at the upwind end, and compress at the downwind
end. Respuesta de un sistema estratificado frente al viento. (a)
Sistema en equilibrio con isotermas en posición horizontal. La
posición del centro del volumen aparece marcada con un punto
negro, y la del centro de masas con un círculo con el interior en
blanco. El centro de masas, como consecuencia de la estratificación queda por debajo del centro de volumen. (b) El viento,
aplicado a la superficie del lago, ejerce una acción desestabilizadora sobre el sistema. La acción estabilizadora le corresponde a la fuerza de la gravedad que actúa sobre el centro de
masas. El diagrama b muestra un respuesta de modo 1 frente al
viento: el metalimnion se mueve hacia la superficie en la zona
de barlovento (c) Respuesta de modo 2 frente a fenómenos perturbadores: las isotermas se abren en la zona de barlovento, y
se comprimen en la zona de sotavento.
(zg – z0) Mg (zm – zT)
ρu2* A3/2 (zm – zg)
(1)
Here, zg is the elevation of the center of volume,
z0 is the elevation of the center of mass, zm is the
elevation of the free surface and zT is the elevation of the thermocline or the center of the metalimnion. The magnitude of the destabilizing
force is related to u*, the friction velocity of the
wind force is, in this case, applied tangentially to
the free surface and is given by the product of τ,
the shear stress τ (presumed uniform over the
lake surface), by the surface area A. The shear
stress, in turn, can be estimated as,
τ = CDρaUa2 = ρ u2*
(2)
where CD is the drag coefficient for wind
(≈ 1.5 x 10-3), ρa is the air density, Ua is the
wind speed, ρ is the water density. The
Wedderburn number W is estimated as
(Imberger and Patterson 1990),
W=
1 g′ h2
L
u2*
(3)
where g′ is the reduced gravity (= gΔρ/ρ), h is the
thickness of the surface mixed-layer and L is
the length of the basin. For LN >> 1, the restoring
force is larger than the forces perturbing the
system, and the tilting of the isotherms is small
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F. Rueda
compared to the thickness of the epilimnion
(Stevens and Imberger 1996). For LN << 1, the
external forcing will generate significant horizontal gradients in the thermal and chemical characteristics of the lake. In the case of wind, deep
hypolimnetic water will reach the surface in the
upwind end. On the other hand, low values of the
Wedderburn number (W ~ O(1) o W < 1) correspond to scenarios in which isotherms open in the
upwind end, while they compress in the downwind end of the lake. This type of response is
known as mode 2, while the behaviour defined by
the condition LN << 1 is known as vertical mode
1 (Fig. 1). For the particular case of systems with
a thin metalimnion, W and LN coincide.
Lake and Wedderburn numbers not only parameterize the large-scale behaviour of aquatic
systems. As shown by Fischer et al., 1979
(among others), the values taken by LN and W
are also indicative of the nature of the dominant
processes leading to mixing in the water
column. For W > 1, mixing is driven by the
direct flux of turbulent kinetic energy through
the free surface (caused by processes like wave
breaking, etc), which in the literature is known
as stirring (Fischer et al., 1979). For W ~ O(1) o
W < 1, on the other hand, mixing is energized by
turbulent fluctuations driven by shear (velocity
gradient) in the water column. Shear driven turbulence and mixing, acting upon horizontal gradients – induced by basin-scale motions for
W ~ O(1) or W < 1 – drive, in turn, longitudinal
dispersion processes that tend to mix the hypolimnetic o metalimnetic water upwelled in the
upwind end of the lake with the surface water
accumulated in the downwind end. Hence,
W ~ O (1) o W < 1 is indicative of vigorous
mixing processes (horizontal and vertical) leading to smeared physico-chemical gradients and
thick metalimnetic layers.
In relatively shallow and enclosed water
bodies, the condition W ~ O (1) o W < 1 occurs
on a frequent basis, and consequently, these
systems tend to exhibit a weak (almost linear)
and intermittent stratification. These are lakes
that mix vertically several times during a year,
i.e. polymictic lakes. In these systems, wind
–its time and spatial distribution– determines,
to a large extent, the transport patterns. The
work of Rueda et al. (2003a), Rueda and
Schladow (2003b) and Rueda et al. (2005a),
suggests that thermal stratification, though
weak and intermittent, may also play a major
role in determining water motions in polymictic systems during the stratified periods (see
section 3). In deeper lakes and reservoirs, stratification is more stable (W >> 1) and the
system undergoes through a unique period
during a year in which the water column mixes
(monomictic lakes) due to surface cooling.
Water motion, in these systems, is almost
totally dominated by internal waves, motions
of all scales and frequencies which cause
isopycnals to oscillate with some well-defined
frequency (Imberger 1998). Internal waves at
the basin scale are directly excited by the wind
forcing acting directly on the free surface.
Once established, basin-scale internal waves
steepen due to non-linear processes and to
shoaling, forming a spectrum of internal wave
motions with frequencies ranging from 10-6 to
10-2 Hz. The energy of the internal wave field
is finally dissipated within the benthic boundary layer (Imberger 1998) through wave breaking or shear driven processes. Section 4 presents some results of work done on the analysis
of basin-scale internal waves in Lake Tahoe, its
temporal and spatial structure. This work suggests that, also in deep lakes, the frequency,
duration and intensity of wind events are factors that need to be taken into account to
understand the internal wave climate, and
hence mixing processes. My interest on basinscale internal waves is justified in that they
provide the main driving force for vertical and
horizontal transport in deep stratified lakes
(Imberger 1998, MacIntyre et al., 1998).
Hence, the first step towards analyzing numerically the dynamics of nutrients and particles in
deep lakes consists on being able to understand
and represent with three-dimensional transport
models the energy and spatial characteristics of
basin scale internal waves.
Upwelling events, either episodic or more
frequent, its intensity and duration control the
exchange of water between basins connected
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37
CIRCULATION PATTERNS
IN POLYMICTIC LAKES
Figure 2. Bathymetry map (in gray) of Clear Lake with contours shown every 2 m. Inset on left locates Clear Lake in
California, USA. Also shown are the three basins of the lake,
connected at the Narrows. Batimetría del Lago Clear, con contornos cada 2 m. En el recuadro de la izquierda se muestra la
localización del lago en California, EEUU. Así mismo, se
muestran los nombres y disposición de las tres subcuencas del
lago, conectadas en el estrecho de ‘Narrows’.
through narrow straights. Section 4 shows the
results of some work done on exchage processes
in semi-enclosed environments, applied to a
particular embayment of Lake Ontario. Reservoirs, are considered open systems, where mass
and energy is exchanged through distinct inflow
and outflow sections. Inflow mixing controls the
pathways of river nutrients in the reservoir and,
in particular, the extent to which these nutrients
can reach the euphotic zone to be used as
resources for primary production. Moreover, the
transport processes that occur at the river inflow
sections, interacting with selective withdrawal
(outflow) and vertical mixing processes driven
by wind or surface cooling, control the average
time a given water parcel remains within the
reservoir. This time scale, known was the
hydraulic residence time, has been proposed in
the literature to explain a wide range of water
quality processes. Section 5, presents results of
work undertaken to estimate the residence times
in reservoirs, as systems with large fluctuations
in water level and inflow-outflow rates.
Mixing and transport in large polymictic lakes
has received little attention in the limnological
and engineering literature. Instead most physical limnology has focused on the dynamics of
deep, seasonally stratified lakes which are characterized by the presence of a vertical thermal
stratification that evolves on a time scale of
months, and free basin-scale internal waves
with time scales on the order of hours to days
(e.g. Antenucci et al., 2000; Hodges et al.,
2000; Pan et al., 2002; among many others).
One of the few attempts to describe transport
patterns in a shallow polymictic lake is the
series of three papers - Rueda et al. (2003a),
Rueda and Schladow (2003) and Rueda et al.
(2005a) - recently published in the peer reviewed literature. The results described in those
publications are the result of a study conducted
in Clear Lake (Fig. 2), during several periods
in 1995, 1997 and 1999. Clear Lake is a
large polymictic lake in northern California
with three sub-basins (Oaks, Upper and Lower
arms) separated by a narrow straight (Narrows).
With an average annual inflow Q estimated to
be 514x106 m3/year (Lynch 1996) the ratio V/Q
(where V is the average lake volume) is approximately two to three years. Hence, advective flow
(through-flow) through the lake is not considered to be a major factor in determining largescale circulation. As with many other large and
polymictic systems, Clear Lake does not support free basin-scale internal waves since the
wave periods are, in general, longer than the
time scales associated with the stratification–destratification cycles. Maximum recorded
temperature differences between the top to the
bottom of the lake (max. depth is c.a. 15m) were
of up to 6oC in 1999 (Rueda et al., 2003). On a
typical summer day the winds are primarily
from the northwest, minimal between midnight
and dawn and reaching a maximum about dusk.
Also, and due to the large extent of Clear Lake
and the irregular nature of its surrounding topography, the wind exhibits a considerable degree
of spatial variability. These two characteristics
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F. Rueda
of the wind field (its marked diurnal periodicity
and its spatial variability) are key features, together with the weak stratification, to explain circulation patterns in the lake.
Wind-stratification interaction in weakly
stratified systems
This research focused on one of the sub-basins
of Clear Lake: the Oaks arm. The evening winds
(blowing along the thalweg of the sub-basin)
encounter a water column, which due to the surface heat flux during the morning and early
afternoon hours, is stratified. The wind acts on
this weakly stratified system to generate strong
horizontal temperature gradients along the axis
of the Oaks Arm (W << 1), which can be of up
to 3oC in hardly 4 km. Deep water rises at the
western end of the arm, near the Narrows. Due
to the weak stratification, rotational effects
modulate the response of the lake to wind forcing, causing upwelling of deep water also on
the north shore of the Oaks Arm. The effect of
the Earth’s rotation on a stratified system being
forced by wind can be analyzed using the theoretical model proposed by Csanady (1982). In
that model, it is assumed that a long-shore wind
is applied impulsively on an initially quiescent
and semi-infinite (i.e. only bounded on one
side) two-layer system of depth H. The interface
will tilt perpendicular to the wind direction and
it will intersect the water surface – complete
upwelling – if the wind impulse I (= u2* t, t being
time) satisfies the condition
I (H – h) / hHc1 ⭓ 1
(4)
where h is the thickness of the upper layer and
c1 is the speed of propagation of perturbations
in the interface (i.e. internal wave speed). For
the Oaks arm, presuming a two-layer stratification with the interface at mid-depth and a density change equal to half the density difference
from top to bottom, the ratio in Eq. 4 was on
average, for the wind events occurring on a 15day period in may 1999, equal to 1.64. Hence,
deep cold water piles up in the northwestern
shores in response to northwesterly winds (see
Figure 3. Transect taken on day 143, 1999 at 16h across the
Oaks Arm. Boat was moving from south to north. Velocity is
given in cm/s in gray scale of each plot. The component of
velocity along the main axis of Oaks Arm is shown on top, and
that in perpendicular direction on bottom of the figure, and
gray scale to left of value of velocity component (positive is
eastward). Transecto de velocidad realizado a las 16h del día
143, 1999 de sur a norte en Oaks Arm. La velocidad se muestra en cm/s utilizando una escala de grises (a la derecha).
Arriba se muestra el componente de velocidad u paralelo al
eje principal de Oaks arm. Abajo se muestra el componente de
velocidad v perpendicular a dicho eje. Valores positivos son
hacia el este y hacia el norte para u y v respectivamente.
circulation patterns in figure 3 captured with an
acoustic Doppler Profiler ADCP). During the
night and early morning hours, when the wind
forcing is negligible, the pressure gradients that
result from the horizontal differences in temperature become the dominant forcing mechanism
in the system, driving currents of up to
10–15 cms-1, westward at the surface and eastward near the bottom (Fig. 4), that try to bring
isotherms to their equilibrium horizontal position. The currents during the relaxation of the
horizontal temperature gradients are affected by
the Earth’s rotation, and westward currents
mainly occur through the north shore.
Residual circulation patterns
The cycle of setup and relaxation of horizontal
temperature gradients driven by wind and modulated by the Earth’s rotation, shown in the previous section, repeats in the stratified lake with a
diurnal periodicity. As a result, a residual cyclonic circulation develops, which facilitates horizontal exchange processes in the lake. A particle
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Figure 4. Transects taken at 14h on day 144, 1999. The gray
scale represents the magnitude (cm/s) of velocity component
along main axis of Oaks Arm (positive is eastward). Transecto
de velocidad realizado a las 14h del día 144, 1999 de sur a
norte en Oaks Arm. La escala de grises muestra el componente de velocidad u paralelo al eje de la subcuenca (cm/s).
Valores positivos son hacia el este.
left in the Oaks Arm would describe counterclockwise trajectories. This circulation pattern
was revealed experimentally, using lagrangian
followers, and numerically, using simulations of
tracer release experiments and particle tracking
techniques. Details of the modelling tools employed (here and throughout this work) can be found
in Rueda and Schladow 2003 or Smith 1997.
Figure 5, for example, shows the sequence of
positions of a particle released near the easternmost end of the Oaks Arm. To describe the particles’ 3-D trajectories, plan-view (2-D) plots have
been used, in which the depth below the surface
of the particle is indicated by the grayscale of the
marker used to show its horizontal position. Each
dot represents the position of a particle, in which
is shown every hour. An increased spacing between positions indicates higher velocities. Figure
5 shows the trajectory of a particle that, after
plunging and rising twice at the downwind end of
the Oaks Arm, is carried in day 142 (1999) by
surface currents to the westernmost end of the
Oaks Arm, and from there it enters the Lower
Arm. The particle plunges following the winddriven downwelling currents close to the mine,
and it rises to the surface at the embayment located at the southwest edge of the Oaks Arm.
Direct wind-driven circulation patterns
In the Lower Arm the particle shown in figure 5
describes several clockwise loops. As shown in
Rueda et al. (2005a), the anticyclonic circulation
revealed by the particle trajectories (and also by
39
observations gathered in 1995 with ADCP) is driven by the spatial variability of the wind field. In
Clear Lake, sheltering by surrounding terrain is
the major source of variability in the wind field.
The topography of the landscape surrounding
Clear Lake is complex and dominated by the presence of Mount Konocti, a dormant volcano
rising over 900 m above the lake level. It is located upwind of the Lower arm in the direction of
the predominant winds. The spatial variability of
the wind field over Clear Lake was analyzed
during two weeks starting on May 18, 1999, using
an array of 14 anemometers. Vorticity is defined
as twice the angular velocity of a fluid parcel and
is used as a measure of the spatial variability of a
flow field and its rotational character. As shown
in figure 6, Mount Konocti modifies the wind
field over the lake and creates areas of negative
vorticity (anticyclonic or clockwise wind circulation) over the Lower Arm. The numerical simulations of Rueda et al. (2005a) show that the anticyclonic circulation patterns observed in the Lower
Arm are not but the fingerprints of the anticyclonic wind circulation existing above it. Furthermore, it was shown that the spatial variability of
winds increases the rate at which the Oaks Arm
exchanges water mass with the rest of the lake,
mainly with the Upper Arm, while it isolates the
Lower Arm from the other two arms by increasing
the recirculation rate (gyre formation) within this
basin. The exchange rates between the Oaks and
the Upper arms, for example, were almost twice
the exchange rates between the Oaks and the
Lower Arms under the spatially variable winds.
Data collected in the last few years (Suchanek,
pers.comm.) has shown that the annual loading
to the sediment of particulate mercury is highest
in the Oaks Arm, followed by the Upper Arm
and then the Lower Arm (in the ratio of approximately 9:5:2). The ratio of loadings in the Upper
and the Lower arms is approximately the same as
the ratio of mass exchange rates of those two
arms with the Oaks Arm, revealed by tracer simulations conducted in Rueda et al., 2005a. The
coincidence of loading and exchange rate ratios
among the different basins suggests a close link
between chemical and hydrodynamic behaviour,
which is driven by spatially variable winds.
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Figure 5. Trajectories followed by one of the groups of particles
released near the eastern of the Oaks arm (Clear Lake) in the particle tracking simulations. Symbols are plotted one hour apart.
The gray scale on the left shows the depth of the particle in the
three-dimensional domain. Trayectoria seguida por una de las
partículas liberadas en ejercicios numéricos de seguimiento de
partículas en el extreme oriental de Oaks Arm (Lago Clear). Los
símbolos representan la posición de la partícula en pasos de tiempo de una hora. La escala de grises representa la profundidad.
BASIN SCALE INTERNAL WAVES
IN DEEP STRATIFIED LAKES
Lake Tahoe (Fig. 7) is located along the crest of
the Sierra Nevada mountain range, between
California and Nevada at an altitude of 1898 m.
It has a maximum depth of 505 m, an average
depth of 313 m, and its surface area is 501 km2.
The lake contains 156 km3 of water, and the
ratio of the lake volume to the average outflow
rate V/Q is approximately 650 years. Typically,
thermal stratification commences around March
and reaches a maximum in August, at which
time the top of the thermocline is approximately
20 m below the water surface. Slow weakening
of the stratification may lead to approximately
isothermal conditions by February in about one
year in four (Jassby et al., 1999). As revealed by
field data collected in Lake Tahoe since 1995
(see Thompson 2000), the temperature structure
of the lake is subject to oscillations (internal
waves) of varying amplitude, occurring at a
range of frequencies. The work of Rueda et al.,
(2003b) analyzes temperature observations
collected during a 25-day period at the end of
1999, to determine the nature of those internal
waves, its frequency and spatial characteristics,
and the factors controlling the energy contained
in the internal wave field. Thermistors were
deployed at different depths and at two stations
in Lake Tahoe (MIDLAKE and INDEX stations
in figure 7). The average temperature profile
during the 25-day period studied in Rueda et al.
(2003b), is characterized by a relatively smooth
gradation in temperatures from the epilimnion,
with an almost uniform temperature of 7.5 oC in
the first 50 meters of the water column, to the
hypolimnion starting at about 150 m below
the water surface and having a uniform temperature of less than 5 oC. The Lake (LN) and
Wedderburn (W) numbers are well above unity
throughout this period, suggesting relatively
small isotherm displacements compared to the
thickness of the surface layer (no upwelling).
Rather than analyzing thermistor records one by
one, the temperature time series collected at different elevations z was used to generate, for
each station, a unique time series of integrated
potential energy IPE, defined as
z1
IPE (t) = 冮 ρ (z, t) g z dz
(5)
z0
It was the time series of IPE (t) that was subject to spectral analysis to identify the frequencies of the most energetic oscillations. As shown
by Antenucci et al., 2000, the IPE not only gives
a clear and concise picture of the frequency content of the internal wave field, but it also represents correctly the relative distribution of potential energy among frequencies.
Basin scale waves: a brief overview
Basin scale internal waves are classified according to the horizontal and vertical structure of the
isotherm displacements they induce. An internal
wave of vertical mode 1 and horizontal mode 1
(V1H1) has only one node in the vertical and in
the horizontal, as shown in figure 8a. Vertical
mode 2 indicates the presence of more than one
node in the vertical, i.e. two interfaces that oscillate with opposite phases compressing and
expanding the metalimnion alternatively (see
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Figure 6. (a) Windfield vorticity constructed from wind
records at 1700 h on day 139, 1999. The triangle represents the
location of Mount Konocti. The effect of an isolated 3-D obstacle in a boundary layer flow is illustrated in panel b. The
flow field past 3-D obstacles is characterized by the presence
of a lee-side separation region and an associated wake. The
wake contains two distinct counter-rotating horseshoe vortices
(see panel a). The near-surface streamlines tend to diverge on
the upstream side and converge on the downstream side of the
obstacle (Baines, 1995). (a) Vorticidad del campo de viento
construida a partir de registros de vientos en 14 estaciones
meteorológicas situadas alrededor del Lago Clear a las 17h
del día 139, 1999. El triángulo blanco representa la localización del Monte Konocti (ver texto). El efecto de un obstáculo
aislado en la capa límite atmosférica se ilustra en el panel (b).
El campo de velocidades del fluido al pasar sobre el obstáculo
está caracterizado por la presencia, en la zona de sotavento,
de una región en que el fluido recircula. En la zona de recirculación existen dos vórtices con signos de rotación contrarios.
Las líneas de flujo tienden a divergir en la zona de barlovento
y a converger en la zona de sotavento (Baines, 1995).
figure 8b, showing a V2H1 mode internal
wave). Internal waves with horizontal modes
two (and above) refer to the presence of two or
more nodes in the horizontal direction (Fig. 8c).
The spatial characteristics and oscillation
periods of individual internal wave modes are
controlled by the density stratification, the geometric properties of the enclosing basin, and, in
the case of large lakes, by the Earth’s rotation
(e.g. Imboden 1990). The Earth’s rotation can
exert a significant influence on the internal
41
Figure 7. Lake Tahoe bathymetry (contoured on 100 m intervals)
and instrument locations (•). Thermistor chain locations are
MIDLAKE and INDEX. USCG indicates the location of the
meteorological station. Batrimetría del Lago Tahoe, con contornos cada 100 m. Se muestra también la localización de las ristras
de termistores (estaciones MIDLAKE e INDEX) y de la estación
meteorológica del Servicio Geológico de EEUU (USGS).
motions of lakes on temporal scales of the order
of the inertial period (~ 19 hours in Clear Lake)
and on spatial scales of the same order as the
Rossby radius of deformation (Λ = c1/f where f
is the Coriolis parameter or inertial frequency).
In Lake Tahoe during the period analyzed in
Rueda et al., (2003b) Λ is c.a. 3 km and smaller
than the with of the basin (19 km), hence rotational effects are significant. Internal oscillations, under those conditions, are either Kelvin
or Poincaré waves. Kelvin waves are long gravity waves with subinertial frequencies (periods
larger than c.a. 17h the inertial period) that are
trapped at the boundaries of the lake. The perturbations travel cyclonically (counterclockwise) around a basin in the Northern Hemisphere
with amplitudes exhibiting an exponential decay
offshore with a scale of O (Λ). The velocity
fields show maxima at the boundaries (where
the motion is rectilinear), and the current vectors in the lake interior rotate cyclonically (see
for example Hutter 1984 or Antenucci and
Imberger 2001). Poincaré modes are also progressive waves. In the Poincaré modes (or rotating internal seiches) the perturbations propagate
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F. Rueda
anti-cyclonically with superinertial frequencies
(periods lower than 17h). As shown by
Antenucci et al. (2000), the isotherm displacements induced by a Poincaré wave measured in
cross section have a standing structure and are,
in general, undistinguishable from those induced by a linear seiche. Poincaré waves, however,
differ from the linear seiches in that the velocity
vectors rotate anti-cyclonically. The maximum
amplitudes of the water velocity induced by the
wave are observed at the center of the basin.
Thus the origin of isopycnal oscillations (either
a linear seiche or a Poincaré wave) can only be
positively confirmed by inspection of the velocity vectors at a single location or inspection of
temperature records from multiple locations.
The internal wave ‘zoo’ in Lake Tahoe
Figure 9 shows the power spectrum of IPE signal, calculated from the temperature records
collected at the INDEX station, exhibiting four
peaks: three peaks in the subinertial range of the
frequency, and one peak in the superinertial
range. The three subinertial modes were identified as vertical mode one Kelvin waves travelling cyclonically around the perimeter of the
lake with periodicities of approximately 128, 57
and 37 h. The longer period Kelvin wave has
horizontal mode 1, while the others have horizontal modes 2 (57h) and 3 (37h). The fourth
internal wave mode was identified as a V1H1
one Poincaré wave having a period of about 17
h, causing transverse oscillations of the isotherms. These results suggest that the internal
oscillations in large lakes may be the result of
the superposition of what Imboden and Wüest
(1995) call a ‘zoo’ of basin-scale internal
modes, rather than individual modes.
Figure 10 shows the spatial structure of IPE,
as simulated by the 3D model, and its evolution
in sequences of plots at quarter-period (T/4)
increments, for the three Kelvin wave modes
identified in figure 9. Kelvin waves are seen travelling cyclonically around Lake Tahoe, with
the shore on the right, and with one (Fig. 10 a,
b, c, d), two (Fig. 10 e, f, g, h) and three (Fig. 10
i, j, k, l) peaks (zones with positive IPE values)
Figure 8. Idealized structure of internal waves. Here, it is
assumed that isotherms (shown in thin line) only oscillate in
the plane shown, i.e. it is assumed that internal waves are seiches (not affected by earth’s rotation). (a) V1H1 mode,
(b) V2H1 mode and (c) V1H2 mode. Estructura idealizada de
ondas internas. Aquí se a asumido que las isotermas (mostradas en línea fina ) solo oscilan en un plano vertical, i. e. se
asume que las ondas internas son secas (no afectadas por la
rotación de la tierra). (a) modo V1H1, (b) modo V2H1 y
(c) modo V1H2.
and troughs (negative IPE values). The spatial
structure of the Poincaré mode is shown in figure 11. The phase propagates anti-cyclonically
around the basin. The behaviour of the IPE
oscillations in the surrounding embayments,
though, does not follow the basin-scale pattern.
For example, in Fig 11 a, b, the wave is causing
upwelling in the western shore in the main basin
(high positive IPE) while close to the shore in
the embayment the IPE signal shows negative
values (downwelling). The IPE in the embayment oscillates with a node along the East-West
direction, contrary to the general behaviour in
the main deep basin. The horizontal structure of
the oscillations is, thus, far from being simple
due to the irregular shape of the basin.
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43
- Interval 3 (from day 2, 2000 to the end of the
record) when the wind again excites the internal waves at the periods previously observed
during Interval 1.
Figure 9. Power spectra of integrated potential energy, IPE, at
the INDEX station (Lake Tahoe). Dashed lines indicate 95 %
confidence levels. Marked in the figure are the periods corresponding to the peaks in the spectrum. Espectro de frecuencias
de las series de tiempo de energía potencial integrada IPE
estimada para la estación INDEX (Lago Tahoe). Las líneas
discontinuas representan el nivel de confianza del 95 %.
Sobre la figura se indican los períodos correspondientes a los
picos de frecuencia.
Interaction of wind and internal waves
Figure 12 shows wind and temperature records
collected in Lake Tahoe, during the 25-day
period studied in Rueda et al. (2003b). Observations at the INDEX station, suggest that three
intervals or sub-periods of internal wave activity can be distinguished:
- Interval 1 (days 350 through 357 in 1999)
with significant internal wave activity at both
subinertial and superinertial frequencies, and
amplitudes of up to 30 m;
- Interval 2 (from day 358, 1999 through day 1,
2000) where the subinertial oscillations
observed in Interval 1 with periods above
24 hours vanish; the diurnal and superinertial
periodicity dominate during this interval;
The evolution of the internal wave energy can
be explained by analyzing the relative phase between the wind and the internal oscillations. A
well known result in classical mechanics is that
the amplitudes of the motions described by a forced harmonic oscillator depend on the energy of
the forcing mechanism, the frequency of the forcing compared with the natural frequency of the
oscillator, and their relative phase (see for example Wilson 1972). The Kelvin wave signal in the
observed IPE time series at the INDEX station
was extracted from the raw IPE signal using
wavelet transforms (e.g. Torrence and Compo
1997) and plotted together with the East-West
wind component at the USCG weather station in
figure 13. An increase in the IPE time series
marks an upwelling at the western shore of the
lake, while downwelling periods correspond to
decreasing trends in the series. Easterly winds
acting at the time of upwelling will act to energize the internal wave (energizing event), as they
act in phase. Easterly winds acting at the time of
downwelling will drain energy from the internal
wave (weakening event). In figure 13 the energizing events are marked with dark gray bars,
while the weakening events are marked with
light gray bars. The decay in the Kelvin wave
amplitude follows a series of weakening events,
while a series of energizing events precedes the
excitation of Kelvin waves in the subperiod 3.
These results suggest that the internal wave climate in lakes is, in general, the result of a complex and subtle interaction of atmospheric and
the own internal wave features (their frequency,
intensity and duration, as well as their spatial
characteristics), which, in turn, is controlled by
stratification, morphology and the Earth’s rotation. The analysis of the internal waves in large
lakes, hence, requires the use of dynamic timevarying (e.g. 3D hydrodynamic models or wavelets) rather than static (classical power spectrum
density, etc) or spatially-simplified (2-D or 1-D
hydrodynamic models) tools.
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Figure 11. First mode Poincaré wave in Lake Tahoe reproduced
in the modeled IPE and presented at 2 hour intervals. Wave
period is approximately 17 h. The arrow on the circle indicates
the direction of time advance. The sense of rotation of the internal wave is anti-cyclonic. Time series of IPE at each horizontal
location are processed as indicated in Figure 10. A black box is
included to display the clockwise phase propagation of the
Poincaré wave. Primer modo de la onda de Poincaré en Lake
Tahoe, según el modelo de simulación. Se presenta la estructura
horizontal del las desviaciones de IPE en relación a su valor
promedio en cada punto. La estructura se muestra cada dos
horas. El período es de aprox. 17h. La flecha en el círculo interior indica la dirección de avance del tiempo. El sentido de
rotación de estas ondas es anticiclónico. Un cuadrado negro se
ha utilizado para poner de manifiesto la propagación a favor de
las aguas del reloj de la onda de Poincaré.
SEMI-ENCLOSED SYSTEMS
IN LITTORAL REGIONS
Figure 10. Three modes of Kelvin waves in Lake Tahoe reproduced in the modelled IPE and presented at quarter period
intervals. Sense of rotation is cyclonic in all cases. (a-d) first
horizontal mode; (e-h) second horizontal mode; (i-l) third
horizontal mode. Time series of IPE at each horizontal location of the model are first de-trended and then band-pass filtered. Ondas de Kelvin: estructura espacial de las excursiones
de IPE sobre sus valores promedio en cada punto, correspondiente a los tres modos (modos horizontal 1, 2 y 3) identificados. La estructura espacial se muestra a intervalos de T/4,
siendo T el período de la onda. (a-d) modo horinzontal 1 H1;
(e-h) modo horizontal 2 H2; (i-l) modo horizontal 3. Los picos
o valles se mueven en contra de las agujas del reloj, siguiendo
la línea de costa.
Under the action of upwelling favourable events
nutrient-rich deepe water reaches the surface in
near-shore regions, hence, inducing changes
in the abundance and structure of the phytoplankton communities in littoral systems, such
as semi-enclosed embayments. Data collected in
Little Sodus Bay (LSB, 43o20’ N – 76o42’30” W,
Fig. 14), a small embayment on the south shore
of LO, is very suggestive in this regard (Fig. 15).
LSB is approximately 4 km long and 1 km wide,
and it is permanently connected to LO through a
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Figure 12. Time series of (a) isotherm depths at the INDEX station, and (b) wind speed and direction in Lake Tahoe. Isotherms in frames
(a) are shown every 0.5 oC. The 5.5 oC has been marked for reference. The period under analysis started on day 350, 1999 (December 16),
and ended on day 9, 2000 (January 9). Series de tiempo de (a) profundidad de las isotermas en la estación INDEX, y (b) velocidad y
dirección del viento en el lago Tahoe. Las isotermas se muestran cada 0.5 oC, y la isoterma de 5.5 oC ha sido marcada como referencia.
El período analizado en Rueda et al. (2003b) comienza en el día 350, 1999 (16 de diciembre), y finaliza el día 9, 2000 (9 de enero).
man-made channel with design length, width,
and depth of 550, 75, and 3 m, respectively.
During summer, upwelling events in LO occur
episodically along its southern shore in response
to strong and sustained easterly winds1 (see
Rueda and Cowen, 2005), and consequently LSB
is exposed to nutrient-rich LO hypolimetic
water. Secchi disk depth measurements were
taken in LSB (Hairston and Doyle, personal
communication) and compared with the time
series of easterly wind impulse I (see section 3
for definition of I) in LO during several months
in 2001. Upwelling events occur if the alongshore easterly wind impulse is larger than a threshold value (marked in figure 1 with crosses),
determined by stratification conditions in the
water column and the magnitude of offshore
(southerly) winds (Csanady 1977). Thermal stratification in LO is presumed constant, characterized by a 14 m deep thermocline and a reduced
gravity (g′ = g Δρ/ρ) of c.a. 1.5x10-3 m s-2, typical of summer conditions (Schwab 1977). The
Secchi depth time series shows cyclic changes,
which are typical of lake environments (see for
example figure 7 in Thomann et al. 1981). Each
1 The Rossby radius of deformation in Lake Ontario during
summer is lower than the average width of the lake (> 100 km),
hence, upwelling occurs in response to along-shore winds
(see section 3).
of the three upwelling events that occurred in the
three summer months of 2001 precede periods
of decreases in water transparency, as indicated
by negative trends in the Secchi depths. In Rueda
and Cowen (2005) it was hypothesized that
these decreases in transparency are the result of
increased biological activity.
Figure 13. Measured IPE signal at the INDEX station for the
first horizontal mode Kelvin wave (as calculated through wavelets) and East-West component of the wind vector. The time
series have been normalized using the maximum value during
the time window analyzed. Vertical dark gray bars mark strong
wind events when the phase of wind and the Kelvin wave results
in the wave being enhanced (also indicated by a vertical arrow
pointing upwards). Light gray bars mark events when the interaction is destructive (downward looking arrow). Series de tiempo de (a) profundidad de las isotermas en la estación INDEX, y
(b) velocidad y dirección del viento en el lago Tahoe. Las isotermas se muestran cada 0.5 oC, y la isoterma de 5.5 oC ha sido
marcada como referencia. El período analizado en Rueda et al.
(2003b) comienza en el día 350, 1999 (16 de diciembre), y finaliza el día 9, 2000 (9 de enero).
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Whether fluctuations in biological activity shown
in the LSB data reflect an endogenous response
of the system or are driven by exogenous (upwelling) influences, it remains to be proved, but what
Rueda and Cowen (2005a and 2005b) showed
was that the hydrodynamic behaviour of LSB
undergo significant changes during upwelling
events in LO. The work of Rueda and Cowen
(2005a and 2005b) used the hydraulic residence
time as a surrogate and single-value description
of the transport characteristics of LSB. The residence time was estimated from the analysis of
simulated tracer release experiments, consisting
on releasing a certain amount of mass m0 of a
conservative tracer at time t0 and at location α in
LSB. The time varying tracer mass m(t) remaining in LSB was then monitored. The quantity
m(t) is found from the spatial integration of the
measured concentration field within the water
body, and its decline over time reflects the net
rate at which tracer leaves the water body. The
rate of mass loss as a function of time r(t) provides the residence time distribution RTD ϕ(t),
1 dm
r(t) = –
(6)
m0 dt
which is the probability density function of the
residence time of the ensemble of individual
particles released at the beginning of the experiments. Equation 6 shows that the RTD has units
of [T]-1. The mean residence time, based on the
first moment of r(t), can then be calculated as
∞
τr = 冮r(t) t dt = –
0
1
m0
∞
冮
0
dm
dt
=
1
m0
∞
冮m(t) dt
(7)
0
If instead of releasing tracer at location α, one
presumes a release in which the whole water body
is initially and uniformly filled with a mass m0 of
a conservative tracer, Eq. 6 yields the flushing
time distribution (see Rueda and Cowen 2005b).
The mean flushing time is estimated as in Eq. 7,
and it provides a bulk or integrative description
(with no spatial dependence) of the transport characteristics of a water body.
As Eq. 6 suggests, the residence or flushing
time in LSB, and in any other aquatic system, is
dictated by he rate at which water parcels leave
Figure 14. Location and bathymetric map of Little Sodus Bay.
Contours are shown every meter, and the 10-m and 7-m isobaths are shown in Little Sodus Bay and Lake Ontario, respectively. The bay is connected to the lake through a channel that
is 550 m long, 75 m wide, and 3 m deep. Localización y batimetría de Little Sodus Bay. Los contornos se muestran a cada
metro, y los correspondientes a 10 y 7 m se muestran en Little
Sodus Bay y Lago Ontario, respectivamente. LSB está conectada al Lago Ontario por un canal de 550 m de longitud, 75 m
de anchura y 3 m de profundidad media.
the system. The processes involved in the transport of water parcels from any point in the embayment until they reach the lake can be classified,
for the sake of clarity, into “exchange” processes,
acting in the channel, and “internal mixing” processes, acting within the embayment. The difference between internal mixing and exchange time
scales can be better understood using two idealized flow reactor models: the continuous stirred
tank reactor (CSTR) and the plug-flow reactor
(PFR − e.g. Levenspiel 1999). In the CSTR internal mixing is considered infinitely fast and the
mean residence time is set by the rate at which
exchange takes place. At the other extreme, in the
PFR the mean residence time is controlled by
the rate particles move within the reactor (internal mixing in our semantics), and the exchange
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Physical processes driving exchange
in the channel
Figure 15. (a) Secchi depth observations (Doyle and Hairston,
personal communication) for LSB in 2001. (b) Series of alongshore impulses of upwelling-favorable wind events (bars), and
threshold values for the alongshore wind impulse to induce a
full upwelling event (crosses). Calculations based on wind
records from buoys 45139 and 45135 (operated by Environment
Canada) in the northern and western ends of LO, respectively.
The arrows in (a) and (b) mark periods when the wind records
suggest that upwelling events could have occurred. (a) Observaciones de profundidad del disco de Secchi (Doyle and
Hairston, comunicación personal) tomadas en LSB durante
2001. (b) Impulso del viento para eventos procedentes del Este
(que provocan el ascenso de las isotermas a lo largo de la costa
sur del Lago Ontario) – barras verticales –, junto a valores
umbrales para que los eventos puedan llegar a producir afloramiento completo (cruces). Los cálculos se han hecho utilizando
los registros de las boyas 45139 y 45135 (operadas por
Environment Canada) en el norte y oeste de LO, respectivamente. Las flechas en (a) y (b) marcan períodos en que los registros
de viento sugieren que fenómenos de afloramiento se producen
a lo largo de la costa sur de LO, frente a LSB.
processes do not set the RTD. Even though
exchange and mixing processes are considered
separately, they need to be jointly considered in
the examination of basin water residence times
since they are inextricably linked in the determination of mass fluxes across the channel (e.g.
Ivey 2004). The density gradients in the channel,
for example, are determined by the density in the
respective basins, which is, in turn, influenced by
the dynamics within the basins themselves.
The estimated mean daily discharge Q from the
contributing watershed into the embayment is
0.15 m3s-1. However, instantaneous discharge
measurements taken in the LSB-LO channel
during several periods of 2001 and 2002 and
under baseline (non-upwelling) conditions were
of up to 400 times larger, suggesting that
through-flow was negligible in comparison with
other exchange mechanisms at the channel scale.
Moreover, discharge measurements were of
oscillatory nature, indicating that the channel
acts both as an inflow and outflow opening. The
most energetic oscillations occurred with 2-3 h
period, which corresponds to the fourth surface
seiche mode in Lake Ontario, inducing water
level oscillations of a few cm range. These
results suggest that barotropic forcing (i.e. water
level oscillations) is the dominant mechanism of
motion under baseline conditions, moving water
either towards LO or LSB with little shear (i.e.
the water at all depths moved in the same direction, see figure 16a). During LO upwelling
events, temperature differences between LO and
LSB (in hardly 500 m) observed during 2002
were of up 16oC. In figure 16b, the temperature
differences along the channel (500 m), during an
upwelling event on day 148 in 2002, are c.a. 6oC
(Fig. 16b). Temperature records collected at different depths in the channel and velocity measurements at the bottom of the channel, suggest
that water from LSB exits the bay through the
surface and LO water enters the bay through the
bottom of the channel (Figs. 16c and d). The
exchange pattern (outflow in the surface, inflow
in the bottom) was persistent, indicating that the
horizontal density gradients, induced in response
to upwelling in LO, became the dominant forcing of exchange between LO and LSB. Furthermore, barolinic forcing is also dictated the by the
internal dynamics of the embayment. As shown
in Rueda and Cowen (2005a), LN and W can frequently reach values lower than 1 suggesting that
isotherm excursions are large compared with the
epilimion thickness and persist in time long
enough, according to the internal hydraulic theo-
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ry (e.g. Armi and Farmer 1986), to modify the
exchange rates through the LSB-LO channel. In
any case, either during base-line or upwelling
conditions, Rueda and Cowen (2005a) show by
scaling analysis that barotropic (water level) and
baroclinic (density driven) forcing, wind stress
applied at the free surface, frictional, and unsteady effects, all need to be considered in the quantitative analysis of exchange processes through
straights in lake environments.
Net exchange rate of mass through
lake-embayment connection
Field data and three-dimensional dynamic simulations for a period of 50 days in 2002, presented
in Rueda and Cowen (2005b), showed that the
net exchange rates between LO and LSB
(the rate at which LSB water is renewed) under
base-line (non-upwelling) conditions is at least
10 to 20 times smaller than during upwelling
Figure 16. Observation of channel processes in 2002. (a) Time series of longitudinal channel velocity profiles during 11/2 h during
base-line (non-upwelling) conditions. Isotachs are shown every 2 cm s-1 (–20 cm s-1 and 14 cm s-1 isotachs are given in the plot),
time is represented along the x-axis and elevation from the bottom on the y-axis. (b), (c) and (d) correspond to upwelling conditions. Time series of (b) temperature at different depths at LO and LSB (near the channel); (c) temperature at 3 depths at the channel; and (d) velocity at the bottom of the permanent connection between Little Sodus Bay and Lake Ontario (negative is flood, i.e.,
toward LSB). Observaciones en el canal de conexión LSB-LO, tomadas en 2002. (a) Series de tiempo de perfiles del componente
de la velocidad del agua paralelo a las paredes del canal (componente Norte-Sur). Los registros corresponden a un período de 11/2
h durante condiciones de base (no afloramiento). Líneas de isovelocidad se muestran cada 2 cm s-1 (los contornos correspondientes a –20 cm s-1 y 14 cm s-1 están marcados). El tiempo se representa en el eje x y la altura sobre el fondo en el eje y. (b), (c) y (d)
corresponden a observaciones tomadas durante condiciones de afloramiento en el Lago Ontario. Son series de tiempo de (b) temperatura a diferentes profundidades en LO y LSB (en las inmediaciones del canal); (c) temperatura a tres profundidades en el
canal; (d) velocidad medida en el fondo del canal (valores negativos indican que el agua se mueve hacia LSB).
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Figure 17. Results of simulated tracer experiments for summer 2002 in LSB. (a) Time series of reduced gravity across the channel joining LSB and LO. (b) Standard deviation of along-channel depth-averaged velocity fluctuations in a 1-day window. Its
mangitude is related to the energy of the water level oscillations (barotropic processes) driving exchange. (c) Volumetric exchange rate for the 11 tracer experiments, given in non-dimensional form and as a percentage of the initial embayment volume.
Resultados de la simulación de experimentos con trazadores durante el verano de 2002, en LSB. (a) Series de tiempo de la gravedad reducida (relacionada con la diferencia de temperatura a ambos lados del canal). (b) Desviación estándar de las fluctuaciones de velocidad promediada en la dirección vertical, en una ventana de tiempo de un día. Su magnitud es representativa de la
energía de las oscilaciones del nivel del agua en LO (procesos barotrópicos) que inducen intercambio de masa entre LSB y LO.
(c) Volumen intercambiado en cuatro días para 11 experimentos simulados de trazadores. La posición de la barra en el eje tiempo indica el momento de la liberación del trazador. El volumen se ha adimensionalizado, utilizando el volumen de agua en LSB
al comienzo del experimento.
conditions (Fig. 17). Baroclinic forcing (associated with density gradients) is, therefore, a much
more effective exchange mechanism than water
level fluctuations in littoral embayments. This
can be better understood if we apply Stommel
and Farmer’s 1952 visualization of exchange
processes driven by water level changes to LSB
inlet, as shown in figure 18a and 18b. Note that
the same fluid mass is not transferred in and out
of the basin on each seiche cycle, the difference
being a net, or residual, mass transfer. The volume of water involved in the exchange, a jet
during flood (filling of LSB) or the potential
flow region during ebb (draining of LSB), will
be, in general (but for the case of very small
embayments), much smaller than the total volume of the embayment. Scaling arguments show
that the average excursion of an individual particle during a typical 2 h seiche cycle is less than
500 m (recall this is the lake-embayment channel
length scale), and considerably smaller than the
length of the basin (4 km). The net exchange rate
is determined by the rate at which the net water
mass introduced during a tidal (event) cycle is
mixed with the rest of the embayment water,
which the numerical simulations indicate is
small. Figure 18c shows, on the other hand, the
circulation pattern induced by large temperature
gradients in the channel, during upwelling
events. During upwelling events gravitational
forces (density differences) bring cold LO water
into the interior of LSB; lake water advanced as
gravity currents and reach the end of the embayment within hours (< 1 day) after penetration
(see Rueda and Cowen 2005). Warmer LSB
water is transported out towards LO through the
surface. The whole basin is, therefore, actively
involved in the exchange process.
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F. Rueda
number of upwelling events in LO, the smaller
the simulated mean residence times in LSB. The
control by upwelling frequency of the transport
rates has strong implications for the study of
spatial patterns of transport in littoral embayments. In LO, for example, the winds are predominantly from the west (Phillips and Almazan
1981); hence, one would expect faster renewal
rates (and different biogeochemical behaviour)
in embayments located along the west and north
shores relative to embayments along the lake’s
southern and eastern shores.
RIVER WATER RESIDENCE TIME
IN RESERVOIRS
Figure 18. Exchange processes in semi-enclosed basins. (a)
and (b) represent Stommel and Farmer’s (1952) visualization of
flood and ebb flows (barotropic forcing) at an inlet, adapted to
the geometry of LSB. (a) The flood flow enters as a confined
jet, while the ebb flow waters (b) are drawn from all around the
mouth in the form of a potential flow to a sink. (c) Basin-scale
circulation in LSB driven by upwelling events in Lake Ontario.
Cold LO water enters as a gravity current following the bottom
and displaces warm LSB waters out through the channel.
Procesos de intercambio en cuencas semicerradas. (a) y (b)
representan la visualización, propuesta por Stommel and
Farmer’s (1952), de los fenómenos de entrada y salida de agua
forzadas por cambios en la superficie libre del agua (mareas o
secas superficiales), particularizada para la geometría de LSB.
(a) La entrada se produce a modo de chorro (‘jet’), mientras
que a la salida (b) el agua es extraída de una región circular
alrededor del canal, como un flujo potencial hacia un sumidero. (c) Circulación a escala cuenca en LSB, forzada por fenómenos de afloramiento en LO. El agua fría de LO entra como
una corriente de gravedad, siguiendo el fondo y desplaza el
agua cálida de LSB, que sale por la superficie.
The simulation results of Rueda and Cowen
(2005b) suggest that the mean residence time TR
(estimated as in Eq.6) in freshwater littoral
embayments depend on the occurrence of upwelling events, the magnitude of the thermal gradients induced across the connections, their
duration, frequency and location in time. In
LSB, for example, TR varied from 15 days to
90 days, depending on those features. Long-term
3D simulations conducted in LSB, furthermore,
suggest that inter-annual variations in the magnitude of the transport time scales in the embayment are related to year-to-year changes in the
number of upwelling events in LO: the larger the
The study of Rueda and Cowen (2005b) showed,
among other things, that widely used approaches
to calculating residence times in dynamic freshwater systems could be wrong. As pointed by
Monsen et al. (2002,) there exists widely spread
misconceptions and confusion among aquatic
scientists on suitable methods for the determination of residence times in dynamic aquatic
systems. In some applications (e.g., Hecky et al.,
1993; den Heyer and Kalff 1998) the computation of transport time has been done without specification of the underlying concept used. In
other cases, the underlying concept and computational steps have been based on an idealized
circumstance that is constrained by critical
assumptions, but the validation (or even recognition) of those assumptions has not always been
considered when applied to a real river, lake,
reservoir or estuary (Monsen et al., 2002). For
example, a widely used expression to estimate
the ‘flushing’ time in reservoirs (Foy 1992,
Sivadier et al., 1994, Straskraba et al., 1995),
consists of dividing the volume of water V stored
in the reservoir by the volumetric flow-rate Q
Tf =
V
Q
(8)
This expression has also been used in this
manuscript as a scale to estimate the importance
of advective processes in driving the hydrodynamics of lakes and reservoirs. However, note that
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51
the name residence time has not been used, purposely, to refer to it. Geyer et al., 2000 defined,
instead, the flushing time as ‘‘the ratio of the
mass of a scalar in a reservoir to the rate of
renewal of the scalar’’, i.e.
Tf =
M
F
(9)
where M is the mass of scalar within the system
and F is the flux of the scalar through the
domain. Both expressions are equal only if one
presumes that the concentration of scalar in the
through-flow is equal to the concentration everywhere within the system. This assumption is
equivalent to considering that reservoirs act as
idealized completely stirred tank reactors CSTR
(Levenspiel 1999), where any introduction of
mass is instantaneously and evenly mixed
throughout the domain. In deep or moderately
deep reservoirs the assumption of instantaneous
mixing is not valid and complex heterogeneus
patterns are usually found (e.g. George 1993).
Therefore, Eq. 8 is fundamentally incorrect.
Stratification develops in the water column on
seasonal time scales and imposes severe restrictions to water movements in the vertical direction. Furthermore, reservoirs are, in general,
highly dynamic systems subject to seasonal and
short-term variations in discharge rates Q
coming in from the watersheds or leaving the
reservoirs, or hence in stored volumes. Mediterranean reservoirs, in particular, experience
large seasonal and inter-annual level fluctuations
as a consequence of their hydrologic behavior
characterized (1) by large variations in runoff
volumes at the inter-annual scales, and (2) at the
seasonal scales by random, scarce and large
inflow events concentrated during the winter
months together with high water demands (withdrawals) concentrated in the summer months. In
highly dynamic systems (such as the Mediterranen reservoirs) it is not clear what values of
Q or V should be entered in Eq. 8. Moreover, it is
not clear at what temporal averaging scales (i.e.
annual, monthly, bi-weekly) should Eq. 8 be
valid. Hence, the estimate provided by Eq. 8 is
not appropriate for this kind of systems (e.g. Toja
1982 or Pérez-Martínez et al., 1991).
Figure 19. Sau Reservoir in North-Eastern Spain. Embalse de
Sau, en el nordeste de España.
Rueda et al. (2005b) reviewed the fundamental
concepts of transport time scales in aquatic
systems as they apply to reservoirs, and presented a physically-based approach to determining
retention time scales in reservoirs. Their approach consisted on simulating tracer releases in the
river using physically-based one-dimensional
transport models. The mean residence time for
river water was estimated, after constructing the
RTD (Eq. 6), following Eq. 7. This approach,
being based on a physically based description of
transport, is fundamentally correct: not only it
takes into consideration the effects of stratification on transport, but it is independent of the
magnitude of level or discharge fluctuations.
Rueda et al. (2006) used this approach (1) to
reveal temporal patterns of variation in average
retention times of river water and (2) to explore
links between hydrodynamic processes and the
transport time scales in a reservoir, taken as
systems with separated inflow and outflow sections. Sau Reservoir (46o46’N - 4o51’E), a canyon type eutrophic reservoir located in NorthEastern Spain was used in that work as a
prototypical example of medium size man-made
lakes (Fig. 19). This is the work described here.
Observations collected during 2003 in Sau
reservoir were used to drive the tracer simulations. A total of 73 inflow releases (one every
5 days in 2003) were simulated, in order to
assess the temporal dependence of mean residence times of river water. The tracer releases
were identified by the day of year when the
release was done. Rueda et al. (2006) show that
the mean residence time TR of river water in
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F. Rueda
Figure 20. Estimated mean residence time for 73 tracer release
experiments simulated with DYRESM. The right scale corresponds to estimates for tracer experiments with release after day
50 in 2003 (solid line). The left scale is for release days before
day 50 (dashed line). Tiempo medio de residencia estimado a
partir de 73 experimentos simulados de liberación de trazadores. Las simulaciones se hicieron con el modelo de transporte
DYRESM. La escala de la derecha corresponde a estimaciones
hechas con trazadores liberados después del día 50, 2003
(línea continua). La escala de la izquierda corresponde a trazadores liberados antes del día 50, 2003 (línea discontínua).
the reservoir undergoes temporal changes at a
range of scales (Fig. 20), from seasonal to
almost-diurnal, associated to the complex interplay of transport processes that determine the
fate of river water within the reservoir. The
equal-weight average value of TR for the
73 experiments conducted is 75 days similar to
that obtained by dividing the average volume by
the average through-flow in 2003 (76 days).
This similarity is, however, deceiving. The
inflow weighted average value of TR, considered
a more correct estimate of average yearly TR,
is almost 30 % less than the V/Q estimate
(76 days). Moreover, by changing the withdrawal elevation the V/Q estimate does not change,
but the average value of TR does: the inflow
weighted average TR for 2003 is 65 days.
The pathways followed by river water entering
the reservoir on any given day t0 from the inflow
point to the outlet was analyzed, as they determine the value of TR. The pathways are controlled
by: (1) the transport and mixing processes that
occur at the inflow section, which determine the
elevation, relative to the outlet, at which river
water penetrates in the reservoir (intrusion layer)
and (2) the interplay of advective and turbulent
diffusion (mixing) processes which determine
the vertical migration of the river water layer
towards the outlet elevation. The most energetic
mixing processes occur close to the surface, driven by wind or surface cooling, and hence the
probability that mixing processes participate in
determining the fate of river water in the reservoir will depend on the depth of intrusion.
Vertical advection of any given layer formed by
intrusion of river water on t0 is the result of the
net extraction of water separating the intrusion
layer and the outlet, which depends on (1) the
withdrawal history (withdrawal elevation, volume, etc.) and (2) the river water insertion history
after t0. In Rueda et al. (2006), it was presumed
that there were no changes in the withdrawal elevation. The inclusion of these operational changes would have modified the numerical estimates of residence time scales, by changing the
thermal structure and the distance from intrusion
depth to the outlet (e.g. Fontane and Labadie
1981, Casamitjana et al., 2003). However, the
factors governing residence times under changing withdrawal elevations are the same as described in Rueda et al. (2006).
SUMMARY AND CONCLUSIONS
As suggested by Ramon Margalef´s Mandala, the
physical processes of transport and mixing exert
a profound influence on the biogeochemistry of
aquatic systems. Not only they determine the
position of particulate and dissolved substances
contained in the water at any given time, but they
also contribute to determine the environment in
which biogeochemical reactions occur (pH, dissolved oxygen, light, etc). This work has reviewed the behaviour of a range aquatic systems,
trying to illustrate that the mechanisms of transport at play in any given environment are determined by both the properties of the external forcing and the characteristics of the system itself
(morphology, in particular). As shown in Eqs. 1
and 2, both features (those of the external forcing
and intrinsic of the system) are taken into
account in the definition of the non-dimensional
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Lake LN and Wedderburn W numbers. They provide a first order description of the physical
behaviour of lakes of small and medium size in
response to perturbation (wind or hydrologic)
events. Scaling arguments, based on length scales (the Rossby radius of deformation Λ compared with the average width of a system) or on
time scales (the ratio of the volume V and the
average through-flow Q) have also been used to
parameterize the influence of the Earth’s rotation
or advective processes on the hydrodynamics of
lakes and reservoirs. These scaling arguments,
based on LN, W, Λ and TR provide a simple and
first order framework in which one can sketch
and classify the physical behaviour of aquatic
systems. A number of systems has been analyzed
in this work to explore a wide range of possible,
but not all, combinations existing in this four
dimensional domain. Transport, however, in
lakes and reservoirs is a dynamic feature and,
therefore, the exact details of how fluxes of mass
and energy into the systems (their magnitude,
frequency, exact location in time, their spatial
variability) determines the large scale circulation
in these systems. In this revision some of those
‘second order’ effects, not captured in the first
order analysis based on simple scales or nondimensional parameters, have been presented.
In Clear Lake, and in particular in one of its
sub-basins (the Oaks Arm) the interaction of
a strong, directional and periodic wind with a
weak stratification has been examined. Clear
Lake is polymictic and exhibits a weak and intermittent stratification, in part, due to being shallow (maximum depth of 15 m) and subject to
strong winds. LN and W are, in general, much
lower than 1 and Λ is of the same order as the
averaged width of the Oaks Arm. The interaction
of wind, stratification and the Earth’s rotation,
occurring with a diurnal periodicity, generates a
net cyclonic residual circulation. The spatial
variability of the wind is also responsible for
creating circulation in shallow lakes and, therefore, establishing the exchange rates among subbasins. In Lake Tahoe, we have explored the
internal waves processes which occur in lakes
with LN >> 1 and W >> 1. Rotational effects
(Λ < 19 km, the width of Lake Tahoe) bring in
53
new ‘creatures’ in the internal wave ‘zoo’: internal waves are either Kelvin or Poincaré waves.
The spatial characteristics are complex and inherently three dimensional. Also, it has been shown
that the energy contained in the internal wave
field varies in time and is a strong function of the
relative phase between wind and the internal
waves themselves. In contrast to Clear Lake and
Lake Tahoe, which are enclosed freshwater
systems where hydrodynamic processes are driven by surface (mechanical or heat) fluxes, the
other two cases presented (Little Sodus Bay and
Sau reservoir) are semi-enclosed or open systems,
where energy and mass fluxes occur at either unique or distinct inflow and outflow sections: TR in
these two cases is < 1 year, compared with 3 years
or 500 years in Clear Lake and Lake Tahoe respectively. In LSB, a semi-enclosed system, there
is a unique forcing section: the channel between
LO and LSB. It is shown there that the net
exchange rate of mass across the channel is largely determined by the occurrence of upwelling
events in LO and their characteristics (magnitude
of temperature gradients induced along the channel, duration, frequency and location in time). In
Sau, with distinct inflow and outflow sections, the
estimates of residence times of river water were
shown to be dependent on the interplay between
processes resulting from the water exchange at
those sections, as well as the vertical mixing rates.
The residence time estimates are subject to variations in time, on a wide range of scales: from seasonal to scales lower than a week.
Through the exploration of transport patterns
in four different systems, it has been shown that
physics in lakes and reservoirs is far from simple. Transport is not only inherently three-dimensional, but also subject to a considerable variability on a wide range of scales. Transport patterns
vary according to characteristics of the external
forcing (which can be arguably thought as stochastic) and the own and fixed characteristics of
the water body. The range of transport patterns
can be, therefore, large. However, although complex, it is possible, through scaling analysis together with the use of advanced numerical and
observational techniques, to determine (leaving
some margin for observational and numerical
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F. Rueda
error) the particular transport patterns that occur
in response to a given forcing scenario. It is plausible to think that the links between transport
patterns and the biogeochemical behaviour of
systems could be explored using similar techniques. However, this exploration poses a serious
challenge, since it would require that physical
and biological sampling be carried on similar
spatial and temporal scales.
ACKNOWLEDGEMENTS
The modelling work carried in Sau reservoir
was done using CWR-DYREMS one dimensional transport model. The author would like to
thank the Center for Water Research CWRUniversity of Western Australia for making the
model available for use in this work. Also, I
want to thank my advisors, staff and research
colleagues from the University of CaliforniaDavis and Berkeley, Stanford University, United
States Geological Survey, Cornell University,
Universidad de Granada and Universidad de
Barcelona, for his invaluable help in conducting
the research mentioned in this work. This research was conducted with support from several
agencies and projects: the Spanish Ministry of
Education and Science who supported the
author during his PhD studies at the University
of California-Davis, the U.S. Environmental
Protection Agency (R819658 & R825433) sponsored Center for Ecological Health Research
(CEHR) at UCDavis, the U.C. Toxic Substances
Research and Teaching Program, the
Ecotoxicology Lead Campus Program at UC
Davis, the National Science Foundation (grant
OCE-9906924), a grant from the U.S. Environmental Protection Agency’s Science to Achieve
Results (STAR) program, National Science
Foundation (NSF) Biocomplexity in the
Environment Program (OCE-0083625) and
National Science Foundation (NSF), CTS0093794, and the Office of Naval Research
(ONR) N00014-98-1-0774, Spanish Ministry of
Science and Technology (project REN20012185-C02-02) and Water Supply Agency ATLL
(contract for the study of Sau reservoir).
In particular, I would like to express my gratitude to Joan Armengol for this kind invitation to
submit this contribution to this special issue of
Limnetica in homage to Ramon Margalef.
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por utilización selectiva de agua a distintas profundidades. Revista de Obras Públicas: 223-231.
TORRENCE, C. & G. P. COMPO. 1997. A practical
guide to wavelet analysis. Bulletin of the American
Meteorological Society, 79(1): 61-78.
WILSON, B. W. 1972. Seiches. In: Advances in
Hydrosciences. V. T. Chow (ed.): 1-94. Academic
Press, New York.
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Limnetica, 25(1-2): 57-70 (2006)
Post-Little Ice Age warming and desiccation of the continental
wetlands of the aeolian sheet in the Huelva region (SW Spain)
Arturo Sousa1,3, Leoncio García-Barrón2, Julia Morales1 and Pablo García-Murillo1
1Department
of Plant Biology and Ecology, University of Seville, C/ Profesor García González, 2, 41012
Sevilla, Spain. [email protected]; [email protected]; [email protected]
2Department of Applied Physics II, University of Seville, Avda. Reina Mercedes s/n, 41012 Sevilla, Spain.
[email protected]
3Corresponding author
ABSTRACT
During the last few decades, studies have been performed and evidence has been found concerning the importance of the climatic period known as the “Little Ice Age” (mid 15th century through late 19th century). However, most of the studies have
been focused to more northern latitudes and, therefore, scarce studies have still been made on the Mediterranean latitudes. In
this paper, an analysis is made of the effects of the post-Little Ice Age warming and on its consequences upon the continental
aquatic ecosystems of the Doñana coastal area and its surroundings. The results of such analysis evidence that the end of this
period –climatically more benign in our latitudes– implied the start of an irreversible regression and disappearance of a large
part of the most typical wetlands in the SW of the Iberian Peninsula. The significant impact of the human exploitation of natural resources in the area has masked the effect of this recent climatic change. Furthermore, when compared with those from
other latitudes, the results of this analysis evidence the global or supraregional features of the impact caused by the post-Little
Ice Age warming. Additionally, these results are useful for indicating which will be the future changing trends in the wetlands
under study as a result of global warming.
Key words: Little Ice Age, global warming, wetlands, lagoons, peat-bogs, Doñana, Huelva, peatlands, aeolian sheets.
RESUMEN
En las últimas décadas se ha estudiado y puesto en evidencia la importancia del período climático conocido como Pequeña
Edad del Hielo (mediados S. XV hasta finales del S. XIX). Sin embargo la mayoría de los estudios se han centrado en latitudes
más septentrionales, por lo que todavía son escasas las investigaciones sobre latitudes mediterráneas. Este trabajo analiza los
efectos del final de la Pequeña Edad del Hielo (post-Little Ice Age warming), y las consecuencias que tuvo sobre los ecosistemas acuáticos continentales del litoral de Doñana y su entorno. Los resultados de este trabajo desvelan que la finalización de
este período –climáticamente más benigno en nuestras latitudes- supuso el inicio de la regresión y desaparición de forma irreversible de gran parte de los humedales más singulares del SW de la Península Ibérica. El gran impacto que tuvo la explotación de los recursos naturales de la zona por parte del hombre, ha ocultado el efecto de este cambio climático reciente.
Asimismo los resultados de este análisis, al compararlos con otras latitudes, ponen en evidencia el carácter global o supraregional del impacto del final de la Pequeña Edad del Hielo. Además estos resultados sirven para indicar cuales serán las tendencias futuras de cambios, en estos humedales, como consecuencia del Calentamiento Global.
Palabras clave: Pequeña Edad del Hielo, Calentamiento Global, humedales, lagunas, lagunas turbosas, Doñana, Huelva, turberas, Manto Eólico Litoral.
INTRODUCTION. THE LITTLE ICE AGE
A climatic period that took place approximately
between 1430 and 1850 (Pita, 1997), characterised by the severity of its winters, is known as
the Little Ice Age and was initially studied
because of the advancements and retreats of the
glacial moraines.
The Little Ace Age (hereinafter LIA) concept
was originally defined by Matthes in 1939
(Grove, 1988) as an epoch of renewed but
moderate glaciations that followed the warmest
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part of the Holocene, when he studied the Sierra
Nevada (California, USA) glaciers. Therefore,
its original definition arose from the field of
glaciological (and not purely climatic) studies.
Although, initially, this climatologically colder
phase has been studied and acknowledged especially in the Alpine Glaciers (Grove, 1988; Le
Roy Ladurie, 1991), it also implied a significant
advancement in the European, North American
and Asian glaciers. Specifically, in the Iberian
Peninsula, it has been studied in the Pyrenean
glaciers (Mateo García & Gómez Ortiz, 2000).
During the LIA, the presence of temperatures
between 1 and 3°C lower than the current ones
has been confirmed in the North Atlantic Ocean
(at about latitude 50°). This cooling trend disappeared in the mid 19th century and was substituted by a new warming process that, with slight
fluctuations, persists until our days (Pita, 1997).
One of the problems posed by this period,
known as exceptionally cold at a global level, is
the difficulty encountered in establishing its
time limits (Font Tullot, 1988; Rodrigo et al.,
1995; Sousa, 2004). As it has been pointed out
by several authors (Grove, 1988; Le Roy
Ladurie, 1991; Rodrigo et al., 1999), probably,
part of the problem rests on the fact that the LIA
itself involved rather a series of frequent fluctuations than a uniform block. However, most of
the authors assert that it ended, approximately,
within the second half of the 19th century.
The fact that it was not a single block and that
warmer intervals arose among the dominating
cold periods does also hinder its interpretation.
This is why Rodrigo et al. (1995) consider that
the concept of LIA must be used with care,
since it cannot be considered as a uniform or
constant climatic phase, in so far as the time
scale is concerned (Rodrigo et al., 1999). Thus,
we are dealing with a climatic period involving
a series of more or less marked fluctuations and,
therefore, the general uniformity of its understanding will depend, to a large extent, on the
time scale under consideration (Sousa, 2004).
The first great reviews on this period –as
from a purely climatic perspective– were carried
out in the late 80s and early 90s (Grove, 1988;
Pfister, 1992). And, precisely, Spain was one of
the 3 countries worldwide where the effects of
the LIA were least known (Grove, 1988). Until
then, only a few essentially descriptive studies
had been made (Rodrigo et al., 1999), such as
those by Font Tullot (1988).
Fortunately, in the mid 90s, the first doctoral
theses dealing exclusively with this period in
Spain started to be developed (Barriendos &
Martín-Vide, 1998). The recent studies performed in Spain reveal that the LIA was characterised by the fact that the increase of aridity
results from the interannual variability of rainfall and from the frequency of several extreme
events, rather than from persistent droughts
(Rodrigo et al., 1995). In short, in diverse studies (Barriendos & Martín-Vide, 1998; Rodrigo
et al., 1999 and 2000) with some subtle differences, three periods within the LIA were
detected as specially humid in the South of
Spain: 1570-1630, 1780-1800 and 1830-1870.
Therefore, the LIA in Andalusia was a specially wet period (thus differing from other that
in more-northern latitudes), even if dry periods
occurred among these three humid episodes.
After the LIA, as of the late 19th century, rainfall in Andalusia has been progressively
decreasing and has only been interrupted by
positive anomalies in the 1960s (Rodrigo et al.,
2000), as it was proved for the observatories in
the SW of Spain by Sousa (2004).
Currently, in the most northern latitudes of
the North Hemisphere, an anomalous warming
has been taking place as compared with the last
three centuries. This trend must be partially
attributed to the recovery of the LIA, but also to
a recent increase of the thermic level. Flannigan
et al. (1998) do also point out an increase in
temperature in the Northern Hemisphere after
the termination of the LIA.
This corresponds with the results of the analyses performed by García Barrón (2002a and
2002b) at observatories in Huelva, which show a
decrease in the spring rainfall and an increase of
the minimum temperatures during the 20th century. They also serve to explain the presence of
last humid peak of the LIA at the end of the 19th
century, as well as the increase in dry years to the
detriment of humid years since the end of the
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Post-Little Ice Age effects on continental wetlands
59
Figure 1. Location of the study area, where the extraordinary number of existing wetlands can be observed. Localización del área
de estudio, donde destaca el extraordinario número de humedales existentes.
19th century, according to the series of rainfall
data of the observatory in San Fernando (Cadiz).
This aridisation, loss of softness or Atlanticity
in the climate since the late 19th century seems to
have led to a global desiccation process in the
local marshy formations (Sousa & GarcíaMurillo, 2002 and 2003; Sousa, 2004), comparable with the one detected by Granados (1987) and
Granados et al. (1988) in the Doñana National
Park. This might also be related to the erosional
processes detected by Devereux (1982) at the
ravines (gullies) in Algarve, Portugal.
AREA OF STUDY. THE WETLANDS OF
THE EASTERN COASTAL AREA OF
HUELVA
The area under study is limited to the continental
humid formations (lagoons and small creeks)
located on the eastern coastal area of Huelva
(between the mouths of the Tinto and Guadalquivir rivers). As can be seen in Figure 1, the
main substrate of these marshy areas is the
Coastal Aeolian Sheet. Therefore, we are referring to a vast coastal area (~44,000 hectares)
located at the southwest of the Iberian Peninsula,
within the Andalusia Region and, more precisely,
in the Province of Huelva (including the municipalities of Almonte, Moguer, Palos de La
Frontera and Lucena del Puerto).
The eastern coastal area is very rich and
diverse in its formations and wetlands, which
are distributed within three natural regions protected by the regulations in force: the Las
Madres and Palos lagoons Natural Area, the
Doñana National Park and the Doñana Natural
Park (Fig. 1). It was only within the boundaries
of the Doñana Natural Park (in its west sector
known as Abalario) that we have studied the
evolution of about 300 small lagoons (both
peat-bogs and seasonal lagoons).
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Sousa et al.
Figure 2. Segregation and division process in the swamped areas of the peridunal lagoons in the Doñana National Park, due to the
advancement of live dunes between the late 19th century and the present. Proceso de segregación y división de las zonas encharcadas de las lagunas peridunares del Parque Nacional de Doñana, debido al avance de las dunas vivas, desde finales del S. XIX
hasta nuestros días.
In addition to the Abalario humid complex
–which would take up the most central area in
this study– the various types of lagoons in the
Doñana National Park were analysed, although
the research work was especially focused on the
peridunal ponds located inside the Doñana
Biological Reserve (Brezo, Charco del Toro,
Taraje, Zaillo, Dulce and Santa Olalla lagoons).
Also under analysis was the genesis of the
coastal lagoons in Palos de la Frontera (Primera
de Palos, La Jara and La Mujer), as well as the
large Las Madres peat-bog, in Moguer, all of
them located within the Las Madres and Palos
lagoons Natural Area.
In spite of the heterogeneity of these continental wetlands, when viewed as from the limnological, hydrogeological and vegetation
points of view, they are all located within a
common geological substrate: the coastal
Aeolian Sheet of Huelva. The coastal Aeolian
Sheet (hereinafter MEL, in its Spanish
acronym) is mostly composed by quaternary
sandy sediments produced by the successive
appearance of several dune fronts.
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61
Table 1. Synthesis of the materials and methods used to rebuild the evolution of wetlands of the Aeolian sheets. Síntesis de los materiales y
métodos empleados para reconstruir la evolución de los humedales del MEL.
Material and methods used
Concrete features
Field data (flora)
Transects of perilagoon vegetation
Vegetation units
Consultation aerial photographs
Satellite images analysed
1998, 1999, 2000, 2001, 2004 and 2005
57 transects and 6 samplings (total 1632.8 m)
Yes
1956, 1984, 1987, 1988, 1991, 1992, 1994, 1996, 1998 and 2000
LANSAT-TM (1986), LANSAT-TM (1987), LANSAT-TM (1990),
and SPOT (1989)
1956, 1998 and 2000 (depending on the specific marshy area)
All the municipalities in the area under study (fundamentally
from the 18th century, although also some data from the 19th
century and from the first half of the 20th century)
All the municipalities in the area under study (fundamentally
the Land Registry of Marqués de la Ensenada, although also
some data from the first half of the 20th century)
17th through 20th centuries
Hygrophytic vegetation maps
Historical floristic inventories
Historical forestry inventories
Data obtained from files and other documentary
sources
Historical cartography
Hypsometric maps
Microrelief
Topographic profiles based on microrelief
Conductivity
> 70 maps and historical navigation charts from the 2nd to
the 20th centuries (although the essential ones are as from the
18th century)
1:50000 and 1:10000 scales
> 2,750 topographic heights interpolated at a 1:10000 scale
Several
Yes
THE DISAPPEARANCE
OF THE WETLANDS
In order to attain a reconstruction of the evolution of the MEL’s continental wetlands, a mixed
method must be followed. The method is
initially based in its reconnaissance and characterisation by means of field work. This allows to
characterize the typology of the wetlands, as
well as the phreatophytic vegetation related to
them. This field phase serves as a basis for the
photointerpretation and cartography –by means
of aerial photographs and satellite images– of
the marshy areas at a highly detailed scale.
A combination of these data with the documental sources available in the files allows for a
very accurate reconstruction of the evolution of
the wetlands under study. To go even further
back in time, the aforementioned results must
be compared with the historical documentary
and cartographic sources. Finally, the analysis is
completed with a study of the area microrelief
(by means of a manual interpolation of the topographic heights at a 1:10000 scale), which
allows to reveal the situation of these wetlands
at the end of the 19th century and, with lesser
accuracy, in the early 17th century. For further
details on the concrete sources and methods
developed on the overall area under study, please see Sousa (2004), Sousa & García-Murillo
(1999, 2001, 2002 and 2003) concerning the
Abalario area, or Sousa & García Murillo
(2005) regarding the Doñana National Park. A
summary of them is shown in Table 1.
The retreat of the continental wetlands in SW
Spain cannot be interpreted as if they were a
single homogeneous unit as far as their behaviour and evolution are concerned. This is
because their ecological, limnologic, waterfeed
and, especially, territorial features are significantly different. This is why we have grouped
them as belonging to the Doñana National Park,
to the Las Madres and Palos lagoons Natural
Area or to the Doñana Natural Park.
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Sousa et al.
Figure 3. Retreat of the peridunal lagoons in the Doñana
National Park, between the late 19th century and the present.
Regresión de las lagunas peridunares del Parque Nacional de
Doñana desde finales del S. XIX hasta nuestros días.
The peridunal lagoons in the Doñana
National Park
At least between the early 18th century and the
late 19th century, these lagoons were altered by the
advancement pulses of the MEL’s Aeolian Unit 5
(Rodríguez-Ramírez, et al., 2005). This turned the
large original swampable area into two large lagoon areas: Santa Olalla and Charco del Toro;
(Sousa, 2004; Sousa & García Murillo, 2006). In
turn, these two large lagoons were, again, split into
the current small ponds. Therefore, the current
string of peridunal lagoons has been generated
based on the advancement of the front of the coastal active dune systems and by silting and splitting
the 2 large original lagoons between the late 19th
century and the early 20th century (Fig. 2).
It is important to note that —as already mentioned by Granados (1987)— these dune advancement pulses are associated with the climatologically driest periods. As it was later
demonstrated by Rodrigo et al. (1994 y1999)
and Barriendos & Martín-Vide (1998), these
dry periods are located among the three abovementioned humid peaks of the LIA.
Further to these changes, the drawing of
water by the Matalascañas tourism centre, along
with the centuries-old occurrence of fires in the
Figure 4. Different phases in the genesis of the Palos de la Frontera and Las Madres lagoons, since the initial plugging of the drainage of
the original creeks [taken from Sousa (2004) modified]. Diferentes fases de formación y génesis de las lagunas de Palos de la Frontera y
Las Madres, a partir de la obturación inicial del avenamiento de los arroyos originales [tomado de Sousa (2004) modificado].
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region (Granados et al., 1986; Granados,
1987;), have affected the original vegetation
irreversibly (Sousa, 2004; Sousa & García
Murillo, 2005). Thus, its area has been reduced
until the current situation was reached (Fig. 3).
Las Madres and Palos lagoons Natural Area
The case of this lagoon complex, located between
the municipalities of Palos de la Frontera and
Moguer, is indeed different. Its transformation
and genesis, from the original Atlantic creeks
(Menéndez & Florschütz, 1964), is due to intrinsic factors of the Spanish Atlantic coast (Sousa,
2004). This dynamics prompted the plugging of
the mouths of these creeks, which were turned,
first, into lagoons (Zazo et al., 2000) and, later
on, into the coastal lagoons that we can find
nowadays. The microtopographic reconstruction
of the area makes it evident that, at the end of the
19th century —during the last wet period of the
LIA— these lagoons were all interconnected
(Fig. 4). This temporary interconnection among
them is confirmed after an analysis of the historical cartographic sources of that time (Sousa,
2004; Fernández Zamudio et al., 2005).
The differences between the great Las
Madres peat-bog and the rest of the Palos lagoons (Jara, Mujer and Primera de Palos) refer to
the intrinsic morphometric and area features of
their hydrological basins.
The lagoons of the Doñana Natural Park
(Abalario sector)
In the case of the numerous lagoons of the
Abalario humid complex, the evolution has been
very different from the former ones. Thus,
during the second half of the 20th century, the
forestry activities in the region (especially
the application of monocultures of eucalyptus),
prompted the desiccation of many lagoons. This
effect was especially significant upon the
Rivatehilos peat-bogs, which went from 178 to
only 30 lagoons during the 1956-1987 period.
These lagoons, located on peat soils and originally occupied by communities of Erico ciliarisUlicetum (minoris) lusitanicus (García Murillo
63
& Sousa, 1999), underwent an intensive desiccation process that led to the mineralisation of all
the retained organic matter. This desiccation,
caused by the lowering of the phreatic level
(Sousa & García Murillo, 1999 and 2003, Trick
& Custodio, 2003), prompted the substitution of
the above-mentioned heather communities by
less-stenohydric hygrophytic bushes (Erico scopariae-Ulicetum australis).
However, when the evolution of these humid
formations is historically reconstructed, it can
be seen that this desiccation process is prior to
the human activities in the region. Actually, it is
a reduction process that started, at least, in the
early 17th century, after the iciest period –and
the most humid in our latitudes– of the LIA.
Since then, 17th century, the surface covered
by the Rivatehilos peat bogs has been decreasing
at mean rate of 1.2 hectares/year. When the third
and last humid peak at the termination of the
LIA (late 19th century), this rate is doubled and
the lagoon area decreases at a mean rate of 2.4
hectares/year. The forestry activities throughout the Abalario region accelerate this process
exponentially (mean rate 43.6 hectares/year)
during the second half of the 20th century, masking the previous climatic changes (Fig. 5).
On the other hand, the seasonal lagoons
in Abalario, due to their epigean waterfeed,
have been less affected by the reduction of the
phreatic level due to the implantation of large
masses of eucalyptus.
Since these are lagoons that depend only on
the rainfall and on the surface runoff, they
reflect very rapidly any change in the climatic
trends. Thus, after the last wet period of the LIA,
in the late 19th century, a significant reduction
starts to take place in the Laguna de Invierno
(covering almost 400 hectares and 5 km long)
and in other smaller seasonal lagoons, at a mean
rate of 5.2 hectares/year. The disappearance of
this large seasonal lagoon –known as Laguna de
Invierno (Winter Lagoon)– is highly relevant
because, in Valverde’s opinion (1885-1888 and
1880), it was the largest in the province of
Huelva. In fact, currently, in Andalusia, its area
would only be exceeded by the Laguna de
Fuente de Piedra (Málaga).
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Figure 5. Evolution of the peat-bogs in Rivatehilos, showing the disappearance of some large lagoons and the reduction of others
linked to recent forestry activities. Evolución de las lagunas turbosas de Rivatehilos, con la desaparición inicial de las grandes
lagunas y la reducción de otras vinculada a las actividades forestales más recientes.
Table 2 shows a comparison of the mean annual
reduction rates for the two lagoon types under
study within the Doñana Natural Park.
Figure 6 shows a comparative version of the
evolution of both types of lagoons at the
Doñana Natural Park (Abalario Sector), between the 17th and the 20th century.
We have referred to different continental
wetlands; nevertheless, similar processes have
also been found in the linear wetlands of the
region. This would be the case of gullies draining on the right margin of La Rocina Creek
(Sousa & García Murillo, 1998 and 2000; García
Murillo & Sousa, 1999; Sousa, 2004) or of the
interesting Atlantic creeks, many of which disappeared before the start of the 20th century
(Sousa & García Murillo, 1998; Sousa, 2004).
Most of them are currently in a process leading
to the full disappearance, in spite of the fact that
they lodge exceptionally valued floristic com-
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65
Table 2. Mean annual reduction rate of the lagoons in the Doñana Natural Park (Abalario sector), in hectares/year. Tasa media anual de disminución de las lagunas del Parque Natural de Doñana (sector Abalario), en ha/año.
17th (≅1630)–19th (≅1869)
19th (≅1869) - 1956
1956-1987
1.2
0.1
2.4
5.2
43.6
*
Rivatehilos peat-bogs
Seasonal lagoons in Abalario
* Non significant because, practically, there is no reduction. Only a slight increase is noticed due to the fact that some peatbogs start to behave as seasonal lagoons.
Figure 6. Comparative evolution of the peat-bogs and seasonal lagoons in the Doñana Natural Park. Evolución comparada
de las lagunas turbosas y temporales del Parque Natural de
Doñana.
munities with an Atlantic influence, as it is the
case of the Sphagnum inundatum moss (García
Murillo et al., 1995), among other taxons.
Until now, we have outlined the evolution of
these wetlands and mentioned some of the factors hidden behind these changes. In the following section, we are going to provide a more
thorough description of the reasons for this
retreat, using these results as a basis for analysing the effects of the ending of the LIA in the
SW of Spain. What were the features of this climatic period that could affect these coastal
wetlands so significantly?
REASONS FOR THE RETREAT
OF THE WETLANDS
As it was mentioned above and as shown in different documents (Sousa & García Murillo,
1998, 1999 and 2001), human activities have
conditioned and affected the wetlands of the
MEL of Huelva very intensively. However, as
shown by Sousa (2004), the studies on the usage
of the region make it evident that human activi-
ties were not relevant until the second half of the
20th century. This can be followed up very clearly, in the case of the wetlands of the Doñana
Natural Park, by analysing usage and the occupation by forestry monocultures.
In the case of the Las Madres and Palos lagoons Nature Area, the greatest impact is also
recent and related to the exploitation of peat,
the growing of strawberries and reforestation
(Márquez, 1986 and 1992; Garrido, 1996;
Sousa, 2004; Fernández Zamudio, 2005;
Fernández-Zamudio et al., 2005).
On the other hand, studies by Granados et al.
(1986) and Granados (1987), reveal the important role played by fire as a differentiating factor
at the Doñana National Park. The application of
GIS to these data, along with the support of historical cartography, has also revealed the severe
impact caused by fires upon the original phreatophytic vegetation of the peridunal lagoons in
the Doñana Biological Reserve (Sousa, 2004;
Sousa & García Murillo, 2005). Probably,
the impact of fire, along with the decrease of the
phreatic level in the region (due to usage and at
the termination of the LIA), are the basic reasons
for the retreat of a good number of peat-bogs in
the Aeolian Unit 2 of the Doñana National Park.
However, all these human activities are
not enough to explain the general retreat of
the MEL lagoons in Huelva and, especially, that occurred until the start of the second half of the 20th century.
Hydrogeological studies (Trick & Custodio,
2003) reveal an important reduction of the
phreatic level in connection with great forestry
monocultures (pines and eucalyptus). However,
when an analysis was made of the historical
sequence in the disappearance of the peat-bogs
of Rivatehilos since the 17th century (Sousa &
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Figure 7. Representation of the absolute frequency of Very Dry, Dry, Normal, Wet and Very Wet years at the observatory of San
Fernando (Cadiz). Representación de la frecuencia absoluta de años Muy Secos, Secos, Normales, Húmedos y Muy Húmedos en el
observatorio de San Fernando (Cádiz).
García-Murillo, 1999 and 2003), it was confirmed that the process is prior to the reforestation
(although intensified by the latter). Which are
the reasons for this decrease of the phreatic
levels, prior to the most intensive human activities in the region under study?
A comparative analysis of the reduction in the
area of the seasonal lagoons and peat-bogs of
Abalario between the 17th and the 20th centuries
reveals a clear relationship with the climate.
Thus, a decrease in the number of humid years
and an increase in the number of dry years might
explain the disappearance of the large seasonal
lagoons in the MEL (Sousa & García Murillo,
1999), since the availability of feedwater (which
is epigean) was smaller. This explanation seems
to be fully coherent with the features and changes
detected in the rest of the MEL’s wetlands.
In order to confirm this hypothesis involving a
change in the sequence of dry years vs. wet years
in the late 19th century (and related to the last
wet period of the LIA), an analysis can be made
of the historical rainfall and temperature series.
Among the most complete instrumental series
in the area, the one related to rainfall at the
observatory of San Fernando (Cadiz) is outstanding (Sousa, 2004). The analysis of the
trends is performed by means of a distribution
of quintiles of the sample (Arlery et al., 1973),
following the recommendations of the WMO.
The result obtained (Fig. 7) shows a markedly
wet period at the end of the 19th century, as
there is an increase of the absolute frequency in
the humid and very humid years.
In order to further highlight this climatic
inflection point, which coincides with the termination of the LIA, also the accumulated frequency can be represented. If, furthermore, the
humid and very humid years and the dry and
very dry years are grouped separately (Fig. 8), a
change in the climatic trend appears clearly.
This wet period squares with what Barriendos
& Martín-Vide (1998) dated between 1830 and
1870 for Mediterranean Spain and with what
Rodrigo et al. (1994 and 1999) dated at the end
of the 19th century for Andalusia (after studying
floods and other non-directly climatic sources),
as the third humid peak of the LIA.
This humid phase did also imply an increase
in the spring rainfall (García Barrón, 2002a and
2002b) and greater anomalies in the rainfall
(relative accumulated deviations of the mean
annual rainfall). This increase in both spring
rainfalls and irregularity appears as associated
with the last wet period of the LIA. Contrarily,
as of the 20th century, unmistakably, a decreasing trend of the spring rainfall can be found
(although not of the total annual rainfall).
Another climatic factor to be considered could
be the increase of the mean minimum temperatures, which was detected in several observatories
near the area under study (García-Barrón & Pita,
2004). This increase is not observed in the mean
maximum temperatures or in the mean annual
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67
Figure 8. The representation of the accumulated frequency of wet versus dry years is a clear indication of the climatic inflection
point occurred in the late 19th century: the end of the LIA in the SW of Spain. La representación de la frecuencia acumulada de
los años húmedos frente a los años secos, pone claramente de manifiesto un punto de inflexión climático a finales del S. XIX: el fin
de la LIA en el SW de España.
temperatures, thus differing from the situation in
other more-northern observatories of the North
Hemisphere (IPCC 2001 and even in the north of
Spain (Castro et al., 2005). Figure 9 shows the
increasing trend of the mean minimum temperature at the Huelva station (Huelva).
These data obtained from the instrumental
series confirm the initial hypothesis (previously outlined) concerning the features of
the termination of the LIA in the SW of Spain.
Thus, in addition to characterising this period
as different from that in other more-northern
latitudes (even in the north of Spain), it
explains the reasons for the natural changes
underwent by the Doñana wetlands and their
surroundings, before the start of any significant human activity in the area.
POST-LITTLE ICE AGE AND GLOBAL
WARMING THE FUTURE OF THE
WETLANDS AT THE HUELVA MEL
It may be concluded that the termination of the
LIA (what, in other more-northern latitudes of
Europe has been named as the post-Little Ice
Age warming) had different effects in southern
latitudes as compared with those (far better
known) of the more-northern latitudes.
Thus, the winter severity feature of the LIA
in more-northern latitudes was characterized in
the southern Iberian Peninsula by increased
rainfall (Rodrigo et al., 2000). The fact is that
the climate change in the Iberian Peninsula may
be related to precipitation rather than to temperature (Pfister et al.,1999).
Figure 9. Since the early 20th century, a rise takes place in the mean minimum temperatures of the area under study. Desde principios del S. XX se produce un incremento en las temperaturas medias de las mínimas en el área de estudio.
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Sousa et al.
This interpretation agrees with the studies on
the rainfall anomalies in Andalusia (16th-20th
centuries) performed by Rodrigo et al. (1994,
1995 and 1999), as well as with the climatic
reconstruction studies made by Barriendos
& Martín-Vide (1998) on the basis of the
hydrological levels and floods in the rivers of
the Spanish Mediterranean basin (Cataluña,
Valencia, Murcia and Baleares).
This is why, as opposed to the situation in other
more-northern European and Spanish latitudes, in
the SW of Spain, the LIA brought along:
• A decrease in the frequency of wet years and,
on the other hand, an increase in the number of
dry years as from the late 19th century. This
decrease in the number of wet years implied a
decrease in the spring rainfall during the 20th
century. In fact, this implies an aridisation or,
at least, more marked Mediterranean conditions, as opposed to the oceanic and Atlantic
features of the Huelva coastal region.
• This effect has been reinforced and sustained
by a gradual and constant rise of the mean
minimum temperatures that, in turn, produced an increase in the evapotranspiration
rates (thus favouring the reduction of the flooded area and the retreat of the most demanding hygrophytic vegetal communities).
Age warming– would be the result of the post19th century warming that –in our opinion– was
due to the summation of the termination of the
LIA and the start of the global warming.
Anyhow, these results make it evident that, if
global warming is as intensive as predicted by the
various models, the Doñana aquatic ecosystems
and their surroundings will be affected negatively. This alteration will depend on which are the
most affected climatic variables (important changes in the amount or seasonal distribution of the
rainfall appears as more critical than changes in
temperature). There will also be an incidence of
the various biological and waterfeed features
involved in the regimes of the different types of
wetlands, even if everything seems to indicate
that the ones more liable to be affected are those
associated with an epigean waterfeed and with a
more stenohydric vegetation (such as the areas
with peat soil), as well as the floristic elements
whose distribution shows Atlantic features.
ACKNOWLEDGMENTS
These studies were supported in part by
the project Doñana 2005 from the Spanish
Environment Ministry.
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Sometimes simultaneously and sometimes
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Limnetica, 25(1-2): 71-80 (2006)
Aquatic macrophytes in Doñana protected area (SW Spain):
An overview
P. García Murillo1, R. Fernández Zamudio1, S. Cirujano2 & A. Sousa1
1Departamento
de Biología Vegetal y Ecología. Facultad de Farmacia. Universidad de Sevilla. Apdo. 874.
E-41080 Sevilla. Spain. [email protected], [email protected] & [email protected]
2Real Jardín Botánico, CSIC. Plaza de Murillo 2. E-28014-Madrid. Spain. [email protected]
ABSTRACT
A big portion of the Doñana protected areas corresponds to wetlands; in them aquatic macrophytes are the main primary producers and play also other important ecological functions. Nevertheless, they are inconspicuous organisms and their importance in these ecosystems does not seem to be well reflected in the bibliography about this natural area. This paper reviews the
most significant information gathered about this group of organisms in this protected area, provides an updated catalogue of
this group of plants, and offers some considerations related with this topic.
Key words: Doñana, aquatic macrophytes, aquatic vegetation, SW Europe.
RESUMEN
Una gran parte de los espacios protegidos de Doñana corresponde a humedales, en ellos los macrófitos acuáticos son los
principales productores primarios, realizando además otras importantes funciones ecológicas. Sin embargo, son organismos
poco conspicuos y su importancia en estos ecosistemas no parece estar reflejada en las publicaciones existentes relativas a
este espacio natural. Este artículo recopila la información más significativa sobre este grupo de organismos en este espacio
protegido, proporciona el catálogo actualizado de este grupo de vegetales y ofrece algunas consideraciones relativas al tema.
Palabras clave: Doñana, macrófitos acuáticos, vegetación acuática, SW Europa.
INTRODUCTION
A large portion of the Doñana protected area
(Fig. 1) is composed of wetlands. In these ecosystems, aquatic macrophytes are responsible for
most primary production and also play an important role in increasing ecosystem structures or
recycling nutrients and elements. Aquatic macrophytes are, therefore, key elements in this paradigmatic natural area. Moreover, flora is one of
the best natural sources of information regarding
current and potential conservation in any natural
place. The scarce number of studies on this conspicuous group of organisms is thus surprising.
When the last Doñana Floristic Catalogue
was published twenty-five years ago (Castoviejo
et al., 1980), it was quite thorough at that time;
however, the bulk of new floristic records since
then have been aquatic plants (as shown in this
Figure 1. Location of Doñana protected areas. Localización
de los espacios protegidos de Doñana.
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García Murillo et al.
paper). This indicates how little is known about
aquatic plants in this area.
This is coupled with the fact that some significant environmental events have occurred in
this same time period, an increase in intensive
agriculture in the area, increased tourism in bordering areas, overexploitation of ground waters,
invasions of exotic species, global warming, etc.
All of these issues point to the need to update the information related to the aquatic
macrophytes in the Doñana protected area,
which is the aim of this paper.
FIRST STEP: CATALOGUING
Unlike the cases of other places of great natural
value in the southern Iberian Peninsula, such as the
Sierra Nevada or Sierra de Grazalema, which were
prospected by botanists during the nineteenth century or before, the first studies on Doñana´s flora
appear much later in the mid-twentieth century. In
1945 C. Vicioso, an Aragonese botanist, published
a list of taxa collected in the south part of the
Huelva Province, but among these references there
was no data on macrophytes. It was not until 1967,
when references to aquatic macrophyte were
found, that seven aquatic macrophytes were
cited in an invertebrate catalogue (Mazaranov,
1967) for Guadalquivir Marshes (belonging to
Doñana). Cabezudo later began the systematic
study of flora in this preserved area, including
18 species of aquatic macrophytes in his studies
(Cabezudo, 1974; 1975; 1978, and Galiano &
Cabezudo, 1976). In 1981, some years later, the
brilliant work of Castroviejo et al. (1980) completed the information compiled by Cabezudo, nearly
completing the list of vascular plants in Doñana
National Park. In subsequent years this Catalogue
has changed very little, with the exception of
aquatic plants that have contributed some important new records to the Doñana Catalogue:
Althenia orientalis (García Murillo & Talavera,
1986), Callitriche lusitanica (Pizarro, 1990),
Zannichellia obtusifolia (Talavera et al., 1986,
Lemna trisulca, and Spirodella polyrhyza (García
Murillo et al., 1991), etc. (Table. 1).
In terms of the “other” plant groups included in aquatic macrophytes, i.e., Charophyta and Bryophyta, data on these are more
scarce and inconsistent.
The first records of Charophyta were noted by
Corillion (1961), who included two of Doñana’s
Charophyta species in his work on southern
Spain and North Africa charophytes: Chara connivens and Nitella flexilis. Some years later,
Comelles (1982) and Sánchez (1984) added two
more taxa to the list: Tolypella hispanica and
Chara fragifera, respectively. Almost one decade
later García Murillo et al. (1993) added nine
new records to Doñana’s charophyte catalogue.
Finally, the most recent records are on Chara
vulgaris var. oedophylla and Tolypella salina,
referenced by Espinar et al. (1997).
In the Bryophytes group, there are two
papers on the Riella genus (Cirujano et al.,
1988 and 1992); Riccia fluitans and Ricciocarpos natans were mentioned by Rivas
Martínez et al. (1980) and Sphagnum inundatum by García Murillo et al. (1995).
Table 1 shows the complete and current catalogue of Doñana’s submerged macrophytes. It
includes 74 taxa (21 more than those related in
1993 by García Murillo et al.) of which 46 are
Spermatophyta (62 %), 3 Pteridophyta (4 %), 6
Bryophyta (8 %) and 19 Chlorophyta (26 %).
Besides, this table points to the first floristic
record of each taxon.
Table 1. Catalogue of aquatic macrophytes of Doñana protected areas. Catálogo de los macrófitos acuáticos de los espacios protegidos de Doñana.
TAXA*
FIRST RECORD
CHLOROPHYTA
Characeae
Characeae
Characeae
Chara aspera Deth. ex Willd. var. aspera
Chara canescens Desv. & Lois.
Chara connivens Salmz. ex A. Braun
García Murillo, Bernués & Montes, 1993
Corrillion, 1961
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Aquatic macrophytes in Doñana
Table 1. Continued. Continuación.
CHLOROPHYTA
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Characeae
Chara fragifera Durieu
Chara fragilis Desv.
Chara galioides DC.
Chara hispida L.
Chara vulgaris L. var. vulgaris
Chara vulgaris L. var. contraria
(A. Braun ex Kütz.) J. A. Moore
Chara vulgaris L. var. oedophylla
(Feldman) R. D. Wood
Chara vulgaris L. var. longibracteata (Kütz.)
J. Groves & Bullock-Webster
Lamprothamnium papulosum (Wallr.) J. Groves
Nitella flexilis (L.) C. Agardh
Nitella hyalina (DC.) C. Agardh
Nitella tenuisissima (Desv.) Kütz.
Nitella translucens (Pers.) C. Agardh
Tollypella glomerata (Desv.) Leonh.
Tollypella hispanica Nordst. ex T.F. Allen
Tollypella salina Corillion
Sánchez, 1984
Fernández Zamudio et al. (2006)
Van Vierssen et al., 1982
Corrillion, 1961
Comelles, 1982
Espinar et al., 1997
Riccia fluitans L.
Ricciocarpos natans L.
Riella cossoniana Trabut
Riella helicophylla (Bory & Mont.) Mont.
Riella notarisii (Mont.) Mont.
Sphagnum inundatum
Rivas-Martínez et al., 1980
Cirujano et al., 1992a
Cirujano et al., 1988
Cirujano et al., 1992a
García Murillo et al., 1995
Azolla filiculides Lam.
Isoetes velatum A. Braun subsp. velatum
Marsilea strigosa Willd.
Cobo et al., 2003
Galiano & Cabezudo, 1976
Apium inundatum L.
Oenanthe fistulosa **L.
Thorella verticillatinundata** (Thore) Briq.
Oenanthe globulosa** L.
Carum verticillatum** (L.) Koch
Eryngium corniculatum** L.
Eryngium galiodes** Lam
Callitriche truncata Guss. subsp. occidentalis
(Rouy) Schotsman
Callitriche lusitanica Schotsman
Callitriche stagnalis Scop.
Callitriche brutia Petagna
Callitriche obtusangula Le Gall
Ceratophyllum demersum L.
Scirpus fluitans L.
Allier & Bresset, 1975
Cabezudo, 1975
Cabezudo, 1974
Cabezudo, 1975
Cabezudo, 1974
Cabezudo, 1978
Castroviejo et al., 1980
Espinar et al., 1997
BRYOPHYTA
Ricciaceae
Ricciaceae
Riellaceae
Riellaceae
Riellaceae
Sphagnaceae
PTERIDOPHYTA
Azollaceae
Isoetaceae
Marsileaceae
SPERMATOPHYTA
Apiaceae
Apiaceae
Apiaceae
Apiaceae
Apiaceae
Apiaceae
Callitrichaceae
Callitrichaceae
Callitrichaceae
Callitrichaceae
Callitrichaceae
Ceratophyllaceae
Cyperaceae
Pizarro, 1990
Mazaranov, 1967
73
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SPERMATOPHYTA
Elatinaceae
Elatinaceae
Elatinaceae
Halogaraceae
Halogaraceae
Hydrocharitaceae
Juncaceae
Lemnaceae
Lemnaceae
Lemnaceae
Lemnaceae
Lemnaceae
Lentibulariaceae
Lentibulariaceae
Nymphaeaceae
Nymphaeaceae
Polygonaceae
Potamogetonaceae
Potamogetonaceae
Potamogetonaceae
Potamogetonaceae
Potamogetonaceae
Potamogetonaceae
Ranunculaceae
Ranunculaceae
Ranunculaceae
Ranunculaceae
Ruppiaceae
Ruppiaceae
Zanichelliaceae
Zanichelliaceae
Zosteraceae
Elatine alsinastrum L.
Elatine macropoda Guss.
Elatine hexandra (Lapierre) DC.
Myriophyllum alterniflorum DC.
Myriophyllum spicatum L.
Hydrocharis morsus-ranae L.
Juncus heterophyllus Dufour
Lemna gibba L.
Lemna trisulca L.
Lemna minor L.
Spirodella polyrhiza (L.) Schleiden
Wolffia arrhiza (L.) Horkel ex Wimmer
Utricularia exoleta R. Br.
Utricularia australis R. Br.
Nuphar luteum L.
Nymphaea alba L.
Polygonum amphibium L.
Potamogeton natans L.
Potamogeton polygonifolius Pourret
Potamogeton lucens L.
Potamogeton trichoides Charm. & Schlecht.
Potamogeton crispus L.
Potamogeton pectinatus L.
Ranunculus tripartitus DC.
Ranunculus peltatus subsp baudotii
(Godron) Meikle ex C. D. K. Cook
Ranunculus peltatus subsp saniculifolius
(Viv.) C. D. K. Cook
Ranunculus peltatus Schrank subsp fucoides
(Freyn) Muñoz Garmendia
Ruppia maritima L. var. maritima
Ruppia drepanensis Tineo
Althenia orientalis (Tzvelev)
García Murillo & Talavera
Zannichelllia obtusifolia Talavera,
García & Smith
Zostera noltii Hornem
Mazaranov, 1967
Mazaranov, 1967
Cabezudo, 1975
Mazaranov, 1967
Van Vierssen et al., 1982
Cabezudo, 1978
Mazaranov, 1967
García Murillo, 2000
Cabezudo, 1975
Mazaranov, 1967
Cabezudo, 1978
Mazaranov, 1967
Mazaranov, 1967
Cabezudo, 1978
Allier & Bresset, 1975
Pizarro, 1993
Cirujano et al., 1992b
Cabezudo, 1978
García Murillo & Talavera, 1986
Talavera et al., 1986
* Some taxa refereed to Doñana Protected Areas have been related with incorrect identifications (as Callitriche palustris L.,
Hippuris vulgaris L., Zannichellia palustris L. or Zannichellia peltata Bertol.) in other cases they correspond with synonyms
(as Ranunculus baudotii Godron; Ruppia maritima subsp. drepanensis L.(Tin.) Maire & Weiller or Utricularia gibba L.).
** The juvenile form of these species show morphological, anatomical and physiological characters corresponding with aquatic macrophytes.
SECOND STEP: AQUATIC MACROPHYTES
AND ENVIRONMENTAL FACTORS
At the end of the 1970s, González Bernáldez
directed a series of studies on the relationship between the plants in Donaña and the environment
(see García Novo, 1997). In this context, there
was practically no mention of aquatic plants, with
just one study found on the marsh’s vegetation
(Allier & Bresset, 1977). In 1980, Rivas Martínez
et al. published an excellent work on the vegetation in the Doñana National Park, in which they
carried out a detailed phytosociological analysis
of the different communities of plants in this protected natural area. Nevertheless, despite the
superior quality of the research done, the informa-
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tion on aquatic plant communities was insufficient (as can be deduced from the chorological
changes after 1980, included in Table 1).
In the 90s, in response to researches by
Montes on Donaña’s aquatic ecosystems, new
data appeared on the ecology of aquatic plants in
the area: data was published on their biomass
(Duarte et al. 1990); the main factors (flooding
time and salinity) controlling the distribution of
these plants in Doñana’s wetlands were identified (Bernués, 1990; Duarte et al., 1990); and a
study was done on the marsh’s seed bank (Grillas
et al., 1993). Santamaría (Santamaría, 1995;
Santamaría & Hootsmans, 1998; Santamaría et
al., 1995; 1996), under the direction of Montes
and with samples collected from Doñana, also
carried out a series of studies on the autoecology
of Ruppia drepanensis Tineo, one of the most
abundant underwater macrophytes in the Doñana
salt marsh. In this same period, Serrano & Toja
(1995), working in sand lagoons of Doñana,
related the presence of some aquatic macrophytes with other ecological parameters.
And finally, more recent works by Espinar
(2004) and Espinar et al. (2002) have contributed valuable information on the salt marsh
aquatic macrophytes in relation to their environment and to communities of helophytes.
Studies on Seed Dispersal
At the turn of the century, and as a result of the
multidisciplinary approach taken by the Doñana
Biological Station to environmental processes
and with the involvement of Santamaría, a
series of studies appeared on the role of birds in
the passive transport of organisms. To be sure, a
large number of these studies focused on
the dispersal of aquatic macrophyte seeds by
aquatic birds (Charalambidou et al., 2003;
Green et al., 2002; Figuerola & Green, 2002
and 2004; Figuerola et al., 2002; 2003 and
2005). These researches have been consolidated
as a line of work, which is currently being
carried out in the Doñana Biological Station
under the direction of Green, with outstanding
results. Likewise, Espinar et al. (2004) have
recently published studies in this area.
75
Aquatic Plants and Climate Change
As stated in the introduction to this paper, a site’s
flora is one of the best natural sources of information on that area. This fact is even more perceptible in aquatic plants since their reaction to
environmental changes (due to their particular
physiology) is much faster and precise. Based
on this premise, recent studies carried out in the
eastern part of Doñana National Park and in the
park’s lagoons have shown how useful diachronic
studies on the presence and distribution of aquatic plants can be in detecting climatic changes
within relatively recent timeframes. The work of
Sousa (2004) and Sousa & García Murillo (1998;
1999; 2003 and 2005) illustrate this fact and find
an explanation for the processes of aridization
and desiccation of the coastal wetlands of Huelva
Province by linking these processes to the end of
the Little Ice Age, using –among other things–
the presence of certain species of aquatic plans,
the distribution of vegetation in the wetlands,
and their changes over time.
THIRD STEP: REGARDING
CONSERVATION
The uniqueness of the diverse species of aquatic
macrophytes found in south-western Europe has
been pointed out by some authors (Cook, 1983;
García Murillo, 2003; Montes & Martino, 1987);
the majority of these taxa are located in the
Doñana area. The work of Cirujano et al. (1992b)
is noteworthy here in its ranking of Spanish
wetlands based on the presence of certain species
of macrophytes; the Doñana salt marsh was ranked second among all wetlands considered.1
In addition, in the late 1990s, the Andalusian
Regional Environmental Agency (later, the
Environmental Council) started a line of research
aimed at identifying the biology of the plant species most at risk from a conservation standpoint.
1 In the work cited, the authors did not consider the Doñana
lagoons or the changes in flora, which occurred after the work’s
publication, which clearly would have significantly increased
the ranking of this area.
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The most salient outcomes of these studies were
the “Andalusian Plant Species Red List (Decreto
104/1994; Ley 8/2003)” and two volumes compiling the most significant information on the species selected (Blanca et al., 1999-2000). In contrast to previous Spanish “Red Lists” (ICONA,
1987), this one included a significant number of
aquatic species, as well as Althenia orientalis,
Hydrocharis morsus-ranae, Utricularia exoleta,
Wolffia arrhiza, Marsilea strigosa, and Thorella
verticillatainundata (García Murillo, 2000;
Silvestre, 2000), found in the Doñana area.
This was followed, as proposed by the Ministry
of Environment, by numerous studies intended to
update the “Red List” (ICONA, 1987) nationally.
These studies were compiled in the AFA Project
(Bañares et al., 2003), including the red list and
the most relevant data from the research done on
the different species, although it was not possible
to study some of the species listed. The AFA
red list includes two aquatic macrophyte species found in Doñana (Utricularia exoleta and
Hydrocharis morsus-ranae).
The surveys and research carried out for the
Red Lists generated numerous articles which
highlighted the state of some species of aquatic
macrophytes. Included among such articles
were those of Cirujano et al. (1998) and García
Murillo et al. (2000) on species found in the
Doñana protected area.
Exotic Organisms
The early 1980s brought the detection of the first
exotic organisms in Doñana (García Murillo et
al., 2004b). The first of such invaders was the
American crawfish (Procambarus clarckii) and
its spectacular proliferation. The ability of
P. clarckii to physically transform its environment
and alter the availability of resources for other
species in the aquatic ecosystems in which it was
introduced, deeply concerned environmentalists
and scientists. Its effect upon macrophyte communities was tremendous, given that they are
its principal food source. After the initial period
of crawfish expansion, numerous Doñana
macrophyte communities were simply dwindling,
with some species wiped out due to the activity of
this animal (Bravo et al., 1993; Duarte et al.,
1990; García Murillo et al., 1993). The numerous
studies on P. clarkii, undertaken by the UAM
(Universidad Autónoma of Madrid) research
team headed by Montes, have emphasised the fact
that it is now a key element in most of the aquatic
ecosystems in Doñana and a significant control
factor when it comes to aquatic macrophyte
populations in this protected area (Bravo et al.,
1993; Gutiérrez-Yurrita et al., 1998).
Likewise, the Azolla filiculoides species –a
floating pteridophyte native to the New World–
began to appear in the Doñana marsh in the
early part of the 21st century (Cobo et al., 2003
and García Murillo et al., 2004a). In just a couple of years, its presence has extended over
nearly the entire marsh, forming carpets sometimes reaching 10cm thick, which can be clearly
seen from the RBD (Doñana’s Biological
Reserve) plane used for bird surveys. These carpets prevent the sun’s rays from reaching the
water below, thereby making it impossible for
submerged macrophytes (nearly all present in
this area) to develop. They also increase
eutrophyzation since they can fix nitrogen, and
their respiratory activity consumes the oxygen
in the water below (García Murillo et al. 2004a).
Just as with the American red crawfish, the
changes in the aquatic ecosystems of Doñana
attributable to Azolla filiculoides may be dramatic. The attempts to control it have, to date, been
futile (García Murillo et al., 2004b).
Finally, in December of 2004, the tropical
neophyte Pistia stratiotes was found in some
irrigation canals located in the Doñana Park in
the area of Sanlúcar de Barrameda covering
3Km of canals (García Murillo et al., 2005a).
Thanks to the quick intervention of the
Andalusian Regional Council on the Environment –faced with the risk posed by this new
invasion to the Doñana protected areas– and the
low temperatures in January 2005, the Pistia
carpets were eliminated (García Murillo et al.,
2005b). Nevertheless, the risk continues to exist
since some of the Pistia plants sampled had flowers and seeds, and it is well-known that the
seeds of this species can remain functional for
long periods of time buried under the water.
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CONCLUSIONS
In the time span since the publication of the last
floristic catalogue on the Doñana area, numerous references whose primary or secondary
objective is the study of Doñana’s aquatic
macrophytes can be cited, although they are few
in comparison to existing information on other
in the same area.
With regard to the catalogue on aquatic
macrophytes, we believe it is completed except
for the addition of new exotic plants whose
effects encompass this natural area (Cobo et al.,
2003), and the withdrawal of others –cited by
20th-century researchers– which have disappeared due to the deterioration in water quality and
the pressure of the environment surrounding
this protected natural area.
The regional administration –and to a lesser
extent the national one– has taken this situation
into account, by including some of the aquatic
macrophytes in its red list of species in danger
of extinction. Nevertheless, the number of taxa
which should be included in the list is greater
(see Cobo et al., 2002).
Moreover, the importance of this area for the
conservation of aquatic macrophytes is evident
since it contains numerous species of aquatic
macrophytes, many of which are limited-area
species (“endemic species”), relatively uncommon in aquatic plants.
Along the same lines, a conflict has arisen in
the area’s flora conservation-management, fully
affecting the group of aquatic plants: many of
the aquatic macrophyte species found in Doñana
cover wide areas of distribution (in theory);
however (in reality) these areas are largely fragmented, with similarly fragmented populations.
These species should be included in the red
lists, since their vulnerability is quite high, a
fact which has been ignored by environmentalists when it comes to the (theoretical) distribution areas of these species.
As studies are concerned on aquatic
macrophytes and how they relate to environmental factors, while there are a number of quality
studies on this subject, more work on the basic
aspects of this relationship would be of value
77
(e.g., how macrophytes relate to nutrients, factors determining macrophyte distribution, studies on succession, etc.). Finally, attention must
be called to the deterioration of Doñana’s waters,
the effects of which operate on two levels:
1. By causing the disappearance of so-called
“difficult environment” specialist plants that
had taken refuge in this natural setting (such
as plants from bogs or oligotrophic wetlands)
2. By facilitating the invasion of exotic species,
some of which have great potential for habitat
modification and its consequences.
A question remains: Is there still time for us
to comprehend the full complexity of the native
aquatic systems of Doñana or are the transformations detected in recent years the beginning
of an irreversible process that will profoundly
change this place?
ACKNOWLEDGMENTS
These studies were supported by the Project
Doñana 2005 from the Spain Environmental Ministry.
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Limnetica, 25(1-2): 81-94 (2006)
Effects of ultraviolet radiation on aquatic bryophytes
Javier Martínez-Abaigar*, Encarnación Núñez-Olivera, María Arróniz-Crespo,
Rafael Tomás, Nathalie Beaucourt, Saúl Otero
Universidad de La Rioja. Complejo Científico-Tecnológico, Madre de Dios 51, 26006 Logroño (La Rioja), Spain
*Corresponding author: [email protected]
ABSTRACT
The depletion of the stratospheric ozone layer as a result of anthropogenic activities increases the ultraviolet-B (UV-B) irradiance at ground level. This may lead to harmful biological consequences affecting photosynthetic organisms. Mountain streams are especially exposed to a UV-B increase, and bryophytes play a key ecological role in them. In this paper, the effects of
enhanced UV-B radiation on photosynthetic organisms in general and on bryophytes in particular are described. Hereafter,
some results obtained by our group on the effects of UV-B on bryophytes from mountain streams are presented. Laboratory
and field experiments show that these effects depend on the species, the environmental factors (such as temperature), and the
origin of the samples (sun or shade conditions, low or high altitude). Among the variables measured, the maximum quantum
yield of photosystem II (Fv/Fm) and the level of UV-absorbing compounds seem to be the most responsive to enhanced UV-B,
but no variable responded in the same manner in every species. The potential use of aquatic bryophytes as bio-indicators of
changes in ambient UV-B radiation would require an adequate selection of both variables and species. Promising variables are
Fv/Fm, the concentration of UV-absorbing compounds (especially if they are analyzed individually) and DNA damage, whereas the liverwort Jungermannia exsertifolia subsp. cordifolia has been revealed to be a good bio-indicator species. Globally, the
responses of aquatic bryophytes to UV-B radiation and their protecting systems are still poorly characterized, and thus further
study is required under both controlled and field conditions.
Keywords: aquatic bryophytes, mosses, liverworts, ultraviolet-B (UV-B) radiation, mountain streams, bio-indicators.
RESUMEN
La degradación antropogénica de la capa de ozono estratosférico provoca un aumento de la radiación ultravioleta-B (UV-B)
en la superficie de La Tierra. Esto puede causar consecuencias biológicas nocivas en los organismos fotosintéticos. Los arroyos de montaña están especialmente expuestos al aumento de UV-B, y los briófitos desempeñan un papel ecológico crucial en
estos ecosistemas. En el presente artículo, se describen los efectos de un aumento de radiación UV-B sobre los organismos
fotosintéticos en general y sobre los briófitos en particular. A continuación, se presentan algunos resultados obtenidos por
nuestro grupo de investigación sobre los efectos de la radiación UV-B en briófitos de arroyos de montaña. Los experimentos
realizados tanto en campo como en laboratorio muestran que dichos efectos dependen de la especie considerada, de los factores ambientales (como la temperatura) y de la procedencia de las muestras (aclimatadas a condiciones de sol o sombra, provenientes de baja o elevada altitud). Entre las variables analizadas, el rendimiento cuántico máximo del fotosistema II (Fv /Fm)
y el nivel de compuestos absorbentes de radiación UV parecen ser las que mejor responden a un aumento de UV-B, pero ninguna variable responde de la misma manera en todas las especies. El uso potencial de los briófitos acuáticos como bioindicadores de cambios en los niveles naturales de radiación UV-B requiere una selección adecuada tanto de las variables analizadas como de las especies empleadas. Fv /Fm y la concentración de compuestos absorbentes de radiación UV (en especial si
éstos son analizados individualmente), junto con los daños en el ADN, parecen ser las variables más prometedoras en este
campo, mientras que la hepática Jungermannia exsertifolia subsp. cordifolia podría resultar una buena especie bioindicadora.
Desde un punto de vista global, las respuestas de los briófitos acuáticos a la radiación UV-B, y los mecanismos protectores
que utilizan para hacerle frente, están todavía poco caracterizados, y en consecuencia se necesita una mayor investigación en
condiciones controladas y en campo.
Palabras clave: briófitos acuáticos, musgos, hepáticas, radiación ultravioleta-B (UV-B), arroyos de montaña, bioindicadores.
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Martínez-Abaigar et al.
ULTRAVIOLET RADIATION AND
ITS EFECTS ON PHOTOSYNTHETIC
ORGANISMS
Ultraviolet (UV) radiation induces many harmful effects in all living organisms, including
humans. UV-C radiation (<280 nm) is ecologically not relevant since it is absorbed by atmospheric oxygen and ozone. However, both UV-B
(280-315 nm) and UV-A (315-400 nm) penetrate to the biosphere and have significant biological effects, although only UV-B is absorbed by
the stratospheric ozone layer. Biological responses to UV radiation are highly dependent on
wavelength, and thus the biologically effective
UV (UVBE) can be calculated. UVBE encompass UV-A and UV-B, but, given the logarithmic increase of its effects with the decrease in
wavelength, it is often dominated by UV-B,
especially at its shorter wavelengths. Thus,
most studies on the effects of UV radiation
have dealt with UV-B.
UV-B irradiance at ground level depends
on a number of factors, such as latitude, season, hour of the day, altitude, presence of
clouds or aerosols, and surface reflectivity
(Björn, 1999). In addition, ozone depletion as
a result of anthropogenic emissions of halogenated carbon compounds leads to an increase in UV-B. In mid-latitudes, the ozone loss
has led to a 6 to 12 % increase in UV-B radiation above 1980 levels, and predicted changes
show the ozone layer will remain vulnerable
to further depletion in the near future
(McKenzie et al., 2003). Consequently, studies on the effects of ambient and elevated
UV-B irradiances are increasingly important.
In humans, an excessive exposure to UV-B
causes acute and chronic damage to eyes and
skin, including sunburn and cancer, and compromises the immune system (Vanicek et al.,
1999). In photosynthetic organisms, increased UV-B may cause diverse damage in the
photosynthetic apparatus: pigment degradation, photoinhibition, and decreases in quantum yield, photosynthetic rate, and the activity of the Calvin cycle enzymes (Jansen et
al., 1998). Also, DNA alterations, oxidative
damage, and changes in mineral absorption
can occur. This may lead to alterations in
growth and development. However, some
controversy about the ecological relevance of
these effects still persists (Fiscus & Booker,
1995; Allen et al., 1998; Searles et al.,
2001a). At the ecosystem level, UV-B can
affect decomposition, nutrient cycling, and
trophic interactions (Caldwell et al., 1998).
Photosynthetic organisms may develop a
number of protection and repair mechanisms
against the adverse effects of UV-B (Jansen
et al., 1998): production of UV-absorbing
compounds (flavonoids, phenyl-propanoids,
mycosporine-like aminoacids, etc.), antioxidant and photo-protective mechanisms, and
DNA-repairing processes.
Much of the research regarding the effects
of UV-B on photosynthetic organisms has
focused on terrestrial environments, especially
using crop plants, whereas aquatic ecosystems
have received less attention. The vast majority
of the studies concerning aquatic ecosystems
have dealt with marine phytoplankton and
macroalgae (Figueroa & Gómez, 2001; Day &
Neale, 2002; Häder et al., 2003; Helbling &
Zagarese, 2003), while the photosynthetic
organisms from freshwater ecosystems have
been less studied in line with their minor contribution to the global biomass and primary
production of aquatic systems. However,
rivers and lakes have an outstanding ecological importance as local systems and, because
of their lower depth compared to marine
systems, they are highly exposed to the harmful effects of UV-B radiation. In lakes, the
penetration of UV radiation and its effects on
phytoplankton have been the most studied
topics (see for instance Villafañe et al., 1999;
Huovinen & Goldman, 2000; Laurion et al.,
2000), but macrophytes have also been occasionally considered (Rae et al., 2001). In
rivers, scarce work has been done (Rader &
Belish, 1997a, 1997b; Kelly et al., 2003), probably due to intrinsic methodological problems derived from their strongly dynamic
environmental conditions (depth, discharge,
water velocity, chemistry, etc.).
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Ultraviolet radiation and aquatic bryophytes
ULTRAVIOLET RADIATION
AND BRYOPHYTES
To our knowledge, around 65 papers have been
published on the effects of UV-B radiation on
bryophytes, among which only 49 contain original data (Table 1). Half of these are strictly
bryological, whereas the remaining ones study
bryophytes together with other photosynthetic
organisms, such as vascular plants or lichens.
The research on this topic has focused mainly
on terrestrial and semi-aquatic bryophytes
from Antarctic habitats and circumpolar heathlands and peatlands. The most used species
belong to mosses: several Sphagnum species,
Hylocomium splendens (typical from forest soils), and Polytrichum commune (typical
from a wide range of acid habitats in damp to
wet situations). Liverworts have been notably
less studied than mosses, whereas no hornwort
has been investigated.
Diverse methodological approaches have
been applied. Studies have been conducted
under both field and controlled conditions, and
in the latter case both in the laboratory and in
the greenhouse. The manipulation of UV-B has
included the two main experimental options in
the context of UV-B research (Rousseaux et
al., 2004): exclusion experiments using filters,
and supplementation using lamps to simulate
ozone depletion. The duration of the experiments has been diverse, from a few hours of
UV-B exposure (usually under controlled conditions) to several years (under field conditions). The bryophyte responses have been
assessed using morphological and, especially,
physiological variables: colour, symptoms of
cell degradation, ultrastructural damage, sclerophylly, reproductive effort, growth (both in
length and dry mass), photosynthesis and respiration rates, chlorophyll fluorescence variables, photosynthetic pigment composition
(chlorophylls, carotenoids), DNA damage (presence of thymine dimers and other photoproducts), protein and glucid concentrations,
mineral elements content, and the appearance
of UV-absorbing compounds which could
serve as a protective mechanism.
83
The results obtained are conflicting, since UVB radiation has been found to either stimulate,
to depress, or to have no effect on bryophyte
performance. Several studies have found a
growth reduction in bryophytes when exposed
to UV-B (Sonesson et al., 1996; Gehrke et al.,
1996; Markham et al., 1998; Gehrke, 1998,
1999; Ballaré et al., 2001), but this effect seems
to depend on the species considered, the experimental design and other additional factors such
as water availability and CO2 concentration.
Other harmful effects (chlorophyll degradation,
reduction in photosynthesis rates and Fv/Fm)
are even less clear, since contradictory results
have been found. In addition, the increase in
UV-absorbing compounds, which represents the
most usual response of vascular plants to
enhanced UV-B (Searles et al., 2001a), has only
been manifested occasionally in bryophytes
(Markham et al., 1990; Ihle & Laasch, 1996;
Newsham et al., 2002; Martínez-Abaigar et al.,
2003a). Beneficial effects of UV-B radiation on
bryophyte growth have also been reported
(Johanson et al., 1995; Searles et al., 1999;
Phoenix et al., 2001), which further complicates the global interpretation of the results. This
controversy contrasts with intuitive thoughts
that bryophytes would be strongly sensitive to
UV-B radiation, because of their structural simplicity and the consequent lack of defenses
commonly found in higher plants: thick cuticles, epicuticular waxes, epidermis (sometimes
with several cell layers), hairs on leaf surfaces,
etc. It must be taken into account that bryophyte “leaves” are mostly mono-stratified and lack
air spaces, which dramatically reduces the
radiation pathway and thus its attenuation.
Thus, bryophytes (with the exception of thalloid forms, which have been understudied in
relation to UV radiation), could only acquire
chemical and metabolic defenses through, for
instance, UV-absorbing compounds, antioxidant mechanisms, and repairing systems of
DNA and photosynthetic machinery. However,
the present knowledge on UV-absorbing compounds in bryophytes suggests that this mechanism does not occur in most bryophytes
(Arróniz-Crespo et al., 2004), and the rest of
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the mechanisms have hardly been tested regarding UV-B radiation. Globally, the responses of
bryophytes to UV-B radiation and their protect-
ing systems are still poorly characterized, and
thus further study is required both under controlled and field conditions.
Table 1. Original papers on the effects of UV radiation on bryophytes. Key for “Used Species”: L, liverwort; M, moss. Key for “Ambient”:
T, terrestrial; P, peatlands; A, aquatic; R, rivers or streams; L, lakes. Key for “Type of Experiment”: F, Field; G, greenhouse; L, laboratory;
E, exclusion of UV-B radiation; S, supplement of UV-B radiation; N, samples exposed to natural levels of solar radiation; VSh, very short
duration (less than 1 day); Sh, short duration (1-30 days); M, medium duration (longer than one month and shorter than 6 months); Lo, long
duration (6 months - 1 year); VLo, very long duration (longer than 1 year); ?, undetermined duration; H, historical study (comparison of
samples over a prolonged period). Key for “Variables used”: A, alterations in DNA; Fl, chlorophyll fluorescence; FlS, fluorescence spectra;
G, growth; H, hydric relations; M, morphology; Mt1, primary metabolites (glucids, proteins, lipids); Mt2, secondary metabolites, including
UV-absorbing compounds; N, mineral nutrients; Ox, variables of oxidative stress (peroxide content, lipid peroxidation, ascorbate, superoxide dismutase, peroxidase, catalase); P, photosynthesis; Ph, phenology; PP, photosynthetic pigments; PS1 and PS2, activity of photosystems I
and II, respectively; R, respiration; Rf, reflectance indices; Sc, sclerophylly; U, ultrastructure. Artículos originales relacionados con los
efectos de la radiación UV sobre los briófitos. Clave para “Especies utilizadas”: L, hepática; M, musgo. Clave para “Ambiente”: T, terrestre; P, turberas; A, acuático; R, ríos o arroyos; L, lagos. Clave para “Tipo de experimento”: F, campo; G, invernadero; L, laboratorio; E,
exclusión de radiación UV-B; S, suplemento de radiación UV-B; N, muestras expuestas a niveles naturales de radiación solar; VSh, duración muy corta (menos de 1 día); Sh, duración corta (1-30 días); M, duración media (mayor de 1 mes y menor de 6 meses); Lo, duración
larga (6 meses - 1 año); VLo, duración muy larga (mayor de 1 año); ?, duración indeterminada; H, estudio histórico (comparación de
muestras a lo largo de un periodo prolongado). Clave para “Variables utilizadas”: A, alteraciones en el ADN; Fl, fluorescencia de clorofila; FlS, espectros de fluorescencia; G, crecimiento; H, relaciones hídricas; M, morfología; Mt1, metabolitos primarios (glúcidos, proteínas,
lípidos); Mt2, metabolitos secundarios, incluyendo compuestos absorbentes de UV; N, nutrientes minerales; Ox, variables de estrés oxidativo (contenido de peróxido, peroxidación de lípidos, ascorbato, superóxido dismutasa, peroxidasa, catalasa); P, fotosíntesis; Ph, fenología;
PP, pigmentos fotosintéticos; PS1 y PS2, actividad de los fotosistemas I y II, respectivamente; R, respiración; Rf, índices de reflectancia;
Sc, esclerofilia; U, ultrastructura.
Reference
Arróniz-Crespo et al.
(2004)
Ballaré et al. (2001)
Barsig et al. (1998)
Björn et al. (1998)
Conde-Álvarez et al.
(2002)
Csintalan et al. (2001)
Gehrke (1998)
Gehrke (1999)
Gehrke et al. (1996)
Huiskes et al. (1999)
Huiskes et al. (2001)
Huttunen et al. (1998)
Used species
Ambient
Type of
Variables used
experiment
Chiloscyphus polyanthos (L), Jungermannia exsertifolia
subsp. cordifolia (L), Marsupella sphacelata (L),
Scapania undulata (L), Brachythecium rivulare (M),
Bryum alpinum (M), Bryum pseudotriquetrum (M),
Fontinalis antipyretica (M), Palustriella commutata (M),
Philonotis seriata (M), Polytrichum commune (M),
Racomitrium aciculare (M), Rhynchostegium riparioides (M),
Sphagnum flexuosum (M)
Sphagnum magellanicum (M)
Polytrichum commune (M)
Aulacomnium turgidum (M), Dicranum elongatum (M),
Hylocomium splendens (M), Polytrichum commune (M),
P. hyperboreum (M), Sphagnum fuscum (M)
Riella helicophylla (L)
A (R)
F, N
Mt2, Sc
P
P
T, P
F, E, VLo
G, S, M
F, S, M-VLo
G, Mt2
Mt1, Mt2, PP, U
G, H
A (L)
L, E, VSh
Fl, Mt2, P, PP, R
Dicranum scoparium (M), Leucobryum glaucum (M),
Mnium hornum (M), Pellia epiphylla (L),
Plagiomnium undulatum (M), Plagiothecium undulatum (M),
Polytrichum formosum (M),
Sphagnum capillifolium (M), Tortula ruralis (M)
Sphagnum fuscum (M)
Hylocomium splendens (M), Polytrichum commune (M)
Hylocomium splendens (M), Sphagnum fuscum (M)
Sanionia uncinata (M)
Dicranum sp. (M), Hylocomium splendens (M),
Polytrichum commune (M)
T
L, S, Sh-M
Fl, FlS, Mt2
P
T, P
T, P
T
T
T, P
F, S, VLo
F, S, VLo
F, S, VLo
F, E, Sh
G, S, ?
G, M, Mt2, P, PP, R
G, M, Mt2, PP
G, H, Mt2, PP
Fl
M
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85
Reference
Huttunen et al. (2005a)
Huttunen et al. (2005b)
Ihle (1997)
Ihle & Laasch (1996)
Johanson et al. (1995)
Lewis Smith (1999)
Lovelock & Robinson
(2002)
Lud et al. (2002)
Used species
Dicranum scoparium (M), Funaria hygrometrica (M),
Hylocomium splendens (M), Pleurozium schreberi (M),
Polytrichum commune (M), Polytrichastrum alpinum (M),
Sphagnum angustifolium (M), S. capillifolium (M),
S. fuscum (M), S. warnstorfi (M)
Hylocomium splendens (M), Pleurozium schreberi (M)
Conocephalum conicum (L)
Conocephalum conicum (L)
Hylocomium splendens (M)
Bryum argenteum (M), Bryum pseudotriquetrum (M),
Ceratodon purpureus (M)
Bryum pseudotriquetrum (M), Ceratodon purpureus (M),
Grimmia antarctici (M)
Lud et al. (2003)
Markham et al. (1990)
Markham et al. (1998)
(2003a)
(2003b)
Montiel et al. (1999)
Newsham (2003)
Newsham et al. (2002)
Newsham et al. (2005)
Niemi et al. (2002a)
Bryum argenteum (M)
Marchantia polymorpha (L)
Jungermannia exsertifolia subsp. cordifolia (L),
Fontinalis antipyretica (M)
Andreaea regularis (M)
Sanionia uncinata (M), Cephaloziella varians (L)
Cephaloziella varians (L)
Sphagnum angustifolium (M), S. papillosum (M),
S. magellanicum (M)
Niemi et al. (2002b)
Sphagnum balticum (M), Sphagnum papillosum (M)
Núñez-Olivera et al.
(2004)
Núñez-Olivera et al.
(2005)
Phoenix et al. (2001)
Post & Vesk (1992)
Cephaloziella exiliflora (L)
Prasad et al. (2004)
Riccia sp. (L)
Rader & Belish (1997a) Fontinalis neomexicana (M)
Robinson et al. (2005) Grimmia antarctici (M)
Robson et al. (2003)
Robson et al. (2004)
Rozema et al. (2002)
Schipperges & Gehrke
(1996)
Searles et al. (1999)
Searles et al. (2001b)
Searles et al. (2002)
Sonesson et al. (1996)
Sonesson et al. (2002)
Taipale & Huttunen
(2002)
Takács et al. (1999)
Ambient
Type of
experiment
Variables used
T, P
N, H
M, Mt2
T
T
T
T
T
N, H
L, S, VSh
L, S, VSh-Sh
G, S, ?
F, E, M
M, Mt2
Mt1
Fl, Mt1, Mt2, P
G, Ph
G
T
F, N, ?
Mt2, PP, Rf
T
F, L, E, S, VSh-VLo
T
T
T
A (R)
A, G, Fl, M,
P, Mt2, PP
F, E, S, VSh-Sh A, Fl, Mt2, P, PP, R
N, H
Mt2
G, S, M
G, M, Mt2, Ph
L, S, M
Fl, Mt2, P, PP, R, Sc
A (R)
L, S, M
G, M
T
T
T
T
P
F, S, ?
F, N, M
F, N, Sh-M
F, N, E, M
F, S, M
Fl
Mt2, PP
Fl, Mt2, PP
Mt2, PP
G, Mt2, PP
P
A (R)
F, S, M
L, S, M
A (R)
L, S, Sh
T
T
T
A (R)
T
F, S, VLo
F, N, Sh
L, S, VSh
F, E-S, M
F, E, VLo
P
P
T
T, P
F, E, VLo
F, E, VLo
F, E, ?
F-L, S, M- VLo
G, Mt2, PP
Fl, G, Mt1, Mt2,
P, PP, R, Sc
Fl, Mt1, Mt2, P,
PP, R, Sc
G, H
M, Mt2, P, PP, U
Ox, PP, PS1, PS2
G
Fl, H, M, Mt2,
P, PP, Rf
G, M
G, M
G, Mt2
G, H, P
Dicranum elongatum (M), Sphagnum fuscum (M)
Hylocomium splendens (M), Pleurozium schreberi (M)
P
P
P
T
P
T
F, E, Lo
F, E, VLo
F, E, VLo
L, S, M
F, S, VLo
F, S, M
G, Mt2, PP
G, M, Mt2
G, M, Mt2, PP
G, P
G, H
Mt2
Dicranum scoparium (M), Leucobryum glaucum (M),
Mnium hornum (M), Pellia epiphylla (L),
Plagiothecium undulatum (M), Polytrichum formosum (M),
Tortula ruralis (M)
T
G, S, Sh-M
Fl
Tortula ruralis (M)
Hylocomium splendens (M), Sphagnum fuscum (M)
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ULTRAVIOLET RADIATION AND
AQUATIC BRYOPHYTES FROM
MOUNTAIN STREAMS
Before our studies, only two other ones had been
conducted on the effects of UV radiation on
truly aquatic bryophytes (see Table 1), despite
the interest of how diverse organisms from different habitats respond to this environmental factor. In Conde-Álvarez et al. (2002), samples of
the thalloid liverwort Riella helicophylla from a
saline lake were cultivated throughout a natural
daily light cycle under two radiation treatments:
solar radiation (UV + photosynthetically active
radiation or PAR) and solar radiation deprived of
UV (PAR treatment). There were significant differences between the two treatments in the
maximum quantum yield of photo-system II
(Fv/Fm), the effective quantum yield of photosynthetic energy conversion of PSII (ΦPSII), the
electron transport rate (ETR) and the initial
slope of ETR vs. irradiance curve (all higher in
PAR plants than in UV+PAR plants throughout
the day), photosynthetic capacity (higher in PAR
plants only at noon), chlorophyll a (lower in
UV+PAR only at 11.00), and phenolic compounds (higher in UV+PAR only at 13.30). No
differences between treatments were found in
dark respiration, photochemical quenching, and
carotenoid concentration, and only slight ones in
non-photochemical quenching (higher in
UV+PAR only in the morning). Thus, UV radiation (particularly UV-B) caused some damage to
the photosynthetic apparatus. Recovery of inhibited photosynthesis took place in the afternoon,
therefore solar UV radiation did not cause irreversible damage in the short term. Rader &
Belish (1997a) carried out a ten-week field
experiment in which samples of the moss
Fontinalis neomexicana were transplanted from
a reference site to both a shaded and an open
section of a mountain stream and were irradiated
with enhanced levels of UV-B radiation. The
transplants from the open site showed an important, although non-significant, reduction in dry
biomass with respect to those growing under
ambient conditions. However, the moss in this
experiment failed to grow in any site and under
any treatment condition, and there was a loss of
material in all samples from the beginning to the
conclusion of the experiment, which casts doubt
on the significance of the results.
For several reasons, we circumscribe our research interest to the effects of UV-B radiation on
aquatic bryophytes from mountain streams.
Firstly, these ecosystems might be particularly
exposed to the effects of UV-B radiation, since
1) the biologically active UV-B radiation increases between 5 % and 20 % per 1000 m altitudinal increase (Björn et al., 1998); 2) UV-B radiation can easily reach the organisms because they
live at relatively low depths or even emersed, and
UV-B radiation can penetrate into the oligotrophic waters typically occurring in mountain streams; and 3) the low temperatures which prevail
during most of the year may limit the development of protection and repairing mechanisms
against UV-B radiation. Secondly, bryophytes
are the most abundant primary producers in
mountain streams and are also important in
nutrient cycles and food webs (Bowden et al.,
1999; Núñez-Olivera et al., 2001). This domination suggests that they can withstand present
levels of UV-B radiation, but the underlying
physiological mechanisms are unknown and the
structural protections against UV-B are lacking
(as it was mentioned above for bryophytes as a
group). Thirdly, bryophytes have an outstanding
bio-indication ability in a number of pollution
processes and environmental changes (see a
recent review in Ah-Peng & Rausch de
Traubenberg, 2004), which could suggest their
potential use as bio-indicators of changes in UVB levels. And finally, the scarcity of studies
existing on this particular topic recommends
increasing our knowledge on it, especially considering the present social interest on the causes
and consequences of global climate change.
In our work, we have conducted both laboratory and field studies. In the laboratory, we have
cultivated bryophytes under enhanced UV-B
simulating a 20 % ozone depletion, with the aim
to characterize bryophyte responses to higher
than present UV-B levels. The applied biologically effective UV-B (UV-BBE) was 0.67 W m-2,
equivalent to an exposure of 9.6 kJ m-2 d-1,
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which was calculated using the generalized plant
damage action spectrum of Caldwell (1971). We
established three general radiation regimes
(PAR, PAR + UV-A, and PAR + UV-A + UV-B)
to distinguish the effects of UV-A and UV-B
radiations separately. Bryophyte responses were
analyzed in terms of sclerophylly, the photosynthetic pigment composition, the rates of net
photosynthesis and dark respiration, some variables of chlorophyll fluorescence, and the UVabsorbing compounds. In addition, we occasionally measured DNA damage (through the
appearance of thymine dimers), protein concentration, length growth and morphological
symptoms (both macro- and microscopic). The
duration of the experiments was diverse, from
3 days to 4 months. We have concentrated our
studies on two species, the moss Fontinalis
antipyretica and the foliose liverwort Jungermannia exsertifolia subsp. cordifolia (hereafter
J. cordifolia), which were always collected from
streams between 1300 and 2000 m altitude.
The two bryophytes mentioned above responded differently to the enhancement in UV-B
radiation under controlled conditions, while UVA radiation had a scarce biological effect
(Martínez-Abaigar et al., 2003a), as it occurred
in other experiments using bryophytes (Niemi et
al., 2002a, 2002b). The samples of the moss
which were irradiated with UV-B showed, with
respect to the control, decreases in the chlorophyll and carotenoid concentration, the chlorophyll a/b quotient, the chlorophylls/phaeopigments ratios, the net photosynthesis rates, the
light saturation point, Fv/Fm and ETR. They also
showed increases in the sclerophylly index
(“leaf ” mass per area) and the dark respiration
rates. The majority of these changes were indicative of plant stress and some of them had been
previously found in bryophytes exposed to
enhanced UV-B radiation. However, the UV-Birradiated samples of the liverwort only showed a
decrease in Fv/Fm, which might be the most sensitive physiological variable to UV-B, together
with a 20 % increase in the concentration of UVabsorbing compounds. This defense mechanism,
rarely described in bryophytes, would enable the
liverwort to have a higher tolerance than the
87
moss against UV-B radiation, at least under the
specific experimental conditions used. Also,
the increment of UV-absorbing compounds
in the liverwort might be a useful ecophysiological tool in the bioindication of UV-B.
The different response of the two studied species to UV-B radiation was also evident in their
morphological features (Martínez-Abaigar et
al., 2003b). When exposed to enhanced UV-B,
the moss showed brown colour, depressed
growth, development of the central fibrilar body
in the cells, chloroplast disappearance and presence of protoplasts progressively vesiculose,
vacuolized and finally hyaline. These symptoms
are little specific and have been described in
several pleurocarpous mosses as a response to
diverse processes of senescence and stress (both
natural and anthropogenic). The uniquely specific response of the moss to enhanced UV-B was
a colour change in the cell walls, from yellow to
orange-brown. In contrast, the exposed samples
of the liverwort looked healthy and their macroand microscopic appearances were quite similar
to those of control samples.
In the laboratory, we also examined the
influence of temperature (2 ºC vs. 10 ºC) on
the physiological and morphological responses
of Fontinalis antipyretica and Jungermannia cordifolia to enhanced UV-B (Martínez-Abaigar et
al., 2003b; Núñez-Olivera et al., 2004). The
influence of temperature on the effects of UV-B
radiation depended on the species: the higher the
UV-B tolerance, the lower the influence of temperature. Also, different morphological and
physiological variables showed varied responses
to this influence. Particularly, the lower temperature used in our study enhanced the adverse
effects of UV-B radiation on several important
physiological variables (Fv/Fm, growth and chlorophylls/phaeopigments ratios) in the UV-B-sensitive F. antipyretica, but not in the more UV-Btolerant J. cordifolia. Thus, the adverse effects of
cold and UV-B radiation were apparently additive in the moss (probably because the development of protection mechanisms was limited by
cold), whereas this additiveness was lacking in
the liverwort. We conducted a Principal Components Analysis (PCA) for both species using
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the physiological data obtained in the experiments and confirmed their different response to
the concomitant action of UV-B and cold.
Another environmental factor that can influence the response of different species to UV-B
radiation is their previous field acclimation to
sun or shade conditions (Núñez-Olivera et al.,
2005). Shade samples of Fontinalis antipyretica
were more sensitive to the UV-B treatment than
sun samples, and Fv/Fm was the physiological
variable which better discriminated both types of
samples, since it decreased 42 % in the shade
samples and only 27 % in the sun samples at the
end of the culture period. In Jungermannia cordifolia, controls and UV-B-treated samples were
not significantly different in either the sun or the
shade samples. PCAs for each species, ranking
the physiological results along the culture
period, strongly supported these points. In conclusion, the shade samples were more sensitive
to UV-B than the sun samples, but only in the
more UV-B-sensitive species.
We also determined that the sensitivity of
bryophyte species to artificially enhanced UV-B
could be tested without having to cultivate the
samples for a long period. A continuous UV-B
exposure of 78 h reproduced the differences in
the responses of Fontinalis antipyretica and
Jungermannia cordifolia, which had been previously found in longer experiments lasting 3682 days (Núñez-Olivera et al., 2005). For this
short-term test, several culture conditions,
which were known to accelerate the appearance
of damage, were imposed: high ratio UVB/PAR, continuous UV-B exposure, and cold
temperature. This type of fast test may therefore
be used instead of long-duration tests to evaluate the UV-B tolerance of bryophytes.
The different nature of the protection mechanisms between mosses and liverworts, which
had been previously pointed out for Fontinalis
antipyretica and Jungermannia cordifolia under
laboratory conditions, was tested in a field survey conducted for 14 aquatic bryophytes from
mountain streams, 10 mosses and 4 liverworts
(Arróniz-Crespo et al., 2004). The diverse species showed significantly different levels of
methanol-extractable UV-absorbing compounds
(MEUVAC) and also different forms in their
absorbance spectra in the UV band. The high
levels of MEUVAC and the clearly hump-shaped spectra in the UV-B and UV-A wavelengths
(280-400 nm), which were found in the liverworts, contrasted with the low levels and non
hump-shaped spectra generally found in the
mosses (except for Polytrichum commune).
Thus, the accumulation of MEUVAC might
represent a frequent and constitutive a protecting mechanism against UV-B radiation in liverworts, but not in mosses.
In another field experiment, we tested the
effects of a natural altitudinal gradient of ultraviolet-B (UV-B) radiation, from 1140 to 1816 m
altitude, on the physiology of 11 populations of
Jungermannia cordifolia (Arróniz-Crespo,
2005). Several physiological variables showed
significant linear relationships with altitude:
global MEUVAC levels, the concentrations of
two phenolic derivatives, ETRmax and NPQmax
increased with altitude, whereas photoinhibition
percentage and respiration rates decreased. This
was also confirmed by a PCA, since most of
these variables represented significant loading
factors ordinating populations by altitude. The
characteristics shown by high-altitude populations may confer tolerance to high UV-B levels,
and the specific response to UV-B of the two
phenolic derivatives suggests that they could be
used as indicators of the spatial changes in UVB radiation. In addition, the concentrations of
these two phenolic derivatives increased specifically under artificially enhanced UV-B radiation
(unpublished results). An interesting question
remains to be solved: could these compounds be
used as indicators of temporal changes in UV-B,
which could be related to ozone depletion?
CONCLUSIONS AND PERSPECTIVES
1. Our results demonstrate that the effects of
UV-B radiation on aquatic bryophytes depend
primarily on the species, and thus they do not
constitute a single functional group in this
respect. The different responses to UV-B are
revealed not only by changes in colour or in
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key physiological variables, such as growth,
chlorophyll concentration, photosynthesis
rates or chlorophyll fluorescence parameters,
but also by variables responsible for protecting mechanisms, such as the concentration of
UV-absorbing compounds. In particular, the
constitutive presence and/or inducible enhancement of UV-absorbing compounds depend
strongly on the species and, outstandingly, on
the type of bryophyte (moss or liverwort)
considered.
2. It is recommendable to evaluate UV-B sensitivity in sufficiently prolonged experiments,
however short-term (72 h) tests may render
comparable results.
3. The responses of aquatic bryophytes to UV-B
radiation depend not only on specific genetic
factors, but also on environmental factors (such
as temperature) and the origin of the samples
(sun or shade conditions, low or high altitude).
The effects of these factors depend on the species: in the UV-B sensitive ones, both cold and
previous shade acclimation may exacerbate the
harmful effects of enhanced UV-B.
4. Among the variables measured under laboratory conditions, the maximum quantum yield
of photo-system II (Fv/Fm) and the level of
UV-absorbing compounds seem to be the
most responsive ones to enhanced UV-B.
However, no variable responds in the same
manner in every species, which still limits
our global comprehension on the effects of
UV-B on bryophytes. The positive thing here
is that a “UV-B syndrome” may be identified
by the treatment of physiological data of control and UV-B-exposed samples through multivariant analyses (such as PCA), since both
types of samples usually appear clearly separated in the generated plots.
5. The noteworthy variability of the results
reported in the literature on the effects of UV
radiation on bryophytes may be due to the
above-mentioned diversity of species, environmental factors, variables and experimental
conditions used in the different studies. Thus,
it is necessary to take into account the methodological approaches to appropriately interpret the results obtained.
89
6. The use of aquatic bryophytes as bio-indicators of changes in UV-B radiation requires an
adequate selection of both variables and species. Promising variables regarding this point
are Fv/Fm, because of its sensitivity to UV-B,
and the concentration of UV-absorbing compounds, due to its remarkable specificity of
response. The analysis of individual UVabsorbing compounds may have a stronger
ecological and physiological relevance than
the usual global analysis of these compounds,
since each one may respond in a different
manner to UV-B. Thus, a previous identification of the compounds occurring in the different species is clearly needed. A third variable
potentially useful for bio-indication purposes
could be the evaluation of DNA damage caused specifically by UV-B radiation. Regarding the species selection, our results point at
Jungermannia cordifolia because of its good
responsiveness to UV-B, availability of
healthy biomass throughout the year (if populations are selected properly), and wide distribution range over mountain streams of the
northern hemisphere.
7. Two phenolic derivatives from Jungermannia
cordifolia which absorb UV radiation have
shown their ability to increase in response to
a field spatial gradient of UV-B. The future
combination of laboratory and field works
studying the behaviour of these compounds
under different UV treatments, and particularly the field assessment of their seasonal
and inter-annual variations in response to
temporal changes in ambient UV-B, could
allow for the development of a protocol of
bio-indication of the potential increase in UVB radiation due to ozone depletion.
ACKNOWLEDGEMENTS
We are grateful to the Ministerio de Educación
y Ciencia of Spain and the Fondo Europeo de
Desarrollo Regional (FEDER), and to the
Government of La Rioja (Consejería de
Educación, Cultura, Juventud y Deportes), for
their financial support through the Projects
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REN2002-03438/CLI, CGL2005-02663/BOS
and ACPI 2003/06. Saúl Otero and María
Arróniz-Crespo benefited from grants of the
Spanish Ministerio de Educación y Ciencia.
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Limnetica, 25(1-2): 95-106 (2006)
Ecological studies in Alto Guadalquivir wetlands: a first step towards
the application of conservation plans
Francisco Guerrero*; Gema Parra; Francisco Jiménez-Gómez; Carlos Salazar;
Raquel Jiménez-Melero; Andrea Galotti; Enrique García-Muñoz; Mª Lucía Lendínez
and Fernando Ortega
Departamento de Biología Animal, Biología Vegetal y Ecología. Facultad de Ciencias Experimentales.
Universidad de Jaén. Campus de las Lagunillas s/n. E-23071 Jaén
*Corresponding author: [email protected]
ABSTRACT
This paper reviews the most recent studies carried out in the Alto Guadalquivir wetlands. Data on wetland inventory and classification in typologies, faunal and floral community values are presented as well as the effects that agricultural pollutants have
on some aquatic species. These results support the need for a correct wetland policy that allows for the conservation of these
aquatic ecosystems.
Keywords: Natural wetlands, artificial wetlands, Alto Guadalquivir, agricultural impacts, conservation.
RESUMEN
Este trabajo revisa los estudios más recientes llevados a cabo en los humedales de la comarca del Alto Guadalquivir. Se presentan datos sobre el inventario de humedales y su clasificación en tipologías, los valores más relevantes de sus comunidades
de flora y fauna y los efectos que ejercen los contaminantes agrícolas sobre algunas especies. Estos resultados apoyan la
necesidad de una correcta política de humedales que permita la conservación de estos ecosistemas acuáticos.
Palabras clave: Humedales naturales, humedales artificiales, Alto Guadalquivir, impactos agrícolas, conservación.
INTRODUCTION
Wetlands are among the most important ecosystems on earth and constitute a major feature of
the landscape in almost all parts of the world
(Mitsch & Gosselink, 2000). In Mediterranean
landscapes, wetlands, as in other drylands, represent a broad variety of natural resources with
environmental, social, and scientific values and
play an important role in the maintenance of biological diversity (Williams, 1999). Despite their
great intrinsic values, these ecosystems have
been subjected to strong human influence
(Naveh & Lieberman, 1994), and are suffering
an accelerating process of degradation leading to
the disappearance of many of them (Hollis,
1995; Brinson & Álvarez, 2002). Casado and
Montes (1995) indicate that more than 60 % of
Spanish wetlands have disappeared in the last 50
years and, in spite of the recent acceptance of the
ecological importance of wetlands, this trend has
not yet been reversed (Amezaga et al., 2002).
Andalusia presents rich ecological wetland
diversity, including the rarest ecosystem in
Western Europe (Molina et al., 2002), which
represents one of the most important wetland
districts in Spain. Inside this group, little is
known about the extent and condition of the
wetlands in the Alto Guadalquivir region. Thus,
as a consequence of the lack of attention given
to wetlands and the scarcity of limnological
information, this paper reviews the recent studies carried out in the Alto Guadalquivir
wetlands, in order to encourage the expansion of
wetland policies beyond the identification and
conservation of individual wetland sites.
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STUDY AREA
The Alto Guadalquivir region comprises the
totality of the Jaén province and the eastern area
of the Córdoba province, in southeastern Spain.
This region covers the upper reaches of the
Guadalquivir River Basin and encompasses a
great environmental and physiographic diversity. The climate is Mediterranean, characterised
by hot dry summers and cold dry winters, with
mild humid springs and autumns. Another peculiarity is the existence of a high inter-annual
variation in precipitation, with periods of wet or
dry cycles. Agriculture, particularly olive plantations have been carried out for centuries in the
area and currently constitute the major land use
and the principal economic activity in the area.
WETLAND INVENTORY
AND CLASSIFICATION
Wetland inventory and classification in typologies are necessary prior to carrying out limnological studies, in order to collect further management-oriented information, which can provide
the basis for many important measures to conserve wetlands (Finlayson et al., 1999). The Alto
Guadalquivir has been considered a poor
wetlands region in the past, but recent studies
have demonstrated the existence of a large number of aquatic ecosystems. The first results obtained showed the existence of at least 400 wetlands
(see Ortega et al., 2003; Ortega et al., 2005;
Guerrero et al., unpublished data). The data from
this wetland inventory must be regarded as provi-
Figure 1. Four examples of wetlands in the Alto Guadalquivir region. A: Laguna Honda (natural wetland); B: Laguna de Brujuelo
(natural wetland); C: Chillar salina (artificial wetland) and D: Torrequebradilla irrigation pond (artificial wetland). Cuatro ejemplos de humedales de la comarca del Alto Guadalquivir. A: Laguna Honda (humedal natural); B: Laguna de Brujuelo (humedal
natural); C: Salina de Chillar (humedal artificial) y D: Balsa de Torrequebradilla (humedal artificial).
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Studies in Alto Guadalquivir wetlands
sional, but it represents the first study based on
field data, which provides a global view to the
Alto Guadalquivir wetland conservation.
An initial classification allows us to divide the
wetlands in two typologies: natural (around one
hundred of them) and artificial wetlands (Fig. 1).
Most of the natural wetlands are situated in endorheic basins and they are strongly dependent on the
hydrological budget, so limnological and ecological characteristics have a large seasonal and interannual variability. The most common natural
wetlands are temporary shallow lakes with a small
surface area, especially “lagunas” or steppe lakes.
Wetlands are not randomly distributed throughout
the Alto Guadalquivir region and are situated between 250 and 1600 m.a.s.l. Three main areas
(Campiña Norte, Campiña Sur, and La Loma),
close to the Guadalquivir River, can be distinguished according to the number of wetlands, representing up to 75 % of the total wetlands in the
region. The rest of the wetlands are distributed in
other six mountain areas (Sierra Morena, Sierra
Sur, Sierra Mágina, El Condado, Sierra de Cazorla,
and Sierra de Segura). In a second classification
using a genetic-functional criterion, we detect the
presence of other six wetlands categories. The predominant type in mountainous areas is karstic, and
in the countryside areas the principal type is associated to calcareous-marl soils (Ortega et al., 2003).
Artificial wetlands are principally represented in the area by inland salinas (solar evaporation salinas) and irrigation ponds, anthropogenic ecosystems, which present also important
scientific, social, and cultural values (we did not
include reservoirs neither man-made pool for
livestock use, which are abundant in Sierra
Morena). The inventory of inland salinas numbers nearly one hundred, ranging in size from
0.1 to 4.5 ha. They are situated between 250 and
1050 m.a.s.l., and principally on the right side
of the Guadalquivir River (Guerrero et al.,
2004), because the origin of inland salinas is
especially linked with Keuper facies areas.
Inland salinas are classified, based on morphological and geographical criteria, in two typologies (Quesada, 1996): inland mountain salinas
(situated normally above 600 m.a.s.l., with
small dimensions) and inland countryside sali-
97
nas (< 600 m.a.s.l., with a larger surface area
situated across the riversides).
Irrigation ponds are the most recent artificial
wetlands in Alto Guadalquivir and they have been
proliferating in recent years as a consequence of
the intensification of agricultural practices. Therefore, they are also more abundant in the areas
close to the Guadalquivir River (Campiña Norte,
Campiña Sur, and La Loma), in which olive agriculture predominate. We can distinguish two categories according to the existence of a sedimentation basin (Ortega et al., 2005); those containing
this basin present a richer diversity. They are permanent wetlands and represent an important refuge for the flora and fauna during dry periods,
when natural wetlands do not store water.
The most important human impacts affecting
the Alto Guadalquivir wetlands are agricultural
practices, including aquifer over-exploitation, tillage, drainage and dredging, and alteration of water
regime (Ortega et al., 2003). These activities together with the economic death of inland salinas lead
to a progressive abandonment and deterioration of
both natural and artificial wetlands. As a result,
many wetlands have disappeared and sometimes
only some unidentifiable debris on the ground serves as a witness to their presence in the past.
FLORA AND VEGETATION
Wetlands comprise a marginal belt of phreatophytes and/or helophytes and an open-water
area with different types of primary producers
(i.e. submerged macrophytes). We studied these
communities in both natural and artificial
wetlands (see Ortega et al., 2001; Ortega &
Guerrero, 2003; Ortega et al., 2004a; Salazar et
al., 2003) and the results obtained show the
existence of a high floristic diversity.
A recent study performed on 140 Andalusian
wetlands in 2004 (Ortega et al., unpublished
data) demonstrated that the Alto Guadalquivir
region is rich in species of submerged macrophytes, with the presence of 12 charophyte algae
and 23 aquatic plants: 1 bryophyte, 1 pteridophyte and 21 angiosperms, which represent
the 55 %, 50 %, 33 % and 70 % of the recorded
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Guerrero et al.
Figure 2. Macrophytes’ species richness (charophytes, angiosperms, bryophytes and pteridophytes) in the Andalusia (grey)
and Alto Guadalquivir (black) wetlands. Data from a study performed in 2004 (Ortega et al., unpublished data). Riqueza específica de macrófitos (carófitos, angiospermas, briófitos y pteridófitos) en humedales de Andalucía (gris) y de la comarca del
Alto Guadalquivir (negro). Datos obtenidos de un estudio realizado en 2004 (Ortega et al., datos no publicados).
species in this study respectively (Fig. 2). The
most significant submerged macrophyte species
are the charophytes Chara connivens, Ch. fragilis, Ch. aspera, Ch. vulgaris longibracteata,
Lamphrothamnium papulosum, Nitella flexilis
and Tolypella hispanica; the aquatic angiosperms Althenia orientalis, Miriophyllum alterniflorum, Najas marina, Potamogeton pectinatus, P. natans, P. pusillus, Ranunculus peltatus,
Ruppia drepanensis, R. maritima, Zannichellia
contorta, Z. obtusifolia, Z. pedunculata and Z.
peltata; and finally, the aquatic bryophyte Riella
helicophylla, and the pteridophyte Isoetes setaceum. From the list mentioned above, Althenia
orientalis and Zannichellia contorta are of
national and regional interest, since they are
included in the Red List of Spanish Vascular
Flora (VV.AA., 2000), the Atlas and Red Data
Book of Vascular Threatened Flora of Spain
(Bañares et al., 2003) and the Red List of
Vascular Flora of Andalusia (Cabezudo et al.,
2005) under the IUCN threat category of vulnerable (VU), and Isoetes setaceum is included in
the Red Data Book of Endangered Wild Flora of
Andalusia (Blanca et al., 2000) with the same
category. Furthermore, Zannichellia peltata and
Z. pedunculata have recently been catalogued in
the mentioned Andalusian Red List as vulnerable (VU), whilst Potamogeton natans, P. pusi-
llus, Ruppia drepanensis and R. maritima appear
as species with deficient data (DD). Finally,
these species are included in three natural habitat types of community interest, and one of them
is a priority type (coastal lagoons).
The studies on phreatophyte communities in
natural and artificial wetlands show the presence
of at least 175 taxa and 40 phytosociological
associations, 26 of which are included in Annex I
of Habitat Directive 92/43/EEC. We also detected
the existence of at least 12 habitats of community
interest, three of which are priority types (mediterranean salt steppes, mediterranean temporary ponds and inland salt meadows). The most
significant phytocoenoses are perennial helophytic communities (Typho-Schoenoplectetum glauci, Acrocladio cuspidati-Eleocharitetum palustris, Bolboschoenetum maritimi), ephemeral
helophytic vegetation (Preslio-Eryngietum corniculati, Damasonio alismae-Crypsietum aculeatae), halophilous riparian shrublands (Elymo
repentis-Tamaricetum canariensis), and a great
variety of halophilous pastures and grasslands: Polypogono maritimi-Hordeetum marini,
Parapholido incurvae-Frankenietum pulverulentae subas. spergularietosum tangerinae, Suaedo
splendentis-Salsoletum sodae, Polypogono maritimi-Centaurietum spicati, Suaedo splendentisSalicornietum patulae, Aeluropodo littoralisJuncetum subulati, including the remarkable
association Limonio quesadensis-Lygeetum sparti, endemic to the eastern Guadalquivir Basin
(García-Fuentes et al., 2001). It is also important
to note the presence of taxa such as Limonium
quesadense, an endemic species included in the
Red List of Spanish Vascular Flora, Atlas and
Red Data Book of Vascular Threatened Flora of
Spain and the Red List of Vascular Flora
of Andalusia as an endangered species (EN),
together with Puccinellia fasciculata, which is
considered as a species with deficient data (DD)
in the Red List of Spanish Vascular Flora.
PLANKTON COMMUNITY
Since knowing all wetlands is very difficult to
attain, the studies on the plankton community in
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natural wetlands were reduced to Laguna
Honda. This is a hypersaline ecosystem situated
in an endorheic basin (south-western Jaén),
whose origin is connected with karstic phenomena in gypsum and saline materials of the
Trias (Castro et al., 2003). This saline lake provides excellent conditions for ecological studies
because is characterised by fluctuations in water
level at different time scales. Thus, environmental factors play a major role in shaping the
plankton community, and consequently changes
in the biological communities are related to
physical and chemical characteristics. We outline the characteristics of the plankton community during two periods with different hydrological budgets and consequently different
behaviour of the system. The low precipitations
during a dry cycle (1994-95) caused the lake to
dry in summer, with an increment of salinity.
This is the principal factor controlling the dynamics of plankton community in these hypersaline ecosystems, with the present number of species tending to decline as total salt content
increases (García & Niell, 1993). As a consequence, plankton assemblages were very different between periods, with the dominance of
the halotolerant pico-nanoplankton species (Dunaliella viridis; Dunaliella salina; Tetraselmis
apiculata; Hantzchia amphioxys; Oscillatoria
lutea and the dinoflagellate Gymnodinium excavatum) during the dry period (Guerrero &
Castro, 1997; López-González et al., 1998), and
a considerable increase of phytoplankton species richness (Jiménez-Melero, unpublished
data) during the rainy cycle (1998-99 and 200102), especially diatoms (i.e. Fragilaria capucina, Navicula cocconeiformis and Nitzschia
reversa). The harpacticoid benthonic copepod
Cletocamptus retrogressus together with turbellarian species and the ciliate Fabrea salina,
were the most important zooplankton species
during dry cycles (López-González et al.,
1998). Zooplankton richness increased in the
wet period, with new species of a more planktonic behaviour that were not collected during the
previous cycle (Daphnia mediterranea; Moina
salina; Alona sp.; Cyclops sp.; Arctodiaptomus
salinus and Hexarthra fennica) (Castro, 2004;
99
Jiménez-Melero, unpublished data). As we can
see, the plankton community responds with
usual changes to a salinity gradient occurring
during a hydrological cycle and during a large
inter-annual time scale. During rainy years,
the plankton community ranges from limnogenic α-hypersaline to β-hypersaline type, whereas as the salinity increases the community changes towards a community similar to that of the
γ-type and δ-type waters (sensu Por, 1980).
The plankton community of artificial
wetlands (inland salinas) is fundamentally
represented by pico-nanoplanktonic species and
by the anostracean crustacean Artemia sp.,
which are mainly present in the water accumulation basins of these ecosystems. In order to
measure biomass, abundance, and the functional composition of the pico-nanoplankton community, scarcely studied in this type of ecosystems, we have applied automated analyses
techniques such as flow cytometry and microscopic image analysis. These techniques allow
for a fast and objective characterization of heterotrophic and phototrophic components of this
biota, and permit the link between single cell
properties and community organization (Platt,
1989). In a study of about thirty inland salinas,
the phytoplankton community showed densities ranging between 1 and 1000 cells/ml,
mostly small flagellates and by benthic diatoms, such as Nitzschia spp., Amphora ovalis
and Cymbella spp. Heterotrophic bacteria reached a maximum of 2.4 10 7 cells/ml and a
minimum of 2.2 10 6 cells/ml, ten/fold more than typical marine concentrations and similar to those of freshwater systems. Episodic
blooms of Dunaliella spp. (up to 1500 cells/ml)
were occasionally detected under summer conditions. The flow cytometry analysis showed
two bacterial populations characterized by the
DNA content (low and high DNA populations).
Similar results have been related to the existence of two functional groups and to different
metabolic activity (Gasol et al., 2000). The size
abundance spectrum shows a non-linear behaviour characterized by discontinuities between
bacteria and phytoplankton community (Fig. 3)
(Galotti et al., in press).
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Figure 3. Size abundance spectrum of the pico and nanoplankton community of two Alto Guadalquivir salinas (San
Carlos and Los Vélez) (modified from Galotti et al., in press).
Espectro de abundancia de la comunidad de pico y nanoplancton en dos salinas del Alto Guadalquivir (San Carlos y
Los Vélez) (modificado de Galotti et al., en prensa).
OTHER FAUNA GROUPS
Invertebrates: Coleoptera
The invertebrate communities of hypersaline environments are of interest for fauna studies. The
most interesting species is the endemic hydraenidae coleopteran Ochthebius glaber (Montes &
Soler, 1988), which coexists in the studied area
with other congeneric species such as O. notabilis,
O. delgadoi, O. corrugatus, O. quadrifossulatus,
O. andalusiacus and O. dentifer (Andrés Millán,
personal communication, 2002). O. glaber is
usually associated with hypersaline waters in the
arid lands southeast of the Iberian Peninsula in
the provinces of Córdoba, Jaén, Murcia and
Albacete (Castro, 1997; Montes & Soler, 1998;
Millán et al., 2002). We analysed the presence of
this species in 35 inland salinas of the Alto
Guadalquivir region and the results showed that O.
glaber was present in 80 % of them, with maximum abundances during the summer-autumn
period. Therefore, the conservation of wetlands
and more specifically of inland salinas has special
relevance in the protection of this endemic beetle.
Waterbirds
The Alto Guadalquivir wetlands are important
sites for migratory waterfowl and shorebirds
because it is located in a region where water is
scarce in summer. We studied waterbirds over
a period of eight years and registered a total of
71 species (see Ortega et al., 2004b; Ortega et
al., 2005). The most common species were
Tachybaptus ruficollis, Ardea cinerea, Anas
platyrhynchos, Anas strepera, Aythia ferina,
Gallinula chloropus and Fulica atra. The data
obtained shows a total of 18 breeding, 17 wintering, 25 passing, and 11 resident bird species, which mainly use natural wetland and
irrigation ponds. From these, twenty-seven
species are included in the Red Data Book of
Spanish Vertebrate (Blanco & González,
1992) and twenty-nine in the Red Data Book
of Endangered Vertebrates of Andalusia
(Franco & Rodríguez de los Santos, 2001).
From the latter, 5 are catalogued as critically
endangered species (CR: Ardeola ralloides,
Marmaronetta angustirostris, Aythia nyroca,
Fulica cristata and Chlidonias niger), 3 as
endangered (EN), 7 as vulnerable (VU), and
14 species as near threatened (NT) or with
deficient data (DD). Lastly, a total of 25 species are also included in Annex I of the Bird
Directive 79/409/EEC. It is important to note
the presence of the species Fulica cristata,
which constitute the first site record in the
region, and was part of a recent introduction
programme in this area of Andalusia. This project permits the introduction of an incipient
colony during the breeding season in the Alto
Guadalquivir region, which has lead to the
local reproduction of this species with a total
of 40 and 34 flying chicks in 2004 and 2005,
respectively (Ortega et al., unpublished data).
Finally, we also analysed the movements of
greater flamingos Phoenicopterus ruber in
order to discover its dispersal routes along the
Alto Guadalquivir wetlands. Greater flamingos
move among wetlands and therefore, the connectivity between the breeding site and foraging
sites may be an important determinant of
wetland use (Amat et al., 2005). Our results
using readings of bird rings seem to confirm
that this species use the Alto Guadalquivir
wetlands in their foraging and dispersal movements (Ortega, unpublished data).
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IMPACT ACTIVITIES: AGRICULTURAL
PRACTICES
Regular procedures in agriculture generate
unquestionable environmental impacts (Vitousek
et al., 1997; Guerrero et al., 2002, 2005). Some
of these impacts are related to a decrease in water
quality, owing to the use of toxic products such as
herbicides, fungicides or pesticides, which induce changes in ecological characteristics. In recent
years, ecotoxicological studies have substantially
increased in order to measure the ecological
effect of agricultural pollution on aquatic ecosystems (Van Dam et al., 1998). The results obtained
could be useful in catchment management plans
in areas where a specific pesticide is used.
Therefore, an important area of interest in research is how intensive agriculture practices affect
the ecological functions and values of the Alto
Guadalquivir wetlands, especially on the tolerance response of different species to pesticides.
Ecotoxicological test, used to predict levels of
contaminants that will cause minimal harm to the
aquatic environment, have been carried out in the
laboratory using aquatic species representative
of these wetland ecosystems. The toxicological
tests have been developed not only to assess
the lethal effect but also the sublethal effects
that could represent a more realistic view of the
consequences of pesticide use. The chosen species were Arctodiaptomus salinus (Copepoda:
Calanoida) and the amphibians Bufo calamita
and Rana perezi (Amphibia: Anura).
The copepod experiments were carried out
with adult females and egg sacs, while the
amphibian experiments were performed with
eggs and tadpoles. Two different toxic substances extensively used in olive cultivation, dimetoate and copper, were used in the toxicological
tests. The results obtained in the toxicological
tests show that adult females, nauplii, egg sacs
of the calanoid copepod, and amphibian larvae
were all negatively affected by exposition to
copper and dimetoate.
The lethal concentration obtained for adult
copepods (24-h LC50) was rather lower than the
regular dose of pesticide used in olive agriculture,
with values between 0.82-1.37 mg Cu/l and 3.30-
101
4.15 mg dimetoate /l. These results also reflect the
negative effect on Arctodiaptomus salinus secondary production as a consequence of an increase
in females and nauplii mortality and by a reduction in hatching rate (Parra et al., 2005).
Amphibians have been widely advocated as
excellent biological indicators of environmental health, because they are particularly vulnerable to environmental change (Wake, 1991).
The results obtained show that amphibian
populations are sensitive to pesticides during
embryonic and larval development in aquatic
habitats. Copper exposure affects growth with
an increase in development time and mortality;
therefore the use of this pesticide in agriculture
has also a negative effect on the maintenance
of amphibian populations in wetlands. The
LC50 values obtained after 96-h were between
0.17-0.32 and 0.34-0.39 mg Cu/l for Bufo calamita and Rana perezi, respectively (GarcíaMuñoz, unpublished data). We also detected
important epithelia damages in both species,
which probably affect osmotic equilibrium,
energy expenditure, and lead to a decrease in
the activity of the immune system. A comparative sensitivity test between both species performed at 0.3 mg Cu/l, showed a different tolerance level, with a survival percentage at 96-h
of 80 % and 23 % in Rana perezi and Bufo
calamita, respectively. This implies a different
resistance to copper, Rana perezi being more
tolerant to pollution. Therefore, the legal restrictions in copper use must take into account
the most sensitive species in order to protect a
wider range of species in the community.
CONSERVATION PLANS
The demand for increased agricultural production in the Alto Guadalquivir region implies an
intensive overexploitation on the water resources,
which in combination with other factors has been
the major cause of the loss of wetlands. This
situation could be reversed and in this sense, the
Andalusian government is operating a programme of wetland management and conservation
(Andalusian Wetlands Plan). The plan defines the
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environmental policy on wetland matters in order
to conserve the integrity of the ecosystem, promoting its use and preserving, now and for the
future, their ecological, socio-economical, and
historical-cultural functions (VV.AA., 2002).
Although a total of 147 wetlands are inventoried
in this plan, only 9 wetlands of the Alto
Guadalquivir region are included. Furthermore,
in the framework of preservation in Andalusia,
only four of these wetlands are currently protected areas (Laguna Honda, Chinche, Conde and
Grande), and the first three are also included on
the Ramsar List of Wetlands of International
Importance. Despite this lack of attention, these
ecosystems are significant components of the
Andalusian natural environment, and urgent conservation priorities are necessary.
One of the first steps essential for the protection and conservation of wetlands is the recognition of their values. The results obtained by our
research group in relation to the high ornithological and floral community values, and the use of
the Habitat Directive on the presence of rare species, allow us to propose a list of another nineteen wetlands. Together with the ones mentioned
above, these areas must be included as Important
Bird Areas (I.B.A.), and Sites of Community
Importance (S.C.I.), and constitute the proposal
of the wetland network “Lagunas del Alto
Guadalquivir” (Ortega et al., 2004b). The preservation of this wetland complex along the
Guadalquivir River is very important at the regional level, not only for the population of migratory
waterbirds, but also to allow for the biotic connection among wetlands, which contributes to the
migration of other aquatic species and to the
maintenance of local aquatic biodiversity.
We also have analysed the spatial relationships between wetlands and the public network
of livestock paths in the province of Jaén
(Madero et al., 2004). Our focus was to assess
possible benef its for wetland conservation
that might arise from the introduction of environmental criteria in defining wetland restoration priorities within the program that the
Andalusian government is carrying out for
the recovery of an effective public ownership
of the livestock paths.
In addition, we have evaluated the land use in the
catchment areas of natural wetlands. The results
show that agriculture is the most important land
use in the region (Ortega et al., in press). We
recognize that there is a growing conflict between
agricultural development and conservation in
Spain, especially in the Alto Guadalquivir region,
as a consequence of the intense growth of olive
cultivation. A crucial aspect of this problem is the
adverse effect on aquatic communities of the
generalised use of pesticides, fungicides, and
herbicides in wetland catchment areas that can
enter the aquatic ecosystem via point or nonpoint sources. Consequently, laboratory toxicity
tests are necessary to permit the regulation and
enforcement of the legislation, which takes into
account the most sensitive species. However,
tolerance response tests cannot be easily used as
tools for the evaluation of the state of a wetland.
Thus, it is necessary to look for other methods
that can provide quick and efficient information
relating to the effects of pesticides on these ecosystems. The use of biomarkers that reflect subcellular or cellular responses can give us this
information. These studies are of special interest
in applied limnological research because the
identification of biomarkers could be used to
assess the toxic effects on wetland biota, and
applied to the management of the ecosystem. Our
future goals will be focused on the study of these
biomarkers and the application of ecological
models based on individuals (Grim & Railsback,
2005), in order to determine the relationships
between individual traits and system dynamics in
order to understand the ecosystems and be able to
protect them right now. Only if agricultural development is balanced by environmental considerations can an effective restoration of the majority
of wetlands in this region be successful.
ACKNOWLEDGEMENTS
This work was supported by the Comisión
Interministerial de Ciencia y Tecnología (CICYT
Projects PB98-0307 and REN2001-3441C0201).
We would like to thank Adel el Amran for his
help with phytoplankton identification.
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Limnetica, 25(1-2): 107-122 (2006)
Groundwater-mediated limnology in Spain
Miguel Álvarez-Cobelas
Instituto de Recursos Naturales, CSIC, Serrano 115 dpdo., 28006 Madrid, Spain, [email protected]
ABSTRACT
Despite the obvious relationship between groundwater and surface waters, only a few studies have indirectly addressed the
effects of groundwater on Spanish limnosystems, although there are many suggesting that such effects may be important. I present here an overview reporting such effects, which affect water balance, motions of lake layers, underwater light climate, conservative hydrochemisty and nutrients, and community structure and dynamics, some of which are time-delayed. Future
Spanish studies on groundwater-mediated limnology will be fostered if regionally oriented hydrogeologists meet locally oriented limnologists and they exchange their knowledge and perform joint research. These efforts will result in updating and
expanding the ideas of regional limnology that Naumann and Margalef have promoted.
Keywords: water renewal, Mediterranean limnosystems, regional hydrogeology, landscape, regional limnology
RESUMEN
A pesar de la relación obvia entre las aguas subterráneas y las superficiales, sólo algunos estudios han considerado indirectamente los efectos que las primeras pueden tener sobre los limnosistemas españoles, aunque haya numerosas sugerencias de
que tales efectos puedan ser importantes. Aquí presento una breve revisión de dichos efectos, los cuales afectan al balance
hídrico, a los movimientos de las capas lacustres, al ambiente luminoso subacuático, a la hidroquímica conservativa y los
nutrientes y a la estructura y dinámica de las comunidades biológicas, y algunos de los cuales presentan desfases temporales.
En el futuro, los estudios de limnología relacionada con las aguas subterráneas serán más interesantes cuando los hidrogeólogos de enfoque regional se mezclen con los limnólogos de enfoque local y ambos grupos intercambien conocimientos y realicen investigaciones conjuntas, lo cual dará como resultado la actualización y expansión de las ideas de la limnología regional
que Naumann y Margalef emitieran en su día.
Palabras clave: renovación del agua, limnosistemas mediterráneos, hidrogeología regional, paisaje, limnología regional
RESUM
A pesar de la relació òbvia entre les aigües subterrànies i les superficials, tan sols alguns estudis han considerat indirectament
els efectes que les primeres poden tenir sobre els limnosistemes espanyols, encara que existeixen nombrosos suggeriments que
indiquen que aquests efectes poden ser importants. Aquí presento una breu revisió d’aquests efectes, els quals afecten al
balanç hídric, als moviments de les capes lacustres, a l’ambient lluminós subaquàtic, a la hidroquímica conservativa i als
nutrients i a l’estructura i dinàmica de les comunitats biològiques, i alguns dels quals presenten desfasaments temporals. En el
futur, els estudis de limnologia relacionada amb les aigües subterrànies es veuran impulsats si els hidrogeòlegs d’enfocament
regional es barregen amb els limnòlegs d’enfocament local i ambdós grups s’intercanvien coneixements i realitzen investigacions conjuntes. Això tindrà com a resultat l’actualització i expansió de les idees de la limnologia regional que Naumann i
Margalef varen emetre el seu dia.
Paraules clau: renovació de l’aigua, limnosistemes mediterrànis, hidrogeologia regional, limnologia regional.
Colc, Senyor Déu, la Figura Carnal,
Els cels fecunds que il.lustren l’Oceà,
Els rius subtils eixits d’un ull llunyà,
El Pic, la Vall i el Pla; l’Ordre Cabdal
J.V. FOIX (1936)
MARGALEF (at the 13th SIL Congress in Finland): Les régions calcaires à érosion karstique sont très développées en Espagne... Je distingue en Espagne deux types principaux de lacs permanents dans ces régions, qu’on
pourrait définir respectivement par la provenance superficielle ou profonde des eaux... (in Stankoviç, 1958)
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M. Álvarez-Cobelas
INTRODUCTION
Noel Hynes’ (1975) seminal paper, which was
published after his conference at the 19th SIL
Congress, held in Canada, certainly opened a
new research field and a new scale of observation for limnological topics, which became more
fruitful some 20 years later with the advent of
GIS methodologies (Johnson & Gage, 1997).
Hynes disclosed basin effects on in-stream ecological dynamics, and hence an astonishing view
of terrestrial landscape influences on aquatic
environments was fostered.
Thomas Winter’s efforts have not been as successful as Hynes’. His attempts to promote the
importance of groundwater in limnology, in
spite of being sustained over time, have not
reached a receptive and widespread limnological audience, as deserved by the topic. Anyway,
he is to be credited for his efforts (Winter, 1981,
1988, 1995, 1999, 2003) to convince limnologists of the paramount significance of groundwater for freshwater ecology. This is particularly
true in Mediterranean environments because the
seasonal rainfall usually recharges aquifers that
then experience delayed discharge during dry
periods, thus supplying water to many limnosystems that would be dry otherwise (ÁlvarezCobelas et al., 2005a).
Spain has 370 aquifers (Table 1), spreading
over some 167,000 km2 (roughly 30 % of overall
peninsular area) in all basins. The effects of
groundwater on limnosystems concern the quantity and quality of freshwater, which in turn affect
most limnological features. One hundred and one
wetlands bigger than 10 Ha are acknowledged to
be influenced by groundwater in Spain (Table 2).
Despite these figures, only a few recent studies
report the important role that groundwater plays
in Spanish limnosystems (Table 3), and most of
that recognition is in passing only. Obviously, this
poor outcome arises from the almost non-existent
relationships between Spanish ecologists and
hydrogeologists, that pioneering studies by
González Bernáldez (1992a) attempted to foster,
but whose efforts were stopped by his untimely
death. However, the intimacy between surface and
groundwater in Spain has long been recognized in
popular culture, a proof of which is the high number of topographic names related to groundwater
upwelling throughout Spain (http://pci204.cindoc.csic.es/tesauros/Toponimo/Toponimo.htm)
Figure 1. Long-term flooding in Las Tablas de Daimiel wetland, a National Park since 1973. Since 1986 groundwater supply was
discontinued due to aquifer overexploitation. Later, flooding was partly artificially induced by water transference from the nearby
Tajo basin, though high rainfall and groundwater discharge from a nearby aquifer (Campo de Montiel) have been responsible for
the 1997-1998 peaks. Data source: Tablas de Daimiel National Park and Álvarez-Cobelas (unpublished data). Inundación a largo
plazo de Las Tablas de Daimiel, que desde 1973 es Parque Nacional. Desde 1986 no hay aportes de aguas subterráneas debido a
la sobreexplotación del acuífero. Después, la inundación se ha conseguido en parte de manera artificial, mediante trasvase desde
la cuenca del Tajo, si bien las elevadas pluviosidad y descarga desde un acuífero cercano (el del Campo de Montiel) han sido responsables de los máximos en 1997 y 1998. Fuente: Parque Nacional Tablas de Daimiel y Álvarez-Cobelas (datos inéditos).
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Table 1. Overall area of aquifers in Spanish hydrographic basins.
Data source: Martín Pantoja et al. (1994). Extensión de los acuíferos en las cuencas hidrográficas españolas. Fuente: Martín Pantoja
et al. (1994).
Catchment/
Cuenca
Northern
Duero
Tajo
Guadiana
Guadalquivir
Guadalete-Barbate
Southern
Segura
Júcar
Ebro
Inner Catalonia
Balearic Islands
Total
Number of aquifers/ Permeable areas (km2)/
Nº acuíferos
Zonas permeables
24
21
12
13
48
12
46
32
52
45
30
35
370
7,009
53,623
15,961
11,960
13,811
1,486
3,138
8,603
24,782
16,770
6,463
3,618
167,224
and the richness of limnosystem names in
Spanish (González-Bernáldez, 1992b) and Mediterranean areas (Alvarez-Cobelas et al., 2005a).
Groundwater effects on water quantity act
upon stream discharge and water balance in
lakes and wetlands. Sometimes, groundwater
inflows may even trigger motions of water and
solutes within limnobasins. Also, groundwater
supply may affect emergent plant structure and
dynamics in wetlands. Aquifer water quality
impinges on the chemistry of limnosystems,
either conservative or not, thus indirectly changing ionic composition and nutrient concentrations that shape inland water metabolism, sensu
Golterman (1975), and biota.
This overview will briefly describe the studies on the interaction between groundwater and
freshwater systems in Spain to end up with
some prospects for the discipline. Hyporheic
effects, however, will not be dealt with here.
WATER BALANCE EFFECTS
These are the effects having paramount importance for limnosystems. They have often been
addressed, albeit without frequent quantification.
Lake and wetland levels and stream discharges
are increased by groundwater inputs in many
areas of Spain, those inputs being the sole ones in
109
some seasons. Thus groundwater enhances water
availability and renewal for many Spanish limnosystems and enables rich biota to thrive.
The oldest known process of groundwater
supply to a Spanish limnosystem is perhaps that
of the Upper Guadiana River (HernándezPacheco, 1932), which relied on a very old idea
(Pliny the Elder, 1st century AC; 1998 edition) of
an underground stream passage. Upper
Guadiana has been believed to infiltrate North of
Lagunas de Ruidera and upwell in Ojos del
Guadiana, some 50 km west. This idea, of strong
metaphoric power over time (any person, fact or
process appearing and disappearing in Spain is
thought to be like the Guadiana River in modern
speaking and the mass media), was scientifically
discredited 30 years ago (Torrens et al., 1976),
but it is still receiving strong affection, maybe
because it is cited in Cervantes’ Don Quixote
(2nd part, chapter XXIII). Anyway, long-range
groundwater has upwelled in a large spring,
called Ojos del Guadiana (“Guadiana Eyes”),
until 1986 when such an upwelling has been over
because of aquifer overexploitation for irrigation
purposes (Álvarez-Cobelas & Cirujano, 1996).
The significance of the groundwater supply at
the Ojos del Guadiana spring for the nearby
Tablas de Daimiel wetland has been very high. A
roughly constant volume of water has emerged
for decades, flooding an area that in the fifties
extended over more than 200 km2 (ÁlvarezCobelas & Cirujano, 1996). For many years, the
wetland experienced a seasonal pattern of flooding (Fig. 1) that began changing in the seventies,
when man-made activities strongly impacted the
wetland, a National Park by then. That groundwater of low ionic strength has merged with
intermittent stream water of higher salinity to
provide a mixed environment, very suitable for a
high species richness biota (Álvarez-Cobelas &
Cirujano, 1996). In 1986 the spring no longer
upwelled, and flooding of this wetland has lacked a predictable pattern since then (Fig. 1).
Other wetlands where groundwater effects on
water availability have been reported are those of
Duero basin (Ávila-Valladolid; Rey-Benayas,
1991), La Safor (Valencia; Rodrigo et al., 2001)
and L’Empordá (Girona; Quintana, 2002).
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M. Álvarez-Cobelas
Figure 2. Groundwater percentage of water inputs (subaquatic and shoreline springs) to the Colgada lake (Ruidera lake complex,
Central Spain) in 2003 and 2004. Inputs of subaquatic springs comprise 47-89 % of overall groundwater entering the lake.
Unpublished data. Porcentaje de los aportes hídricos de origen subterráneo (fuentes subacuáticas y manantiales litorales) a la
laguna Colgada (complejo lacustre de Ruidera, Centro de España) en 2003 y 2004. Las entradas subacuáticas representan un 4789 % del total de las aguas subterráneas que entran al lago. Datos inéditos.
Groundwater flows into lakes were recognized
more recently. Margalef reported groundwater
delivery in the Banyoles Lake in the same presentation of Stankoviç (1958), the evidence dating
back to the early 20th century (Vidal-Pardal,
1960). In the eighties the study of groundwater
upwelling in Banyoles started, as related to the
sediment resuspension process (Roget &
Casamitjana, 1986). Entering mostly through the
southern basins of the lake, groundwater inflows
ranged 1.6-5.1•104 m3 d-1, being 4-fold of surface
inflows (table 2 of Casamitjana & Roget, 1993)
and lacking strong seasonality.
The Ruidera Lakes, comprised of 18 flowthrough, seepage basins, have also been known to
experience groundwater supply for long, but its
measurement has started very recently (CEDEX,
1997). More than a half of overall water supply
received by the Colgada Lake (one of the largest
basins of that lake complex) is of groundwater
origin (Fig. 2), amounting to 7-14•104 m3 d-1 in
2003-2004. A slight seasonality is obvious in the
groundwater supply of Colgada Lake in that picture, taken in years of high water availability,
but such a seasonality wanes as rainfall
becomes lower than the annual average (450 mm;
Álvarez-Cobelas, unpublished data). These lakes
are strongly dependent on the interplay among
rainfall, surface- and groundwater, which
–when sufficient- enhance lake connectivity,
albeit showing delays of up to nine months from
strong rainfall to surface connection of all lakes
(Álvarez-Cobelas, unpublished data). As a result,
their water renewal time happens to be highly
variable (Table 4), and this has overwhelming
effects on limnological features of Ruidera
lakes (Álvarez-Cobelas et al., submitted).
Since groundwater regional flow is a very
complex phenomenon (Tóth, 1963), it is not
surprising that some lakes experience sudden
net groundwater inflow, irrespective to rainfall
seasonality. This has been observed in the
Table 2. Wetlands larger than 10 Ha connected to aquifers in Spain.
Data source: Martín Pantoja et al. (1994). Humedales de más de
10 Ha conectados a acuíferos en España. Fuente: Martín Pantoja
et al. (1994).
Catchment
Northern
Duero
Tajo
Guadiana
Guadalquivir
Guadalete-Barbate
Southern
Segura
Júcar
Ebro
Inner Catalonia
Balearic Islands
Total
Number of
wetlands
4
16
2
35
6
7
1
11
12
7
101
Overall wetland
area (Ha)
77
574
24
2.068
125
1.923
10
2.458
6.745
370
14.374
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111
Table 3. Spanish limnosystems for which groundwater effects have been reported since the birth of the journal Limnetica. Limnosistemas
españoles para los cuales se han referido efectos de las aguas subterráneas desde el comienzo de la publicación de Limnetica.
Site/Lugar
Major catchment/
Cuenca Hidrográfica
Banyoles lake
Inner Catalonia
Jarama lakes
Tajo
Ruidera lakes
Doñana sandy ponds
and marshes
Duero sandy wetlands
SW Madrid wetlands
L’Empordá wetlands
(“aiguamolls”)
La Safor wetland
Tablas de Daimiel wetland
Chícamo river
Sea-side intermittent
streams (“ramblas”)
Albufera de Valencia wells
Pyrenean springs
Guadiana
Guadalquivir
Reference/Referencia
Duero
Tajo
Inner Catalonia
Casamitjana & Roget (1993), Colomer et al. (2001),
García-Gil et al. (1996), Roget & Casamitjana (1987),
Serra et al. (2002, 2005)
Álvarez-Cobelas et al. (1999, 2005c), CEDEX (2001),
Domínguez (2002), Himi (2001), Roblas & García-Avilés (1999)
Álvarez-Cobelas et al. (2006)
de Castro & Muñoz-Reinoso (1997), Muñoz-Reinoso (1995, 1996),
Sacks et al. (1992), Suso & Llamas (1993)
Rey-Benayas (1991), Rey-Benayas et al. (1990)
González-Besteiro (1992)
Quintana (2002), Quintana et al. (1998)
Júcar
Guadiana
Segura
Segura
Rodrigo et al. (2001, 2003)
Álvarez-Cobelas et al. (2001)
Vidal-Abarca et al. (2000)
Moreno et al. (1995)
Júcar
Ebro
Alonso & Miracle (1987), Miracle et al. (1995), Soria (1993)
Roca (1990), Roca & Baltanás (1993), Roca & Gil (1992),
Roca et al. (1992), Sabater & Roca (1990, 1992)
Campillos lake complex (Málaga; Benavente &
Rodríguez, 1997). Local, intermediate, and
regional flows are also responsible for patterns
of inundation of sandy ponds in Doñana
National Park (Sacks et al., 1992; MuñozReinoso, 1996) and in SW ponds of the Madrid
aquifer (González-Besteiro, 1992).
Southeastern intermittent and permanent streams may also have groundwater discharge and
recharge (Moreno et al., 1995; Vidal-Abarca et
al., 2000), in some places at the scale of tenths
of metres (Vidal-Abarca et al., 2000). The
Upper Guadiana River and its tributaries in the
Lagunas de Ruidera Natural Park receive discharged groundwater, but they have not been
computed as yet. It is very likely that many
other Spanish streams are fed by groundwater,
but the SAIH (acronym for the Spanish Network
for River Discharge Measurement) has not been
designed to study that process.
OTHER EFFECTS ON LIMNOPHYSICS
Groundwater entering the Banyoles Lake (Girona)
has been shown to resuspend sediments, also
enhancing lutocline motions upwards up to 20 m
in periods of average groundwater inflow (Casamitjana & Roget, 1993). Since that groundwater
has a nearly constant temperature (17-19 ºC)
throughout the year, a hydrothermal plume is often
developed, also entraining cold hypolimnion water
into the base of the seasonal thermocline of the
lake. The plume spreads laterally at the level of
neutral buoyancy, thus behaving as a horizontal
baroclinic intrusion (Colomer et al., 2001). This
plume experiences a strong seasonality because it
is controlled by thermocline dynamics; i.e. it is
constricted to the hypolimnion during stratifying
periods but it spreads throughout the whole water
column when Autumn thermal circulation is fully
established (Serra et al., 2002). A similar phenomenon has been found in Las Madres lake
(Madrid), with groundwater inflow of roughly
constant temperature (13-14 ºC), promoting a lutocline of a much shorter behaviour (ÁlvarezCobelas, unpublished data), that delays complete
mixing of the whole water column for more than a
month (Álvarez-Cobelas et al., 2005c). Double
diffusion processes, associated with groundwater
inflows, have also been reported for both lakes
(Serra et al., 2005; Álvarez-Cobelas et al., 2005c).
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M. Álvarez-Cobelas
Figure 3. Daily total phosphorus concentration per unit of surface present during 93 consecutive days at the 1991 transition of late
stratification-to-early mixing in Las Madres lake (Central Spain). Unpublished data. Concentración diaria del fósforo total por
unidad de superficie presente durante 93 días consecutivos de la transición de la estratificación tardía al inicio de la mezcla de
1991 en la laguna de Las Madres (Centro de España). Datos inéditos.
EFFECTS ON LIMNOCHEMISTRY
Conservative hydrochemistry is one of the main
targets of the study of springs. Roca (1990) has
undertaken a thorough study on 207 Pyrenean
springs in Huesca and Lleida, attempting a
typologic synthesis on a wide range of limnological features. Two complex factors are responsible for most of the observed hydrochemical
variability: solubilization plus water residence
time (42 % explanation of overall variance) and
geochemical substrate (12 %), but there is a
group of springs, seemingly heavily dependent
on the hypogean environment, where expected
hydrochemical patterns do not hold. Also, seaside intermittent streams of Murcia may have
groundwater inflows upstream, comprised of
magnesium chloride salts (Moreno et al., 1995).
Springs surrounding Albufera de Valencia
lake have high ionic strength and nitrate content,
but are of oligo-mesotrophic nature, thus reflecting the low concentrations of phosphorus and
ammonia in the groundwater they come from
(Soria, 1993); this result is striking because in
recent decades the lake has become hypertrophic
as a result of sustained wastewater inputs over
time (Romo et al., 2005), but springs partially
feeding it are not. That oligo-mesotrophic status
is also common in springs of La Safor wetland
(Valencia), which also drain to hypertrophic
areas of the marsh (Rodrigo et al., 2001).
Groundwater conductivity and the length of
groundwater flow are linearly correlated
with each other in the vicinity of sandy wetlands of the Duero basin (r2 = 0.65 p < 0.01;
Rey-Benayas et al., 1990), that relationship
being of interest for plant community structure
in those wetlands (see below).
In seepage lakes located in alluvial plains,
groundwater flows may result in depleting inlake areal phosphorus concentration during
autumn circulation (Fig. 3). The lake setting in
the plain, viz. being close to the river or not, may
change seston sedimentation patterns (Fig. 4). In
Campillo lake, a hypertrophic seepage lake located very close to Jarama river, phosphorus sediTable 4. Water renewal time (in years) of Ruidera lakes. Data source: Álvarez-Cobelas et al. (in press and unpublished). Tiempo de
renovación del agua (en años) de las lagunas de Ruidera. Fuente:
Álvarez-Cobelas et al. (en prensa y datos inéditos).
Lake
Conceja
Tomilla
Tinaja
San Pedra
Lengua
Santos Morcillo
Colgada
Rey
Cueva Morenilla
Year
2000
2001
2004
9.14
10.30
11.19
12.38
14.47
32.50
262.12
111.21
130.10
4.08
4.60
5.00
3.91
2.00
3.96
270.31
114.69
132.56
0.10
0.12
0.11
0.15
0.14
0.12
0.11
0.13
0.14
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mentation rates are not much higher (0.047 ±
0.018 mg P m-2 d-1) than those in Las Madres
lake (0.034 ± 0.020 mg P m-2 d-1), a mesotrophic
seepage lake away from riverine influence. It is
well known that river Jarama in the area of
Campillo lake acts recharging the aquifer and the
lake (Himi, 2001); so riverine-driven groundwater inflow to Campillo lake might reduce phosphorus sedimentation. Total seston sedimentation
has been shown to be high close to the hydrother-
113
mal plume of Banyoles lake, reaching values up
to 25 g m-2 d-1 (Serra et al., 2005). Such a silt
deposition can be observed embedded with biocalcarenites in littoral sediments of that lake.
Groundwater flows appear to be the main
cause of the differential confinement of nutrients in L’Empordá wetland (Girona; Quintana
et al., 1998), whereby nitrate is easily washed
out but phosphorus accumulates, thus increasing eutrophication of this wetland.
Figure 4. Chlorophyll-a in the mixed column and sedimentation rate of total phosphorus in two Central Spanish lakes. Both lakes lie on
an alluvial plain, but Las Madres lake is a mesotrophic seepage lake complex, solely influenced by the alluvial aquifer, whereas
Campillo lake is a hypertrophic seepage lake, influenced by groundwater strongly-dependent upon a heavily polluted stream (Jarama
river). Scale for chlorophyll for the Campillo lake is the one on the right. Groundwater movement in the vicinity of the lakes is also
shown in the lower pictures (Himi, 2001). Unpublished data. Clorofila “a” en la capa de mezcla y tasa de sedimentación del fósforo
total en dos lagos del Centro de España, situados en una llanura aluvial. Las Madres es un conjunto de cubetas mesotróficas, influidas
únicamente por el acuífero, mientras que El Campillo es un lago hipertrófico, afectado por aguas subterráneas muy dependientes de las
entradas de un río muy contaminado (el río Jarama). La escala de la clorofila para el lago de El Campillo es la de la derecha. También
se representa, en las figuras inferiores, el flujo del agua subterránea en las proximidades de los lagos (Himi, 2001). Datos inéditos.
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114
M. Álvarez-Cobelas
EFFECTS ON THE LIMNOBIOTA
Obviously, they can be more clearly seen in
environments of groundwater upwelling, say,
springs. In addition to the plethora of
Margalef ’s biological observations in the forties
and fifties in North and Central Spain (for
example see his 1946, 1948, 1949, 1950, 1952
and 1955 papers), there have been some studies
on the biota of springs that address this topic
more recently. In his Ph.D. thesis, Roca has
carried out a thorough study of Pyrenean
springs (see Roca, 1990) that has enabled him
and his coworkers to outline the main effects of
groundwater discharge on spring assemblages
of diatoms, turbellarians, mayflies, water mites
and ostracods. Ionic content and current velocity appear to be the dominant factors affecting
distribution of the 198 diatom species recorded
in these springs (table 2 of Sabater & Roca,
1990), where peculiar environmental factors,
such as dim light, high temperature and salinity,
could be responsible for the 40 % variability of
diatom distribution left unexplained by the two
most important factors. Roca et al. (1992)
recorded four species of turbellarians in some
50 springs, suggesting that interspecific competition explain the presence of single species in
every spring, water temperature and current
velocity being the most important abiotic factors shaping the distribution of the observed
species. 22 taxa of hydracnellae have been
found by Roca & Gil (table 1 of their 1992
paper), with substrate rocks and water renewal
as the main factors explaining their distribution.
21 ostracod species have been reported in the
same survey (table 1 of Roca & Baltanás, 1993),
but high mineralization and low flow enhance
ostracod species richness in those springs.
Barquín & Death (2004) have studied the
contrasting faunas of some springs and nearby
rivers in Cantabria (N Spain), showing a switch
between species richness and density: invertebrate density is higher in springs, but species
richness is higher in streams. Predation effects
by the amphipod Echinogammarus strongly
shape communities in springs. Chlorophyll-a, as
a surrogate for periphyton biomass, is much
higher in springs than in streams and it is the
factor showing the highest explanatory power,
albeit weak, of the faunal patterns observed.
Phytoplankton biomass, as ascertained by
chlorophyll-a, also experiences the contrasting
effect of the position of lakes in Madrid alluvial
plains. Seepage lakes located close to nutrientrich rivers develop higher phytoplankton biomass
(47.9 ± 24.6 µg Chl-a L-1 in 1995) than seepage
lakes (2.5 ± 1.2 µg Chl-a L-1) farther away from
such rivers (Fig. 4), despite the fact that neither is
fed by riverine water via surface flow. In the
Figure 5. Chlorophyll-a concentration in La Safor wetland in
September 2000. White circles represent areas of lower chlorophyll-a, which increases along with the gray scale of the circles. Data from Rodrigo et al. (2001). Concentración de la
clorofila “a” en el humedal de La Safor en Septiembre de
2000. Los círculos blancos representan las zonas de menor
concentración, la cual se incrementa siguiendo la escala de
grises de los círculos. Datos de Rodrigo et al. (2001)
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115
Table 5. Indicator species organisms on recharge and discharge wetlands, from the Madrid aquifer; discharge wetlands are fed by groundwater.
Data source: González-Besteiro (1992). Especies de organismos indicadores en humedales de recarga y descarga del acuífero de Madrid; los
humedales de descarga se alimentan con aguas subterráneas. Fuente: González-Besteiro (1992).
In recharge wetlands
Cyzicus grubei
Chirocephalus sp.
Triops cancriformis
Alona azorica
Moina brachyata
Mixodiaptomus incrassatus
In discharge wetlands
Local flow
Regional flow
(alkaline chemistry)
Regional flow
(mixed chemistry)
Ranunculus sp.
Daphnia obtusa
Aeschna mixta
Ischnura pumilio
Lestes virens
Chara vulgaris
Lymnaea peregra
Lymnaea truncatula
Chydorus sphaericus
Simocephalus vetulus
Ruppia drepanensis
Alona salina
Arctodiaptomus salinus
Arctodiaptomus wierzejskii
Cletocamptus retrogressus
“ullals” (freshwater springs) of La Safor wetland,
Rodrigo et al. (2001) found lower chlorophyll-a
concentrations when light was not limiting
(September) than in the remaining wetland
(Fig. 5). The ecological stability of those springs
is higher than that of other areas of the wetland
(Rodrigo et al., 2001). Groundwater upwellings
importantly contribute to spatial heterogeneity
that drives the species richness recorded in that
wetland (128 phytoplankton and 126 zooplankton species; Rodrigo et al., 2003), which is higher than that in other wetlands, such as La Safor,
experiencing strong polluting inputs.
“Ullals” surrounding Albufera de Valencia
lake can be considered as refuges for rare species of zooplankton. Alonso & Miracle (1987)
recorded some interesting rotifers and crustaceans in those springs, such as the copepods Eucylops graetieri, Microcylops rubellus
major and Horsiella brevicornis, the cladoceran Dunhevedia crassa, and the rotifers
Asplachnopus multiceps, A. hyalinus, and
Euchlanis dapidula. The species richness of
rotifers in Albufera springs is very remarkable,
because another study (Miracle et al., 1995)
recorded 107 species (see their table 1), including many taxa only known to occur in very
restricted geographical areas (Lecane pides,
Erignata saggitoides, Dicranophorus hercules,
Paradicranophorus hudsoni).
Crustacean, dragonfly, molluscan, and plant
species indicating distinct flows of groundwater
have also been recorded in Central Spain by
González-Besteiro (1992), whose recorded biological species enabled her to distinguish between recharge and discharge wetlands, the latter
being fed by local or regional (either alkaline or
mixed) groundwater flows (Table 5).
González-Besteiro’s results are but one outcome of a research line relating biological species and groundwater, which has been explored
by the González-Bernáldez school, also providing evidences of that relationship in sandy
wetlands of Central and Southwestern Spain.
Long-range groundwater flow (i.e. regional)
results in the development of halophytic species, such as Juncus subulatus, Limonium costae, and Suaeda vera, that thrive in sulphate and
chloride-rich Duero wetlands; local flows promote community compositions rich in
glycophytic (Mentha suaveolens, Juncus inflexus, Poa trivialis) and alkalinophytic plants
(Festuca arundinacea, Juncus acutus, J.
gerardi) (Rey-Benayas et al., 1990). Furthermore, the depth of the hydraulic head and
groundwater chemistry interact to produce four
different plant communities in discharge
wetlands of the Doñana National Park: 1st) higher mineralized discharges (conductivity >
1500 µS cm-1) are related with Tamarix canariensis and Juncus acutus; 2nd) mid-mineralized
discharges (conductivity ≈ 500 µS cm-1) promote Scirpus holoschoenus and Juncus maritimus
growth; 3rd) mid-mineralized, chloride-rich discharges are indicated by Juncus acutiflorus and
J. effusus; and 4th) lower mineralized inflows
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M. Álvarez-Cobelas
Figure 6. Annual cover (vertical bars, left-hand scale) of the
two main species of emergent macrophytes in Las Tablas de
Daimiel National Park and average flooding area in the previous
year (circles, right-hand scale) during the period 1945-2002.
Groundwater supply ended in 1986 because of aquifer overexploitation. Data published in Álvarez-Cobelas et al. (2005b).
Cobertura anual (barras verticales, escala de la izquierda) de
las dos principales especies de macrófitos emergentes en el
Parque Nacional Tablas de Daimiel y superficie inundada promedio en el año anterior (círculos, escala de la derecha) durante el periodo 1945-2002. Dejó de haber aportes de aguas subterráneas en 1986 debido a la sobreexplotación del acuífero.
Datos publicados en Álvarez-Cobelas et al. (2005b).
(conductivity < 300 µS cm-1) result in two distinct plant communities mediated by the depth
of the water table (Muñoz-Reinoso, 1995).
The outstanding, albeit delayed, effects of
groundwater discharge into surface wetlands
could be supported by the counterexample of
Las Tablas de Daimiel wetland. When the
underlying aquifer contributed waters for wetland flooding, the dominant emergent vegetation
was comprised by the cut-sedge Cladium mariscus, whereas this plant dominance was substitut-
ed by the reed Phragmites australis once that
groundwater effect was over (Fig. 6; ÁlvarezCobelas et al., 2005b). Groundwater flooding
provided longer hydroperiod and less fluctuating
hydrolevels for many areas of the wetland that
partly benefited Cladium. Later, the cover extent
of that plant greatly diminished, being substituted by Phragmites, much better adapted to the
higher water fluctuations that arose when
groundwater supply was discontinued as a result
of aquifer overexploitation by growing irrigation
(Álvarez-Cobelas et al., 2001).
Bacteria and fish are also suspected to suffer
from indirect effects of groundwater inflows.
García-Gil et al. (1996) suggest that the differential development of suspended layers pushed
upward in different basins of Banyoles Lake by
groundwater inflows, change underwater light
climate, thus affecting seasonal dynamics of
autotrophic sulphur bacteria in each basin.
Serra et al. (2002) report that the hydrothermal
plume may reduce the vertical habitat suitable
for fish, that are restricted to areas above
the plume, because suspended silt impairs the
light climate that those fish (Perca and Rutilus)
need to feed on their preys.
FUTURE PROSPECTS
The preceding pages have attempted to outline
many studies dealing with the effects of groundwater on Spanish limnosystems. Despite its fragmentation, those studies demonstrate that
groundwater happens to act upon many different
features of aquatic environments, but results are
very preliminary as yet. That groundwater and
the solutes they transport impinge on surface
waters is an obvious outcome of the water cycle.
What is not as obvious is the strong tie that
groundwater may develop with surface waters in
Mediterranean environments (Álvarez-Cobelas et
al., 2005a), where rainfall seasonality and frequent unevenness pose very strict limits to surface water availability, but whose limits can be
partly circumvented on account of groundwater
inflows into limnosystems. This could be another
facet of the supply-side ecology that Margalef
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(1980: 191) advocated when he talked about
“external energy”. Turbulent flow is of course
external or “exosomatic”, as he has later coined,
but solutes and materials transported by groundwater can also enhance (or depress) biological
productivity in freshwaters (i.e. they are energycontrollers), groundwater itself supplying a suitable environment when surface water is depleted.
Anyway, all these contentions are preliminary
enough to satisfy our scientific thirst. In its present state most Spanish limnologists work on a
very local basis (a given lake, reservoir, stream
or wetland), whereas most Spanish hydrogeologists work at a regional scale (a more or less
large aquifer), and both scales hardly match.
However, most Mediterranean limnosystems are
very small, their basins having a paramount
influence on their ecology (ÁAlvarez-Cobelas
et al., 2005a). Spanish limnological studies on
basin effects, that obviously include groundwater effects as I have already shown, are very
scarce. Also, groundwater effects happen to be
dictated by the hydraulic head field in the
surroundings of the studied limnosystem, often
implying the interplay of regional and local
groundwater flows (Winter, 1999). It is then
time both, for Spanish limnologists to gain a
landscape approach and for Spanish hydrogeologists to gain a local focus. Another drawback
for improving such a relationship is the expensiveness of some hydrogeological techniques. In
order to study the groundwater field surrounding a given lake or wetland, it is necessary to
have nested piezometers around it, but these
devices exist very infrequently and are expensive to build. In sandy basins they can be profitably substituted by hand-operated piezometers
that are reasonably cheaper (Winter et al.,
1988), but in calcareous basins environmental
managers (because many valuable limnosystems
occur in environmentally-protected areas) must
deal with the expenses needed to build piezometers in numbers enough to cover the whole local
area to be surveyed.
The superposition of regional and local flow
systems associated with surface water bodies
results in complex interactions between groundwater and surface water and climate in all land-
117
scapes, often including effects of limnosystem
topographic setting. Such a complexity promotes very different patterns of water and solute
transport into and out of the limnosystems,
impinging on water renewal and exchange of
materials. Those processes have been documented in glacial, dune, coastal, karst, and alluvial
environments of U.S.A. (Winter, 1999), but
none in Spain, despite some obvious similarities
of local climate and geological substrate.
Among the many features resulting from the
connection of limnetic and groundwater environments in Spain that have not been explored
as yet, are the responses of limnosystems to
transient conditions of local and regional
groundwater flow. Such as the effects of transpiration of emergent vegetation on the transport of
groundwater solutes into playa lakes; the effects
of changing aquifer geometry (3-D values of
transmissivity and permeability) on limnosystem functioning; the effects of short, middle
and long-term (including those of climate change) variability of rainfall in aquifer recharge
from and discharge into limnosystems; the
effects of man-made impacts on aquifer quantity and quality of water that may affect limnosystems later; and many more. Some of them
are specifically known from other places
(Winter, 1999; Winter & Rosenberry, 1995;
Webster et al., 1996; lter.limnology.wisc.edu);
others have even been modeled (Nield et al.,
1994; Smith & Townley, 2002), but most of this
knowledge has been achieved in cold temperate
areas of the world, quite different from Spanish
ones. It is clear that the advancement of groundwater-mediated limnology in Spain should rely
on the match between landscape-oriented limnology and site-oriented hydrogeology that must
result in updating and expanding Naumann
(1932) and Margalef (1951) ideas of regional
limnology. There is a world ahead to be gained.
ACKNOWLEDGEMENTS
The life and work of Ramón Margalef has
always been an inspiration to me. His very
talented scientific endeavors and pursuits in very
Limnetica 25(1-2)02
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M. Álvarez-Cobelas
hard times for the ever-lasting illiterate Spain
have always been very encouraging and, not
having been one of his students, I owe my dedication to limnology to one of his papers published in Jano: Medicina y Humanidades (a
Spanish Medicine Journal) sometime in the
seventies. This overview is an outcome of my
interaction with some Spanish botanists, limnologists and hydrogeologists over time, such as
Santos Cirujano (Real Jardín Botánico de
Madrid, CSIC), Carmen Rojo and María Antonia
Rodrigo (Instituto Cavanilles de Biodiversidad,
Univ. Valencia), Youssef Himi (Spanish Ministry
of Home Affairs) and Esperanza Montero (Dept.
Geodinámica, Univ. Complutense de Madrid),
and it is a pleasure to acknowledge them the
countless hours of talking about anything, even
groundwater-mediated limnology. Juan Soria
(Confederación Hidrográfica del Júcar), Javier
Vidal (Univ. Girona), Carmen Rojo, María
Antonia Rodrigo and José Barquín (Univ.
Cantabria) have sent me papers and information.
Carmen Rojo read carefully a preliminary draft
of the manuscript. And last but not least, María
Antonia Rodrigo translated the abstract into
Catalan, as an homage to Ramón Margalef native language.
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Limnetica, 25(1-2): 123-134 (2006)
A functional approach to the ecology of Atlantic Basque streams
Arturo Elosegi1, Ana Basaguren2 & Jesús Pozo3
Department of Plant Biology and Ecology, Faculty of Science and Technology, University of the Basque
Country, POBox 644, 48080 Bilbao, Spain
e-mail: [email protected], [email protected], [email protected]
ABSTRACT
Most research on stream ecology is focused on structural characteristics of stream ecosystems, while less effort is being dedicated to the study of their functional attributes. The laboratory of stream ecology, at the University of the Basque Country, has
been researching streams from an ecosystem perspective, including both structural and functional properties. Here we review
the research done so far. Basque streams running to the Atlantic Ocean are short, steep, and flashy, and tend to show large spatial variations as a result of changes in geology and soil use. Where they exist, riparian forests limit the growth of primary producers and are a source of coarse particulate organic matter, an important food resource for consumers. The trophic structure
of benthic macroinvertebrate communities changes downstream with resource abundance, although temporal variations of
both resources and consumers are not coupled. Gut content analyses show the diet of some species to change with instar development, and thus, warn against the indiscriminate assignation of trophic categories. Floods are important disturbances, affecting primary producers, consumers, and ecosystem processes. Other important disturbances are changes in riparian vegetation,
which can profoundly affect the food resources of stream communities. This basic knowledge has been used to develop new
tools to assess stream functional impairment, based in two pivotal functions: litter breakdown and whole stream metabolism.
Both eutrophication and changes in riparian vegetation affect the use of leaf litter, and thus impact stream functioning making
litter breakdown a promising tool for stream monitoring. On the other hand, whole stream metabolism is affected by many
human impacts, and can be measured continuously in modern gauging stations, thus allowing for almost real-time monitoring
of ecosystem functioning. We hope these and other functional properties will be built into routine monitoring schemes, which
will therefore look at both the structural and functional sides of stream ecosystems.
Keywords: Basque streams, structural and functional properties, leaf breakdown and metabolism processes.
RESUMEN
La mayor parte de los trabajos sobre ecología de ríos se ha centrado en características estructurales, y se ha dedicado menos
esfuerzo a estudiar los atributos funcionales de los ecosistemas fluviales. El laboratorio de ecología de ríos, en la Universidad
del País Vasco, ha estado estudiando arroyos desde una perspectiva ecosistémica, incluyendo propiedades estructurales y funcionales, y mostramos aquí una revisión de los trabajos llevados a cabo. Los ríos vascos que drenan al Océano Atlántico son cortos,
de fuerte pendiente, y torrenciales, y tienden a mostrar fuertes variaciones espaciales a consecuencia de cambios en geología y
usos del suelo. Cuando existen, los bosques riparios limitan la capacidad de crecimiento de los productores primarios, y son
fuente de materia orgánica particulada gruesa, un importante recurso alimentario para los consumidores. La estructura trófica
de las comunidades de macroinvertebrados bentónicos cambia aguas abajo con la abundancia de recursos, aunque las variaciones temporales de recursos y consumidores no están sincronizadas. El análisis de contenidos digestivos muestra que la dieta de
algunas especies cambia con el grado de desarrollo larvario, y en consecuencia, previene contra la asignación indiscriminada
de categorías tróficas. Las riadas son perturbaciones importantes que afectan tanto a productores primarios como a consumidores, o a procesos a nivel de ecosistema. Otras perturbaciones importantes son los cambios en la vegetación riparia, que pueden
afectar profundamente los recursos alimentarios de las comunidades fluviales. Este conocimiento básico se ha utilizado para
desarrollar nuevas herramientas de evaluación del estado de los ríos, basadas en dos funciones claves: la descomposición de
hojarasca y el metabolismo del río. Tanto la eutrofización como los cambios en la vegetación de ribera afectan el uso de la hojarasca y, por tanto, impactan en el funcionamiento del ecosistema, lo que hace de la descomposición de hojarasca una herramienta prometedora para el seguimiento de los ríos. Por otra parte, el metabolismo del ecosistema fluvial es afectado por
muchos impactos humanos, y puede ser medido en continuo en estaciones de aforo modernas, lo que permite seguir el funcionamiento del ecosistema casi a tiempo real. Esperamos que estas y otras propiedades funcionales se integrarán en los protocolos
de seguimiento rutinario, que así tendrán en cuenta tanto el lado estructural como el funcional de los ecosistemas fluviales.
Palabras clave: Ríos vascos, propiedades estructurales, propiedades funcionales, descomposición de hojarasca, metabolismo.
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INTRODUCTION
For many years, research on stream ecology has
focused mainly on the structural characteristics
of ecosystems , like water chemistry, physical
habitat or abundance, and structure of biotic
communities (Plafkin et al., 1989; Rosenberg &
Resh, 1992; Rosgen, 1996; Fairweather, 1999;
Maddock, 1999). In contrast, functional attributes of stream ecosystems received less attention,
and were often inferred from structural characteristics, like the widespread use of invertebrate
functional feeding groups (Merritt & Cummins,
1996) to indirectly gather information on stream
function (e.g. Rawer-Jost et al., 2000). This bias
toward ecosystem structure and neglect of ecosystem function is reflected in most streammonitoring studies. These studies are typically
based on biotic or other structural indices, or in
the case of indices of biotic integrity (IBI, Karr,
1981), infer stream dysfunctions from structural
attributes of fish or invertebrate communities.
Even the European Water Framework Directive
(WFD, 2000/60/EC), the most recent and ecologically sound framework for Community action
in water policy for protection and restoration of
aquatic ecosystems, relies heavily on structural
stream characteristics, probably because standard functional metrics are still lacking.
However, some effects of disturbances could go
unnoticed if only community composition is
measured, especially those resulting in longterm responses in the ecosystems.
According to Young et al. (2004), including
functional indicators together with structural
variables for regular monitoring of river health
has several advantages: Firstly, the evaluation of
both structural and functional components of an
ecosystem gives a wider and more complete
view of ecosystem health. Secondly, functional
variables give a true measure of ecosystem integral status. As an example, leaf litter decomposition is influenced by hydrology, water temperature, nutrient concentrations, etc., and many
kinds of organisms are involved (e.g., bacteria,
fungi, invertebrates, fishes).
The group of Stream Ecology at the Faculty
of Science and Technology, University of the
Figure 1. Climate diagram of San Sebastian for the period
1992-2003. Data from Eustat (www.eustat.es). Diagrama climático de San Sebastián para el periodo 1992-2003. Datos de
Eustat (www.eustat.es).
Basque Country, is researching streams in the
Basque Country and Cantabria from an ecosystem perspective since 1988. Our main goal is to
know the basic structure and function of streams
running to the Gulf of Biscay, and how they are
affected by human activities in the basin. Here
we review some of the functional characteristics
of these streams, with special emphasis on the
transport and processing of materials and on
production.
PHYSICAL SETTING
The Basque mountains are relatively low (highest elevations ca. 1500 m.-a.s.l.) and are composed of sedimentary rocks of Jurassic, Cretaceous,
and Tertiary age (reefal limestone, sandstones
and marls), along with some granitic, basaltic,
and metamorphic outcrops, uplifted during the
Pyrenean Orogeny. They are very close to the
coast and thus, valleys are narrow, and rivers
short (ca. 40 km), steep, and straight. The climate of the whole area is under the influence of the
Atlantic Ocean, and therefore, is relatively mild
and humid (Fig. 1). A twofold gradient in rainfall
is usually observed: a southward increase in rainfall as the moist oceanic air rises over the mountain slopes, and superimposed, an eastward
increase toward the corner of the Gulf of Biscay.
As a result, average annual rainfall ranges from
below 1500 mm in the area of Bilbao to over
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Ecology of Atlantic Basque streams
Figure 2. Weekly solar irradiation at Ordizia during 2003.
Open: measured data. Closed: estimated irradiation under a
closed canopy. Shaded area represents the period when the
forest has leaves. Irradiacion solar semanal en Ordizia durante 2003. Open: datos medidos. Closed: irradiación estimada
bajo un dosel cerrado. El área sombreada representa el periodo en que el bosque tiene hojas.
3000 mm in the mountains near the east end of
Guipuscoa (Anonimous, 1995). Irradiance shows
the typical pattern of mid latitudes, but is much
affected by weather (Fig. 2). As a result of heavy
rainfall and short, steep valleys Basque streams
are quite flashy (Fig. 3). Around 1.8 million people live in this area (density = 400/km2), mainly
in cities. Both heavy (steel and paper factories)
and light industry (machine-tool, car components) are abundant in one of the most industrialised Spanish regions. Most rural areas are covered by intensive tree plantations of Monterey
pine (Pinus radiata D. Don) and blue gum
(Eucalyptus globulus Labill.), harvested in short
rotations with heavy machinery. Native forests
consist mainly of beech forests on the wet
uplands; some oak stands on the lowlands, and
some holm oak forests on karstic outcrops. There
are also a number of farms and dairies, most of
them growing cattle and sheep. As a result of all
these activities in such a small territory, most
streams were severely polluted after the mid
20th-century. Over the last 20 years, clean production systems and domestic and industrial
water treatment plants have been implemented in
most basins, resulting in a significant improvement of water quality (Arluziaga, 2002).
125
Figure 3. Hydrograph of the Agüera stream in a wet and in a
dry year. The stippled line marks the level of bed-moving floods according to Elósegui & Pozo (1998). Hidrograma del río
Agüera en un año húmedo y otro seco. La línea discontinua
marca el nivel de las riadas que remueven el sedimento según
Elósegui & Pozo (1998).
Nowadays, some rivers are still severely degraded, whereas others have recovered and have
been chosen for reintroduction programs, as it is
the case of salmon (Galera & Antón, 2001).
Apart from chemical pollution, many Basque
streams are severely affected by canalisation and
degradation of river margins and floodplains
(Basque Government, 2003).
THE AGÜERA AS A MODEL STREAM
The Agüera, a stream at the border between
Biscay and Cantabria, has been studied over the
last 17 years by the group of Stream Ecology at
the University of the Basque Country. It is a relatively well-preserved stream and thus, well suited
as a model for the natural functioning of Basque
streams (Elosegi et al., 2002). In the Agüera,
large spatial variations in water chemistry reflect
changes in geology and soil uses, and from a
temporal perspective, seasonality is relatively
small, main changes being related to relatively
unpredictable floods, or to diel variations
(Elósegui & Pozo, 1994; Elósegui et al., 1997).
Riparian forests are present along most of the
Agüera stream, with the exception of some rea-
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Figure 4. Mass-balance of water (top) and dissolved phosphorus (bottom) along the Agüera stream on July 10, 1990. Numbers are
sampling sites along the stream axis; vertical bars from the top represent inputs from the tributaries, and arrows on the bottom
represent the net balance of other inputs and outputs. In the case of water, inputs are mainly groundwater. In the case of phosphorus, most inputs come from sewers, and outputs reflect in-stream nutrient withdrawal (self-purification). Note the differences between both variables, and the important retention of phosphorus in the mid reaches of the Agüera stream. Balance de masas de agua
(arriba) y fósforo disuelto (abajo) a lo largo del río Agüera el 10 de julio de 1990. Los números representan estaciones de muestreo a lo largo del eje principal, las barras verticales de arriba la contribución de los afluentes, y las flechas de abajo el balance
neto de otras entradas y salidas. En el caso del agua las entradas son fundamentalmente de agua subterránea. En el caso del fósforo, la mayor parte de las entradas proviene de colectores, y las salidas reflejan la retirada de fósforo en el cauce (autodepuración). Observense las diferencias entre ambas variables, y la importante retención de fósforo en tramos intermedios del Agüera.
ches close to villages (Elosegui, 1992). Therefore, periphytic biomass tends to be low except
at open reaches affected by sewage water, where
primary production can be high in long periods
of base-flow (Elosegui & Pozo, 1998). Izagirre
& Elosegi (2005) showed that at closed reaches
the cycle of growth and abscission of riparian
tree leaves control periphytic biomass, whereas
discharge is the main temporal controller at
open ones. As a result of contrasting rainfall
patterns, inter-annual differences can be very
large. In periods of large periphytic biomass,
mid reaches of the Agüera stream can be highly
retentive (Fig. 4), thus resulting in active selfpurification (Elosegui et al., 1995).
One important food source for consumers in
forested streams, as is the present case, is allochthonous organic matter that enters the stream as
coarse particulate organic matter (CPOM), fine
particulate organic matter (FPOM) or dissolved
organic matter (DOM). Both FPOM and DOM
are also good indicators of intermediate disturbances and recovery in streams of the basin (Pozo
et al., 1994; González & Pozo, 1995).
Invertebrate communities tend to change their
trophic structure along the main reach as a result
of changes in resource abundance. Shredders
and gatherers dominate the headwaters, and
gatherers or scrapers, depending on shading and
organic pollution, dominate the mid reaches
(Riaño et al., 1993; Basaguren et al., 1996;
González et al., 2003a). Nevertheless, the trophic structure of benthic invertebrate communities also changes throughout the year because of
floods (Basaguren et al., 1996). From a trophic
point of view, strict specialist invertebrates are
rare. Gut content analyses indicate changes in
diets with development, the main food in early
instars being fine detritus, while other food
types increased in importance in more developed
instars (Riaño, 1998; Basaguren et al., 2002, Fig.
5). Thus, the grouping of all individuals of a
population into a single trophic category might
result in an oversimplification leading to imprecise trophic characterisation of the community
(Basaguren et al., 2002). On the other hand, density and biomass of macroinvertebrates increase
from headwaters to low reaches (Riaño et al.,
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Figure 5. Changes in diet with instar development of
Hydropsyche siltalai in the Agüera stream. Further information in Basaguren et al. (2002). Cambios de la dieta con el
desarrollo larvario de Hydropsyche siltalai en el río Agüera.
Más información en Basaguren et al. (2002).
1993; Basaguren et al., 1996; González et al.,
2003a), and communities tend to be most diverse
at mid and low reaches of the Agüera stream
(González et al., 2003a).
The frequency, timing, and intensity of floods
in such a flashy stream can have strong effects on
benthic community resistance (Imbert et al.,
2005). A comparison of the taxonomic structure
of winter invertebrates at several sites during
several years showed large inter-annual variability (Fig. 6), with floods being an important factor in controlling temporal changes of benthic
invertebrates. This fact, together with the small
seasonality, would explain the low synchrony of
invertebrate biological cycles in the Agüera stream (Basaguren et al., 2002; González et al.,
2003b). Annual production of invertebrate species studied in the Agüera system fell well inside
the values reported in the bibliography, and spatial patterns on production are mainly determined
by in situ biomass (González et al., 2003 b, c).
As mentioned before, riparian forests cover
most reaches of the Agüera stream, and therefore,
the dynamics of allochthonous organic inputs are
of great importance to the ecosystem. In the
Agüera basin indigenous deciduous species coexist with vast plantations of exotic species (pine
and eucalyptus), thus making it an ideal setting
to assess the impact of vegetation changes on
stream ecology. Eucalyptus plantations change
the timing of litter inputs, which tend to peak in
127
Figure 6. Structural similarity of winter invertebrate communities. Cuchillo, Salderrey, and Jornillo are 1st-order streams
surrounded by deciduous forests; Jerguerón and Peñalar, 1storder streams surrounded by eucalyptus plantations; and
Agüera 7 and Agüera 9, 3rd-order reaches with mixed vegetation. Riparian vegetation does not seem to affect the taxonomic structure of macroinvertebrate communities. Note that
large inter-annual variability makes a reach more similar to
another reach sampled the same year than to itself sampled in
different years. Similaridad estructural de las comunidades
invernales de macroinvertebrados bentónicos. Cuchillo,
Salderrey y Jornillo son arroyos de orden 1 rodeados de bosques caducifolios, Jerguerón y Peñalar arroyos de orden 1
rodeados de plantaciones de eucaliptos, y Agüera 7 y Agüera 9
tramos de orden 3 y vegetación mixta. La vegetación de ribera
no parece afectar a la estructura taxonómica de las comunidades de macroinvertebrados. Observese que la fuerte variabilidad interanual hace que un tramo sea más similar a otro
muestreado el mismo año que a sí mismo en años diferentes.
summer, unlike in the deciduous forests, and,
because summer low flows do not scour leaves
downstream, they increase benthic storage of
coarse particulate organic matter (Pozo et al.,
1997; Molinero & Pozo, 2004), which may facilitate its use by consumers and decomposers.
However, because of its toxic oil glands, eucalyptus leaves are a poor resource for large streams
detritivores (Canhoto & Graça, 1999), and could
therefore affect stream invertebrate communities.
Furthermore, because eucalyptus plantations are
harvested in extremely short rotation times, the
impact of logging and road-building operations
can also severely affect streams in areas where
eucalyptus is abundant (Graça et al., 2002).
Most results on leaf litter decomposition processes in the Agüera stream have been already
reviewed in a former work (Elosegi et al.,
2002). In short, among the deciduous species,
Alnus glutinosa (L.) Gaertner is processed faster than other species in the region such as
Quercus robur L. or Castanea sativa Miller
(Molinero et al., 1996; Pozo et al., 1998), a
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consequence of its higher quality (i.e. nutrient
content) and soft texture (Pozo 1993; Molinero
et al., 1996; Molinero & Pozo 2002). On the
other hand, decomposition of the exotic
Eucalyptus globulus can be fast if dissolved
nutrients, particularly phosphorus, are high,
although leaf colonisation by decomposers (e.g.
fungi) and detritivores (e.g. macroinvertebrates)
can be delayed with respect to alder leaves
(Basaguren & Pozo 1994). So, the impact
of eucalyptus plantations on stream communities is far from clear. New essays on this matter
focus on applied science, aiming to design a
tool for the assessment of stream ecological status based on litter decomposition. Some of the
results are mentioned below.
EXPANDING THE MODEL:
NEW APPROACHES IN STREAM
MONITORING
Structure and function are the two sides of the
ecosystem coin, and both are necessary to keep
healthy ecosystems. Until now, most studies that
monitored stream health relied heavily on structural components of the ecosystem. Recently
there has been an awareness of the need to
incorporate functional components into stream
monitoring (Gessner & Chauvet 2002; Young et
al., 2004). The group of Stream Ecology has
been recently involved in two projects aiming to
develop functional tools to be included in stream monitoring. These projects are called RIVFUNCTION and METATOOL.
RIVFUNCTION (www.ladybio.ups-tlse.fr/
rivfunction/) was a project funded by the EU,
involving 12 partners in 9 European countries,
from Portugal to Sweden, and from Ireland to
Romania. Its goals were to develop and disseminate a methodology based on litter breakdown
to assess stream functional impairment. Litter
breakdown is a pivotal function in stream ecosystems, as they are highly dependent on allochthonous organic inputs; furthermore, litter
breakdown is sensitive to many stressors such as
eutrophication, pollution, acidification, or changes in riparian vegetation, and the relatively
easy and cheap measurements make it a good
potential candidate for a monitoring tool.
RIVFUNCTION focused on two kinds of
impacts that could affect stream function: changes in riparian vegetation and eutrophication,
which can differ from country to country. To
study the impact of changes in riparian vegetation, we selected 5 site pairs with similar chemistry but contrasting riparian vegetation (deciduous forests vs. eucalyptus plantations), and to
assess the impact of eutrophication on litter
breakdown, we selected 5 site pairs similar in
size and water chemistry but differing in nutrient
status. During October and November of 2002,
freshly-fallen alder leaves were picked in the
field, air dried, and 5.00 ± 0.25 g of leaves were
enclosed in coarse (10 mm) and fine (0.5 mm)
mesh bags, and incubated in the streams from
December 10th, 2002 to January 10th, 2003
(vegetation series), and from December 27th,
2002 to January 24th, 2003 (eutrophication
series). The end of these experiments was established at the time when an estimated loss of
50 % of the material in coarse bags was reached
at the reference sites. The remaining material was
given in terms of ash-free dry mass (AFDM,
550 ºC, 4 h). Results showed that at nutrient-enriched sites microbial decomposition is enhanced,
but invertebrate consumption is not; on the other
hand, eucalyptus plantations slow down invertebrate breakdown, but do not have a significant
effect on the role of microbes (Fig. 7). Thus, leaf
breakdown is a process sensitive to anthropogenic impacts on streams, but before using it to
assess stream functional integrity, target values of
loss rates must be defined in a broad variety of
reference environmental conditions.
METATOOL (METAbolism and periphyton of
Cantabrian streams: a TOOL to assess the state of
river ecosystems) is a project funded by the
Spanish Department of Science and Technology,
with researchers from the University of the
Basque Country. Its objectives are to measure
primary production and respiration of Basque
streams, to study periphytic communities and
their role in stream metabolism, to identify environmental factors governing metabolism, and to
create a software to calculate metabolism auto-
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Figure 7. Ash-free dry mass remaining in alder litterbags after 4 weeks of exposure. Reference (R) streams are in grey, impacted
streams (I) in black, and asterisks mark significant differences. Left, coarse mesh bags; right, fine mesh bags. Masa seca libre de
cenizas remanente en bolsas de hojarasca de aliso tras 4 semanas de exposición. Arroyos de referencia (R) en gris, impactados (I)
en negro. Los asteriscos marcan diferencias significativas. Izquierda, bolsas de malla gruesa, derecha, bolsas de malla fina.
matically from continuous oxygen monitoring.
The Governments of the provinces of Biscay and
Guipuscoa are involved in this project, as they
are responsible of the so-called hydrometheorological net, consisting in a series of dischargegauging stations where several physical and chemical parameters are monitored continuously.
From continuous data on discharge, water
temperature, and dissolved oxygen, we calculate
whole ecosystem metabolism with the single
station method (Odum, 1956), the re-aeration
coefficient being calculated by the nighttime
method by Hornberger & Kelly (1975). These
calculations are performed with RIVERMET©,
an Excel-based software created for that purpose (Izagirre et al., in press), which can be freely
downloaded at www.ehu.es/streamecology.
Channel geomorphology is described from
transversal transects, where total channel width
and entrenchment, water depth, substrate category (silt, sand, gravel, pebble, cobble, boulder
or bedrock), canopy cover (vertical projection),
and water velocity at mid-depth (A.OTT
Kempten Z30 current meter) are measured.
Additionally, we measure bank and thalweg
slope with a clinometer (Silva clino master) and
a laser-beam level, and estimated changes in
depth, wetted channel width, and water velocity
resulting from changes in discharge with
hydraulic modelling software HecRas 2.2.
Results show that there are large differences
in periphytic biomass in Basque streams, summer maxima ranging from below 2 to almost
200 gAFDM/m2 (Fig. 8). Periphyton is especially abundant at the mid and low reaches of
the Oria stream, but its biomass shows no significant correlation to stream metabolism.
Summer gross primary production in Basque
streams ranges from negligible to almost 20
gO2/m2d, and respiration to 35 gO2/m2d. All
Basque streams seem to be heterotrophic most
of the time, and both gross primary production
and ecosystem respiration are clearly related to
the ecological status of the reach (Izagirre et al.,
in prep.). Nevertheless, stream metabolism
shows high temporal variability, as it is strongly
affected by floods, by changes in insolation, as
well as by increases in water turbidity, which
can often occur due to human activities. Thus,
whole stream metabolism offers a powerful tool
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to assess stream function continuously, but
some care must be taken in order to discern
natural variability from human impacts.
BEYOND THE MODEL: HOT TOPICS IN
MANAGEMENT OF BASQUE STREAMS
Basic knowledge on the functioning of ecosystems is essential to improve the management
and status of our environment. Here we briefly
discuss some topics drawn from the research so
far done in Basque streams, that have potential
implications for environmental management.
One of the aspects emerging from our studies
is the high variability of the Agüera, and by
extension, of Basque streams. Several sources of
variability interact in complex patterns, and affect
the different variables in a distinct way. Thus, all
characteristics of stream ecosystems, from water
chemistry to invertebrate production, are subject
to large changes, at temporal scales that can range
from diel to inter-annual. This variability makes
it difficult to characterise streams and to measure their response to any environmental stress.
Although it is a well-known fact, it has not been
incorporated into monitoring studies as much as
it should (Elosegui et al., 1997).
Streams are metabolically active reactors, not
mere conduits. Riparian and in-channel nutrient
withdrawal can be among the most important
services rendered by streams in the future, as
the costs of ever greater reduction of nutrient
loads into streams rise sharply. Thus, it is necessary to gain insight into the stream characteristics that control retention capacity. There has
been extensive research on the role of riparian
forests on nutrient retention (Haycock et al.,
1993; Sabater et al., 2003), but no guidelines
exist as to the extent, characteristics, and management these forests should have in Basque streams. On the other hand, there is less knowledge
on the factors governing in-channel nutrient
retention, and especially, on how to optimise it.
Our studies also highlighted the important
role of in-stream woody debris structuring the
channel, creating habitats and retaining sediments and organic matter (Díez et al., 2000),
and being a source of nutrients and energy
(Díez et al., 2002). This role was pretty well
known in the American Northwest, but much
less evidence existed of its importance elsewhere, and in particular in European streams.
Streams have been devoid of large wood during
centuries, and strategies to enhance the amount
and role of wood on stream function are necessary to restore streams (Kail & Hering, 2005).
The existing literature shows that wood is
always important, even in reaches where it is
extremely rare (Elosegi & Johnson, 2003), and
therefore efforts should be made to introduce it
where it is absent. This should obviously be
made with caution, as floating wood can block
or damage bridges and other structures.
Furthermore, people tend to perceive dead
wood as an undesired feature of stream channels (Piégay et al., 2005), so any management
activity meant to raise the amount of wood in
streams should also target public awareness.
Our work on organic matter stressed the role
of riparian forests on many aspects of stream
ecology, as well as some important characteristics of these forests, like species composition
and size structure. For example, the retention
rate of alder leaves is higher than the rates of
other tree species (Larrañaga et al., 2003),
which may have implications when alders are
removed from riparian areas. Furthermore, the
loss of alder trees can reduce the storage of benthic particulated organic nitrogen (Molinero &
Pozo, 2002, 2004). Nevertheless, there is still a
long way from our research to management guidelines. Many laws state that it is mandatory to
keep riparian forests, but they most often give
few clues about their characteristics, behind
some minimum width, which is hardly based on
ecological research. These laws should incorporate concepts like forest diversity, integrity, connectivity, size structure and management, and
metrics to evaluate them should be built into
methods to assess stream quality. For instance,
QBR (Munné et al., 1998), the method to assess
the quality of riparian forests most widely used
in Spain, and also applied to Basque streams
(Basque Government, 2003), totally neglects
important aspects of riparian forest ecology, like
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Figure 8. Periphytic biomass (top), gross primary production (mid) and ecosystem respiration (bottom) of Basque streams in summer 2003. Biomasa perifítica (arriba), producción primaria bruta (medio) y respiración del ecosistema (abajo) de ríos vascos en
verano de 2003.
the vigour of seedling growth, even or uneven
size- or age-distribution, abundance of dead
wood, or presence of snags.
Connectivity is important for riparian forests,
but also for in-channel communities. Basque
streams show high temporal variability, and are
subject to frequent disturbances, both natural
and artificial. The most frequent natural disturbances are spates, but droughts can also affect
small tributaries (Otermin et al., 2002).
Artificial disturbances like toxic spills, or siltation from forest clear-cut or highway construction, are also frequent in densely populated
areas like the Basque Country. Both natural and
artificial disturbances affect stream biota, and
the communities can only recover through recolonization, which can be negatively affected by
the many barriers present in Basque streams,
like dams, polluted areas, or reaches affected by
water extraction. The Province Governments of
Biscay and Guipuscoa are nowadays making a
big effort to build fish ladders, remove non-operating dams, and ensure in-stream flows, but the
problem of stream connectivity is far from solved. The sharp decline of some endangered species (e.g., Pyrenean desman, Álvarez et al.,
1985; González-Esteban et al., 2003) in a period
of improvement of water quality point to the
fact that stream habitats are still being degraded,
and to the need of considering population dynamics in the whole stream network. In this context of spatial relationships, we must stress the
tight relationship between human activities in
the basin and stream status. This has been well
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Elosegi et al.
known for a long time, but its implications are
often neglected in real management. Aspects
that should be incorporated into landscape planning include restrictions to forestry and roadbuilding activities in sensitive basins (e.g., feeding reservoirs), localisation of farms with large
amounts of livestock, effects of impervious
areas on basin hydrology, and so on.
ACKNOWLEDGEMENTS
The group of Stream Ecology at the University of
the Basque Country has been funded through
the years with several projects by our university (UPV 118.310-0067/88; UPV 118.310-EA
154/92; UPV 118.310-EA 043/93; UPV 118.310EA 113/95; UPV 118.310-EC 23/97; UPV118.310G14/99; 9/UPV00118.310-14476/2002), by the
Basque Government (PIGV 8924), by the
Spanish Government (DGICYT PB 92-0459;
DGICYT PB 95-0498; DGESIC PB98-0151;
MCYT BOS2003-04466), and by the European
Union (EVK1-CT-2001-00088). Many people
took part in the research, via post-doctoral grants
(Bosco Imbert), with doctoral grants (Raúl
Bañuelos, Joserra Díez, Javier Galán, Esther
González, Jose Manuel González, Oihana
Izagirre, Santiago Larrañaga, Aitor Larrañaga,
Jon Molinero, Ainhoa Otermin, Pilar Riaño,
Joseba Santiago), through MsC Thesis (Xabi
Arana, Arantza Arechederra, Lide Aristegi,
Txomin Bargos, José Barquín, Ainhoa Beltrán de
Nanclares, Elsa Milena Cabrera, Nicola Crosby,
Joxemari Gonzalez, Arantzazu López de
Luzuriaga, Javier Pérez, Amelia Rubio), as well
as many assistant students. Our thanks also to the
Confederación Hidrográfica del Norte, Instituto
Nacional de Meteorología, Basque Government,
and the Province Governments of Biscay and
Guipuscoa, for data and support.
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Limnetica, 25(1-2): 135-142 (2006)
Past, present and future of la Albufera of Valencia Natural Park
J. M. Soria García
Departament de Microbiologia i Ecologia. Facultat de Ciències Biològiques. Universitat de València.
46100 – Burjassot (Valencia). Spain. E-mail: [email protected]
ABSTRACT
The Albufera of Valencia is a highly eutrophied coastal lagoon. Its ecological importance is recognised internationally. The
complexity of the system functioning, as well as its location in a highly populated environment (marshland surrounded by rice
crops), makes difficult the recovery of the system, in spite of the works carried out by the different organisms. The Department
of Environment is developing an integral study to determine which are the necessary performances in order to make possible
the sustainable development of the area as well as the conservation of the ecosystem. This and other studies aim to determine
the current state of the problem and the necessary main works to solve it. With the current information, and keeping in mind
the result of the last evaluation, it is necessary to contribute a bigger volume of water of good quality to the lagoon. This water
will maintain the general quality of the lagoon and also, the arrival of the polluting substances will be eliminated.
Keywords: Albufera of Valencia, Water management, Eutrophication, Coastal lagoons
RESUMEN
La Albufera de Valencia es una laguna costera altamente eutrofizada. Su importancia ecológica está reconocida internacionalmente. La complejidad del funcionamiento de este sistema, unida a que está ubicada en un entorno altamente poblado y
rodeada de un marjal dedicado al cultivo del arroz, hace difícil la recuperación del mismo a pesar de las acciones en este sentido que realizan los diferentes organismos que tienen competencias sobre la Albufera y su entorno. Para determinar cuáles
son las actuaciones necesarias para mantener el desarrollo sostenible de la zona y procurar la mejora del ecosistema, el
Ministerio de Medio Ambiente ha encargado recientemente un estudio integral. El fin de este y otros estudios es llegar a conocer el estado actual del problema y sugerir las acciones necesarias para solventarlo. Con la información que se tiene en la
actualidad y teniendo en cuenta el resultado de la última evaluación se concluye que es necesario que a la Albufera se le aporte un mayor volumen de agua de buena calidad que será quien mantenga la calidad general del lago y además, se elimine la
llegada de las sustancias contaminantes.
Palabras clave: Albufera de Valencia, Gestión del agua, Eutrofización, Lagunas costeras
INTRODUCTION
La Albufera of Valencia is a coastal lagoon placed in the Mediterranean coast line south from
Valencia about 7.5 kilometres from Turia river
mouth. The lagoon is surrounded by marshlands
mainly devoted to rice crops and orchards, scattered country houses and coast line resorts, conforming a stunningly beautiful landscape, under
human pressure. The marshland, the lagoon and
the sandy dunes that detach the lagoon from the
Mediterranean Sea were awarded the category
of Natural Park by the Valencian Autonomous
Government by decree 89/1986. Since 1990 this
natural park has been included among the
“international important wetlands” record established by the Ramsar agreement dated
February 1971. It was also recognized as a special area for bird protection (ZEPA) since 1991.
The set of small subterranean water springs or
“ullals” had been protected by decree 96/1995
that gave green light to the “Natural resources
management plan” of La Albufera hydrographical basin. This wetland shows a very knotty
state, where several interests interact: agriculture, fishing and hunting, the town planning construction, neighbouring villages industrial plans,
the tourism pressure, recreational or conservati-
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J. M. Soria
ves interests, making up a not yet solved problem. The maintenance of natural areas (landscape sites, reserve areas) and water management policy are very controversial matters, as it
is currently happening in La Albufera.
The goal of this paper is to describe the problem, departing from its roots and summing
up the current intervention proposals to improve its situation.
METHODS
The data about the past and present limnological
estate have been drawn both from existing bibliography and studies carried out by the author. The
on-coming plans for La Albufera the work made
by TYPSA for the Confederación Hidrográfica
del Júcar (River Júcar Basin Authority) have also
been considered (TYPSA, 2005).
Figure 1. The Albufera of Valencia Lagoon and the rivers and gullies in the surrounding area. They are indicated the Turia and
Jucar river, the Poyo gully and the minor gullies: B, Picassent - Beniparrell; C, Hondo - Tramusser; D. Berenguera y E, Agua –
Alginet. ARJ indicates the way of the Royal Channel of the Jucar River. Croquis de La Albufera de Valencia y los cauces naturales
en su entorno y su cuenca hidrográfica. Se indican el río Turia y Júcar, el Barranco de Poyo y los barrancos menores: B, Picassent
- Beniparrell; C, Hondo – Tramusser; D. Berenguera y E, Agua – Alginet. ARJ indica el trazado de la Acequia Real del Júcar.
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The Albufera of Valencia Natural Park
The Albufera morphometry
The most recent morphometric (TYPSA, 2005)
and hydrologic parameters of La Albufera, are
the following:
Average depth: 0.9 m a.s.l.
Lagoon perimeter: 23.9 km
Surface of the free water sheet: 22.3 km2
Surface covered by vegetation. 2.0 km2
Water volume: (minimal at 0.0 masl) 17.2 hm3
Water volume (highest ordinary) 27.0 hm3
Basin area: 917 km2
Natural Park area: 211.2 km2
Water contribution volume (2004): 170 hm3
Average water renewal rate (2004): 7.39 times=
49, 5 days.
THE HYDRAULIC OPERATION
Five natural streams flow into La Albufera lagoon: Poyo-Torrent-Massanassa Gully, Picassent –
Beniparrell Gully, Hondo – Tramusser Gully,
Berenguera Gully and Agua - Alginet Gully
(Fig. 1). Before the expansion of the irrigation
in the surrounding shores (until the construction
of the second phase of Acequia Real del Júcar
and Acequia Mayor in Sueca in XVIII century)
the surface waters flowing to the lagoon were
scarce. The rivers Turia and Jucar, which were
originating the system, did not contribute on an
steady basis to its hydric balance, because they
only flew into the lagoon in case of flood. La
Albufera was receiving the water which was left
from the irrigation channels of Favara and Oro
(coming from Turia river)
The sea connection of La Albufera has been
controlled by man in the last centuries throughout
the building of artificial channels called “golas”.
Prior to XVIII century the lagoon connected with
the sea by means of just one 200-m wide “gola”,
allowing water flowing in both directions depending on the weather conditions. The water was
more or less salty because its broad and scarcely
controlled sea channel that allowed the mixing
of sea and continental water. This system made
possible the fishing use and yielded an important
137
salt crop as well. The lagoon level fluctuations
began to be modified in order to allow for bigger
catches. For such a purpose the channel would be
opened or closed depending on the fish reproductive cycles. By then, rice production was still
scant (Rosselló, 1995).
Later (XIX and XX centuries) the connections between La Albufera and the sea began to
increase. New “golas” were ditched, up to the
three which are existing nowadays (Perelló,
Perellonet and Pujol Nou). Improvements in the
water circulation was the answer to the ancient
solid matter sedimentation problem, an expression of the materials dragged by flowing water
to the lagoon. That could entail a natural barrage on the water exit to the sea causing catastrophic consequences on the neighbouring crops
(property of land owners and nobility). This,
together with the need to regulate and manage
the flooding periods, brought in the need of
taking into consideration the problem of drainage and the widening of the sea connections.
Irrigation development increased drastically
the incoming of surface water. The expansion of
irrigation in the lowlands of Valencia, from Turia
river, Jucar river, Acequia Real and Sueca and
Cullera Main irrigation channels derived to the
lagoon a high amount of fresh water which before flew into the sea, a tendency that has been
increasing in the last years. The most representative civil work was the second phase of the
“Acequia Real del Júcar” started out in 1767 by
the Duke of Hijar and Señor of Sollana for the
purpose of irrigating his own land and those of
royal property which were at the borders surrounding the lagoon. This channel allowed the transformation on unirrigated areas (those spanning
between Magro river and Poyo gully) into irrigated land. New channels were opened along XVIII
and XIX centuries in La Ribera Baja (Acequia
Mayor of Sueca in 1798) and the Huerta of
Valencia (extension of Favara irrigation channels
and that from Francos and Marjales towards the
lagoon). In all the number of streams flowing
into La Albufera multiplied its number by ten in
slightly more than a century.
As a consequence, agrarian colonisation
increased during the XVIII century the lagoon,
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J. M. Soria
and lead to the expansion of the rice crops. This
supposed a spectacular increase in the water
intake, the reduction of the lagoon dimensions
and the opening of new sea exits. One of the
main worries of those who have been in charge
of the lagoon management has been the functioning of its hydrologic system, aiming to control
the water level needed in the rice crops.
The hydrologic management model is reflected
in the two entities that embody the conflict between both historical activities linked to La Albufera:
fishing and agriculture. On one hand the “Comú
de Pescadores de la ciudad de Valencia” put forward the fishing activity as a mean to take advantage of the natural resources, confronting it to the
farmer’s aggressions. On the other hand, the
“Junta de Desagüe” which took over competences on the hydrological system management and,
up to the present day, still controls the “golas”
without any environmental sensibility.
Nowadays fresh water flows into the lagoon
through 64 spots, of which 5 are mouths of its
hydrographical basin and the rest are channels
which flow into La Albufera, mainly carrying
irrigation waters from the fields, as well as
urban and industrial outflows. The irrigation
waters coming from the Turia system are derived through a channel network from Quart,
Mislata, Favara and Oro irrigation channels and
its water quality is poor. This is particularly
true for the two last ones, which stem from the
highly polluted Repartiment reservoir, that
receives untreated industrial wastes coming
from Paterna and the industrial estate “Fuente
del Jarro”. The irrigation waters from the
Acequia Real are more relevant and of better
quality. Since water contribution to the whole
system has been linked to agriculture, the contribution coming from the rivers has been
decreasing throughout the last twenty years due
to a better water management. The volume of
water contribution has been of 280 hm3 in 1988
and 170 hm 3 in 2004. During the severe
drought of year 1995 the volume was just 120
hm3 (Soria et al., 2005), being the minimum
flow fixed by the Jucar Basin Hydric
Management Plan (Plan Hidrológico del Júcar)
(CHJ, 1997) of 100 hm3.
Another important aspect when studying the
hydrological changes produced in La Albufera
of Valencia is the urban and industrial development. Both urban and industrial development
have been characterized by accelerated growth,
producing a chaotic situation that impairs its
control and spatial planning. This constitutes a
severe problem for the lagoon, because of the
high polluting capability of these industries and
the lack of wastewater treatment plants.
Altogether, 428 123 inhabitants and 206 industries were censed in 1989, which join an unknown number of sources which are pouring
untreated toxic waste to municipal sewers
(Soria and Vicente, 2002).
WATER QUALITY
La Albufera is a coastal lagoon characterised by
more or less muddy waters, depending on the
suspended matter it receives. Due to its shallowness, it had an extraordinary development
of emerging aquatic plants; air photographs
(1956) and contemporary film documentaries
(Anwander, 1957) showed that vegetation covered the largest part of the lagoon surface. This
vegetation disappeared in the so called channels
or paths used by the boats to sail from port to
port. Water quality in La Albufera and its
surroundings impaired since the 40’s, with a
severe turndown from the 60’s on. In the 70’s
aquatic plants had already disappeared, just
remaining the reed in the shores. Main causes of
this were both urban development and the industrialization of the basin and its surroundings.
Between 1970 and 1980 the lagoon shifted
from an oligotrophic to an hypertrophic state
(Vicente and Miracle, 1992). Nowadays it is an
hypertrophic system caused by the excessive
intakes of allocthonous organic material and
inorganic nutrients, mainly nitrogen and phosphorus compounds (Soria et al., 2002). This
causes the overwhelming growing of cyanobacteria all throughout the year (Villena and Romo,
2003). The temporal rain distribution in its basin
and the irrigation contributions have effects on
the water conductivity (Soria et al., 2000,
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139
Table 1. Limnological variables before (1986–88) and after (1997–2000) sewage diversion from Albufera of Valencia lagoon. Probability
values (P) refer to Wilcoxon’s test on monthly means (modified from ROMO et al. 2005).Variables limnológicas antes (1986-88) y después
(1997-2000) de la disminución de los vertidos a la Albufera de Valencia. Los valores de probabilidad P se refieren al test realizado sobre las
medias mensuales de Wilcoxon (modificado de ROMO et al. 2005).
Variable
Conductivity (µS cm-1)
Average annual temperature (ºC)
pH
Secchi disc depth (m)
Nitrate (mg N l-1)
Ammonia (mg N l-1)
Soluble Reactive Phosphorus (mg P l-1)
Total Phosphorus (mg P l-1)
Chlorophyll a (µg l-1)
1986-88
1784 ± 389
18.7 ± 6.9
8.8 ± 0.4
0.21 ± 0.08
0.94 ± 0.61
0.99 ± 1.66
0.17 ± 0.25
0.49 ± 0.2
269 ± 68
2005), but this is mainly affected by the water
flow management through the opening and closing of the “golas” floodgates (Soria and
Vicente, 2002). Finally, water quality is influenced by both the amount of nutrients and organic
matter carried by several channels that produces
the eutrophic estate of the system.
The substitution of rice crops by more profitable and intensive crops (orchards) are related to
the water pollution used to irrigate the fields.
These turned out to be true decantation reservoirs
(Vicente and Miracle, 1992). The most serious
outcome derived from the shift in crops has been
the reduction of the marshland (wetland surrounding the lagoon shore) by means of drainage and
filling works, thus shrinking the wetlands that
confer La Albufera its value as Natural Park.
Generally speaking, agrarian transformation
meant a highly negative drawback due to agrochemical pollution. Nowadays a better coexistence between agricultural activity and natural
values preservation is meant. Therefore, in order
to preserve rice crops (which are a cushion area
against pollution), several grants have been
issued aiming to decrease the pesticide pollution
in the lagoon. Moreover, farmers are compensated for hypothetical damage in crops caused by
protected birds (Las Provincias, 2004).
Erosion is another of the lagoon big issues,
since it is fundamentally related to human activity.
Different soil classes, forest practices and arsons
set up the amount of actual eroded and erodable
1997-2000
1878 ± 306
19.4 ± 6.2
9.0 ± 0.4
0.27 ± 0.15
0.97 ± 1.00
0.81 ± 0.49
<0.01
0.34 ± 0.14
180 ± 53
P Wilcoxon
0.67
0.21
0.05
0.59
0.75
0.78
<0.01
0.03
<0.01
soil. Furthermore, waters in contributing channels
carry on a high amount of anthropic sediments
and eroded materials. So, all in all, this erosion
means a big amount of material which is alien to
the lagoon (Sanjaume et al., 1992).
La Albufera receives undepurated sewage
waters from urban origin, together with those of
water treatment plants. Moreover, dumping of
highly polluted sewage through gutters occurs
mainly under rainy weather. Nowadays, the
West Main Sewer dumps sewage through its
gutters even in times of dry weather (TYPSA,
2005), and its pumping devices show clear
deficiencies (Soria, personal observation).
All these contributions have negatively contributed to the trophic quality of the lagoon in the
last fifteen years (Romo et al., 2005). Although
phosphorus concentration has decreased since
the 80’s, nowadays it stands between 0.49 until
0.34 mg P l-1; chlorophyll average stands between 269 µg l-1 until 180 µg l-1 (Table 1). During
some clean phases it has gone under 10 µg l-1 for
a few days only to restore, also in a short span of
time, to the actual average values (Miracle &
Sahuquillo, 2002), which are clearly unacceptable to keep the ecosystem quality.
Due to the high phytoplankton growth, pH
stands at very high levels, seldom below 8.5,
being the most frequent values of 9.5, even 10.
That raises the rate of a toxic ammonia compound in water in a very dangerous way
–mainly for fish– (Blanco et al., 2003). It is not
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J. M. Soria
only that the pH contributes to the ammonia
compounds balance, but also that high temperatures provoke a rise in the rate of non ionized
ammonia in water. Moreover, fluctuations in
the dissolved oxygen as a result of the high production make animals more sensible to its toxicity (Vicente & Miracle, 1992). A further consequence of the high micro-algae concentration
is the high amount of native organic matter,
whose breakdown can not be completed in an
aerobic way, so that, as well as producing
nightly oxygen shortage in water it remains
sedimented organic matter, part of which
decomposes in an anaerobic way freeing H2S.
BIODIVERSITY
Phytoplankton population growth is well described by Villena & Romo (2003a and 2003b).
Its main characteristic is the presence of cyanobacteria and chroococcales along with diatoms
and chlorophyta. Zooplankton production is
very reduced, and the hardly edible phytoplankton is recycled mainly by detritic way. There
has been a loss of plankton filtering species
capable of regulating the phytoplankton concentration, and consequently also of water quality and transparency. There is a higher contribution of Rotifera nowadays (Oltra et al.,
2001). Native fish fauna is composed mainly by
sea mullet (Mugil cephallus), sharpnose mullet
(Liza aurata), carps , silverside fish and pumpkinseed sunfish (Blanco et al., 2003).
In the middle 80’s (1984, 1987) “clear water
phases” in which the phytoplankton population
drastically decreases was observed by Miracle &
Sahuquillo (2002). During 10 to 20 days water
appears less muddy and in some spots, the lagoon bottom can even be seen. This phenomenon is
produced by a substitution of the usual cyanobacteria plankton by other micro algae (diatoms
and others) and the outcome of filter-feeders like
Daphnia magna, capable of a very efficient filtration upon the algae population. This period of
clear waters is probably related with the flowing
of clean waters coming from the rice fields drainage after its initial flooding in January –
February. The “clear water phase” does not show
a regular pattern, either temporally or spatially in
the lagoon. The clear phases are an evidence of
the recovering capability of the whole system
provided that the quality and quantity of water in
the tributary channels improves.
Nowadays some 250 bird species find shelter
in the park and around 100 of them also reproduce there. The most numerous group during winter are the anatidae, ranging from 40.000 to
60.000 individuals. Worth mentioning is the redcrested pochard (Netta rufina), which attains
about 17.000 individuals. Other common anatidae are the mallard (Anas platyrhynchos), the
pochard (Aythya ferina) and the shoveler (Anas
clypeata). Smaller number of individuals are
pintails (Anas acuta), wigeons (Anas Penelope)
and teals (Anas crecca). Ardeidae, such as grey
herons (Ardea cinerea) little egrets (Egretta garzetta) and cattle egrets (Bubulcus ibis), are common during the winter and during nestling time
attract international attention due to the high
number of nests in the area (more than 2500).
Significantly, the nestling species in the area of
La Albufera ranks among those in the “List of
Wetlands of International Importance” following
the numerical criteria issued by Ramsar for several bird species, according to Valencian
Ornitology Society data (2005).
The lagoon fauna (both fish and benthic) have
undergone modifications throughout the years, as
a result of several actions taken upon the system.
Eel and bass have shown sharp decreases, while
others are now extinct, such as the endemic
“samaruc” Valencia hispanica and “fartet”
Aphanius iberus. Other fishes as the loach and
the chub are possibly restricted to some channel
areas (Blanco et al., 2003). Several endemic species among the invertebrates are now considered
extinct: Dugastella valentina and Palaemonetes
zariquieyi (known as “gambetes”) and the bivalvia Unio turtoni valentinus (“petxinot”), which
can only be found in protected areas (Dugastella
valentina and Palaemonetes zariquieyi) and possibly in some irrigation ditches (Unio turtoni
valentinus) (Sánchez, 1991).
The flora plays an important role within the
Park structure, setting up the placement and
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diversity of both bird and fish species living
there. Nonetheless in the last years flora diversity
has decreased because of the rise of pollution and
the building construction along the coast line.
Flora biodiversity is now reduced to the presence
of riparian in channels irrigation ditches and the
lagoon shores: common reed, common cattail,
yellow iris, and bulrush. The surface of the lagoon occupied by macrophytes has drastically reduced the former abundance of Myriophyllum,
Chara, Ceratophyllum, Potamogeton and Nymphaea alba. Ranunculus aquatilis grows only in
rice crops during the flooding period (Sánchez,
1991). Variety and abundance of “clean water”
species in the future would be an expression for
the wetland water quality improvement.
ACTIONS TO BE TAKEN ACCORDING
TO THE STUDY OF SUSTAINABLE
DEVELOPMENT
Some essential actions are planned in the near
future in order to improve the ecosystem quality
(TYPSA, 2005), aiming to obtain a sensible
improvement in 15-20 years time (Phillips, 2005):
a) Water. The ideal volume of water flowing
into the lagoon should be more than
200 Hm3 of water of good quality (less than
0.1 mg P l-1 phosphorus), in order to prevent euthrophication (Romo et al., 2004).
This would mean that contributions to the
lagoon should come directly from reservoir
waters. This water could be distributed
among the rice fields in an ascending way,
thus reversing the current flow scheme,
later on it could be derived to other basins
for agricultural purposes. In this sense the
option of sending the amount of water left
towards the Vinalopó basin would be highly
advisable. The spare waters should be then
directed towards the sea.
b) Erosion. Lessening the erosion means taking
actions such as the reforestation in the
watershed, updating agricultural practices,
arson control and clear actions leading to
recover affected spots.
141
c) Agricultural areas. It would be desirable to
obtain higher economic profits derived from
rice crops and other varieties, cutting off the
fertilizer and pesticide pollution hazards at
the same time, taking into account that
waters left from the flooding could be used
for other agricultural practices.
CONCLUSION
Regardless of the efforts carried out by National,
Regional and Municipal administration to preserve and improve the environmental state of La
Albufera, after 20 years of investment and struggle, the eutrophy of the system has been scarcely
reduced. Main problem is the west park boundaries urbanization which has overwhelmed the
existing infrastructures, plus the slow pace in
taking up unavoidable actions and the inappropriate management of flowing waters. To all this
we must add the reduction of clean waters contribution coming from irrigation.
The achievement of a desirable situation, between mesotrophic and oligotrophic, is only possible if good quality waters, needed to keep the
environmental qualities, flow in. Also the arrival
of polluting substances directly linked to eutrophy
must be cut in short time. The seldom appearance
of clear water phases during last years is a manifestation that improvement is possible.
BIBLIOGRAPHY
ANWANDER, CH. 1957. Entre el agua y el barro:
Estampas de la Albufera de Valencia. Writted by
Christian Anwander & Alfredo Marquerie.
Documental film performed by NODO. 16 min.
BLANCO, S., S. ROMO, M. J. VILLENA & S.
MARTINEZ. 2003. Fish communities and food
web interactions in some shallow Mediterranean
lakes. Hydrobiologia, 506 (1-3): 473-480.
CHJ (Confederación Hidrográfica del Júcar). 1997.
Plan Hidrológico del Júcar. Documento nº 2.
Normativa. (On line). Available in World Wide Web
http://www.chj.es/web/pdf/NORMATIVA.pdf
LAS PROVINCIAS. 2004. Territorio sólo paga la
mitad del daño causado por aves en arroz. En
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Albufera de Valencia (On line). 12 november
2004. Available in World Wide Web:
http://www.albufera.com/portal/modules.php?na
me=News&file=article&sid=800
MIRACLE, M. R. & M. SAHUQUILLO. 2002.
Changes of life-history traits and size in Daphnia
magna during a lear-water phase in a hypertrophic
lagoon (Albufera of Valencia, Spain). Verh.
Internat. Verein. Limnol., 28: 1203 – 1208.
OLTRA, R., M. T. ALFONSO, M. SAHUQUILLO &
M. R. MIRACLE, 2001. Increase of rotifer diversity after sewage diversion in the hypertrophic
lagoon,
Albufera
of
Valencia,
Spain.
PERIS, T. 1991. La problemática génesis del
Segundo tramo de la Acequia Real del Xúquer.
Investigaciones Geográficas 9: (On line).
Available in World Wide Web: http://www.cervantesvirtual.com/servlet/SirveObras/013716304556
15946322257/invg_11.pdf
PHILLIPS, G., A. NELLY, J. PITT, R. SANDERSON & E. TAYLOR. 2005. The recovery of a very
shallow eutrophic lake, 20 years after the control
of effluent derived phosphorus. Freshwat. Biol.,
50 (10): 1628–1638.
ROMO S., M. R. MIRACLE, M. J. VILLENA, J.
RUEDA, C. FERRIOL & E. VICENTE. 2004.
Mesocosm experiments on nutrient and fish effects
on shallow lake food webs in a Mediterranean climate. Freshwat. Biol., 49 (12): 1593-1607.
ROMO, S., VILLENA, M. J., SAHUQUILLO, M.,
SORIA, J. M., GIMENEZ, M., T. ALFONSO, E.
VICENTE & M. R. MIRACLE. 2005. Response of
a shallow Mediterranean lake to nutrient diversion:
does it follow similar patterns as in northern shallow lakes? Freshwat. Biol., 50 (10): 1706-1717.
ROSSELLÓ, V. M. 1995. L’Albufera de Valencia.
Publicacions de.l’Abadia de Montserrat. 190 pp.
SANCHEZ, J. 1991. Plan Especial de protección del
Parque Natural de la Albufera. Conselleria de
Medi Ambient. Generalitat Valenciana. 148 pp.
SANJAUME, E., F. SEGURA, M. J. LÓPEZ & J.
PARDO. 1992. Tasas de sedimentación en
L’Albufera de València. Cuad. de Geogr., 51:
63–81.
SORIA, J. M., E. VICENTE & M. R. MIRACLE.
2000. The influence of flash floods on the limnology of the Albufera of Valencia lagoon (Spain).
Verh. Internat. Verein. Limnol., 27: 2232-2235
SORIA, J. M. y E. VICENTE. 2002. Estudio de los
aportes hídricos al Parque Natural de la Albufera
de Valencia. Limnetica, 21(1-2): 105-115.
SORIA, J. M., M. R. MIRACLE & E. VICENTE.
2002. Relations between physico-chemical and
biological variables in aquatic ecosystems of the
Albufera Natural park (Valencia, Spain). Verh.
Internat. Verein. Limnol., 28: 564-568
SORIA, J. M., M. SAHUQUILLO y M. R. MIRACLE. 2005. Relaciones entre las aportaciones a la
zona regable del río Júcar y la conductividad
de la Albufera de Valencia. Limnetica, 24 (1-2):
155-160.
SVO (Societat Valenciana Ornitología). 2005. Sitios
de interés: La Albufera de Valencia. (On line).
Available in World Wide Web: http://www.ctv.es/
USERS/miguel-peris/LaAlbufera.htm
TYPSA. 2005. Estudio para el desarrollo sostenible
de l’Albufera de Valencia. Confederación
Hidrográfica del Júcar. (On line). Available in
World Wide Web: http://www.albufera.com.es.
VICENTE, E. & M. R. MIRACLE. 1992. The coastal lagoon Albufera de Valencia: An ecosystem
under stress. Limnetica, 8: 87-100.
VILLENA M. J. & S. ROMO. 2003. Temporal changes of cyanobacteria in the largest coastal Spanish
lake. Algological Studies. 109: 593-608.
VILLENA, M. J. & S. ROMO. 2003. Changes in the
phytoplankton of a shallow Mediterranean lagoon
(Albufera of Valencia, Spain) after nutrient diversión. Hydrobiologia. 506: 281-287
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Limnetica, 25(1-2): 143-154 (2006)
Limnology of natural systems for wastewater treatment. Ten years
of experiences at the Experimental Field for Low-Cost Sanitation
in Mansilla de las Mulas (León, Spain)
Eloy Bécares
Area de Ecología. Facultad de Biología. Universidad de León. 24071 León. Spain
Telf: 987291568 Fax: 987291409, e-mail: [email protected]
ABSTRACT
The first experimental field in Spain for low-cost wastewater treatment was constructed in 1998 in Mansilla de las Mulas
(León). The project was funded by the Diputación de León and was run scientifically by the Department of Ecology at the
University of León until 1999. The objective of the field was to compare performances and to study the biological fundamentals of systems such as constructed wetlands and algal ponds which had been adapted for wastewater treatment in rural areas.
The experiences carried out on constructed wetland systems proved that macrophytes had a significant role in the treatment of
diluted wastewaters. The role of vegetation varied depending on the technology used, being important for organic matter removal in free water surface systems, whereas plants were only significant for nutrients and faecal bacteria in subsurface flow
systems. There were no differences among plant species in the performance of free water surface systems. Algae-based
systems such as high rate algal ponds proved to be a highly efficient technology for wastewater disinfection. Studies on decay
and inactivation of faecal bacteria and parasites (helmiths and protozoan oocysts) demonstrated for the first time that the
physico-chemical conditions created by the algae are powerful mechanisms for pathogen destruction. Wastewater treatment
plants are technologically-confined ecosystems in which limnological studies should be further encouraged both examining
basic knowledge on natural species and processes and leading to a better understanding of the biological foundations for their
design and operation.
Key words: wastewater, natural systems, constructed wetlands, high rate algal ponds, macrophytes, algae, nutrients, faecal
bacteria, helminths, protozoa.
RESUMEN
La Universidad de León estableció en 1998 un proyecto de investigación con la Diputación Provincial cuya principal materialización fue la construcción del primer campo experimental en depuración de bajo coste en nuestro país. El objetivo del
campo experimental fue el de estudiar, a escala experimental, diferentes procesos de bajo coste, entre ellos los humedales
construidos y los lagunajes de alta carga. El presente trabajo resume los principales resultados de las investigaciones realizadas después de diez años de experiencias. Las plantas demostraron tener un papel significativo en humedales que tratan aguas
residuales diluidas. Dicho papel fue diferente dependiendo de la tecnología. En sistemas de flujo sub-superficial las plantas
demostraron influir significativamente sobre la eliminación de nutrientes y bacterias fecales, sin embargo en sistemas de flujo
superficial dicho efecto fue insignificante, siendo sin embargo importante sobre la eliminación de materia orgánica. La especie de planta utilizada demostró no ser una variable importante en el funcionamiento de los sistemas de flujo superficial. Los
experimentos realizados en lagunajes de alta carga demostraron que las condiciones físico-químicas creadas en dichos sistemas eran responsables de la inactivación del 50-90 % de los organismos patógenos e indicadores fecales estudiados. Los tratamientos biológicos de aguas residuales deben entenderse como ecosistemas tecnológicamente confinados en los que el
conocimiento de su estructura y funcionamiento es aún excesivamente escaso en comparación con la importancia que representan para el medio ambiente.
Palabras clave: aguas residuales, sistemas naturales, humedales, lagunajes de alta carga, algas, macrófitos, nutrientes, bacterias fecales, parásitos, protozoos.
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INTRODUCTION
The biology of wastewater treatment plants is an
important aspect of applied limnology. It includes
a wide variety of aquatic organisms, from viruses
to fishes, and their interactions and processes in
relation to the fate and degradation of conventional and industrial water pollutants. Unfortunately,
limnologists have paid much less attention to this
field than chemical or civil engineers, and even
today much of the biological research in this field
is not performed by limnologists.
Wastewater treatment plants are technicallyconfined ecosystems in which design and operational variables have to be carefully controlled
for the adequate selection of the organisms responsible for pollutants removal. Wastewater
treatment plants are biotechnological reactors
working under the same principles of other reactors (e.g. antibiotics production) with some
important differences. First, they are biologically
open systems. The continuous input of propagules is an important process for the system. The
second aspect is that the quantity and quality of
food (i.e. wastewater) is highly variable.
Wastewater treatment was also of interest for
Ramon Margalef, who described them as “forced
ecosystems” in which an intensification of processes occurs with a production in excess of their
organisms (Margalef 1983). These systems have
contributed to scientific knowledge in important
aspects of microbial ecology and taxonomy, for
instance of filamentous bacteria, nitrogen and
phosphorus biological removal or protozoan ecology and taxonomy (Seviour & Blackall 1999).
Wastewater treatment systems can be classified depending on their electric requirements and
maintenance costs in one of two types. The “conventional systems” (activated sludge, biofilters)
require a substantial number of pumps, aerators
or other devices for their functioning. The second
types are “low-cost systems” (trickling filters,
stabilization ponds, constructed wetlands). In the
latter type, minimum electric requirements and
maintenance costs are necessary, although a
much larger surface is required for their installation. These characteristics make low-cost systems
a sustainable option in most rural areas. The re-
search on their biological foundations is an
important objective for applied limnology. As the
scientific coordinator of the activities carried out
at the Experimental Field for Low-cost sanitation
in Mansilla de las Mulas from 1988 to 1999, here
I present the main outcomes of the research
carried out by the group during that time.
THE EXPERIMENTAL FIELD
In 1988 I was asked to propose ideas on how to
manage the funds that the Diputación Provincial
de León was planning on investing in wastewater
treatment studies. The knowledge gained at that
time in one of the first masters in Spain on
Environmental Engineering at the Cantabria
University, the limnological background of the
Ecology Department at León University, and
the fact that the Diputacion is an administrative
body focused on promoting the development of
rural areas, gave me the idea of proposing the
construction of the experimental field, the first
in Spain, to compare performances, economic
costs and the biological fundamentals of lowcost wastewater treatment plants. No studies
were available at that time on low-cost wastewater treatment systems and specially, on the biological basis of processes occurring in them. The
location of the field was decided among villages
surrounding the city of León on the following
criteria: i) the proximity to the University of
León, ii) the availability of a municipal area
large enough for construction, iii) the vicinity to
the area of a sewage system with enough flow
for the experiments, and iv) the absence of
industrial or heavily loaded wastewaters.
Climatic conditions in León province are a limiting factor for the design of wastewater treatment
plant. Low temperatures in winter (mean temperature of the coldest month: 5 ºC) means systems
must be larger than in other parts of Spain.
Moreover, wastewater is much diluted in our rural
areas as a consequence of the deficient construction of sewers and the entrance of water from urban
wells. Also, wastewater is recycled for use, either
directly from the sewer or indirectly by using water
from polluted rivers and streams. Thus, the study of
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Limnology of wastewater treatment
145
Figure 1. A view of the experimental field. High rate algae ponds in the bottom left. Stabilization pond in the middle and pilot
tanks for constructed wetlands research on the top. Vista del campo experimental. Lagunajes de alta carga (parte inferior), lagunaje convencional de maduración (centro) y tanques piloto para experimentación con macrófitos acuáticos (parte superior).
the use of low-cost systems is warranted. Having in
account previous considerations, the systems selected for experimentation fulfilled the following
requirements: i) to be low cost systems with regard
to operation and management, ii) they should be an
adequate technology considering the climatic conditions in León province, iii) they should priorize
the study of systems with good pathogen removal
efficiency since wastewater reuse is a common
practice in rural areas. I outlined the construction
project with advice from the Departments of
Environmental Engineering at the Universities of
Cantabria and Polytechnic University of Barcelona
and the Fondation Universitaire Luxembourgeoise.
The construction project was finally designed in
collaboration with the engineer from the Diputacion, César Roa. I decided first on a macrophytebased system due to the scientific background on
aquatic plants available at the Department of
Ecology, through the expertise of Drs. Camino and
Margarita Fernández-Aláez. The experimental layout was based on that previously built by Radoux
(Radoux & Kemp 1982) in Viville (Belgium), and
adapted to our particular conditions (Bécares 1989,
Bécares et al. 1989). Other pilot-plant systems considered were a conventional facultative pond and a
high rate algal pond (Fig. 1).
APPLICATION OF CONSTRUCTED
WETLANDS FOR WASTEWATER
TREATMENT
Constructed wetlands are based on the use of
plants for the treatment of wastewater. Depending on the type of plants, the systems can
include submerged plants (limnophytes, e.g.
Myriophyllum spp., Ceratophyllum spp.), emergents (helophytes, Typha spp., Scirpus spp.,
Phragmites spp.) or floating plants (pleustophytes, Lemna spp., Eichhornia spp., Salvinia
spp.). Because of the presence of surface water,
the systems are classified as free-water surface
(FWS) systems, in which a small depth of water
(approximately 25-30 cm) is maintained on the
top of the soil, or subsurface flow (SSF) when
water is a few centimetres below the soil level.
According to flow type the systems are classified as horizontal or vertical flow systems. The
combination of plant species and flow types has
produced different technologies.The most common combine FWS with horizontal flow or SSF
with horizontal or vertical flow. Detailed descriptions on these technologies can be found in
several books and manuals (Hammer 1989,
Moshiri 1993, Reed et al. 1995, Kadlec &
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E. Bécares
Knight 1996, Vymazal et al. 1998, Young et al.
1998, Kadlec et al. 2000). An up-to-date
synthesis on these systems can be found in
García J. et al. (2004). The first experiences in
Spain on constructed wetlands for wastewater
treatment were carried out in Murcia in the 80’s
using Phragmites in small tanks (Moreno
1989), followed by our work in Mansilla de las
Mulas. At present there are about ten different
groups in Spain working on the subject. The
experiments conducted by García and collaborators are probably the most widely embracing
(e.g. García J. et al. 2003, 2004)
Have plants a significant role in constructed
wetlands for wastewater treatment?
The removal of organic and inorganic matter
and bacteria from wastewater carried out by
macrophytes has been explained through several mechanisms, such as sedimentation, mechanical filtration or nutrient assimilation by
plants. Their roots may also serve as substrates
for attached bacteria degrading and taking up
nutrients and organic carbon (Brix, 1995). The
latter process is favoured by oxygen release
into the rhizosphere (Gersberg et al., 1986),
and by plant exudates (Stengel, 1985). Despite
this evidence, there is still controversy about
the mechanisms of functioning of macrophytes
for wastewater treatment in constructed
wetlands. Some researchers have found wastewater treatment is improved in the presence of
macrophytes (Rogers et al. 1991, Farahbakhshazad et al. 1995), while other studies
have not detected significant differences in
treatment results between planted and unplanted systems (Tanner et al., 1995). Nevertheless,
comparisons between studies are difficult
because they utilize diverse aquatic plant species, wastewaters and flows. One of our first
objectives was therefore to test if plants had
any significant role under the climatic conditions and wastewater characteristics in the province of León. Experiments were carried out in
both FWS and SSF systems. Further details on
the experiments and results are available in
Ansola (1994) and García (2002).
Organic matter and nutrient removal
in subsurface flow systems
SSF pilot-scale tanks 0.6 m3 in volume and with
a surface of 1.1 m2 were planted with Scirpus
lacustris growing on siliceous gravel. These
systems were compared with treatments in
which S. lacustris was grown in hydroponic culture, and with only gravel. The average diameter
of gravel was 6 mm, and porosity was calculated
as 35 %. Two replicates of each treatment were
used and all received the same wastewater.
Hydroponic cultures were designed to work
under plug-flow hydraulic regime by means of
deflectors, and plants were supported using a
small roll of garden net of 2 cm pore. Details on
the experiments and results are available in Soto
et al. (1999a, 1999b, 2000).
Organic matter and suspended solids
Plants do not take up organic matter but foster
the growth of organisms in the rhizosphere
which could help on such task (Brix 1994). In
our experiments, removal rates for biological
oxygen demand (BOD5) and suspended solids
(SS) were slightly higher in summer than winter
but no statistically significant differences were
found when comparing winter/summer or planted/unplanted conditions. In accordance with
other authors (Tanner et al. 1995), absence of
differences between tanks could be due to two
reasons. First, solids (organic matter) accumulate in the first third of the system, mostly at the
head of the tanks (Tanner & Sukias 1995).
Therefore, efficiency is potentially independent
of plant presence. Second, BOD loads were very
low, below 3 g/m2 d. Tanner et al. (1995) found
this value to be the limit for detecting significant plant effects on BOD removal.
Nitrogen. Ammonia
A strong positive correlation was found between
ammonia surface loading and removal rate (Soto
et al. 1999a, 1999b). Removal rate was significantly higher (P<0.05) in planted tanks and also
higher in summer than winter. Planted tanks removed 50 % more ammonia than unplanted under
winter conditions (0.145 g/m2 d.), and about 63 %
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more in summer conditions (0.297 g/m2 d). This
means that plant activity in summer was only responsible for 13 % of the ammonia uptake (i.e.
0.152 g/m2 d), while other mechanisms unrelated
to plant uptake but linked to plant presence, were
responsible for the rest of the removal observed.
Nitrogen. Nitrate
Nitrification effectively occurred in both planted and unplanted tanks. Unplanted tanks were
marginally more effective than those with
plants in the removal of nitrate (differences
were not statistically significant). The ratio
C:N (as BOD5:TN) was 2.42 in winter, and
3.55 during summer, high enough for denitrification to occur (Radtke, 1995). Nitrogen accumulation by plants was calculated in about
30 % of the TN removed into the planted
system, including the N accumulated in submerged parts, while the remaining 70 % was
likely removed by the microbial pool and lost
through denitrification (Soto et al. 1999a).
Phosphorus
It is generally accepted that the mechanisms for
phosphorus removal are more related to gravel
surface processes, such as physical adsorption and
chemical precipitation by Ca or Fe (see, for example, Nichols, 1983; Richardson, 1985; Faulkner
and Richardson, 1989), than to biological processes, such as uptake by plants and microorganisms.
However, and similarly to that found by Faulkner
and Richardson (1989), the comparison between
planted and unplanted systems showed that the
removal rate of reactive phosphorus (SRP) and the
removal efficiency were higher in planted than
unplanted tanks. According to this, macrophytes
were directly involved in the removal of 29 % of
TP in winter and 47.3 % in summer (i.e. 0.024 and
0.057 g/m2 d removed, respectively). These differences were statistically significant in summer.
The effect of plants in free-water surface flow
systems
In systems with subsurface water flow, substrate
hydraulic conductivity is an important design
parameter. Wastewater interacts directly with
147
the rhizosphere, and roots have additional functions apart from being the physical support for
the biofilm. Nevertheless, in FWS systems, the
hydraulic conductivity of the gravel bed and
therefore the role of the rhizosphere is negligible (Kadlec and Knight, 1996). The main role of
macrophytes is to provide additional surface for
the development of a biofilm on the submerged
parts of plants. Comparing planted and unplanted tanks, Ansola et al. (1993, 1995) proved that
planted tanks were significantly different from
control plots with regards to DBO, COD and
total phosphorus, but no differences were found
between planted and unplanted tanks with
regards to nitrogen forms (ammonia, nitrates,
organic nitrogen) nor phosphates. Differences
between FWS and SSF in the type of substances
removed (nutrients in SSF, organic matter in
FWS) are evidence that plants have a passive
role in the superficial-flow systems, acting as a
physical support for bacterial growth on their
submerged leaves and stems.
Mechanisms responsible for bacteria removal
in constructed wetlands
Similarly to previous researches (Gersberg et
al. 1990a, Rivera et al. 1995, Tanner et al.
1995, Loveridge et al. 1995), from differences
between planted and unplanted systems it can
be concluded that macrophytes also play an
active role in the removal of microorganisms
from wastewater. Rooted biofilms provide a
better substrate than gravel surfaces for microbial activity (Loveridge et al. 1995). Constructed wetlands are more efficient than conventional systems in the removal of bacteria
but generally less efficient than stabilization
ponds (García & Bécares 1997). The comparison of planted and unplanted systems generally
showed a higher rates of bacteria removal in the
presence of plants, although results were highly
variable and were dependent on plant type,
hydraulic design and wastewater characteristics
(Hammer 1989, Tanner et al. 1995).
Several mechanisms have been proposed, and
on occasion demonstrated, to be responsible for
bacteria removal. Oxygen production and bacte-
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E. Bécares
rial activity in the rhizosphere (Brix 1987,
1997), sedimentation, filtration and adsorption
(Gersberg et al. 1989, Williams et al. 1995) are
commonly cited mechanisms in the literature.
Decamp & Warren (1998) and Rivera et al.
(1995) have pointed bacterivory as a key mechanism in CW, as microfaunal densities and predatory activity were higher in the presence of
plants in their experiments. Excretion of antibacterial compounds by plants is another controversial mechanism frequently cited in the
literature, but not yet clearly proved. Plants
modify the soil microenvironment and probably
release substances which enhance the development of specialized bacterial on its rhizosphere
(Hatano et al., 1993). Commonly cited papers,
such as those by Seidel (1955, 1976), Gopal &
Goel (1993), Ottová et al. (1997) and others
have not been able to prove a direct effect of
plants on bacteria inhibition but have found that
plant presence is related to higher bacteria
reductions. There is evidence showing that some
plants produce secondary metabolites with antibacterial properties (e.g. Dellagreca et al.
2001), although their role in wastewater treatment has still to be proved.
An important part of our studies at the
Experimental Field were focused on determining and quantifying the mechanisms involved
in bacteria and pathogen removal by constructed
wetlands. Studies comparing the role of Scirpus
lacustris in planted and un-planted subsurface
flow tanks (García et al. 1999, Soto et al. 2000)
showed that planted tanks were more efficient at
removing microbes (up to 99.9 %) than unplanted tanks. There were statistical differences between planted and control conditions (p<0.01)
for total coliforms, faecal streptococci and total
heterotrophs, and removal rates were higher in
summer than winter. Antibacterial activity
potentially exerted by plants was not detected by
using filtered effluents from planted tanks as
dilution water for total bacterial growth in the
influent (García et al. 2004). Predation, as a
potential mechanism for bacteria removal was
also evaluated in these systems (García et al.
2004). Decamp et al. (1999) found a higher
ciliate abundance and predatory activity in plan-
ted than unplanted gravel wetlands. In our tanks,
results showed that the abundances of ciliates
and flagellates, the only bacterivorous organisms found, were much lower than reported in
other similar systems (Decamp et al. 1999,
Panswar and Chavalparit 1997). Differences in
abundance in planted and unplanted tank were
statistically significant for flagellates but not for
ciliates. The presence of plants increased protozoan abundance, which could be another reason
for the higher bacteria removal in planted
systems (García et al. 2004).
Do the plant species matter?
It is well known that different plant species have
different resource requirements and rates of matter processing. A question here was if there were
differences among helophytes with regard to
their processing efficiency on nutrients or organic matter pollution. Scirpus lacustris, Typha
angustipholia, Iris pseudacorus, Phragmites
australis and a control without plants were
simultaneously grown in small tanks with surface flow hydraulics. Ansola et al. (1994) did not
find differences among plant species for all
variables studied with the exception of phosphates for Iris pseudacorus, which presented significantly higher removals compared to Typha and
Phragmites. In other experiment (Ansola 1994),
compared three-way combinations of plants
(Scirpus-Iris-Phragmites, Typha-Iris-Scirpus,
Typha-Scirpus-Phragmites) growing in separate
tanks. Results showed no differences among
combinations of plants for nutrients or organic
matter. Concerning faecal indicators and pathogen removal, López & Bécares (1993) found that
Scirpus lacustris had higher bacteria removal
rates than the aforementioned species, as shown
by most faecal indicators used.
ALGAE-BASED SYSTEMS FOR
WASTEWATER TREATMENT
A high rate algal pond (HRAP) was the algaebased system selected for testing under the aforementioned climatic conditions in León. HRAP
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is a low-cost wastewater treatment system
designed to achieve two goals: the secondary
treatment of wastewater and the production of
algal biomass. The HRAP is a combination
of intensified oxidation pond and algal reactor.
Algae supply the oxygen needed for the bacterial degradation of organic matter, while bacteria supply mineral compounds excreted to the
algae for their nutrition. The HRAP is characterized by shallow depths, mechanical mixing and
short residence times, implying smaller surface
areas than those of conventional stabilization
ponds (Fallowfield & Garret 1985). HRAP
systems were first designed for wastewater
treatment by Oswald (1963, 1988), and thereinafter were used in other parts of the world.
Particulary worthy of notice are the projects by
Shelef in Israel (Shelef & Azov, 1987). The first
experiences in Spain were carried out in the
extinct Centro de Investigaciones del Agua in
La Poveda (Arganda del Rey), where studies on
algal biomass production were carried out in
outdoor pilot plants (Velasco et al. 1988).
Studies on this topic continued into the early
nineties at the Polithecnic University of
Barcelona (UPC; García J., 1996). Our experiences on HRAP systems followed the experimental layout and pilot plant characteristics previously developed at the UPC.
The complete description of performances
and details on the experiments and their seasonal and diurnal patterns are described in
González 1999 and González et al. (1994,
1999, 2000, 2001) and Araki et al. (2000,
2001). Experiments tested several hydraulic
detention times (10, 5 and 3 days) under both
summer and winter conditions. A water velocity of 15 cm/sec could be achieved in the
system by using paddle wheels at a rotational
speed of 3 r.p.m. A depth of water of 30 cm was
maintained constant during the study. The best
results for COD, BOD 5 and solids removal
were obtained with a hydraulic detention time
of 3 days. Algal populations were dominated by
blooms of Monoraphidium contortum followed
by Scenedesmus spp. HRAP was therefore considered an adequate system for the wastewater
treatment in the León area, being potentially
149
appropriate for the treatment of heavily polluted wastewater such as that from pig farms,
which are abundant in the area.
The pathogen removal mechanism in
algae-systems. The role of physico-chemical
conditions
Evidence on disease transmission associated
with raw wastewater reuse, points most strongly
to the helminths and protozoa parasites as the
number one problem, with only limited transmission of bacterial and virus disease (Shuval
1991). The very low doses required to produce
infection (Boutin 1982), their cosmopolitan distribution and their long persistence in the environment (Feachmen et al, 1983) make these
organisms the main problem in wastewater
reuse, mainly in rural areas. Because of their
tough wall, these parasites are also extremely
resistant to disinfection with chlorine, monochloramine and many other chemicals (Campbel
et al. 1995, Fayer et al. 1996). In recent years,
various studies in different countries have shown
that one of the most important waterborne pathogen is Cryptosporidium parvum, a protozoan parasite causing diarroheal disease in a
wide range of vertebrates including humans
(O’Donoghue 1995). International guidelines
strongly recommend the use of low-cost, highly
efficient pathogen removal systems for wastewater treatment and stabilization ponds systems are
the most efficient for their removal (Schwardzbord et al. 1989, Shuval 1991). High rate algal
ponds follow basically the same removal mechanisms than stabilization ponds. Both are highly
efficient according to faecal bacteria indicators
(García & Bécares 1997). However, the precise
mechanisms underlying the removal of parasites
in these systems are still unclear (El Hamouri et
al. 1994). Several experiments were carried
out at the Experimental Field with an aim to discriminating the role of ionic conditions as a
mechanism for helminths and protozoa parasites
decay, excluding other removal factors like sedimentation or predation. Further details on the
experiments and the results obtained are presented in Araki et al. (2000, 2001).
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The viability of Cryptosporidium parvum oocysts
and Parascaris equorum eggs were studied in two
experimental pilot-scale High Rate Algal ponds
(HRAP) working in parallel during hydraulic
retention time (HRT) of 3 and 10 days. Semi-permeable bags of cellulose (14000 daltons pore
size) were used to determine the effect of the
physico-chemical environment on the infectivity
of the oocysts. Semi-permeable bags only allow
the exchange of small ions and water between the
reactor and the eggs. Viability was tested using
the method used by Caseres et al. (1987) for helminths and neonatal NMRI mice inoculated intragrastricaly with treated and untreated C. parvum
oocysts. Results revealed that inactivation of
oocysts by HRAP was higher than in conventional wastewater treatment systems. The HRAP
physico-chemical conditions were responsible for
more than 97 % of the reduction of infection
cases in mice. The lack of differences between the
two retention times tested suggests that oocysts
lose their infectivity shortly after contact with the
water environment. With regard to the helminths,
a 60 % reduction in viability was achieved after
4 days exposure to conditions in the HRAP, reaching a 90 % reduction after 10 days. The effect
of ions and general osmotic conditions on the viability of nematode eggs under the HRAP water
environment was responsible for 50-60 % of egg
mortality. This implies that mortality due to the
ionic environment in the HRAP could be more
important than physical retention and other removal factors potentially involved.
CONCLUSIONS
Main outcomes after ten years of research can
be summarized as follows: i) Plants have a significant effect on wastewater treatment when
treating diluted wastewater, and are a necessary
element of the treatment system. Nevertheless,
their role depends on the type of treatment process. Plants are relevant to nutrient removal but
not to that of organic matter in subsurface flow
systems, whereas the opposite occurs in freewater surface systems, ii) the type of species
used in FWS systems is not important; all spe-
cies had good performances, iii) plants did not
show a direct effect on bacteria removal but
demonstrated an indirect effect through the
creation of an hostile rhizosphere environment
in turn stimulating the growth of bacterivorous
populations, iv) high rate algal ponds are an
adequate technology for wastewater treatment,
specially when bacteria removal is an important
treatment objective, v) physico-chemical conditions created by the metabolism of algae is the
main mechanism responsible for the inactivation and decay of bacteria and parasites cysts.
ACKNOWLEDGEMENTS
This paper is especially dedicated to Félix Soto,
Juan Manuel González and My Smail Araki,
whose hard and enthusiastical work made possible the functioning of the Experimental Field. It
is a tragedy of our university system that reasons
other than scientific excellence very often prevail, driving away good researchers. The
University of León established an agreement
with the Diputación Provincial entitled: Estudio
comparado de diferentes sistemas de depuración
de bajo coste. Asesoramiento y análisis de la eficacia del Plan de Saneamiento de la Provincia
de León. I would also like to thank the official
director of this agreement, Dr. Estanislao Luis
Calabuig, for his good management of institutional relations, and all students and University staff
which were involved in the project, and specially
my thanks to Gemma Ansola, Mercedes García,
Gloria López, María J. López and Camino and
Margarita Fernández-Aláez. This project was
funded by the Diputación Provincial de León
thanks to the good insight of its former President
D.Alberto Pérez, and also thanks to the dedication of D. César Roa and D. Jaime Martino, who
invested in the “frog’s pond” much of their valuable time during the years of the study.
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Limnetica, 25(1-2): 155-170 (2006)
Diversity of patterns and processes in rivers of eastern Andalusia
J. Jesús Casas1, Mark O. Gessner2, Peter H. Langton3, Demetrio Calle4, Enrique Descals5
y María J. Salinas1
1Departamento
de Biología Vegetal y Ecología, Universidad de Almería, 04120-Almería.
of Aquatic Ecology, Swiss Federal Institute of Aquatic Science and Technology (Eawag), and
Institute of Integrative Biology (IBZ), ETH Zurich, Suiza
3Cambridge University Museum, Cambridge, Reino Unido
4Instituto de Enseñanza Media “Los Manantiales”, Torremolinos, Málaga.
5Instituto Mediterráneo de Estudios Avanzados (IMEDEA), CSIC/UIB, Palma, Mallorca
Corresponding autor: J. J. Casas ([email protected])
2Department
ABSTRACT
We document the outstanding diversity of fluvial ecosystems in eastern Andalusia, mostly attributable to the high environmental
heterogeneity of the region. The area’s altitudinal and climatic gradients are among the most pronounced in the Iberian Peninsula,
and together with a concomitant high variability in geological characteristics and human impacts, result in a noticeable heterogeneity of the rivers’ thermal regime, discharge regime and chemical properties. Fluvial communities respond to this spatial heterogeneity with marked qualitative and quantitative changes among rivers and along the upstream-downstream continuum, generally
exhibiting a great decrease in taxonomic and functional diversity as human impacts increase towards the lower reaches.
Discharge fluctuations add heterogeneity on the temporal scale and are an additional essential determinant of biological diversity.
Climatic, geological and hydrological characteristics profoundly affect the structure of the riparian vegetation, which in turn
strongly conditions the community structure of benthic macroinvertebrates and organic matter turnover in fluvial ecosystems.
Key words: rivers, eastern Andalusia, environmental gradients, human impacts, macroinvertebrates, chironomids, riparian
vegetation, leaf litter decomposition, aquatic hyphomycetes.
RESUMEN
Se ilustra la notable diversidad de ecosistemas fluviales de Andalucía Oriental, atribuible a la gran heterogeneidad ambiental
de esta región. Gradientes altitudinales y climáticos de los más pronunciados de la península Ibérica, concomitantes con una
gran variedad de condiciones litológicas y de impactos humanos, acentúan en esta región la heterogeneidad térmica, de caudal y calidad química de los ríos. Las comunidades fluviales responden a esta heterogeneidad espacial con profundos cambios
cualitativos y cuantitativos, y generalmente con una disminución de la diversidad taxonómica y funcional en respuesta al
incremento de impactos humanos hacia los tramos bajos de los ríos. Las fluctuaciones de caudal suman heterogeneidad en la
dimensión tiempo, y constituyen un determinante esencial de los patrones de diversidad biológica. La estructura de la vegetación de ribera responde con grandes cambios a las condiciones climáticas, litológicas e hidrológicas, y ésta a su vez condiciona sustancialmente el marco trófico del ecosistema fluvial.
Palabras clave: ríos, Andalucía oriental, gradientes ambientales, impactos humanos, macroinvertebrados, quironómidos,
vegetación de ribera, descomposición de hojarasca, hifomicetos acuáticos.
INTRODUCTION
Margalef (1960, 1983) highlighted as an essential
characteristic of fluvial ecosystems their large
interfaces with the surrounding landscape. This
means that the climatic, geomorphic and biological setting of river basins strongly condition the
structural and functional traits of river communi-
ties (Hynes, 1970; Allan, 1995). This idea was
behind Gasith & Resh’s (1999) proposal of a
functional convergence of rivers under
Mediterranean climate. The most general characteristic of the Mediterranean climate is the strong
seasonality of precipitation, which tends to be
concentrated in a few events mostly occurring
during winter, and with high variability between
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years. According to Gasith & Resh (loc. cit.), this
pattern determines an annual sequence of predictable flood events (autumn-winter), during which
the fluvial ecosystem is under abiotic control,
followed by periods of stable flow (spring-summer) during which biotic control dominates.
Although broadly speaking the above-mentioned convergence may occur, these authors and
others (e.g. Blondel & Aronson, 1999) also
recognise the great climatic diversity of the
Mediterranean basin, together with the considerable heterogeneity of geomorphic features and
human uses of rivers and surrounding land,
which make generalisations difficult. For instance, both the highest and lowest precipitation in
Europe occur in the Mediterranean basin, namely
in the coast of Montenegro (4640 mm) and
Almería (150-200 mm) (Grove & Rackham,
2001). At almost any spatial scale of observation
the Mediterranean basin appears as a remarkably
diverse mosaic, due to the extraordinary heterogeneity of topography, climate, geology and past
and present human actions on the landscape
(Monserrat, 1998; Blondel & Aronson, 1999).
The present review aims to document, from a
regional perspective, the diversity of ecological
patterns and processes in rivers of eastern Andalusia, one of the most mountainous European
regions. The area includes the Sierra Nevada, the
southernmost high mountain range on the continent, which strongly influences the surrounding
areas. We also aim to document the consequences
of human impacts on these rivers. The results
summarised here come from studies we have
carried out in this region over the last two decades.
ENVIRONMENTAL CHARACTERISTICS
Physical, geological and climatic heterogeneity
Eastern Andalusia includes a wide range of geological and climatic conditions. The Betic mountain
ranges occupy most of the area, shaping an
extraordinarily uneven relief. The mountain massif of the Sierra Nevada stands out with several
peaks above 3000 m a.s.l., but several other sierras also reach altitudes above 2000 m a.s.l. (e.g.
Cazorla, Mágina, Tejeda, Filabres-Baza, Gador).
The geology of the sierras is diverse, including
Palaeozoic metamorphic rocks, particularly in the
upper layers of the Sierra Nevada and FilabresBaza, and mostly calcareous rocks, marble and
vast dolomitic outcrops extending mainly into the
mid to low altitudes of most sierras. A peculiar
trait of the Betic mountain ranges is the high
abundance of calcareous sierras with a varied
karstic relief, also including a gypsum karst
(Durán & López, 1999). This geologic diversity
adds physical heterogeneity to the river courses of
the area. Fluviokarstic canyons, for example, are a
frequent landscape feature, particularly in the
foothills of the sierras. Lowlands made up of continental, and frequently marine, sediments deposited during the Neogene period separated the sierras. These lowlands often comprise terrain
composed of marl rich in sulphate and chloride.
The greatly compartmentalized relief of the
Betic ranges favours high spatial heterogeneity
in precipitation. Mean annual rainfall ranges
from over 1000 mm at the heart of Sierra de
Cazorla, or over 800 mm on the peaks of Sierra
Nevada or Sierras de Tejeda-Almijara, to less
than 200 mm in Cabo de Gata (Capel-Molina,
2000). Cyclonic events from the northwest or
west are the most frequent, and provide most of
the rainfall in the region, particularly on the western mountain slopes. On the eastern sides, especially of the Sierra Nevada owing to its high altitude, the descending cool and dry air tends to
accentuate arid conditions (Föhn effect), e.g. in
the subdesert of Tabernas. Cyclonic events from
the east are less frequent but bring a great proportion of the rainfall to the eastern part of the
region, mainly during storms towards the end of
summer and autumn (Castillo-Requena, 1981).
Heterogeneity of discharge regimes: natural
and human constraints
The annual hydrograph of rivers in eastern
Andalusia generally fits to the Mediterranean
pluvial subtropical type, which is characterised
by very low flow in summer and peak discharge
during winter (Fig. 1a). This regime is distinctly
modified (i.e. peaking slightly during summer)
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Diversity in rivers of eastern Andalusia
157
Figure 1. Discharge regimes in several Betic rivers: a) upper Río Guadalquivir; b) Río Monachil in the Sierra Nevada; c) Río de
Aguas in eastern Almería. Regímenes de caudal registrados en distintas cuencas béticas: a) Alto Guadalquivir; b) Río Monachil en
Sierra Nevada; c) Río de Aguas en el levante almeriense.
by reservoir regulation, particularly in the upper
Río Guadalquivir, and mainly for agricultural
purposes (Fig. 1a) (Calle et al., 1990). The abundance of groundwater sources up-welling from
large karstic aquifers is noteworthy. It tends to
confer rather constant discharge to rivers during
summer. In fact, most permanent reaches in
rivers from the most arid zone of the region
(Almería) are fed by these karstic sources. The
annual hydrograph of the rivers from the Sierra
Nevada merit special attention in that it exhibits a
pronounced snowmelt influence peaking between
March and May (Fig. 1b) (Casas, 1990). The
scarcity of rainfall and extreme concentration in
a few events, together with the high permeability
of the calcareous and/or evaporitic bedrocks, are
responsible for the high percentage of temporary
water courses in the fluvial networks. This ten-
Figure 2. Thermal regimes in several streams located in the Sierra Nevada (Río Monachil at 2100 m a.s.l. and Río Jérez at 1280 m a.s.l.)
and eastern Almería (Río de Aguas at Molinos, 265 m a.s.l., and Perales, 220 m a.s.l.). Río Monachil: maximum and minimum daily temperatures on 5 different dates in the year. Other streams: maximum and minimum monthly temperatures. Régimen térmico en varios ríos de
cabecera de Sierra Nevada (Río Monachil y Río Jérez) y del litoral semiárido almeriense (Río de Aguas). Río Monachil: temperatura máxima-mínima diárias de 5 dias. Restantes ríos: temperaturas máximas-mínimas mensuales.
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dency is most noticeable in the semiarid east of
the region and in most lowland reaches, where
water abstraction for agriculture and human
settlements dries up many river courses (Fig. 1c).
Physico-chemical variability of water quality:
the influence of altitudinal gradients and
geological and human factors
Figure 3. Results of principal component analyses (PCA) of
physico-chemical characteristics of three river ecosystems: a) Río
Monachil from the Sierra Nevada (Casas, 1996), b) the upper Río
Guadalquivir (Calle et al., 1990), and c) Ríos Almanzora and
Aguas in eastern Almería (Vivas et al., 2001). The percentage of
variance explained by the PCA axes is indicated in parentheses.
A third axis is presented for the rivers in eastern Almería.
Resultados del análisis de componentes principales (ACP) para
las características físicas y químicas de las aguas de tres ecosistemas fluviales: a) Río Monachil en Sierra Nevada (Casas,
1996), b) Alto Guadalquivir (Calle et al., 1990), y c) cuencas de
los ríos Almanzora y Aguas en el levante almeriense (Vivas et al.,
2001). El porcentaje de varianza explicada por cada eje se indica entre paréntesis. Para los ríos almerienses se indican también
los resultados del tercer eje extraido por el ACP.
The thermal regime of rivers is considered a primer
determinant of diversity patterns of aquatic communities along altitudinal-longitudinal gradients
(Ward & Stanford, 1982; Jacobsen et al., 1997).
The pronounced altitudinal gradient of the Betic
ranges in such a southern region results in marked
differences in temperature both between and within
river basins (Fig. 2). Streams above tree line in the
Sierra Nevada exhibit the lowest temperatures,
with minimum values around 2 ºC and summer
maxima below 13 ºC (Fig. 2a) (Casas, 1990), not
much unlike other alpine streams (Milner et al.,
2002). The middle and foothill reaches of the rivers
often have closed canopies and higher discharge.
Minimum temperatures do not usually fall below
5 ºC in these reaches, and maximum values
rarely surpass 20 ºC (Fig. 2b). In contrast, summer
temperatures in reaches flowing in the lowlands
often greatly surpass 20 ºC, particularly in the
semiarid eastern area. The Río de Aguas, which
originates at 260 m a.s.l. and less than 15 km from
the Mediterranean coast, illustrates the strong
influence of karstic springs, riparian vegetation
and discharge on the thermal regime of Mediterranean rivers: Seasonal temperature fluctuations
are minor at Molinos, less than 1 km from the source, where ground water is upwelling from the
gypsum karst of Sorbas and the cover by riparian
tress and reeds is about 80 % (Fig. 2c). The site at
Perales, in contrast, has sparse riparian vegetation
and is impacted by water abstraction for irrigation;
it is located only 2 km further downstream but
shows strong annual thermal fluctuations (Fig. 2d).
The impact of reservoir regulation often produces a net attenuating effect on temperature amplitudes. For instance, the upper Río Guadalquivir at
Charco del Aceite, 6 km downstream from the
Tranco reservoir, exhibits relatively narrow diel
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159
Figure 4. a) Percentages of species of the major taxonomic groups of macroinvertebrates present (except Diptera) with different
biogeographical distribution: Transiberic species (present in Europe, the Iberian Peninsula and North Africa); northern species
(transiberic species absent from North Africa); southern species (transiberic species absent beyond the Pyrenees). b) Family richness and c) species richness of the main taxonomic groups of macroinvertebrates combined for the two river basins studied (total)
and for each basin separately. a) Porcentajes de especies (excepto dípteros) según su distribución biogeográfica: Transibéricos,
presentes en Europa, Peninsula Ibérica y Norte de África; Norteños, transibéricos que no alcanzan el Norte de África; Sureños,
transibéricos que no traspasan los Pirineos. b) Riqueza de familias, y c) riqueza de especies, de los principales grupos taxonómicos en las dos cuencas almerienses estudiadas (total) y en cada una por separado.
thermal oscillations and less extreme minima
(10-11, 13-14, 17-21 and 18-21 ºC, respectively,
in December, March, June and August) compared
to the Santo Tome site, located 45 km downstream from the same reservoir (10-15, 9-21
and 7-12 ºC, respectively, in January, April and
November) (Calle et al., 1990).
On the river basin scale, the increase in ionic
content from the headwaters to the lowland reaches is the most evident pattern of spatial variation in the rivers of eastern Andalusia (Fig. 3), as
is commonly found in rivers worldwide (Margalef, 1983; Allan, 1995). This pattern may mainly
be attributable to a marked altitudinal transition
of geological formations: e.g. from siliceous to
calcareous rocks in the Sierra Nevada (Río
Monachil, with electrical conductivity ranging
from 33-358 µS cm-1) (Casas, 1996); from calcareous to marl rocks in the upper Río Guadalquivir
(conductivity ranging from 376-2960 µS cm-1)
(Calle et al., 1990); or from siliceous to evaporitic
rocks in the Río Almanzora (conductivity ranging
from 150-8000 µS cm-1) (Vivas et al., 2001)
(Fig. 3). Evaporation may also be important. In
the intermittent Río de Aguas, for example, several permanent pools remaining during summer
reach conductivity values of over 11 mS cm-1.
In the western catchments, such as in the upper
Rio Guadalquivir (Fig. 3b), the longitudinal gra-
dient is more strongly related to increasing nitrogen concentrations downstream than to ionic content of the river water in general. This is due to
prominent diffuse and point-source contamination
caused by intensive agricultural activities and
abundant human settlements in the olive-growing
areas of Jaén. In this river, discharge shows a net
increase downstream despite important water abstractions, probably due to the high number of relatively large tributaries. In the olive countryside of
Jaén, as in many lowland reaches of western
Andalusia, pollution by suspended solids shows
strong seasonality. During winter-spring, surface
runoff transports clay to the river from the nearby
olive orchards, and there are also frequent inputs
of olive-mill waste (alpechín), which increase the
concentration of suspended solids and oxygen
demand (axis 2 in Fig. 3b). Another notable feature of this river, which is widespread also in other
calcareous sierras, is the high carbonate concentration in the headwaters (Fig. 3b), which makes
calcite precipitation a common event and may
lead to important tuff formation (Calle et al.,
1990; Casas et al., 1994; Casas & Gessner, 1999).
In eastern basins (Almería) the fluvial continuum is normally fragmented as a result of the
arid climate and impoundments built for agricultural irrigation. Lowland reaches only carry
water after marked rainfall events or where
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springs generate isolated permanent or temporary pools. In these fluvial networks, wastewater
pollution does not continuously increase downstream (axis 2 in Fig. 3c), but is rather limited to
a few sites. This is probably because wastewaters
that would normally be discharged into them are
often diverted into the aquifer. Additionally, the
human population density here is lower than in
the western basins. In the eastern semiarid
basins rainfall events are the main source of discharge variability, which is positively related to
nitrate concentrations, probably due to inputs
originating from agricultural soil lixiviation
(axis 3 in Fig. 3c) (Vivas et al., 2001).
MACROINVERTEBRATE DIVERSITY
PATTERNS: NATURAL AND HUMAN
CONSTRAINTS
The Mediterranean basin, particularly the Betic
region, is a biodiversity hot spot on the global
scale (Médail & Quézel, 1999; Myers et al.,
2000). The extraordinary species richness of the
area may be attributable to its present environmental heterogeneity, its special geographical
location, and the diverse climatic and geological
changes in the past, which have led many taxa
from different biogeographical origins to colonise, find refuges and/or radiate (Blondel &
Aronson, 1999). Two fluvial basins in Almería
illustrate the variety of biogeographical origins and distributions of macroinvertebrate taxa
(Fig. 4a) (Vivas et al., 2002; Vivas, 2003). In both
river systems, the dominant taxa have a transiberic distribution (Central Europe-North Africa),
but taxa endemic to the Iberian Peninsula were
also relatively abundant (10 %). The main difference between basins, Almanzora vs. Aguas, lies
in the high percentage of taxa with a northern distribution (i.e. taxa absent in North Africa). These
taxa are present in mountain sites of the Río
Almanzora basin, but are absent in the lowland
basin of the Río de Aguas. Plecoptera, Ephemeroptera, Trichoptera, Odonata and Coleoptera species, which during their larval stages or, in the
case of Coleoptera, during their entire life cycle
require cool water and moderate hardwater condi-
tions which characterise the mountain headwaters
of the Río Almanzora (Fig. 4 b and c). In the Río
de Aguas, on the other hand, the relatively high
temperature and ionic content, plus the abundance
of pools with aquatic vegetation, provide favourable habitats for Heteroptera and Crustacean.
On the basin scale of the rivers Almanzora and
Aguas, macroinvertebrate communities are primarily structured by the longitudinal increases in
salinity and temperature towards the lower reaches (Vivas, 2003) (Table 1), similar to the trend
observed in 12 basins studied across the Spanish
Mediterranean region (Vivas et al., 2002).
Furthermore, in the semiarid basins this longitudinal gradient was also negatively correlated with
discharge and riparian cover by deciduous woody
vegetation, which greatly diminishes towards the
lower reaches. Although the overall taxonomic
richness of benthic macroinvertebrates is not significantly correlated with these longitudinal
changes in environmental conditions, the richTable 1. Spearman correlations of environmental variables and taxonomic richness, with Axes 1 and 2 determined by a canonical correspondence analysis of the environmental variables-macroinvertebrate
taxa matrix. Level of significance, *p < 0.05. Correlaciones de
Spearman de variables ambientales y riqueza taxonómica en lo ejes 1
y 2 extraídos por un análisis canónico de correspondencias ambiente-taxones de macroinvertebrados. Nivel de significación, * p < 0,05.
CCA Axis 1
Environmental variables
Electrical conductivity
Temperature
Discharge
Hydroperiod span
Index of physical impacts
Deciduous vegetation cover
Emergent macrophytes cover
0.85
0.46
-0.44
-0.20
0.14
0.80
0.49
Taxa richness
Total richness
Ephemeroptera
Plecoptera
Trichoptera
Odonata
Coleoptera
Heteroptera
Diptera
Mollusca
Crustacean
0.01
-0.40
-0.61
-0.30
0.44
0.30
0.58
0.36
0.12
0.46
*
*
*
*
*
*
*
*
*
*
*
*
*
CCA Axis 2
-0.10
-0.10
-0.61
-0.75
0.83
-0.02
-0.64
*
*
*
-0.79
-0.65
-0.31
-0.66
-0.75
-0.59
-0.46
-0.39
-0.55
-0.32
*
*
*
*
*
*
*
*
*
*
*
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161
Table 2. Transformed values of basin γ-richness (i.e. γ-richness / ln number of samples), with untransformed values shown in parentheses; average local α-richness; and average basin β-richness calculated as the percentage dissimilarity between pairs of sites (Sørensen index). The altitudinal range and the proportion of temporary study sites are also shown. Betic basins are shaded. Data extracted form GUADALMED-1 database (modified from Vivas, 2003). Riqueza de cuenca, valor transformado (γ-riqueza / ln nº de muestras), y valor absoluto entre parentesis.
α-Riqueza media por localidad y β-riqueza media de la cuenca, calculado como el porcentaje de disimilitud entre pares de localidades (índice de Sorensen). Se muestra el intervalo altitudinal y el porcentaje de localidades temporales estudiadas. Las cuencas béticas se han sombreado. Datos tomados de la base de datos GUADALMED-1 (modificado de Vivas, 2003).
Fluvial basin
Adra
Segura
Llobregat
Besos
Jucar
Turia
Mijares
Soller
Almanzora
Aguas
Pollença
Guadalfeo
γ-Richness
23 (85)
22 (102)
20 (94)
20 (93)
20 (84)
20 (77)
20 (71)
18 (51)
17 (74)
17 (64)
17 (60)
16 (76)
α-Richness
β-Richness
Altitudinal range (m)
% Temporality
45
46
40
37
31
27
30
24
28
34
30
41
0.43
0.34
0.34
0.36
0.35
0.33
0.39
0.40
0.38
0.39
0.41
0.36
2060
1413
1200
556
1505
1520
1500
275
1270
270
325
2980
0
22
7
40
0
21
45
100
83
75
90
19
ness of particular groups changes markedly, with
Ephemeroptera, Plecoptera and Trichoptera
being gradually replaced by species from other
groups towards downstream sites characterised
by high summer temperatures and salinity (Table 1). A second environmental gradient is positively related to the shortening of the hydroperiod
and extent of anthropogenic impact on channel
morphology, which both have negative effects on
total macroinverterate richness and on richness
within specific taxonomic groups (Table 1).
Temporary reaches with short hydroperiods and
close to human settlements are frequently used
by farming vehicles. The resulting high frequency and magnitude of impacts on river channels reduces physical heterogeneity of aquatic
habitats and impairs water quality.
A comparative analysis of macroinvertebrate
family studied among 12 Mediterranean basins
revealed an unusually large heterogeneity of
diversity patterns in the Betic basins. This result
may be mainly due to the wide range of altitudes
and temporality of the studied aquatic habitats
(Table 2) (Vivas, 2003). Accordingly, γ-richness
of macroinvertebrates at the basin scale showed a
significant positive correlation with altitudinal
range and a significant negative correlation with
the percentage of temporary aquatic sites
(rs = 0.72 and rs = -0.78, respectively, p < 0.05),
whereas it was not significantly correlated with
basin surface area as would be expected from the
general species-area relationship. The permanence of fluvial habitats appears to affect local
α-richness positively, as suggested by higher than
average α-richness in the Río Guadalfeo and Río
Adra in the Sierra Nevada, where all sites studied
were permanent. Moreover, the great altitudinal
gradient in these basins appears to favour a relatively high β-richness. It is well known that the
wider the altitudinal range in a basin, the greater
the habitat heterogeneity, with corresponding
positive effects on taxa richness on this scale
(Giller & Malmqvist, 1998; Jacobsen, 2004). On
the other hand, basins with a high percentage of
temporary sites generally exhibited relatively low
values of local α-richness, although this effect
may be somewhat offset by high basin β-richness
(e.g. Almanzora and Aguas) (Table 2). The intermittent discharge regime of Mediterranean rivers,
whether due to natural or human causes or both,
induces fragmentation of the fluvial continuum,
which in turn reduces longitudinal connectivity
(e.g. Gasith & Resh, 1999; Boulton, 2003; Lake,
2003). This process tends to diversify fluvial habitats among reaches that are hydrologically isolated, possibly by favouring a stronger influence of
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Figure 5. Dynamics of chironomid species from March 1996 to March 2000 (fortnightly sampling intervals) in two mountain streams of the Sierra de Albuñuelas (southern Granada) as a function of time elapsed since the last spate. Species number per sample
was transformed according to the richness index proposed by Margalef (1982). The plot of the upper right panel takes into account
the species list of the previous period of stable flow during the long-lasting drought from 1991-1995. For the temporary stream
the first study year (1996, closed circles) was distinct from later years (1997-2000, open circles). Dinámica de las especies de quironómidos durante el periodo marzo-1996 a marzo-2000 (muestreos quincenales) en dos arroyos de cabecera de la Sierra de
Albuñuelas (sur de Granada) en función del tiempo transcurrido desde la última avenida. El número de especies por muestra fue
transformado según el índice de riqueza de Margalef (1982). En el gráfico de nuevas especies colectadas, para el río permanente se
ha tenido en cuenta la lista de especies del periodo previo con caudal estable durante la prolongada sequía (1991-1995); para el río
temporal se ha diferenciado el primer año de estudio (1996, circulos cerrados) de los posteriores (1997-2000, circulos abiertos).
local factors as opposed to upstream basin factors,
and also by suppressing downstream drift.
Chironomid species richness and indicator
value
Diptera, particularly Chironomidae, are an ubiquitous, species rich and the often most abundant group of macroinvertebrates in fluvial ecosystems. However, this group is rarely identified
to species or genus level in ecological studies of
fluvial benthos, due to the difficulty of identifying larvae. Chironomid pupal exuviae are
easier to identify thanks to their more distinctive
morphological features. Using this approach, the
chironomid communities of rivers in the Sierra
Nevada and the upper Río Guadalquivir in the
Sierra de Cazorla were compared in order to
determine the indicator value of this group at
species level. The study revealed high species
richness of chironomids in the two regions (total
of 204 species) and the group’s remarkable value
for the biological classification of rivers (Calle
& Casas, 2006). Two major groups of rivers were
distinguished when using chironomid species as
association criteria, clearly reflecting the two
geographical locations. Obviously, there are pronounced enough differences between the
rivers of both regions, including higher slopes
and altitudes in the Sierra Nevada coupled with
geological and hydrological differences, for
the rivers to harbour distinctive chironomid
communities. Interestingly, however, the sites
studied on the Río Aguas-Blancas (Sierra
Nevada) showed local characteristics close to
those of rivers in the headwaters of the upper Río
Guadalquivir (gentle slope, pluvial discharge
regime and travertine precipitation), but faunistically were closer to the Sierra Nevada rivers.
This suggests a preponderant role of regional
rather than local factors determining chironomid
species distribution. This pattern may well
reflect a strong geographical “proximity effect”,
sensu Hawkins et al. (2000), on the source-sink
dynamics of midge populations, coupled with
considerable aerial colonising capacities of
adults and the strong seasonal dynamics of fluvial ecosystems in eastern Andalusia.
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Temperature and especially the extent of water
pollution were the primary determinants of chironomid community composition within both
regions. Six chironomid species showed a significant indicator value for sites with impaired
water quality (Chironomus riparius, Eukiefferiella claripennis, Paratrichocladius rufiventris,
Rheocricotopus chalybeatus, Cricotopus bicinctus, and Tvetenia calvescens). The indicator
values of these species cover a wide gradient of
water quality impairment and occurrence of the
species did not depend on the regional setting,
whether in eastern Andalusia or elsewhere in
European rivers, suggesting great potential for
biomonitoring purposes (Calle & Casas, 2006).
Chironomid diversity and hydrological
disturbance
The highly irregular rainfall among years is characteristic of Mediterranean climate, determining
extreme discharge fluctuations, which in turn are
considered the key factor for fluvial ecosystem
dynamics (e.g. Boulton et al., 1992; Sabater et al.,
1992). The southern part of the Iberian Peninsula
was affected by drought from 1992-95, ending with
heavy rains in winter 1995-96. These events caused
profound changes in the chironomid communities
in two first-order Betic mountain streams, one permanent and the other temporary. From 1991-95 the
permanent stream showed no peaks in discharge,
which decreased progressively till December 1995,
and the chironomid assemblage was dominated by
species of the subfamily Tanytarsini. The temporary stream had no surface water flow during these
four years. Heavy rains in winter 1995-96 ended
the drought and caused extraordinary discharge
peaks and transport of benthic materials in both the
permanent and temporary stream. Once the river
bed had become more stable the dominant species
were those typically associated with torrential lowtemperature mountain streams, belonging to the
genera Diamesa, Eukiefferiella and Orthocladius
in both streams (Langton & Casas, 1999).
In contrast to several other studies (e.g. Williams, 1996; Rüegg & Robinson, 2004), which
have found lower macroinvertebrate diversity in
temporary aquatic habitats compared to perma-
163
nent ones, the intermittent and permanent stream
showed a similar overall chironomid species richness and mean richness per sample. This
unexpected resemblance may have been due to the
long period of surface water permanence in
the temporary stream (1996-2000) following the
drought in the first half of the 1990s. In addition,
the temporary stream showed a greater temporal
β-richness, a feature that could be attributable to
higher temporal thermal variability and substrate
heterogeneity compared to the permanent stream.
The observed pattern of species richness variation
as a function of time elapsed since the last spate is
in general agreement with the patch dynamics
concept in streams (Townsend, 1989). In both streams, spates favoured records of new species but
hardly affected sample α-richness (Fig. 5). On the
other hand, prolonged periods without spates (>1
year) favoured records of new species in the permanent stream, but in the temporary stream this
led to a disruption of flow an isolation of a few
remaining pools which harboured an extremely
impoverished chironomid assemblage (Fig. 5).
Apart from the similar overall chironomid
γ-richness, the two examined streams also exhibited similar mean values of species rarity and
notable species assemblage complementarity
(20 %). These results corroborate other studies
which have stressed the high conservation value
of temporary aquatic habitats, so frequently
neglected, and the need to include them in plans
for nature conservation. Although headwater
streams in the Mediterranean are typically much
less impacted by human actions than lower river
reaches, agricultural activity in the catchments
of headwaters is increasing, promoted in part by
EU subsidies. In eastern Andalusia, this is particularly true for the expansion of olive groves
and almond orchards, which is likely to exacerbate impacts on temporary stream habitats and
diversity of the associated aquatic fauna.
RIPARIAN VEGETATION, RESOURCE
USE AND THE FLUVIAL FOOD WEB
Fluvial ecosystems are landscape elements characterised by intense interactions between the
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Figure 6. Results of a canonical correspondence analysis (CCA Axes 1 and 2) based on a biotic matrix of tree and shrub species
(percentage cover data) and an abiotic matrix of fluvial environmental variables. Only the most abundant species and the environmental variables significantly correlated with the axes shown. The box in the top left-hand corner shows the ordination of macroinvertebrate functional feeding groups (relative abundance data) resulting from a CCA using the same environmental variables as the
CCA for the vegetation: FI = Filterers, C = Collector-gatherers, SR = Shredders, GR = Grazers, SC = Scrapers, P = Predators.
Resultados de un análisis canónico de correspondencias (ACC ejes 1 y 2) de las especies de árboles y arbustos (datos de cobertura) constreñidas por variables ambientales del río. Se indican las especies más abundantes y las variables ambientales (vectores)
con correlación significativa con alguno de los ejes. Arriba a la izquierda se indica la ordenación de los grupos tróficos funcionales de macroinvertebrados (datos de abundancia relativa) resultante de un ACC utilizando las mismas variables ambientales: FI =
filtradores, C = colectores, SR = fragmentadores, GR = ramoneadores, SC = raspadores, P = depredadores.
river channel and riparian vegetation (Malanson,
1993). The climatic setting, hydrological dynamics and channel structure are of prime importance for riparian vegetation, which in turn determines essential traits of the fluvial benthic
community (e.g. Gregory et al., 1991; Naiman et
al., 2005). A study carried out in the rivers
Almanzora and Aguas (Almería) illustrates how
strong environmental gradients in these river
basins affect the structure of riparian vegetation
(M.J. Salinas & J.J. Casas, unpublished data). As
for macroinvertebrate communities (see above),
the altitudinal-longitudinal gradients determine
the major structural changes of vegetation: a
decrease in richness and percentage cover of
deciduous species downstream and an increase in
emergent macrophytes, phreatophytes and halophytes or halotolerant species. These changes
appear to be related to the hydrological decay of
rivers and a concomitant increase in salinity
towards the lowland reaches under more arid climate (Fig. 6). Although this is a common natural
pattern in arid zones (Jacobson et al., 2000),
water impoundments of Mediterranean rivers
exacerbate the situation, especially in the middle
and lower reaches (Gasith & Resh, 1999). A
second environmental factor is the downstream
increase in physical anthropogenic impacts on
the channel, in parallel to the shorter hydroperiods and concomitant with a reduction in species
richness and percentage stream cover by riparian
vegetation, except by ruderal and nitrophilous
species such as Artemisia barrelieri and the
phreatophyte Tamarix canariensis (Fig. 6).
The trophic structure of the benthic community broadly matched the pattern of riparian
vegetation variation (Fig. 6), in agreement with
the “river continuum concept” (Vannote et al.,
1980), at least in terms of resource-consumer
relationships. However, mechanisms underlying
the pattern in Mediterranean streams may differ
substantially from those proposed by Vannote et
al. (1980). Occurrence of macroinvertebrate
shredders clearly correlated with high vegeta-
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Figure 7. Percentage cover (mean + 1 SE) of two types of riparian vegetation at 3 Sierra Nevada sites and 4 semiarid sites of
Almería. The second y-axis represents benthic coarse particulate organic matter (CPOM > 1 mm) measured in December 1999
(unpublished data). AFDM = ash free dry mass. Porcentajes de
cobertura (valores medios + 1 error típico) de la vegetación
riparia diferenciando estratos (árboles vs. arbustos y helófitos)
para 3 localidades de Sierra Nevada y 4 de la región semiárida
almeriense. El eje “y” derecho representa la materia orgánica
particulada gruesa (CPOM > 1 mm) del bentos medida en
diciembre de 1999 (datos no publicados). AFDM = Peso seco
libre de cenizas.
tion cover by deciduous woody vegetation
and/or emergent macrophytes at sites with permanent surface water flow. Similarly, the relative abundance of scrapers was associated with
permanent headwater sites of the Río de Aguas,
which has low discharge, abundant pool habitats, and intermediate cover by emergent
macrophytes such that ample light reaches the
stream bottom and allows algae and submerged
macrophytes to develop abundantly. At the most
impacted sites (mainly in the middle reaches of
the Río Almanzora), with sparse riparian vegetation and short hydroperiods, the trophic structure was extremely simplified with an overwhelming abundance of invertebrate collectors (Fig.
6). These results suggest that trophic structure
of the benthic macroinvertebrate community
responds sensitively to changes in natural conditions and human disturbances in Mediterranean
rivers. Trophic characterisation of macroinvertebrates may therefore be a useful approach to
biomonitoring of these rivers, as in more humid
temperate regions (e.g. Barbour et al., 1996).
According to Statzner & Higler (1985) and
several other studies (Davies et al., 1994; Schade
& Fisher, 1997), it is unusual to observe linkages
between riparian vegetation structure, benthic
resource availability and macroinvertebrate
functional feeding groups in rivers under arid
climates. This is thought to be due to the very
165
irregular discharge regime (i.e. high spates and
severe droughts), which makes resource availability highly unpredictable. The unusual link we
observed in our two rivers might be attributable
to the absence of spates during the two-year
study. Additionally, or alternatively, it may be
argued that the possible lack of coupling on the
local scale may be masked by the strong environmental gradients at the regional scale.
Inputs of coarse particulate organic matter
(CPOM) from riparian vegetation, mainly leaf litter, frequently constitute the main source of
energy in forested headwaters. Fungi and macroinvertebrate shredders consume and transform
these materials into fine particles, a resource more
readily exploitable by bacteria and invertebrate
collectors (Wallace & Webster, 1996; Gessner et
al., 1999). Several studies suggest that in rivers
under arid climates riparian inputs are quantitatively and qualitatively less important for river food
webs than in temperate forest streams (e.g. Schade
& Fisher, 1997). A comparative study between
mountain headwaters from the Sierra Nevada and
lowland headwaters in the semiarid region of
Almería indicated, however, that despite large differences in riparian cover, the availability of
CPOM for benthic macroinvertebrates may be
relatively important even in semiarid permanent
streams. The standing stock of benthic organic
matter was not significantly different between
regions (Fig. 7), probably due to the extensive
cover of emergent macrophytes in the lowland
streams, which make up for the virtual absence of
a tree canopy. Furthermore, the retention of benthic organic matter in the semiarid region might be
greater than in the Sierra Nevada as a consequence of less frequent discharge peaks.
Aquatic hyphomycete fungi appear to make a
greater contribution to leaf litter decomposition in
the Sierra Nevada streams, as deduced from the
higher reproductive activity and biomass in streams of this region compared to streams in the
semiarid lowland streams (Fig. 8). The diversity
of aquatic hyphomycetes was also greater in the
Sierra Nevada streams. Macroinvertebrates consuming leaf litter (in contrast) were similarly
abundant and diverse in both regions (Fig. 9).
Shredders belonging to the orders Plecoptera,
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Trichoptera and Diptera dominated in the Sierra
Nevada streams, whereas scrapers, primarily the
snail Melanopsis praemorsa, were most important
and occasionally reached extremely high biomass
in the semiarid lowland streams (Fig. 9). This
snail is known to be an efficient consumer of leaf
litter (Chergui & Patte, 1991; Maamri et al., 1997;
Heller & Abotbol, 1997), feeding on leaf species
of even very low food quality. As a result, the
highest decomposition rate of leaf litter was
recorded in our study at Molinos in the Río de
Aguas, coinciding with a very rapid and extraordinarily high colonization of leaves by M. praemorsa (Fig. 9). Therefore, although autotrophic
resources may be abundant because of the virtual
absence of a tree canopy shading these semiarid
headwaters, it is likely that CPOM inputs from
riparian vegetation or emergent macrophytes are
quickly consumed. This is facilitated by a high
degree of trophic generalism of the macroinvertebrate consumers present in theses streams.
Moreover, as such inputs of leaf litter are a donorcontrolled energy source, they may constitute a
trophic alternative when autotrophic sources are
limited as a result of intense grazing and/or the
nutrient limitation of primary production.
Leaf litter dynamics in travertine streams
Calcium carbonate precipitation is a frequent,
sometimes intense, event forming travertine or
tuff in headwaters of the sierras in eastern
Andalusia, where limestone-dolomite bedrock is
abundant. Calcite precipitation may occur when
groundwater rich in carbon dioxide and calcium
is upwelling in the river channel and the outgassing CO2 shifts the CO2-carbonate equilibrium
towards CaCO3, a process that can also be caused
Figure 8. Fungal dynamics on decomposing alder leaves (Alnus glutinosa) in two contrasting streams (Sierra Nevada vs. semiarid
region of Almería). Means ± 1 SE (n = 4) are shown for each sampling date. Sporulation rates (conidia mg-1 leaf dry mass d-1) were
measured in vitro at stream temperature. Ergosterol is a fungal lipid frequently used as a proxy of fungal biomass (Gessner &
Newell, 2002). Dinámica de los hongos, aquaticos que intervienen en la descomposición de hojarasca de aliso (Alnus glutinosa)
en dos ríos con gran contraste tipológico (Sierra Nevada vs. región semiárida almeriense). Se indican valores medios y error típico (n = 4) para cada fecha. Las tasas de esporulación (conidios mg-1 de peso seco de hojarasca d-1) se midieron mediante incubación in vitro a la temperatura del agua del río. El ergosterol es un lípido específico de los hongos que se utiliza como indicador de
la biomasa fúngica en la hojarasca (Gessner & Newell, 2002).
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167
Figure 9. Dynamics of shredders and scrapers on decomposing alder leaf litter in two contrasting streams (Sierra Nevada vs. semiarid
region of Almería). Symbols in the top panels show means ± 1 SE of biomass (n = 4), symbols in the bottom panels show richness, and
relative abundances for the main shredder and scrpaer taxa, as indicated by different shading. Note the extremely high value of scraper
biomass (M. praemorsa) after only 5 days (231 mg per leaf bag). Dinámica de invertebrados (fragmentadores y raspadores) colonizadores de bolsas de hojarasca de aliso en dos ríos con gran contraste tipológico (Sierra Nevada vs. región semiárida almeriense). Los
gráficos de biomasa (mg de peso seco libre de cenizas de invertebrados por bolsa) muestran valores medios y error típico (n = 4). Los
datos de abundancia relativa y riqueza corresponden a los principales taxones de fragmentadores y raspadores representados por diferentes tramas. Nótese que en el gráfico de biomasa para la localidad semiárida el valor de raspadores a los 5 días es 231 mg por bolsa.
by CO2 removal through photosynthesis. Studies
in a travertine stream in the Sierra de Almijara
(Casas et al., 1994; Casas & Descals, 1997;
Casas & Gessner, 1999; Vivas & Casas, 2002)
have highlighted the key role of this process
often exerts in ecosystem energy flow, mainly by
interfering negatively with the action of microbial decomposers and detritivores. Two sites were
studied in this stream, one devoid of riparian
vegetation and experiencing intense travertine
precipitation, the other covered by a canopy of
riparian woody vegetation and reduced travertine
precipitation. Large differences in standing
stocks of benthic CPOM between both sites
reflected the differences in riparian vegetation.
Benthic CPOM stocks were remarkably high at
the vegetated site compared to other studies, probably due to the extraordinary retention capacity
of this type of travertine streams because riparian
inputs typically become encrusted with calcite
and firmly cemented to the stream bottom.
Additionally, benthic CPOM stocks may be
partly due to low decomposition rates. In fact, the
leaf species examined exhibited decay rates
among the lowest recorded in the literature, and
rates were significantly lower at the open site
with massive high travertine precipitation (Casas
& Gessner, 1999). The calcite crust, which forms
on CPOM rather quickly, probably acts as an
effective barrier preventing physical abrasion and
also hindering decomposition by both microbes
and detritivores. Accordingly, the species richness and reproductive activity of aquatic
hyphomycete fungi associated with leaves were
lower at the open site with higher travertine precipitation (Casas & Descals, 1997), as was the
biomass of detritivores (Vivas & Casas, 2002).
These data indicate that in Mediterranean streams that are prone to travertine precipitation,
basal resources such as leaf litter but also algal or
bacterial biofilms on rocks are difficult to exploit
for primary consumers. However, a well-developed woody riparian vegetation may significantly
enhance energy flow in these streams not just by
providing organic matter inputs but also by reducing tuff formation on the stream bed.
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Casas et al.
Conclusions and perspectives
The results presented above illustrate the diversity of fluvial ecosystems in eastern Andalusia.
The pronounced altitudinal, climatic and geological gradients in the area provide a wide range
of natural conditions, superimposed by human
impact. This sitting makes the area are well suited to carry out theoretical and applied studies
on river ecology, particularly comparative investigations. The studies carried out to date have
been mostly descriptive. The hypotheses they
have generated now require experimental testing. Long-term observations are also needed to
provide information on fluvial ecosystem dynamics at large time-scales, especially in view of
the important interannual variability typical of
the Mediterranean climate. In addition, our
results highlight the conspicuous effects of
human impacts on rivers in eastern Andalusia.
This is another piece of evidence documenting
the vast human pressures on aquatic ecosystems,
which are likely to have particularly severe consequences in the Mediterranean where scarce
and irregular rainfall is combined with everincreasing water demands and diverse sorts of
surface water contamination. Headwater reaches
may partly escape from this pressure, since they
are often located in sierras that benefit from
some sort of environmental protection status.
However, even these streams are subject to
human activities in their basins. A variety of farming, forestry and tourist activities are allowed
in natural parks, and together with the overexploitation of groundwater, they may impair the
ecological integrity of these rivers if preventive
measures are not taken. The ambitious environmental objectives of the European Water Framework Directive to secure ecological sustainability of rivers are a step in the right direction,
from which the rives in Andalusia may benefit.
ACKNOWLEDGEMENTS
Professor Ramón Margalef was a friend and
outstanding scientist whose ideas were a guiding force for us and surely will continue inspi-
ring new generations of limnologists. His death
is an irreparable loss to many of us. We are grateful to Joan Armengol for having invited us to
joint in this tribute. A substantial part of the
results discussed in the present contribution has
been funded by the Spanish “Ministerio de
Educación y Ciencia” as part of the following
projects: GUADALMED-1 (HID98-0323-C0504), GUADALMED-2 (REN2001-3438-C0705), and CGL2004-02496.
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Limnetica, 25(1-2): 171-180 (2006)
Significance of atmospheric deposition to freshwater ecosystems
in the southern Iberian Peninsula
Rafael Morales-Baquero1, Elvira Pulido-Villena, Otilia Romera, Eva Ortega-Retuerta,
Jose Mª Conde-Porcuna, Carmen Pérez-Martínez and Isabel Reche
Departamento de Ecología e Instituto del Agua. Universidad de Granada, 18071 Granada, Spain.
1Corresponding autor ([email protected])
SUMMARY
The Iberian Peninsula is close to the Saharan Desert, which is the biggest source of atmospheric aerosols of the World.
Currently, it is recognized that atmospheric deposition of aerosols over ecosystems is a significant source not only of elements
with gaseous phases but also of rock-derived ones. In the last years we have been quantifying the atmospheric flux of elements
and substances of biogeochemical interest on the aquatic ecosystems of the South Iberian Peninsula, and their impact on their
functioning and structure. The results we are obtaining indicate that atmospheric contribution of P and Ca are essential to
explain the functioning of high mountain lakes, and that atmospheric input of organic matter partially supports the pelagic
food web of these ecosystems. In this article we offer a summary of some of the results obtained to date.
Key words: atmospheric deposition; phosphorus; nitrogen; calcium; soluble organic matter; lakes.
RESUMEN
La Península Ibérica está próxima al Desierto del Sahara que es la mayor fuente de aerosoles atmosféricos del Planeta.
Actualmente, se reconoce que la deposición de aerosoles sobre los ecosistemas es una entrada significativa no sólo de elementos con fases gaseosas sino, también, de elementos derivados de rocas. En los últimos años hemos estado cuantificado el
flujo atmosférico de elementos y sustancias de interés biogeoquímico sobre los ecosistemas acuáticos del sur de la Península
Ibérica y el impacto sobre su funcionamiento y estructura. Los resultados que estamos obteniendo indican que los aportes
atmosféricos de P y Ca son esenciales para explicar el funcionamiento de los lagos de alta montaña y que las entradas atmosféricas de materia orgánica sostienen parcialmente las redes tróficas pelágicas de estos ecosistemas. En este artículo ofrecemos un resumen de algunos de los resultados obtenidos hasta ahora.
Palabras clave: deposición atmosférica; fósforo, nitrógeno, calcio, materia orgánica soluble, lagos.
INTRODUCTION
Biogeochemical cycles impose strong restrictions to the organization of the Biosphere
(Margalef, 1997). Therefore, Ecology needs to
understand the causes and consequences of the
global mobilization and distribution of elements. In this sense, the atmosphere is a major
component of the Biosphere (sensu Vernadsky)
as well as a pathway throughout which elements
can be mobilized among ecosystems. However,
we are far from having a complete comprehension of its role in Biogeochemistry.
Traditionally, the atmosphere has been considered the main source of elements with gaseous
phases, such as N, for terrestrial and aquatic ecosystems. By contrast, atmospheric contribution of
rock-derived elements, such as P or Ca, has been
considered of minor relevance, highlighting the
role of weathering inputs and sediment releases
particularly in Limnology. This paradigm has prevailed over decades in studies on nutrient budgets
in fresh-water ecosystems (Rodhe, 1948; Vollenweider, 1968, 1975, 1976). Due to anthropogenic
activity, atmospheric nitrogen deposition in the
Northern hemisphere significantly increased in
the 70s and 80s and its effects on lake chemistry
were intensely studied (Sullivan et al. 1990,
Stoddard et al. 1999), whereas atmospheric inputs
of phosphorus were not considered.
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Currently, it is well established that the atmosphere can mobilize amazing quantities of dust
from the arid areas of the world (Schlesinger,
1997), and the role of the atmosphere as a vehicle
for rock-derived elements has been recently
revindicated (Chadwik et al., 1999). These
authors found that the tropical ecosystems of
Hawaii depend critically on phosphorus supplied
by the atmosphere coming from the Central
Asian Desert. At a global scale, the Sahara
Desert is the largest arid area in the world and,
consequently, it is the origin of the largest loads
of dust to the atmosphere (D’Almeida, 1986).
This dust is transported towards the Atlantic by
the predominant westerly winds and towards the
Mediterranean basin influenced by the presence
of cyclones (Moulin et al., 1997).
Saharan dust contains high quantities of particulate matter, soluble minerals and organic carbon
(Talbot et al., 1986). The effects of dust deposition in aquatic and terrestrial ecosystems are now
receiving attention (Ridame and Guieu 2002;
Okin et al., 2004) and there is a considerable interest in assessing the effects of dust deposition on
marine ecosystems (Guerzoni et al. 1999; Herut
et al., 1999, 2002; Lenes et al., 2001). A particular scientific effort has been done to determine
the availability of dust-derived P to primary producers (Migon and Sandroni, 1999; Ridame and
Guieu, 2002; Markaki et al., 2003). Although the
potential relevance of atmospheric P deposition
for freshwater ecosystems has been previously
exposed (Peters, 1977), the effects of P atmospheric inputs on lake biogeochemistry have been
scarcely studied (Gibson el al, 1995).
Every year, the Iberian Peninsula receives
intrusions of air masses loaded with dust from the
Sahara Desert (Querol et al., 2003). Their deposition rates are poorly known and published data
comes mainly from the Northeast of the Iberian
Peninsula and is linked to rainfall (Camarero and
Catalan, 1996; Avila et al., 1997). Nevertheless
the Iberian Peninsula shows a strong gradient in
the rainfall with minimum values in the Southeast
where the dry deposition is predominant. During
the last years, we have been developing a program to study wet and dry deposition in the
Southern part of the Iberian Peninsula, and
Figure 1. Relationship between TOMS aerosol index (NASA,
Goddard Space Flight Center), as a surrogate of dust in the
troposphere, and the dry PM collected weekly at 1000 m a.s.l.
Despite that the TOMS aerosol index only provides valid
values of dust at altitudes over 2000 m above ground, it is possible to establish a direct connection between the dust content
at medium and high altitudes in the troposphere and the PM
collected at ground level. Dust at these altitudes is mainly due
to massive dust inputs originated in the Sahara Desert (after
Morales-Baquero et al., 2006). Relación entre el índice de
aerosoles TOMS (NASA, Goddard Space Flight Center), indicador del polvo en la troposfera, y el material particulado
(PM) seco recogido semanalmente a 1000 m snm. A pesar de
que el índice TOMS sólo proporciona valores válidos de polvo
por encima de 2000 m sobre el nivel del suelo, se puede establecer una relación directa entre el contenido de polvo en
niveles medios y altos de la troposfera y el PM recogido a
nivel del suelo. En las altitudes citadas, el polvo procede principalmente de las inyecciones masivas originadas en el
Desierto del Sahara (de Morales-Baquero et al., 2006).
their effects on biogeochemistry of high mountain lakes and reservoirs of this area. Our goal is
to quantify the atmospheric inputs of elements
and substances relevant for the biogeochemical
cycles of aquatic ecosystems, as representative
of atmospheric deposition in the Southwest
Mediterranean, and to establish direct links between such inputs and freshwater ecosystem responses. Here, we offer a summary of some of the
more relevant results obtained to date.
DEPOSITION OF PARTICULATE
MATTER (PM)
Data on dry and wet deposition obtained on a
weekly basis during two years at 1000 m and at
2900 m a.s.l. (only ice-free periods), conforms
to a Mediterranean regional pattern with gro-
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Atmospheric deposition in the southern Iberian Peninsula
wing values, as distances from Sahara are shorter and closer to the Eastern latitudes (Goudie
and Middleton, 2001). The mean PM total (dry
+ wet) deposition in Sierra Nevada at 1000 m.
(11.2 g m-2 yr-1 Morales-Baquero et al., 2006)
is higher than the values reported for Catalonia
(Montseny) (5.3 g m-2 yr-1, Avila et al, 1997)
and the Alps (0.2 to 0.4 g m-2 yr-1 Wagenbach
and Geis, 1989; De Angelis and Gaudichet,
1991), similar to the values registered in
Corsica (12-12.5 g m-2 yr-1, Bergametti et al.,
1989; Löye-Pilot et al., 1986) and lower than
those reported for the East Mediterranean area
(36 to 72 g m-2 yr-1 Herut and Krom, 1996).
Although rainfall can washout high quantities
of PM, the contribution of dryfall to the annual
total PM inputs registered in Sierra Nevada
(79 %) was much higher than the wet deposition, emphasizing the importance of dry deposition in areas where the rain is scarce
(Morales-Baquero et al., 2006).
Several evidences point out that Saharan dust
dominates PM deposition in the Southwest
173
Iberian Peninsula: 1) dry PM deposition exhibited
a similar seasonal pattern to Saharan dust export
toward the Mediterranean basin, which is characterized by maximum values particularly during
spring and summer (Moulin et al. 1997); 2) there
was a positive relationship between dry PM deposition and TOMS aerosol index (Fig. 1), which is
a suitable estimator of Saharan dust content in the
atmosphere (Chiapello et al., 1999); and 3) we
registered higher PM deposition at 2900 m
than at 1000 m which is consistent with the dynamics of Saharan dust transport, with maximum
loads mobilized between 1500 and 4000 m
(Talbot et al., 1986). Furthermore, the analysis of
dust deposition depending on the air masses’ origin, determined by using backward trajectories
analysis (HYSPLIT model, NASA) have clearly
shown higher dust deposition when the air masses
come from the South or Southwest rather than
from other directions (Fig. 2). All these results
reveal a significant and regular atmospheric
transport of material from the African Continent
to the South of the Iberian Peninsula.
Figure 2. Synchronous measurements of dry PM deposition in three collectors located in sites up to 40 km distant in the Granada
province and near three studied reservoirs. The values are cumulative data from weekly measurements during spring and summer of
2004 segregated according to the origin of the air masses for each corresponding week. Origins were determined by analysing 5-day
backward trajectories at 3000 m asl using the HYSPLIT model (NOAA, Air Resources Laboratory). It is evident that PM deposition
is higher when air masses arrive from the South or Southwest. Medidas sincrónicas de la deposición seca de material particulado
(PM) en tres colectores situados en localidades separadas hasta en 40 km dentro de la Provincia de Granada y cercanos a tres
embalses en estudio. Los valores son datos acumulados de medidas semanales durante la primavera y verano de 2004, separados de
acuerdo con el origen de las masas de aire en cada semana. Los orígenes se determinaron analizando retrotrayectorias de 5 días a
3000 m snm calculadas aplicando el modelo HYSPLIT (NOAA, Air Resources Laboratory). Es evidente que la precipitación de PM es
más elevada cuando las masas de aire llegan del Sur o Suroeste.
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174
NITROGEN AND PHOSPHORUS
DEPOSITION
Saharan dust contains significant quantities of
phosphorus. In fact, this source accounts for 3040 % of the total atmospheric flux of phosphorus
into the Northwestern Mediterranean (Guieu et
al., 2002). In Sierra Nevada the mean deposition
of total phosphorus (TP) was 513 µmol m-2 yr-1
(Morales-Baquero et al., 2006), that is similar to
deposition attributed to Saharan dust in Corsica
(about 500 µmol m-2 yr-1, Bergametti et al.,
1992), lower than the deposition measured in the
East Mediterranean (1300 µmol m-2 yr-1, Herut et
al., 1999), and higher than in Catalonia (100 µmol
m-2 yr-1-only wet deposition-, Avila et al., 1998).
The P deposition in our studies showed a similar
pattern to PM deposition, with maximum values
in spring and summer. In fact, we found a significant correlation between PM deposition and TP
deposition (n= 107; r= 0.45; p< 0.001).
In contrast, total nitrogen (TN) deposition
(39.6 mmol m-2 yr-1, Morales-Baquero et al.,
2006) was always lower than other sites from the
Mediterranean basin (55 mmol m-2 yr-1 in the
Northwest (Guerzoni et al., 1999) and about
50 mmol m-2 yr-1 in the East (Herut et al., 2002;
Markaki et al., 2003), only inorganic fractions in
both cases), making the anthropogenic impact
over that area evident since N deposition is
mostly linked to anthropogenic activity (Driscoll
et al., 2003). TN deposition did not show a clear
season pattern and was not correlated to PM
deposition. In addition, wet deposition contributed more than 50 % to N inputs, whereas most P
inputs over our study area were linked to dry
deposition (72 %). The differences in the N and P
inputs are also reflected in the molar TN:TP ratio
of the atmospheric deposition, which varies seasonally from values as low as 11.9 in spring or
summer to values >100 in fall or winter.
The atmospheric inputs coming from
Saharan dust appear to affect the biogeochemistry of the high mountain lakes from Sierra
Nevada. The relatively high atmospheric inputs
of P during summers were previously suggested
as the responsible for the enhancement of P
deficiency, as the summer progress, in lakes
Figure 3. Relationships between the TN:TP molar ratio of the
atmospheric deposition and the TN:TP molar ratio of the water
column in La Caldera and Río Seco lakes. Although the
atmospheric TN:TP molar ratio significantly affects both
lakes, the influence on La Caldera nutrient status is steeper
than that on Río Seco (after Morales-Baquero et al., 2006).
Relaciones entre la razón molar TN:TP de la deposición
atmosférica y la razón molar TN:TP de la columna de agua en
las lagunas de La Caldera y Río Seco. Aunque la relación
molar TN:TP atmosférica afecta significativamente a ambas
lagunas, la influencia en los nutrientes de La Caldera es más
acusada que en Río Seco. (de Morales-Baquero, et al. 2006).
with bigger catchments areas (MoralesBaquero et al., 1999). The lakes with relatively
smaller basins reflect the atmospheric N:P ratio
more closely than the inputs from watersheds.
In fact, N:P ratios of atmospheric inputs significantly relate to N:P ratios of lakes, affecting
nutrient status (Morales-Baquero et al, 2006)
(Fig. 3). In addition, we found a direct connection between atmospheric TP input and the response of phytoplankton (Morales-Baquero et
al., 2006) (Fig. 4), demonstrating the importance and bioavailability of the P delivered from
the atmosphere in natural conditions.
CALCIUM DEPOSITION
It is well known that Saharan dust contains high
quantities of calcium carbonate (Löye-Pilot et al.
1986), which significantly increases the pH of
rainwater and its deposition is an important input
of Ca to terrestrial ecosystems (Avila et al.,
1997). The mean total atmospheric inputs of calcium at 1000 m registered in Sierra Nevada
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Figure 4. Relationship between atmospheric deposition of
total phosphorous (TP) and Chlorophyl-a (Chl a) in La
Caldera Lake. Each point represents the weekly input of TP
(dry + wet) against the water column integrated Chl-a values
at the end of the corresponding week. TP deposition explains
66 % of Chl-a variability (after Morales-Baquero et al., 1966).
Relación entre la deposición atmosférica de fósforo total (TP)
y el contenido en clorofila a (Chl a) de la laguna de La
Caldera. Cada punto representa la entrada semanal de TP
(seco + húmedo) frente a los valores de Chl a integrados para
la columna de agua al final de la semana correspondiente. La
deposición de TP explica el 66 % de la variabilidad de la
Chl a (de Morales-Baquero, et al. 2006).
(39.2 mmol m-2 yr-1; Pulido-Villena et al., 2006)
were slightly higher than those reported for
the Northeastern Iberian Peninsula, an area
also influenced by Saharan dust inputs
(24.8 mmol m-2 yr-1; Avila et al., 1997; 1998),
and clearly higher than the reported for Northern
Europe (6.1 mmol m-2 yr-1; Hultberg and Ferm,
2004). Like P deposition, Ca dry deposition was
prevalent (64 % of total deposition), and showed
the same seasonal pattern of PM dry deposition.
Consequently, PM and Ca dry deposition were
correlated (r=0.60; p<0.001; n=106). The influence of Saharan dust in Ca deposition is showed by
the 50 % mean increase when Saharan intrusions
over the Iberian Peninsula occur (Fig. 5).
The atmospheric inputs of Ca are a determining factor for the Ca content in the Sierra
Nevada lakes. We have recorded the Ca concentrations during three ice-free periods in two
lakes: Rio Seco and La Caldera, with and
without superficial outlets respectively. Ca concentration was always higher in La Caldera
Lake (107.6 ± 1.1 µM) than in Río Seco Lake
175
Figure 5. Weekly averages of Ca deposition for 2001 and 2002
at the 1000 m a.s.l. collector. Values are segregated depending
on the existence (62 weeks) or not (42 weeks) of Saharan intrusions (SI) over the Iberian Peninsula (after Pulido-Villena,
2004). Promedios semanales de deposición de Ca recogidos
durante 2001 y 2002 en el colector situado a 1000 m snm. Los
valores se han separado de acuerdo con la existencia (62 semanas) o no (42 semanas) de intrusiones saharianas (SI) sobre la
Península Ibérica (de Pulido Villena, 2004).
(37.4 ± 1.1 µM). Ca concentration in both lakes
showed a significant synchronous dynamics
(r=0.63; p< 0.001; n= 35). This fact suggests a
climatic control, which could also be due to
evaporative processes during summer. Nevertheless, analysing the in-lake variation of the
18O isotope, as a surrogate for evaporation, and
the total direct Ca deposition to lakes, it has
been possible to establish that Ca deposition is
positively affecting the Ca concentration in
both lakes. Furthermore, a mass estimate of Ca
inputs to lakes and basins, realized on an
annual basis, showed that atmospheric inputs
can fully explain the Ca concentrations found
in both lakes (Pulido-Villena et al., 2006).
Therefore, the atmospheric Ca deposition appears to be a key factor to understand the high Ca
content (and related variables, e.g. the acidneutralizing capacity) of some lakes in Sierra
Nevada in comparison with the central Europe
high mountain lakes (MOLAR, 1999).
ORGANIC MATTER DEPOSITION
Among the soil components mobilized as aerosols by the atmosphere, there are important
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Figure 6. Food web interactions in La Caldera lake as deduced from a 13C isotope analysis of their major components. The three
crustacean species appear clearly segregated in their respective food sources. The significance of atmospheric inputs of organic
Carbon is deduced from the high values of ␦13C signature of POM in the lake (after Pulido-Villena et al., 2005). 1) ␦13C of phytoplankton for La Caldera was estimated using a fractionation factor of 20‰ and the values of ␦13C of dissolved inorganic carbon
measured by us. 2) ␦13C values for atmospheric inputs after Eglinton et al. (2002). Interacciones en la red trófica de la laguna de
La Caldera según se deducen de un análisis del isótopo 13C en sus principales componentes. Los tres crustáceos mayoritarios aparecen claramente segregados en sus respectivas fuentes de alimento. La importancia de las entradas atmosféricas de carbono
orgánico se deduce de los altos valores de ␦13C de la materia orgánica particulada (POM) en la laguna (de Pulido-Villena et al.,
2005). 1) El ␦13C del fitoplancton en La Caldera se estimó usando un factor de fraccionamiento del 20‰ y los valores del ␦13C del
carbono inorgánico disuelto medidos por nosotros. 2) Valores de ␦13C de las entradas atmosféricas según Eglinton et al. (2002).
quantities of particulate and water-soluble organic carbon (W-SOC) (Talbot et al., 1986). Its
deposition rates and ecological effects are
poorly known, although it has been reported that
atmospheric wet deposition of W-SOC to oceans
can be similar to the dissolved organic carbon
(DOC) derived from global river discharge
(Willey et al. 2000). The atmospheric deposition
of W-SOC is expected to be very significant in
oligotrophic high mountain lakes where these
compounds have low concentrations (<1 mg l-1)
but play important functions, such as the regulation of ultraviolet radiation attenuation (Laurion
et al. 2000; Reche et al., 2001).
The total summer cumulative atmospheric
deposition of W-SOC collected at 2900 m in
Sierra Nevada was 20-mmol m-2, and about
50 % of this quantity arrived with the dry deposition of PM. Total PM deposition showed a
direct relationship with total W-SOC deposition
(r=0.62; p<0.001; n=33) (Pulido-Villena, 2004).
These rates demonstrate that there is a substantial input of organic carbon, potentially bio-available, from the atmosphere to Sierra Nevada
lakes. In fact, an analysis of the pelagic food
webs, using a stable isotope approach, showed
that this source of carbon might be essential for
the food-webs in these lakes (Pulido-Villena et
al., 2005). Fig. 6 shows the carbon stable isotope
signature of the pelagic food web of La Caldera
Lake. ␦13C of the zooplanktonic community
revealed species-specific differences in their
food sources, probably as a result of an ecological niche segregation. The cladoceran Daphnia
pulicaria relied mainly on bulk particulate organic matter (POM) whose isotopic signature
(␦13C= -24.5 ‰) was heavier than that estimated
for phytoplankton (␦13C= -32.5 ‰). The most
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plausible explanation for this discrepancy is that
POM was composed not only of phytoplankton
(and derived biota and detritus) but also of
terrestrial organic matter, which is usually enriched in 13C. Since the basin of La Caldera Lake
is rocky, the atmospheric inputs of organic matter, mainly derived from Sahara, could explain
the heavy signal of POM in La Cadera. In this
regard, the measured atmospheric input of WSOC in La Caldera may represent as much as
33 % of the dissolved organic carbon concentrations found in this lake (Pulido-Villena et al.,
2005). Furthermore, a recent study by Eglinton
et al. (2002) reported that the isotopic signature
of total organic carbon in atmospheric dust derived from the Sahara Desert is especially heavy
(c.a. -18 ‰) suggesting the presence of biomass
and burning residues derived from predominantly C4 vegetation accumulated in the soils.
The atmospheric inputs of organic matter
can be a source of food not only for indiscriminate filter-feeding animals such as D. pulicaria, but also for bacteria. In fact, a bacteria culture enriched with atmospheric dust
showed higher growth efficiency on atmospheric W-SOC than on lake DOC (Pulido-Villena,
2004). Therefore, the food webs of the high
mountain lakes of Sierra Nevada can be partially supported by a source of energy originated in terrestrial ecosystems from other continents and whose transport is regulated by
global atmospheric circulation patterns.
MICROORGANISMS AS AEROSOLS
Although the microbial component of aerosols
is known since the 19th century, it has not received attention up to the present decade. Aerosols
can mobilize about 1018 cells per year (Griffin
et al., 2002) and these air-transported microorganisms can survive long distances suspended
in dust particles. However, their colonizing abilities and outcompeting success are almost unknown. The deposition, viability, and expansion of these invading microorganisms can
affect the indigenous microbiota, particularly
in remote lakes with high ecological value. In
177
an ongoing project (ECOSENSOR, Fundación
BBVA) we have selected remote lakes from the
Arctic area, Antarctica, Patagonia, and high
mountains to establish microbial biogeography
patterns. We pretend to assess the role of the
atmospheric long-range transport of microorganisms as a dispersal mechanism affecting
microbial biodiversity patterns, since the spatial structure appears to contribute significantly
to lake bacterial composition (Reche et al.,
2005). Some preliminary experiments in our
laboratory have also confirmed the existence of
viable bacteria linked to dust deposition.
CONCLUSIONS
From the results obtained to date, it appears evident that the deposition of elements and compounds mobilized as aerosols by the atmosphere
plays a significant role in the biogeochemistry
of high mountain lakes from the Southern
Iberian Peninsula. This deposition is related to
dust exported from the Sahara desert on an
annual basis, implying a regular intercontinental
transfer of material. The deposition occurs
mostly in dry form, and its effect on terrestrial
and aquatic ecosystems needs to be addressed.
The dryfall of particulate matter is a climatic
variable that, contrary to rainfall, has been scarcely considered. This variable is connected to
global atmospheric circulation patterns such as
the North Atlantic Oscillation (NAO). High positive NAO years involve high dust export from
the Sahara to the Mediterranean basin (Moulin
et al., 1997). The increase in the transport of
aerosols that the recent models of climatic change have predicted, will probably lead to in an
increase in the input of mineral nutrients and WSOC to freshwater ecosystems. In areas, such as
the Mediterranean basin with long periods of
absence of rainfall, the dryfall is continuously
reaching aquatic ecosystems where the soluble
components appear to have consequences, which
we are now beginning to understand.
From Margalef ’s legacy two major ideas have
emanated and inspired this ongoing research.
First, that biogeochemical cycles impose strong
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restrictions to organization of the Biosphere,
and second, that high mountain lakes are the
finest sensors available to detect changes in
the troposphere. We have now evidences that the
atmosphere can supply both gaseous and rockderived elements essential for the biogeochemistry of Sierra Nevada high mountain lakes
that are particularly sensible to variations in that
supply. Since we are very grateful to the work of
Margalef, these pages are a little tribute to him.
ACKNOWLEDGEMENTS
Financial support was provided by the projects
CICYT AMB99-0541 and MCYT REN03-03038.
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Limnetica, 25(1-2): 181-188 (2006)
Physical Limnology in Lake Banyoles
Xavier Casamitjana, Jordi Colomer, Elena Roget and Teresa Serra
Environmental Physics Group i Institut de Medi Ambient. Campus de Montilivi. Universitat de Girona.
17071-Girona. Spain
Corresponding author: [email protected]
ABSTRACT
The main physical long-scale processes occurring in Lake Banyoles are reviewed as a tribute to Prof. Margalef. These processes include the water fluxes below the surface of the lake, the behavior of the sediment in suspension in the basins, the heat fluxes at the surface and at the bottom layers, the internal seiching, the formation of a baroclinic current due to differences in cooling between the two lobes, the mixing dynamics, the meromictic behavior of some of the basins and the formation and
dynamics of hydrothermal plumes
Keywords: physical processes, sediment in suspension, internal seiches, baroclinic currents, hydrothermal plumes
RESUMEN
Los principales procesos físicos de gran escala que tienen lugar en le Lago de Banyotes son revisados como tributo al
Profesor Margalef. Estos procesos incluyen los flujos de agua bajo la superficie del lago, el comportamiento del sedimento en
suspensión en las cubetas, los flujos de calor en la superficie y en el fondo, las secas internas y la formación de corrientes
baroclínicas debido a la diferencia de enfriamiento entre los dos lóbulos, la dinámica de mezcla, el comportamiento meromíctico de algunas de las cubetas y la formación y dinámica de las plumas hidrotérmicas.
Palabras clave: procesos físicos, sedimento en suspensión, secas internas, corrientes baroclínicas, plumas hidrotérmicas.
INTRODUCTION
It was about 1985 when two of the authors of
this paper and their future PhD supervisors
(David Jou and Josep Enric Llebot) visited Prof.
Margalef in his office at the Ecology Department of the University of Barcelona. The authors
wanted to focus their studies on the physical processes occurring in Lake Banyoles and, knowing
of Prof. Margalef ’s scientific interest in the lake,
asked him for his advice. The pioneering work of
Prof. Margalef in Lake Banyoles (Margalef,
1946), had been followed by various studies on
the ecology of the lake (Planas, 1973; Guerrero
et al., 1978; Rieradevall & Prat, 1991), but very
little was known about the physical processes
there. Prof. Margalef was always encouraging
researchers with background in physics to
undertake a study of the subject. After long and
fruitful discussions, we decided to approach
Lake Banyoles by studying the hydrodynamics
of the lake. In 1992, when Prof. Margalef came
to Girona as a member of the committee evaluating one of these PhDs, he let us know that he
agreed with our approach and encouraged us to
pursue physical limnology further. Since then,
many other research studies undertaken by the
Environmental Physics Group of the University
of Girona have been related to the lake, and the
PhD theses of the other two authors of this article were also related with Lake Banyoles.
Although nowadays the focus of the Environmental Physics Group has been broadened to
cover other fields such as reservoirs, wetlands
and oceans (Serra et al., 2003; Roget et al.,
2005), there are still two PhD theses being written about different aspects of Lake Banyoles,
from which some preliminary results are referred to in this paper. The complexity of the lake’s
hydrodynamics, together with recently develo-
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Casamitjana et al.
ped instruments, such as microstructure sensors,
acoustic Doppler current meters, particle size
analyzers, etc., make these new studies seem as
passionate as those in the beginning. Going back
to 20 years ago, we can’t avoid thinking in the
first pioneering researches of Lake Banyoles.
Although some of those papers, written in
Catalan or Spanish, will not be of interest to an
international audience, they might be for local
researchers; the same applies to some of our first
works which have never been published in
English and are reviewed in this paper.
Today we are pleased to review the history of
hydrodynamic studies in Lake Banyoles as a tribute to Prof. Margalef. The main physical processes to be reviewed here are schematised in
figure 1 and include the water fluxes below the
surface of the lake, the behaviour of the sediment in suspension in the basins, the heat fluxes
at the surface and at the bottom layers, the internal seiching, the formation of a baroclinic
current due to differences in cooling between the
two lobes, the mixing dynamics, the meromictic
behaviour of some of the basins, and the formation and dynamics of hydrothermal plumes.
THE SUBTERRANEAN SPRINGS
OF THE LAKE
Lake Banyoles, located in the eastern Catalan
pre-Pyrenees (42°07’N, 2°45’E), is a small
multi-basin lake of mixed tectonic-karstic origin
with a surface area of 1.12 km2 (Fig. 2). The tectonic constraint of the lake forces the groundwater flow through the bottom of the basins in a vertical discharge (Moreno & García-Berthou,
1989). The subterranean springs keep several
meters of sediment in suspension at the bottom of
the conic inflow areas and up to a fairly sharp
sediment interface, known as the lutocline, where
the particle concentration can vary from about
180 mg/l to less than 1 mg/l. The suspended sediment is formed by a mixture of marly and argillaceous materials and they have a nearly constant
temperature throughout the year (19 ºC), being 23 ºC higher that that of the hypolimnetic water in
summer and 8-10 ºC higher in winter. Subterranean water inflow is also saltier than the bulk
water of the lake (Casamitjana, 1989).
The importance of the underground inflow to
the lake’s dynamics did not escape the first re-
Figure 1: Scheme of the main physical processes occurring in Lake Banyoles. Esquema de los principales procesos físicos que tienen lugar en el Lago de Banyoles.
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searchers studying the lake. Calculations based
on the difference between the surface inflows
and outflows had already been done 80 years
ago, estimating a global underground inflow of
around 600 l/s (Mascaró, 1914; Vidal-Pardal,
1925). However, the global estimations did not
account for the different physical, chemical,
and biological properties of the basins
(Guerrero et al., 1978). The groundwater
inflow within the individual basins was first
determined in 1986 based on measurements of
the settling velocity of the particles in the suspensions, their porosity and the cross sectional
183
area (Roget, 1987; Roget et al., 1994). The
most important conclusion from this study was
the location of the main underground inflow in
the southern lobe (BI in Fig. 2), which usually
supplies around 85 % of the total incoming
water by subterranean springs. Casamitjana et
al. (1988) calculated the heat flux from the
basins of Lake Banyoles and estimated that, if
the lake were not heated from below, the heat
budget would increase by around 6.7 107 J m-2
over 7.9 107 J m-2. They also estimated that the
heat through BI was 90 % of the total underground heat coming into the lake.
Figure 2. Bathymetric map of Lake Banyoles. Depth contours are in meters. The lake is composed of six basins (BI to BVI). The
bottom right panel shows a schematic view of basin I of Lake Banyoles (BI) obtained from a seismic profile and its interpretative
section (adapted from Canals et al. 1990). The top right panel shows two echosounding profiles of basin BII for two different
recharge volumes. Sediment in suspension is clearly delimited by an horizontal interface. Mapa batimétrico del lago de Banyoles.
Los contornos de profundidad están en metros. El lago está formado por 6 cubetas. El panel de la parte inferior derecha es una
representación esquemática de la cubeta I del Lago de Banyoles (BI), obtenida mediante perfiles sísmicos, juntamente con un
esquema interpretativo (adaptado de Canals et al. 1990).El panel de la parte superior derecha muestra dos perfiles de ecosonda
de la cubeta BII, para dos flujos diferentes de entrada. El sedimento en suspensión está claramente delimitado por una interfase
horizontal.
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Casamitjana et al.
Different laboratory studies were carried out in
order to simulate the re-suspension of particles
by a flow from below. Roget & Casamitjana
(1987) used a glass column 140 cm high and
10 cm in diameter with a variable water supply
at its bottom and extraction points every 10 cm
along the column to study the fluidisation process. Casamitjana & Roget (1990a) also stated
the importance of the history of the bed on its
departure for the ideal behaviour. Laboratory
studies were also performed by Colomer &
Fernando (1996) using a tank with a jet nozzle
located at the centre of the bottom, beneath the
particle layer. The jet, carrying fluidised particles, penetrated through the particle bed and
emanated into the upper layer. They identified
two different flow regimes. The first one, called
the “continuous re-entrainment” regime, occurring at low values of momentum fluxes, is characterized by no significant deposition of particles from the particle-laden jet onto the bed. In
the second one, called the “full deposition” regime, the particles that were initially entrained
into the jet were found to deposit back onto the
bed with time, thus forming an axisymmetric
particle mound around the inlet jet.
Casamitjana et al. (2000) carried out a similar
experiment but using a conical basin. They identified two regimes: the so-called “jet flow” and
“lutocline flow” regimes; the “full deposition”
regime was not found here. In the “lutocline flow
regime”, the particles were re-suspended forming a well-established interface along the entire
cross sectional area of the cone. The maximum
height to which particles can rise was found to
depend on the momentum jet, the initial height
of the particle bed, the particle diameter, the
Reynolds number of the particle, and the slope of
the conical basin. Results obtained in this experiment show good agreement to what was observed in Lake Banyoles, provided that the sediment
particles aggregate within the suspension.
Particle size distributions of the suspended
sediments in Lake Banyoles have repeatedly
been obtained by different methods (Sanz, 1985;
Roget, 1987). However, the handling and analysing procedures may disrupt the suspended
aggregates, altering the results. To measure the
“real” diameter of the suspended particles, samples of sediment were immediately frozen in situ
by introducing them into liquid nitrogen containers (Casamitjana et al., 1996). From the frozen
sediment, samples were taken and deposited on
a nucleopore membrane and then analysed by a
scanning electron microscope. This technique
showed mean diameters of around 60 µm, which
reveal good agreement with the experiments
carried out by Casamitjana et al. (1996).
The re-suspension of sediments and the formation of the lutoclines can be also simulated
by using a two-dimensional κ-ε model (Colomer
et al., 1998). The results of the model predict
a re-circulation zone below the lutocline and a
severe damping of the turbulent kinetic energy
and effective viscosity at the bottom of the lutocline due to the buoyancy flux. The lutocline
acts as a barrier to the propagation of the turbulent kinetic energy. However, due to the convective processes generated immediately above this
interface, there is a local increase of the rate of
turbulent kinetic energy dissipation (Muñiz,
2000; Sánchez, 2001; Lozovatsky et al., 2005).
Records over the past 20 years show that the
sediment in basin BII (see Fig. 2) usually
remains consolidated at the bottom, with the
lutocline at a depth of approximately 44 m
(Casamitjana & Roget, 1993, Colomer et al.,
1998, Soler et al., 2005). Eventually, the subterranean springs in BII supply water to the lake
at a rate comparable to those in BI. This is possible for high precipitation periods that recharge
the aquifer, which in turn increases the pressure
enough for incoming water to re-suspend the
confined and consolidated sediment at the bottom of BII (Colomer et al., 2002). In this case
the sediment migrates upward and initiates the
fluidisation of the confined bed sediments. The
initiation of the fluidisation usually coincides
with the maximum mean monthly rainfall,
which was about 250-350 mm/month. The rainfall, in turn, is associated with six main atmospheric circulation patterns among the 19 fundamental circulations that emerged in an earlier
study that focused on significant rainfall days in
Mediterranean Spain. They comprise a wide
variety of flows over the Iberian Peninsula, with
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marked seasonal distributions and a clear distinction between Atlantic and western Mediterranean disturbances (Soler et al., 2005).
THERMAL STRUCTURE, SEICHING
AND BAROCLINIC CURRENTS
The thermal structure of Lake Banyoles can be
predicted by using a one dimensional lake
model, like DYRESM, and inserting the inflow
directly to the bottom of the water column
(Casamitjana et al., 1993). Mean daily measurements of global radiation, downward long-wave
radiation, wind speed, air temperature and water
surface temperature were used to calculate a
new equation for the non-radiative fluxes
(Colomer et al., 1996). The predicted water temperature profiles are in good agreement with the
observations in basin BI. In basins BIII and BIV
the evolution of the temperature inversion in the
hypolimnion, due to meromixis (Casamitjana &
Roget, 1986) and the rate of mixed layer deepening is also well predicted. The model shows
how the groundwater intrusion greatly reduces
the extent of summer stratification.
The first vertical modes of the internal seiches in Lake Banyoles were first calculated by
Besalú et al. (1988) using a one-dimensional
model. Although by that time high vertical
modes had rarely been described in the literature, it was found later using a multi layer twodimensional model, that second vertical modes
dominate the internal wave field in Lake
Banyoles in spring (Roget, 1992) and their
amplitudes were measured to be up to 2 m.
Furthermore, in an unexpected result, their
period was not correlated to that of wind forcing (Roget et al., 1993a) as it was believed to
always be the case for high vertical modes.
Even considering the small fetch of the lake in
this direction, due to the importance of the seabreeze regime (Roget et al., 1997), a persistent
first vertical transversal mode was also found in
the southern lobe. This transversal mode corresponds to the third horizontal for the southern
lobe and coincides with the seventh of the
whole lake. Coupling between oscillations of
185
other density interfaces of the lake was also
described in Roget (1992).
Because of the different thermal inertia of the
two main lobes of the lake (with mean depths of
10 and 18 m) and the different incoming heat
flux through the underground sources located in
them, it was found that in winter, denser water
of the shallower northern lobe was plunging into
the deepest lobe forming a bottom current, with
velocities of up to 12 cm/s observed in the central part of the lake. This current redistributes
water between the two lobes and replaces the
water of the northern lobe about every 5 days
(otherwise its residence time would be about
one year). Due to its magnitude (around 20000
l/s), it dominates the circulation of the lake
(Roget et al., 1993b). This current was affected
by the wind pattern and deflected towards the
southwest due to the bottom topography, but
also due to the Coriolis force. This is described
in Roget & Colomer (1996) where it is found
that the cross slope of the isotherms agrees with
the geostrophic balance.
DYNAMICS OF THE HYDROTHERMAL
PLUME
The difference in temperature between the sediments in suspension and the hypolimnetic water
immediately above them leads to the formation
of a turbulent convective plume immediately
above the lutocline (Colomer et al., 2001). The
plume develops upwards until it reaches the
level of neutral buoyancy, spreading laterally as
gravity current. The estimated values of the
Rossby number showed that rotation affects the
plume development and that the plume develops
to a maximum height limited by the strong temperature gradient of the seasonal thermocline.
When there is no inflow through basin BII, a
second stationary thermocline develops at a
depth of around 22 m in this basin. Casamitjana
& Roget (1990b) found that this thermocline is
only destroyed in the coldest months of January
and February or whenever the fluidization of
basin BII occurs. In this last case, the lutocline
migrates upward, the secondary thermocline is
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Casamitjana et al.
destroyed and a new hydrothermal plume is formed in basin BII (Colomer et al., 2003).
As a major characteristic of convection from
an isolated source, the plume entrains particles
from the lutocline and carries a suspension of
clay and silt particles upward with particle volume concentrations of ∼ 5-10 µl L-1. Because of
the temperature inversion at the lutocline, the
plume is negatively buoyant. As a result, it
moves upward and in the absence of a thermal
stratification background or in the presence of a
weak one (as in the mixed lake period), it reaches the surface waters and then spreads laterally, with the consequent change of water quality caused by an increase of the suspended
particle concentration (Serra et al., 2002a). It
can be expected, then, that the suspended particles change the clarity of the water, which might
imply a habitat constraint for fishes by limiting
their feeding opportunities and other visual activities (Serra et al., 2002b).
Two-dimensional temperature and particle
concentration measurements show the fate of
the hydrothermal plume and its associated turbidity current and reveal its seasonal development
(Serra et al., 2005). Silt particles transported by
the plume have been used as tracers to determine the maximum and equilibrium heights of the
plume. When the lake is stratified, the vertical
transport of sediment is confined to the lake
hypolimnion, since the thermocline limits the
vertical propagation of the plume. In contrast,
when the lake water column is mixed, the plume
reaches the surface of the lake. The field measurements have been compared with models for
thermal convection from finite isolated sources.
Measurements of flow velocity at the source of
the hydrothermal plume (i.e. the rim current
velocity) indicate that cold hypolimnetic water
is entrained by the plume. In the zone where the
turbidity current develops, sedimentation rates
measured from sediment traps vary between 10
and 25 g m-2 d-1, and result from continuous silt
particle sedimentation from the turbidity
current. Sedimentation rates in traps are higher
for stations situated close to the source than for
those further away (< 5 g m-2 d-1). Moreover, the
results demonstrate that double diffusive sedi-
mentation from the turbidity current was dominant over grain-by-grain settling, causing a
mixed distribution of sediments in the region
where the turbidity current spreads. The deposition of silt particles could explain the occurrence of silt layers interbedded with biocalcarenites
in the littoral zones of the lake and the stratigraphy identified by seismic profiles and cores
taken from the lake floor. Therefore, all of the
mentioned results demonstrate that the presence
of the plume and the turbidity current affects the
sedimentary records in the lake.
At the upper interface of the fluidised bed at
the base of the plume, particle concentration and
salinity (decreasing upwards) have the opposite
effect of that of temperature (also decreasing
upwards) on their contribution to the vertical
density distribution, and the corresponding density ratio is a little bit greater than one. This condition, plus the fact that particle and salt diffusivities are lower than thermal diffusivity, makes
the double-diffusive convection possible within a
diffusive regime. At present, this is studied
according to microstructure measurements
recorded in basin BII, when a second thermocline at the entrance of the conic underground
spring region exists, greatly isolating this region
from the rest of the lake. In these circumstances,
the upper interface of the fluidised bed is steplike for the three scalar fields, with well-mixed
turbulent convective layers separated by diffusive interfaces (Sánchez & Roget, 2005).
ACKNOWLEDGMENTS
We would like to thank the various collaborators
involved in different ways in Lake Banyoles’
research though the years: David Jou, Josep
Enric Llebot, Josep Pararols, Marianna Soler,
Ramon Julià, Xavier Vila, Esperança Gacia,
Rafael Juanola, Jaume Piera, Romualdo
Romero, John Alan Ross, Geoffrey Schladow,
H. Joe Fernando, Xavier Sánchez, Javier Vidal,
and Josep Pasqual. The authors are grateful to
captain Joan Corominas for the outstanding help
and support in the field campaigns in Lake
Banyoles. Most of the funds were provided by
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the Catalan and Spanish Governments, through
different agencies and projects.
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Limnetica, 25(1-2): 189-204 (2006)
Complex interactions in microbial food webs: Stoichiometric
and functional approaches
Presentación Carrillo1, Juan Manuel Medina-Sánchez2, Manuel Villar-Argaiz2,
José Antonio Delgado-Molina2 and Francisco José Bullejos2
1Instituto
del Agua, Universidad de Granada. 18071 Granada, SPAIN
Departamento de Ecología, Facultad de Ciencias, Universidad de Granada, 18071 Granada, SPAIN
1,2
ABSTRACT
The food web structure in some high mountain lakes deviates from the established tendency of high heterotrophic bacteria: phytoplankton biomass ratios in oligotrophic ecosystems. Thus, the microbial food web in La Caldera Lake is weakly developed, and
bacteria constitute a minor component of the plankton community in terms of abundance, biomass and production. Autotrophic
picoplankton is absent, and heterotrophic microbial food web is weakly developed compared to a grazing chain dominated by
calanoid copepods and a phytoplankton community mainly composed of mixotrophic flagellates. In order to explain the singular
food web structure of this lake, functional, stoichiometric and taxonomical approaches are followed to assess, on various temporal
and spatial scales, the relevance of stressful abiotic factors (ultraviolet solar radiation and P-limitation) on the structure and functioning of this ecosystem. P-availability was the main factor controlling the algal biomass whereas bacterial P- limitation was a
transient phenomenon. The algae-bacteria relationship was predominately commensalistic. In contrast to algae, full-sunlight
radiation had no negative effect on bacterial growth but rather enhanced bacterial dependence on the carbon released by algae.
The prevalence of the commensalistic-mutualistic relationship and the development of a more complex microbial food web were
related to the stoichiometry of algae and bacteria (N:P ratios). The microbial food web only developed at balanced algal and bacterial N:P ratios, with the appearance of ciliates after a nutrient pulse. However, mixotrophic algae dominated the planktonic community under P-deficit conditions, and they were the main factor controlling bacterioplankton. Their regulatory effect has a dual
nature: (i) a resource-based control, where bacteria depend on the photosynthetic carbon released by algae, i.e., a commensalistic
interaction (“without you I cannot live”); and (ii) a predatory control, where bacteria is a prey for mixotrophs (“with you I die”).
Hence, the niche of microheterotrophs (nanoflagellates and ciliates) is occupied by mixotrophs, and there is a resulting simplification of the planktonic structure. With respect to the carbon cycle, mixotrophic bacterivory constitutes a “by-pass” for the flux of
C towards the grazing chain, precluding the development of a complex heterotrophic microbial food web. Mixotrophs thereby
improve the energetic transfer efficiency in high mountain lakes through a reduction in the number of trophic levels. Antagonistic
effects of UVR x P interactions on the algae-bacteria relationship were caused by an enhancement of dual (resource and predation) control. Based on these results, an alternative model for the flux of C in autotrophic high mountain lakes has been proposed.
Key words: Carbon flux, high mountain lakes, microbial loop, nutrient limitation, stoichiometry, ultraviolet radiation.
RESUMEN
La estructura de la red trófica en algunos lagos de alta montaña, se aleja de los patrones establecidos para ecosistemas oligotróficos que proponen el predominio de la red trófica microbiana sobre la cadena de pastoreo. Así, en la laguna de La Caldera las
bacterias son el componente minoritario de la comunidad planctónica en términos de abundancia, biomasa y producción. El
picoplancton autótrofo está ausente y la red microbiana heterotrófica se encuentra escasamente desarrollada frente a una cadena
de pastoreo dominada por copépodos calanoides y algas mixotróficas. Para comprender los mecanismos que determinan esta
estructura trófica hemos seguido diferentes aproximaciones de análisis: funcional, estequiométrica y taxonómica sobre distintas
escalas espaciales y temporales, en relación con los principales factores de estrés abiótico (radiación ultravioleta y limitación
por fósforo) que controlan el funcionamiento de los ecosistemas de alta montaña. Nuestros resultados indican que la disponibilidad de fósforo, de forma generalizada, controla la biomasa algal y de manera transitoria la bacteriana, estableciéndose entre
ambas comunidades una relación comensalista. La radiación solar completa no afecta negativamente el desarrollo de las bacterias y si el de las algas y potencia la relación de dependencia por el carbono orgánico (comensalismo) entre algas y bacterias. El
predominio de la relación comensalista-mutualista y el desarrollo del bucle microbiano esta relacionado con la estequiometría
(razón N:P) de algas y bacterias. Así, sólo cuando la razón N:P de algas y bacterias es equilibrada para crecer, un pulso de
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Carrillo et al.
nutrientes permite el desarrollo del bucle microbiano. En condiciones naturales de déficit de P, sin embargo, existe un predominio
de “algas” con metabolismo mixotrófico. Las algas mixotróficas ejercen un efecto regulador dual sobre las bacterias que denominamos Ni contigo ni sin ti, (i) control por depredación, donde las bacterias son consumidas por algas mixotróficas (“contigo
me muero”), (ii) control basado en los recursos estableciéndose una relación de dependencia de las bacterias sobre del carbono
liberado por las algas (“sin ti no puedo vivir”). La mixotrofia supone un simplificación en la cadena trófica microbiana, donde
los mixótrofos ocupan el nicho potencial de nanoflagelados y ciliados. Desde un punto de vista energético implica un cortocircuito en el flujo de energía y un incremento en la eficiencia de transferencia energética en ecosistemas ultraoligotróficos y con alta
dosis de radiación ultravioleta (RUV). Los efectos de la interacción entre RUV y pulsos de P tienen un efecto antagónico sobre la
interacción alga-bacteria, intensificando la interacción comensal-depredadora. A partir de los resultados obtenidos proponemos
un modelo alternativo de flujo de energía para ecosistemas autotróficos de alta montaña.
Palabras clave: Bucle microbiano, Estequiometría, Flujo del carbono, Limitación por Nutrientes, Lagos de Alta Montaña,
Radiación Ultravioleta
INTRODUCTION
The food web constitutes one of the most complex conceptual phenomena in modern Biology
(Pimm et al., 1991). This complexity is increased by the “nodes” of the food web, through
which energy and materials flow, are formed by
individual organisms of diverse species, each
individual is a complex biochemical system, and
every species is the product of ongoing evolutionary change (Holt, 1995). Nevertheless, according to Margalef, knowledge of the relations
among elements is more important to the understanding of a system than knowledge of the precise nature of its constituents (Margalef, 1992).
It is therefore crucial to study the interactions
that take place in the food web –an important
issue in the emerging science of “Biocomplexity” (Michener et al., 2001)–, and to develop
appropriate concepts or key variables for this
purpose. Trophic level (tropho-dynamic view,
Lindeman, 1942) and body size, which control
many of the physiological properties of organisms and influence the trophic structure of the
communities (Rodriguez, 1999), are among two
of the traditionally considered variables. Recently, Sterner & Elser (2002) proposed stoichiometry as a tool for ecological analysis in order to
relate the elemental composition of organisms
(carbon [C], nitrogen [N] and phosphorus [P]) to
their growth rates and resource availability
(Biological stoichiometry; Elser et al., 1996).
Furthermore, stoichiometry permits the examination of the balance of energy (usually characterized in carbon terms) with multiple chemical
elements (e.g., nitrogen and phosphorus) in the
organisms, and of the relationship between ecological interactions and biogeochemical cycles
(Ecological stoichiometry; Sterner & Elser,
2002). More recent studies of food webs suggest
that the supposed complexity of the food web is
more apparent than real (Elser & Hessen, 2005).
On the grounds of biosimplicity, they proposed
the consideration of the Darwinian paradigm in
food web models alongside elemental composition and the laws of thermodynamics. This multiple approach avoids the bias implied by the use
of a single tool to study the complexity of food
webs. As these authors put it, “If your only tool is
a hammer, then every problem looks like a nail”.
This multiple approach has been adopted in the
food web analysis presented in this review.
WHAT WE KNOW AND WHAT WE DO
NOT UNDERSTAND
Food webs in aquatic ecosystems function via
the channeling of energy and the flux of materials among diverse organism assemblages
organized in two chains: the classical grazing
chain (phytoplankton-zooplankton-fish), traditionally considered the main pathway for the
flow of energy within these ecosystems; and the
so-called “microbial loop”, (bacteria-heterotrophic nanoflagellates [HNF] - ciliates). The
latter is parallel to and converges with the grazing chain via zooplankton consumption of
bacteria, HNF, and ciliates (Pomeroy, 1974;
Azam et al., 1983). The clearly established
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Complex interactions in microbial food webs
191
¿Why does the microbial loop fail to develop
in La Caldera Lake?
Figure 1. Inter-annual variability of the food web structure of
the plankton community in La Caldera Lake. AUT, phytoplankton; BAC, bacteria; HNF, heterotrophic nano-flagellates; CIL,
ciliates; ZOO, zooplankton. Variabilidad interanual de la estructura de la red trófica de la comunidad planctónica de la laguna
de La Caldera. AUT, fitoplancton; BAC, bacteria; HNF, nanoflagelados heterotróficos; CIL, ciliados; ZOO, zooplancton.
dominance of the microbial communities in oligotrophic ecosystems (del Giorgio & Gasol,
1995; Gasol et al., 1997; Vadstein, 2000) derives from their ability to consume organic matter and mineral nutrients at higher rates compared to algae (Biddanda et al., 2001). The
dominance of the microbial community in the
metabolic balance of oligotrophic ecosystems
invites an in-depth study of: the functioning of
this compartment, the effects of variations in
abiotic factors (i.e. ultraviolet radiation [UVR],
nutrient availability) on interactions among its
components, and the propagation or attenuation
of stress-factor effects throughout the food
web. The biological simplicity of the high
mountain lakes in Sierra Nevada National Park,
which is mediated by strong physical regulation
and extreme oligotrophic conditions, makes
them an ideal setting for observing and modelling the structure and functioning of the microbial food web. Thus, on an inter-annual scale,
the food web structure of one of these ecosystems, La Caldera Lake (Fig. 1), deviates from
the trends established for oligotrophic ecosystems, with scant microbial loop development,
frequently represented only by bacteria. The
first question that arises is then,
Algae-bacteria interactions have been shown to
be the key link between the classical grazing
chain and the microbial food web, and the coexistence of both trophic levels is a “sine qua
non” condition for the persistence of the ecosystem (Daufresne & Loreau, 2001). Different
analyses in freshwater and marine ecosystems
of the relationship between the biomass/production of bacteria and algae (Cole et al.,
1988; Teira et al., 2001; Carrillo et al., 2002)
have shown a direct relationship between these
organisms. The commensalistic algae-bacteria
relationship based on bacterial dependence on
organic C released by algae has provided the
key explanation of some empirical trends.
Accordingly, it is assumed that bacterioplankton is not limited by mineral nutrients
(Chrzanowski & Grover, 2001) and establishes
an indirect mutualistic interaction with algae,
whereby algae provide carbon to sustain bacterial growth and bacteria recycle mineral
nutrients to support algal growth (Aota &
Nahajima, 2001). This mutualistic interaction
is often established in lakes with negligible
allochthonous C input, absence of anoxic (or
hypoxic) hypolimnion (see discussion in Pace
& Cole, 1994), or a high N:P inorganic ratio
(Lee et al., 1994), all characteristics of high
mountain lakes. Nevertheless, preliminary studies in La Caldera Lake showed a decoupling
between the two communities (Reche et al.,
l996). Since the algae-bacteria relationship is
critical to an understanding of microbial loop
development, deeper analysis of the nature of
this interaction is required.
Daufresne & Loreau (2001) were the first to
propose that the alga-bacteria relationship
varies according to stoichiometry of C and
nutrient content in the matter transferred from
primary producers to decomposers. Because
the composition of the released organic matter
and the nutritional status of the algae are
intrinsically linked (Obernosterer & Herndl,
1997; Caron et al., 2000; Sterner & Elser,
2002), organism stoichiometry may be a useful
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tool to understand algae-bacteria interactions,
since it yields data on nutrient demands and
nutrient limitations to growth. Thus, according
to the elemental C:N:P cell content and the
availability of substrates, bacteria or algae will
be limited by one element or another. For
example, if algae are P-deficient (high C:P),
the organic matter released will be characterized by a high C:P ratio, leading towards a Plimitation for bacteria. The higher nutrient
requirements of bacteria in terms of biomass,
C:P and N:P ratios are met by the immobilization of mineral nutrients (Daufresne &
Loreau, 2001). Therefore, competition results
from a high C:nutrient ratio of the matter transferred between algae and bacteria. Thus, the
algae-bacteria relationship can be changed
from mutualism to competition for mineral
nutrients, especially in oligotrophic environments (Aota & Nahajima, 2001; Cotner &
Biddanda, 2002; Joint et al., 2002). Because of
the dynamic nature of the algae-bacteria interaction, the relative importance of mutualism
versus competition must be studied, as well as
the individual and interactive effects of stressfactors (UVR and P-limitation) on this interaction. These fundamental ecological questions
will be addressed in this review, allowing us to
propose an alternative model for the microbial
food web in high mountain lakes.
Figure 2. Bacterial [3H] TdR incorporation rates in the presence versus the absence of algae under different light treatments. Error bars indicate means ± SD, n = 3. Valores de la
tasa de incorporación de [3H] TdR por bacterias en presencia
vs. ausencia de algas bajo diferentes tratamientos de luz
(UVB+UVA+PAR, UVA+PAR, PAR, Oscuridad). Las barras
indican medias ± SD, n=3.
THE MICROBIAL LOOP DOES NOT
DEVELOP. DAZZLED BY THE LIGHT?
The high mountain lakes at Sierra Nevada receive
a high ultraviolet radiation (UVR) flow due to
their altitude (3000 m) and proximity to the subtropics. On the other hand, the low DOC concentration in the water of these lakes (Reche et al.,
2001), a consequence of the scarce development
of terrestrial vegetation in their catchment areas,
allows for a high level of UVR penetration into
the water column. Therefore, the low development of the microbial loop might be related to the
direct negative effects of ultraviolet-B radiation
(<320 nm) (UVB) on these communities. Experimental evaluation of the bacterial activity under
different light qualities confirmed our initial
hypothesis. Thus, UVB reduced bacterial [3H]
TdR incorporation rates by 39-87 % at the surface, although it had no significant effect at intermediate layers (Fig. 2) (Carrillo et al., 2002; Medina–Sánchez et al., 2002). In contrast, either
UVA and photosynthetic active radiation (UVA +
PAR) or PAR alone exerted a stimulatory effect
on [3H] TdR incorporation rates (Fig. 2). This
bacterial response was interpreted as the result of
photorepair mechanisms (Kim & Sancar, 1993;
Vincent & Roy, 1993), which were especially efficient under high light intensities (Carrillo et al.,
2002). This photorepair effect is translated into a
net stimulatory full-sunlight effect (UVB + UVA
+ PAR) on bacterial growth (Fig. 2). Nevertheless,
an adequate supply of photosynthetic carbon from
algae is required to support a net increase in bacterial growth after the recovery of UVB-damaged
bacteria. In fact, in experiments made in the
absence of algae, this potential recovery did not
manifest as enhanced bacterial growth (Fig. 2)
(Medina-Sánchez et al., 2002).
Can photosynthetic C supply be inhibited by
ultraviolet radiation? Our experimental results
showed that UVR inhibits primary production
(PP) but increases excretion of organic carbon
by algae (EOC) (Carrillo et al., 2002). A higher
proportion of this released C is assimilated by
bacteria (% PEA) under UVA + PAR or PAR
rather than under UVB radiation. These data
are compatible with the mechanism proposed
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Figure 3. Annual relationship between Bacterial Production
and Excretion of Organic Carbon (EOC) for the entire water
column (A), and discriminating between depths in 1997 (B).
Error bars indicate means ± SD of bacterial production, n = 3.
Relación anual entre Producción Bacteriana y Excreción de
Carbono Orgánico (EOC) en la columna de agua completa
(A) y discriminando entre profundidades (B). Las barras indican medias ± SD de producción bacteriana, n=3.
by Berman-Frank & Dubinsky (1999) for the
regulation of algae-bacteria interaction in poormineral nutrient ecosystems with high UVR
flow, and also with the stimulatory effect on
bacterial [3H] TdR incorporation rates under
photorepair light (UVA and/or PAR). The consistency of the results obtained (increased net
bacterial activity) using two different methodological approaches suggests a relationship between the consumption of bacterial C and the
ability to channel it to new bacterial biomass.
For this reason, our group proposed the PB:
PEA ratio (as percentage; % CUEb) to quantify
the use of photosynthetic carbon by bacteria.
Interestingly, UVA + PAR or PAR were shown
to increase carbon use efficiency (Carrillo et
al., 2002). Therefore, a net increase of bacterial
production (PB) under high full-sunlight inten-
193
sity is possible because the negative effect of
UVB radiation on bacteria is counteracted by
the positive effect of UVA and PAR (Kim &
Sancar, 1993; Kaiser et al., 1997; Davidson &
Van der Heijden, 2000), and the presence of
higher photosynthetic C availability. A net
positive bacterial response to full-sunlight
would lead to a coupling between PB and PP
(and EOC) (Fig. 3). In fact, when PB measurements were obtained under full-sunlight instead
of the usual darkness incubations (Fig. 4), a
coupled algae-bacteria relationship was verified. From all of the data above, we conclude
that full-sunlight has no negative effect on bacterial development (basis of microbial loop)
and increases bacterial dependence on organic
carbon released by algae (commensalism).
Moreover, the absence of the algal fraction produced a highly significant decrease in PB under
different regions of the spectral solar radiation,
confirming the dependence of bacteria on carbon released by algae (“without you I cannot
live”) (Fig. 2). Further questions arose from
these results: (i) What other factors, besides
light quality, restrict bacterioplankton and
microbial loop development in La Caldera
Lake?; (ii) How prevalent is commensalism
versus competition between algae and bacteria
during the ice-free period? A more detailed
analysis of the results obtained in this ecosystem revealed that (i) the stimulation of bacterial
growth under full-sunlight only occurred when
bacterial elemental composition was balanced
and therefore suitable for growth (N:P ⭐ 20-24
sensu Chrzanowski et al., 1996) (MedinaSánchez et al., 2002) and (ii) algae can regulate
C-release depending on their elemental composition and the nutrient-availability in lake
water (Berman-Frank & Dubinsky, 1999; Villar-Argaiz et al., 2002a). This ability permits
algae to modify the strength of their commensalistic relationship with bacteria and can even
lead to a competitive interaction (Reche et al.,
1997; Villar-Argaiz et al., 2002a). These results indicate that mineral nutrient availability
and algal and bacterial elemental composition
can generate important constraints in the development of the microbial loop.
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Figure 4. Relationship between Bacterial Production and
Excretion of Organic Carbon (EOC) measured at upper depths
in 1997, with the bacterial production data obtained under full
sunlight or dark incubations. Bars indicate means ± SD of
EOC, n = 3. Relación entre Producción Bacteriana y Carbono
Orgánico Excretado (EOC) en la capas superficiales del lago
en 1997, con medidas de producción bacteriana obtenidas en
incubaciones con luz completa y oscuridad. Las barras indican medias ± SD de EOC, n=3.
DEVELOP. IS IT NUTRIENT LIMITED?
Most high mountain lakes in Sierra Nevada are
chronically P-deficient ecosystems (DIN:TP>
40; Carrillo et al., 1996; Morales-Baquero et
al., 1999). This strong limitation may be the
cause of the low bacterial and microbial loop
development. Nevertheless, our studies have
shown that variances in bacterial activity or biomass in La Caldera Lake are not explained by
the concentrations of Total Phosphorus (PT) or
Total Dissolved Phosphorus (TDP) (Reche et
al., 1997; Medina-Sánchez et al., 1999; Carrillo
et al., 2002). Moreover, P-rich atmospheric
inputs (Saharan dust) (Carrillo et al., 1990a;
Villar-Argaiz et al., 2001, 2002a; PulidoVillena, 2004) are largely (60-80 %) incorporated into the algal fraction (Fig. 5) (Villar-Argaiz
et al., 2001; Villar-Argaiz et al., 2002b). This
preferential algal P-consumption has been
demonstrated in natural observations (Fig. 5)
and after experimental nutrient amendments in
the short-term (hours) (Villar-Argaiz et al.,
2002a; Medina-Sánchez et al., 2002), mediumterm (days), and long-term (weeks-months).
These results are consistent with the greater abi-
lity of algae versus bacteria to incorporate limiting nutrients at high P-substrate concentrations
(Tarapchak & Moll, 1990). These mechanisms
undoubtedly offer an ecological advantage for
algae since they can grow by consuming intermittent P-pulses associated with allochthonous
inputs (Cotner & Wetzel, 1992; Duarte et al.,
2000) and P recycled by zooplankton (Carrillo
et al., 1995, 1996; Reche et al., 1997). This
generalized response to enrichment suggests
that P-availability controls algal biomass in La
Caldera Lake but is not a leading factor in the
restriction of bacterial activity or biomass on a
seasonal or experimental scale.
The stimulatory effect of P on primary producers, in some conditions, is propagated throughout the microbial loop, allowing ciliate development. C-flux quantification allowed us to
propose that the complex microbial loop is established by atmospheric nutrient inputs (MedinaSánchez et al., l999; Villar-Argaiz et al., 2002a)
or P-addition, which stimulate algal growth, and
algae-released organic carbon (Medina-Sánchez
et al., 2006). The increase in C-availability
enhances bacterial production (without changes
in bacterial abundance or biomass), enabling
ciliate development (Medina-Sánchez et al.,
Figure. 5. Temporal distribution of phosphorus in particulate
(bacteria, phytoplankton, and zooplankton) and dissolved
(TDP) fractions during 1996 in La Caldera Lake. Arrows indicate humid P inputs in the Lake area. Distribución temporal de
fósforo en las fracciones particuladas (bacteria, fitoplancton y
zooplancton) y disuelta (TDP) durante el año 1996 en la laguna de La Caldera. Las flechas indican entradas atmosféricas
húmedas de fósforo en el área de la laguna.
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1999; Villar-Argaiz et al., 2002a). These results
confirm that nutrient enrichment can lead to a
lengthening of the microbial food chain
(Samuelson et al., 2002; Vaqué et al., 2003).
Therefore, the scarce microbial loop development can at least be partly ascribed to the severe
oligotrophic conditions of this ecosystem.
Nevertheless, propagation of the P-pulse effect
to ciliate development does not always take
place in the system, raising the question as to
what constrains the development of a complex
microbial loop. Microcosm studies showed that
balanced algal (N:P = 16) and bacterial (N:P
⭐ 20-24) elemental compositions are required,
providing further evidence that bacterial elemental composition plays a key role in determining
the nature of the algae-bacteria relationship and
the subsequent development of ciliates.
Thus, when bacteria (N:P ⭐ 20-24) and
algae (N:P ≈ 12) were P-sufficient (i.e., at
thaw), algal P-incorporation rates were higher
than bacterial rates after P addition (Fig. 6).
This P-incorporation resulted in a substantial
enhancement of algal abundance (Carrillo et
al., submitted). P-enriched algae pursuing a net
growth strategy may produce and release
enough organic material to meet the P-requirements of bacteria, obviating the need to compete for inorganic nutrients (Caron et al., 2000).
In this scenario, bacteria would function as a
carbon-link to other trophic levels (ciliate development during thaw experiment) (Fig. 6). In
contrast, growth of P-deficient algae (mid-summer) after nutrient pulse generates algal
“bloom-growth” that limits the release of organic carbon (Villar-Argaiz et al., 2002a), thereby
intensifying the mutualism-commensalism
relationship and producing a tendency to bacterial P impoverishment (to see also Caron et al.,
2000) that ultimately restricts the development
of the microbial loop (Fig. 6).
When bacteria were P-deficient (N:P⭓2024, late ice- free period), bacterial P-incorporation rates were significantly higher than algal
rates (Fig. 6). Bacterial and primary production, were stimulated after the nutrient pulse,
and competition for P was the prevalent relationship between algae and bacteria (Villar-
195
Figure 6. Phytoplankton and bacterial P-incorporation rates in
un-enriched (control) and enriched (N:P16 and N:P5) treatments after short-term incubations [24 h in thaw experiment
and 48 h in mid and late ice-free period experiments] (upper
panel). Ciliate abundance in un-enriched (control) and enriched (N:P16 and N:P5) treatments after 15 days incubation
period in the thaw, mid and late ice-free period experiments
(lower panel). Error bars indicate means ± SD. Asterisks indicate significantly higher values relative to the controls (onetailed t-test). *p<0.05; **p<0.01; ***p<0.001. Tasas de incorporación algal y bacteriana de fósforo en tratamientos no
enriquecidos (control) y enriquecidos en fósforo (N:P16 and
N:P5) [después de 24 horas de incubación en el deshielo y
48 horas en mitad y al final del periodo libre de hielo] (gráfica superior). Abundancia de Ciliados en tratamientos no enriquecido (control) y enriquecido en fósforo (N:P16 and N:P5) al
final de 15 días de incubación en el deshielo, mitad y final del
periodo libre de hielo (gráfica inferior). Las barras de error
son medias ± SD. Los asteriscos indican valores significativamente mayores en relación con el control (t-test).
*p<0.05; **p<0.01; ***p<0.001.
Argaiz et al., 2002a). In spite of the addition of
P, bacterial elemental composition did not reach
the threshold for balanced growth (Fig. 6). To
summarize, the algae-bacteria interaction shifted from mutualism during most of the ice-free
period to competition for available P towards
the end of the ice-free period. Therefore, limitation by mineral nutrients can be a transient phe-
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nomenon that lasts for a shorter time than the
total turnover of the system (Daufresne &
Loreau, 2001). This interpretation is further
sustained by seasonal and inter-annual observations. First, the algal elemental composition is a
good predictor of algal growth rate, whereas the
bacterial N:P ratio is not related to bacterial
growth rate (Fig. 7). This lack of relationship
indicates that bacteria are limited by an element
other than P, despite the strong P-limitation in
La Caldera Lake (DIN:TP>40). Second, there is
a coupled relation between bacterial and primary production based on a direct dependence
on the organic carbon released by algae (see
above) (Fig. 3). The prevalence of commensalism over competition contrasts with the clear
bacterial P-limitation established for other oligotrophic ecosystems (Rivkin & Anderson,
1997; Vrede, 1999; Caron et al., 2000; Carlsson
& Caron 2001) and with the Light Nutrient
Hypothesis (LNH) (Sterner et al., 1997) on the
structure and functioning of aquatic ecosystems. We consider that the main difference with
our approach is that the above authors did not
include irradiance quality effects within the
predictive conceptual framework of the LNH.
This is especially important because we have
shown that UVR favours algae-bacteria commensalism. Therefore, we can affirm that
mutualism-commensalism is more prevalent
than competition between algae and bacteria.
However, given that bacteria can tolerate
UVR-stress and respond positively to a shift in
EOC-availability (basis of the predominantly
mutualistic relationship), the question arises
as to why they do not reach higher development if C exceeds bacterial requirements
(Carrillo et al., 2002; Medina-Sánchez et al.,
2006) and mineral nutrients do not strongly
limit their growth (Villar-Argaiz et al., 2002a).
Another relevant question is why bacterial production, C-assimilation and use efficiencies
show such low or even decreased values
(Carrillo et al., 2002; Medina-Sánchez et al.,
2004) after the addition of P (Medina-Sánchez
et al., 2002; 2006). It appears that other controls are responsible for the scarce abundance
and activity of bacteria in La Caldera Lake.
This paradoxical scenario calls for a shift from
a simple trophic to a metabolic approach, whereby the “algal” community is divided into
strict-autotrophs or mixotrophs.
DEVELOP. ARE ALGAE ANIMALS
OR PLANTS?
The traditional approach to the energy and matter that flows through an ecosystem involves the
classification of organisms into two main trophic groups: osmotrophs/phototrophs (primary
producers) and phagotrophs/heterotrophs (consumers). This basic division does not always
match the reality of the microbial world because
some organisms, such as protista, combine both
Figure 7. Relationship between algal (upper panel) and bacterial (lower panel) growth rate and elemental composition (N:P
ratio) of algae and bacteria, respectively, in nutrient-pulse
experiments for the mid-and late ice free period experiments.
Relación entre la tasa de crecimiento y la composición elemental (razón N:P) de algas (gráfica superior) y bacterias
(grafica inferior), en experimentos de nutrientes pulsados llevados a cabo en la mitad y final del periodo libre de hielo.
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197
Figure 8. Bacterial production measured as BPA, BPB, and BPA+B in light treatments (‘dark’ and ‘full light’) in different
moments during the ice-free period. August and September. BPA, Bacterial production incorporated in algal fraction; BPB, bacterial production not incorporated by algae; BPA+B,bacterial production in both the algal and bacterial fractions. Error bars indicate
means ± SD. Producción bacteriana medida como BPA, BPB y BPA+B en tratamientos de luz (oscuridad y luz completa) en distintos momentos durante el periodo libre de hielo. BPA, Producción bacteriana incorporada en la fracción algal, BPB, Producción
bacteriana no incorporada por las algas; BPA+B, Producción bacteriana en las fracciones algales y bacterianas juntas. Las
barras de error son medias ± SD.
trophic abilities and are designated mixotrophs
(Sanders, 1991; Jones, 1997). By definition, a
mixotroph is an organism in which both photosynthesis and phagotrophy are possible (Sanders, 1991). Although mixotrophic metabolism
is an energetically more demanding strategy
(maintaining both photosynthetic and phagotrophic systems in the same cell) compared with
strictly autotrophic or heterotrophic metabolism, mixotrophy has evolved as an adaptive
strategy for growth in adverse conditions.
Mixotrophs constitute a functional type usually
present in oligotrophic and dystrophic lakes
(Salonen & Jokinen, 1988; Jansson et al., 1996)
and at high latitudes (Duthie & Hart, 1987;
Eloranta, 1989, 1995; Lepistö & Rosenström,
1998), and they are frequently dominant in the
phytoplanktonic community of high mountain
lakes (Sánchez-Castillo et al., 1989; Carrillo et
al., 1990 b, 1991 a, b, 1995; De Hoyos et al., 1998;
Straskrabová et al.,1999) and in marine and freshwater ecosystems (Raven, 1997; Sanders et al.,
2000; Sherr & Sherr, 2002).
The dual nature of mixotrophic (autotrophic
and heterotrophic) metabolism offers an advantage in ecosystems where light or mineral nutrients
are scarce, since the phagotrophic ability enables
supplementation of autotrophic growth with the
extra C and nutrient contents of prey, e.g., picoplankton (Rothhaupt, 1996 a, b; Raven, 1997). In
this way, mixotrophs can constitute a key functional type for the structure of the community
and the flow of energy through the food web.
They not only contribute to the input of autochthonous C into the ecosystem but also, by their
interaction with picoplankton, constitute a
mechanism for the transfer of nutrients and
energy towards upper trophic levels, such as zooplankton (Thingstad et al., 1996; Sherr & Sherr,
2002). On the other hand, it was recently shown
that mixotrophy is an adaptive strategy both to
deficient environmental conditions and also to
excess-induced stress conditions. Thus, MedinaSánchez et al., (2004) found higher bacterial
consumption rates by mixotrophs exposed to
stressing-sunlight (UVB+UVA+PAR) (Fig. 8),
suggesting that bacterivory by mixotrophs allows
them to acquire organic carbon and mineral
nutrient contents from their prey under inhibitory
light conditions (+UVR) for C fixation (Carrillo
et al., 2002) and acquisition of dissolved
nutrients (Hessen et al., 1995; Döhler, 1997).
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With this background, we propose that mixotrophy is an adaptive strategy of “algae” to stress
by ultraviolet radiation (Medina-Sánchez et al.,
2004). Mixotrophy implies a strong predatory
control on bacteria, especially on the most active
bacteria (Sherr & Sherr, 2002).
Since algae act both as predators of, and C
suppliers bacteria (dual control), we estimated
the relative importance of Predation vs. Resource control by defining a new variable. This
variable, designated “algal control”, is the ratio
between the potential mixotrophic consumption
of bacterial production and the photosynthetic
C supplied (i.e. PMCBP:EOC, Medina-Sánchez
et al., 2004). Figure 9 shows the strong inverse
relationship between bacterial production and
“algal control” when the algal community is
dominated by mixotrophs. Therefore, predation
by mixotrophs accounts for the scarcity of bacteria (105 cell/ml, lowest end of the values
recorded for oligotrophic ecosystems; Cotner &
Biddanda, 2002) and the low carbon use efficiency by bacteria.
Hence, a complex regulation is established
between algae and bacteria, namely “neither
Figure 9. Intraannual relationship between bacterial production and ‘algal control’ measured as the ratio between potential mixotrophic consumption of bacterial C and photosynthetic C supplied by algae (i.e. PMCbp:EOC). Dashed lines
indicate 95 % confidence intervals around the fitted regression line (solid line). Variación estacional en la relación entre
la producción bacteriana y el ‘control algal’ medido como
la razón entre consumo de carbono bacteriano por potenciales mixótrofos y el carbono fotosintético aportados por las
algas (i.e. PMCbp: EOC). Líneas discontinuas indican el
intervalo de confianza del 95 % en relación con la línea de
ajuste por regresión (línea continua).
with nor without you” (Medina-Sánchez et al.,
2004). Bacteria are preyed upon by algae (“with
you I die”) and simultaneously depend on the
organic carbon released by the algae (“without
you I cannot live”) (Carrillo et al., 2002;
Medina-Sánchez et al., 2002, 2004). According
to Thingstad et al. (1996), mixotrophs can take
advantage in a situation where phytoplankton
contribute the organic substrate that bacteria
demand, while bacteria can favourably compete
with algae for mineral nutrients (Currie & Kalff,
1984; Cotner & WetzeL, 1992) and finally,
mixotrophs consume these enriched bacteria.
Thus, when sunlight is intense and mineral
nutrients are scarce, mixotrophic algae are able
to grow because they can feed bacteria with
“cheap” organic carbon and consume the
“expensive” mineral nutrients packaged in the
bacteria (sensu Thingstad et al., 1996).
This scenario is consistent with (i) the ability
of algae to regulate photosynthetic C release
depending on their elemental composition
(Villar-Argaiz et al., 2002a); (ii) the ability of
bacteria to grow under high full-sunlight intensities (Carrillo et al., 2002, Medina-Sánchez et al.,
2002); and (iii) the decrease in bacterial production in the presence of algae after a nutrient pulse
(Medina-Sánchez, 2002, 2006). Furthermore,
mixotrophs can retain or release the P contained
in bacteria depending on their predominant nutrition mode (autotrophic vs. phagotrophic) and
nutrient requirements (Rothhaupt, 1997).
These abilities can constitute mechanisms that
modulate the competition with bacteria for mineral nutrients, suggesting a higher complexity of
the algae-bacteria interaction where algae exert
the main control. Knowledge of this complex
algae-bacteria interaction further advances the
understanding of the planktonic structure and
energy flow in this kind of aquatic ecosystem:
1) Thus, mixotrophs can overcome the harmful
effects of UV radiation and displace: (i) strict
autotrophs, less competitive at low concentrations of dissolved mineral nutrients
(Rothhaupt, 1996a; Christaki et al., 1999),
(ii) picoplankton autotrophs, better adapted to
the uptake of dissolved nutrient at low con-
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centrations but more sensitive to the inhibitory effect of UVR (Callieri et al., 2001), and
(iii) heterotrophic nanoflagellates (HNF),
which depend on a higher minimum threshold
of bacterial abundance compared with mixotrophs (Havskum & Riemann, 1996).
2) With respect to the energy flow, mixotrophic
bacterivory acts as a “by-pass” for the flow of
energy and nutrients towards the grazing chain
(Medina-Sánchez et al., 2004). Hence, mixotrophs would occupy the niche of microheterotrophs and produce a weakening of the traditional microbial loop (Medina-Sánchez et al.,
2004). Besides, mixotrophic metabolism may
improve the efficiency for the transfer of energy
towards upper trophic levels, because the number of trophic steps has been reduced due to the
by-pass from the heterotrophic microbial chain
to the grazing chain (Fig. 10) (Rivkin &
Anderson, 1997; Medina-Sánchez et al., 2004).
Considering all of the above, mixotrophy may at
least in part explain the scarce development of
the microbial loop in La Caldera Lake.
199
Nevertheless, several important questions are raised by the fact that mixotrophy is an advantageous
metabolism strategy for a system stressed by UVR
and oligotrophy and that P-atmospheric input of
Saharan origin are relatively frequent: How do
allochthonous nutrient inputs affect mixotrophs?
Do P pulses interact with UVR in modulating this
metabolism? The addressing of these questions
responds to the current demand for research into
the interactive effects of multiple factors on the
functioning and structure of ecosystems.
DEVELOP. INTERACTION OF LIGHT
AND NUTRIENTS?
The interaction between UVR and atmospheric
nutrient inputs can affect species performance
and their relationships in ways that cannot be
predicted from single-factor analysis, since the
effects of two or more factors are non-additive.
The interaction among distinct factors can
Figure 10. Proposed model of carbon flux through microbial food web in oligotrophic clear-water high mountain lakes. The curved
arrows indicate reinforcement of the by-pass from bacteria to mixotrophs as a response to the interactive effect between UV radiation and the nutrient-pulses added. Thickness of arrows indicates the relative importance of C flux. Modelo propuesto de flujo de
carbono en la red trófica microbiana de lagos oligotróficos claros de alta montaña. Las flechas curvas indican refuerzo del corto
circuito desde bacterias a mixótrofos como respuesta al efecto interactivo entre la radiación ultravioleta y los aportes pulsados de
nutrientes. El grosor de las flechas indica la importancia relativa del flujo del carbono.
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Carrillo et al.
either shift the sign of the effect of a single-factor (antagonistic effect) or, on the contrary,
accentuate this effect (synergetic effect).
Therefore, the analysis of multiple factors at
different rates and scales is crucial for building
a realistic model of the functioning of food
webs (Breitburg et al., 1999; Folt et al., 1999;
Xenopoulos et al., 2003; Medina-Sánchez et al.,
2006). The increasing intensity and frequency
of atmospheric material inputs from the Sahara
desert to Sierra Nevada is associated with periodic climatic variations (North Atlantic
Oscillation, NAO) and movements of the Intertropical Zone Convergence (ITZC). Hence,
planktonic communities may be controlled not
only by the “harmful” effect of UVR but also by
the “stimulatory” effect of allocthonous inorganic nutrients. According to our initial hypothesis, P-pulses would reduce the negative effects
that UVR exerts on algae, thereby decreasing
the bacterivory rate (associated with mixotrophy). As a consequence, the nutrient pulses
would modify the algae-bacteria relationship by
shifting it from dual control (with bacterivory)
to commensalism. Our investigations on short
temporal scales showed that:
1) The elemental composition of algae and bacteria plays a key role in the ecological and biological response to light quality variation and
nutrient availability. UVR x P interactions have
an antagonistic effect on primary production,
C-release by autotrophs, and bacterial production on both communities in un-enriched conditions. This antagonistic effect implies the
attenuation or elimination of UVR effects.
2) The interactive effect between solar radiation
and atmospheric nutrient inputs on the algaebacteria link reinforces the dual control that
algae exert on bacteria, i.e., it increases both
the carbon released by algae and the mixotrophic predation on bacteria (MedinaSánchez et al., 2006). The by-pass is therefore enhanced, with consequences for the
development of the microbial loop.
Finally, our results have implications at the
ecosystem level. Thus, the attenuation of the
harmful effects of UVR by P-inputs may smooth
UVR stress, which is relevant in a scenario of climatic change with increased UVB fluxes. This
buffering effect would be more accentuated in
clear-water ecosystems, where the main nutrient
inputs have an atmospheric origin, such as in the
tropical Atlantic Ocean (Prospero & Lamb, 2003,
see also NOAA htpp://toms.gsfc.nasa.gov/aerosol). In the context of carbon flux, this buffering
effect of harmful UVR at the base of both aquatic
trophic webs (grazing chain and microbial loop)
would favor the diversion of a higher proportion
of released carbon flux through biotic webs, trapping carbon that would otherwise be susceptible
to photolysis and be lost to the atmosphere.
In spite of the light thrown by this study on
the processes that regulate the pelagic community of clear-water high mountain lakes, many
areas remain in the shadows and new questions
have arisen, briefly summarized as follows:
1. Do the results obtained over short time scales apply to long-term scales (days, weeks or months)?
2. How do UVR x nutrient interactive effects
propagate along the grazing chain?
3. How universal are the patterns established for
high mountain lakes with high UVR doses?
Can structural tendencies established across
trophic gradients be extrapolated to high-UVRflux oligotrophic ecosystems?
ACKNOWLEDGMENTS
We sincerely acknowledge L. Cruz-Pizarro,
R. Morales-Baquero, P. Sánchez-Castillo and
I. Reche for their contribution to the database.
We thank M. J. Villalba assistance in the field.
We are indebted to the staff of the Radiopharmacy Department of Granada University for
contributing their laboratory and experience.
This study was supported by Acidification of
Mountain Lakes: Palaeolimnology and Ecology
(Al:PE2 EC-ENVIROMENT) Project, Contract
Number EV5V-CT92-0205; MOLAR: Measuring and modelling the dynamic response of
remote mountain lake ecosystems to environmental change; a programme of mountain lake
Limnetica 25(1-2)02
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research, Project EC Contract Number ENV4
CT95-0007); the Spanish Ministry Science and
Technology Project AMB 0996 (to P.C.), Project
REN2001-2840 (to PC). Red UVIFAN: Project
FEDER 1FD97-0824, and the Spanish Ministry
of Environment RPN25/2003 (to PC).
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Limnetica, 25(1-2): 205-216 (2006)
Algae in the motion: Spatial distribution of phytoplankton in thermally
stratified reservoirs
E. Moreno-Ostos1, 2, L. Cruz-Pizarro2, A. Basanta-Alvés3, C. Escot3 & D. G. George4
1
Flumen Research Group. Dept. Ecology. University of Barcelona. Av. Diagonal 645. 08028. Barcelona
(Spain)
2 Water Research Institute. University of Granada. C/Ramón y Cajal, 4. 18071. Granada. (Spain)
3 Aquatic Ecology Station. EMASESA. Avenida Leonardo da Vinci. 41092, Sevilla (Spain)
4 Centre for Ecology and Hydrology. Lancaster Environment Centre. Library Avenue, Bailrigg LA1 4AP
England (UK)
Corresponding Author: [email protected]
ABSTRACT
Phytoplankton spatial distribution patterns in four Andalusian reservoirs with different physical characteristics are described
and evaluated in this work. Both vertical and horizontal distribution patterns are presented in order to demonstrate that thermally stratified reservoirs are dynamic, complex and heterogeneus ecosystems. Vertically, phytoplankton patchiness was
physically controlled by turbulent mixing and light climate and biologically determined by the hydromechanical characteristics
of each microalgae functional group. The horizontal distribution of phytoplankton was the result of the interaction between the
wind-induced advective transport of water masses and the vertical distribution of each algal group. The precise knowledge of
the phytoplankton spatial distribution patterns and their responsible agents constitutes a powerful tool for a limnologicallybased dynamic reservoir management and must be integrated in reservoir water quality monitoring procedures. The use of fastresponse and high-resolution technologies, such as in vivo and in situ spectrofluorimetry, should significantly helps in reaching
this objective.
Keywords: Reservoir, Phytoplankton, Heterogeneity, Patchiness, Physical-biological coupling, Spectrofluorimetry.
RESUMEN
En este estudio se describen y evalúan algunos de los principales patrones de distribución espacial del fitoplancton en un conjunto de cuatro embalses andaluces de diferentes características físicas. Los patrones de distribución vertical y horizontal del
fitoplancton obtenidos demuestran que los embalses térmicamente estratificados constituyen un tipo de ecosistema especialmente dinámico, complejo y heterogéneo. La distribución vertical del fitoplancton estuvo controlada por agentes físicos como
la mezcla turbulenta de la columna de agua y el clima lumínico al que son sometidas las algas así como por factores biológicos como las características hidromecánicas de los distintos grupos funcionales de fitoplancton. La distribución de las microalgas en el plano horizontal surge como resultado de la interacción entre el transporte advectivo de las distintas masas de
agua inducido por el viento y la distribución vertical de cada grupo algal. El conocimiento preciso de los distintos patrones de
distribución del fitoplancton y de los agentes responsables de los mismos constituye una herramienta de especial utilidad para
una gestión dinámica y basada en principios limnológicos de los embalses y debe ser integrado en los procedimientos de control de la calidad del agua embalsada. El uso de tecnologías de respuesta rápida y alta resolución de toma de datos, tales
como la espectrofluorimetría in vivo e in situ, puede facilitar considerablemente alcanzar este objetivo.
Palabras clave: Embalses, Fitoplancton, Heterogeneidad, Manchas, Acoplamiento físico-biológico, Espectrofluorimetría.
INTRODUCTION
As Professor Ramón Margalef pointed out in his
book La Biosfera: entre la termodinámica y el
juego (1980), ecosystems constitute a complex
organisation structured along two main axis:
time and space. Every living organism contributes to the temporal organisation of the whole
system with a characteristic time scale and also
participates in its spatial organisation as a con-
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sequence of their motility, function and requirements. In nature, populations of most of the species exhibit heterogeneus spatial distributions
with organisms agregating where favourable
conditions for growth, reproduction and survival are found or accumulating in regions where
they are passively transported by water in
motion (Begon et al., 1999).
Spatial heterogenity in the distribution of
organisms in their habitats is considered a key
concept for a number of ecological theories
such as competition, species diversity, sucession, evolution, adaptation, parasitism, population genetics, population growth, predator-prey
interactions and social behaviour (Legendre &
Fortin, 1989; Wiens, 1989). In addittion, boundaries between “homogeneus” regions in the
space are a significant structuring factor for the
ecosystems (Allen, 1977; Legendre et al., 1986)
especially relevant for their dynamics (Legendre
& Demers, 1984). Moreover, some studies reveal a notable contribution of spatial heterogeneity
to the ecosystem stability (Huffaker, 1958; May,
1974; Hassel & May, 1974; Neill, 1990).
The interaction between the hydromechanical
properties of each algal functional group (swimming algae, positively-, negatively- or neutrallybuoyant algae) and the physical conditions of
the waterbody (i.e. turbulence, thermal structure, mixing conditions, advective transport and
light climate, among others) is responsible for
the generation and evolution of the vertical and
horizontal algae patchiness (George & Edwards,
1976; George & Heaney, 1978; George, 1981a;
Moreno-Ostos, 2002; Moreno-Ostos, 2004).
In the case of water supply reservoirs, patches
of non-desirable or harmful algae (such as
Cyanobacteria and some Dinoflagellates) can
induce notorious water quality problems
increasing water treatment costs.
An adequate and scientifically-based management of the stored water quality requires a
profound knowledge of the spatial and temporal dynamics of phytoplankton in reservoirs,
taking into consideration their intrinsic heterogeneus character from the appropriate scales
and using modern technologies to obtain data
series under an optimal spatial and temporal
resolution (Moreno-Ostos et al., 2004a;
Moreno-Ostos et al., 2005). This is a particularly relevant topic in the case of the reservoirs located in the Mediterranean arid and
semiarid region, where water is a scarce
resource. In this work we describe spatial distribution patterns of phytoplankton (both in
the vertical and horizontal axis) found in four
Andalusian reservoirs, and analyze the underlying physical-biological mechanisms that create them. Reservoir Limnology and the study
of the spatial distribution of organisms in their
habitat are among the main “Margalef ian”
subjects and, no doubt, the results of our research are inspired on his pioneer studies.
Study site
The spatial distribution of different phytoplankton functional groups is analyzed in four
Andalusian reservoirs, with varying physical
and morphometrical characteristics (Fig. 1 and
table 1). In three of them, we focus on the distribution of phytoplankton groups in the vertical.
In El Gergal reservoir we describe both vertical
Table 1. Main morphometrical and light climate characteristics of the studied reservoirs. Principales características morfométricas y de clima
lumínico de los embalses analizados
Reservoir
Location
Surface
(ha)
Quentar
Bermejales
Béznar
El Gergal
Granada
Granada
Granada
Sevilla
41.6
561.8
170.0
250.0
Volume
(hm3)
13.6
102.6
54.0
35.0
Maximum depth
(m)
100.0
48.0
100.0
35.0
Mean depth
(m)
Water transparency
(m)
32.7
18.3
31.8
14.0
6.04
3.5
2.0
1.7
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and horizontal distribution patterns. All surveys
were conducted during July and August 2003
and 2004, coinciding with the maximum thermal stability of the water column.
Physical determinations
The mixed layer depth (Zmix) was estimated,
from vertical temperature profiles, as the depth
of the maximum thermal gradient. The euphotic layer depth (Z eu) was determined as in
Walker (1980), from Secchi disk depth observations collected with a 20 cm diameter white
disc. Water motions in El Gergal reservoir
were characterized using a Nortek Acoustic
Doppler Current Profiler (ADCP) and surficial
free-running drogues (see George, 1981b and
Moreno-Ostos, 2004).
207
Vertical and horizontal distribution
of phytoplankton
A recently developed fast-response high-resolution in vivo and in situ spectrofluorimetric probe
(bbe Fluoroprobe, Moldaenke. Fig. 2) was used
to examine the vertical and horizontal distribution patterns of different phytoplankton functional groups in the studied reservoirs. The probe
directly measures total Chlorophyll-a concentration in the water and the fraction corresponding
to four different functional groups of algae
(Chlorophyceae, Cyanobacteria, Diatoms and
Cryptophytes). To differentiate functional groups
of phytoplancton the bbe Fluoroprobe uses
5 Light Emiting Diodes (LEDs) for fluorescence
excitation. The LEDs emit pulsed light at selected wavelenghts (450 nm, 525 nm, 570 nm,
Figure 1. Location and morphology of the studied reservoirs 1) El Gergal reservoir; 2) Quentar reservoir; 3) Bermejoles reservoir;
4) Béznar reservoir. (Andalusia map not to scale.) Localización y morfología de los embalses estudiados 1) Embalse de El Gergal
2) Embalse de Quentar; 3) Embalse de Bermejoles; 4) Embalse de Béznar. (El mapa de Andalucía no es a escala.)
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590 nm and 610 nm) while fluorimetric emission
is measured at 680 nm by photomultiplier at an
angle of 90 degrees to the exciting light source
and directly transformed to Chl-a concentration
by the bbe Fluoroprobe 1.4 software. The probe
is also equiped with a temperature sonde so coupled thermal structure and algal distribution profiles/transects can be displayed.
In order to describe the vertical distribution
of phytoplankton, free-falling mode vertical
spectrofluorimetric profiles were collected from
fixed sampling stations located at the maximum
depth points of each reservoir. The horizontal
distribution of algae in El Gergal reservoir, on
the other hand, was characterized from a set of
surficial (1 meter depth) spectrofluorimetric
data collected at a grid of up to 30 GPS-georeferenced sampling stations. Due to the reduced
size of the reservoir and the use of a fast-response probe the time spent to complete a whole
horizontal survey was minimized and the
collected data can be considered as synoptic.
For further details on bbe Fluoroprobe technical characteristics and its applications to study the
phytoplankton spatial dynamics in reservoirs see
Beutler et al. (2002) and Moreno-Ostos (2004).
RESULTS
Vertical patchiness
Figure 2. bbe Fluoroprobe components (modified from bbe
Moldaenke user manual). 1) Screw. 2) Connector. 3) Eyelet. 4)
Screen. 5) Detector window. 6) LED window. 7) Temperature
sensor. 8) Transmission window. 9) Preassure sensor
Componentes del bbe Fluoroprobe (modificado del manual de
usuario de bbe Moldaenke). 1) Tornillería de sujección. 2)
Conector. 3) Arándela de sujección. 4) Pantalla de protección.
5) Ventana de detección. 6) Ventana de LEDs. 7) Sensor de
temperatura. 8) Ventana de transmisión. 9) Sensor de presión.
Figure 3 shows the vertical distribution of temperature, total Chl-a and the biomass corresponding to different algae functional groups in
the studied reservoirs. The described patterns
are closely related to the physical (light climate,
mixing regime) and the biological (composition
of the phytoplankton community) characteristics of each ecosystem.
In Quentar, a deep and meso-oligotrophic
reservoir, Zmix and Zeu are similar in magnitude (16.0 meters and 16.3 meters respectively). As a result, there exists a marked Deep
Chlorophyll Maximum (DCM) located at the
depth of the thermocline, i.e. 16 meters.
Phytoplankton finds in the thermocline an
appropriate site for growth and development
since light levels are adequate for photosynthesis, nutrients from the hypolimnion are frequently entrained into the epilimnion and the
stability is large. As for the composition of
the phytoplankton community, Diatoms and
Cryptophytes –two light-stress tolerant algae
groups- are the most abundant groups. While
Diatoms were mainly located just above the
thermocline, the Cryptophytes (more tolerant to
reduced light levels) were mostly below it.
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Figure 3. Vertical distribution of temperature (left), total Chl-a (centre) and different phytoplankton functional groups (right) in
the studied reservoirs. Horizontal dotted line represent the euphotic layer depth. Distribución vertical de la temperatura (izquierda), Chl-a total (centro) y distintos grupos funcionales del fitoplancton (derecha) en los embalses estudiados. La línea de puntos
horizontal representa la profundidad de la zona eufótica.
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Light attenuation was higher in the mesotrophic
Bermejales reservoir than in Quentar and the
thermocline was shallower (around 10 meters).
Here, Zmix and Zeu also have similar values
(10.0 meters and 9.5 meters respectively). The
vertical profile of Chl- a shows a DCM at 10 m.
Chlorophytes, Diatoms and Cryptophytes were
the main functional groups at the time when the
profile was taken. As in Quentar, the Diatoms
and Cryptophytes were most abundant in the
DCM region. Chlorophytes, on the other hand,
are green algae demanding higher light levels
and were consequently detected only in the
upper layers of the water column.
The thermal structure of Béznar, an eutrophic reservoir, at the time when the profile was
taken, was characterised by a relatively shallow
diel thermocline located 8 meters below the
free surface and a deeper and more stable seasonal thermocline at 18 meters. The bottom of
the euphotic zone was estimated to be around
A)
B)
5.5 meters depth. As a consequence, maximum
Chl-a concentration was found around the diel
thermocline, where settling algae were accumulated. The intense and frequent wind-induced turbulent mixing events recorded in this
reservoir (Rueda, unpublished data) allow the
algae to enter from this main patch to the
euphotic zone thus ensuring their survival
(Margalef, 1983). The vertical distribution pattern was similar for the three phytoplankton
groups found in the reservoir (Chlorophytes,
Diatoms and Cyanobacteria).
The final study case corresponds to El Gergal
reservoir (Zmix=12.0 meters; Zeu=5.0 meters).
In this ecosystem, the dominance of positivelybuoyant Cyanobacteria at the time when the
profile was taken induced the formation of an
intense surficial algal patch mainly composed
by Aphanizomenon sp., while the neutral-buoyant Chlorophytes remained randomly distributed along the water column.
C)
Figure 4. Observed surficial hydrodynamics (A) and simultaneous horizontal variations in water temperature (°C) (B) and Chl-a
(µg l–1) (C) under the influence of a Northeast wind. Note that arrows just show wind and current direction and are not proportional to velocity. Hidrodinámica superficial observada (A) y variación simultánea en el plano horizontal de la temperatura del agua
(°C) (B) y de Chl-a (µg l–1) (C) bajo la acción de vientos de componente Noreste. Los vectores sólo indican las direcciones predominantes adoptadas por el viento y la corriente superficial y no son proporcionales a su velocidad.
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211
Figure 5. Vertical distribution of temperature (°C), Cyanobacteria biomass (µg l–1) and Diatoms biomass (µg l–1) during the late
summer 2003 extensive survey in El Gergal reservoir. Distribución vertical de temperatura (°C), biomasa de Cianobacterias (µg
l–1) y biomasa de Diatomeas (µg l–1) en el embalse de El Gergal durante el muestreo extensivo llevado a cabo a finales de verano
de 2003.
Horizontal patchiness
During summer 2003 and 2004 a series of data
collection campaigns were conducted in El
Gergal in order to describe the horizontal distribution of phytoplankton in a thermally stratified
reservoir and to analyze the mechanisms that
were responsible for the spatial variability of
phytoplankton abundance. In a first survey we
studied the role of wind-induced water movement
on the physical (water temperature) and biological (total Chl-a concentration) horizontal patchiness in the ecosystem. During a second survey,
we focused on the differencial Cyanobacteria and
Diatoms horizontal distribution as induced by the
action of a prevailing North wind. Results from
both surveys are presented below.
The role of wind-induced water movement
on the horizontal distribution of temperature
and total Chl-a
Under the influence of a constant wind of
1.2 ms-1 average module blowing from Northeast (60º) free-running drogues topographical
monitoring revealed the existence of a main surficial water current moving from North to South
and characterised by a marked gyre into the
Cantalobos bay (Fig. 4a).
The horizontal distribution of water temperature was consistent with this hydrodynamic
behaviour and the warmer water masses were
retained into the bay and along the West shore
of the reservoir (Fig. 4b).
Algal biomass acted as a passive tracer of
surficial water movement and remained accumulated in the Cantalobos Bay, generating also
some patches in the West shore associated to the
higher temperature areas (Fig. 4c).
The impact of North wind on the horizontal
distribution of Diatoms and Cyanobacteria
During the survey, wind (blowing from North
with an average module of 2.5 ms-1) was the
only factor inducing movement in the water
mass. Figure 5 shows the vertical distribution of
temperature, Cyanobacteria and Diatoms during
the sampling. Thermal structure was characterised by a marked thermocline at 12 meters depth.
Cyanobacteria were detected in the upper layers
of the water column, while settling Diatoms
were found deeper in the water column and
mainly above the thermocline.
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The hydrodynamical behaviour of the reservoir
under such circumstances is shown in figure 6.
Our ADCP records revealed wind-driven surface
currents from North to South and deeper return
currents above the thermocline. The horizontal
distribution of phytoplankton (Fig. 7) was a
result of the interaction between vertical patchiness and water-mass displacements at different
depths. Positively-buoyant Cyanobacteria were
passively transported by surface currents to the
South region of the reservoir, accumulating
around the dam. By contrast, the negativelybouyant Diatoms were selectively transported
to the riverine region of El Gergal (North) by
the deep recirculating currents, where they
emerged and agregated due to the effect of
upwelling currents and the increasing turbulence
(Moreno-Ostos, 2004). This conveyor belt
hydrodynamical mechanism for the generation
of phytoplankton patchiness has previoulsly
been documented by George & Edwards (1976).
DISCUSSION
The results presented in this paper show that
thermally stratified reservoirs of all trophic states must be considered as spatially heterogeneus
and complex ecosystems, both in the vertical
and horizontal dimensions.
From the clear-water reservoirs characterised
by a Deep Chlorophyll Maximum to the more
turbid ones, in which positively-buoyant algae
agregate in the upper layers of the water
column, phytoplankton adopts a whole gradient
of heterogeneus (patchy) vertical distribution
patterns. Under non-regulated hydraulic conditions (i.e absence of selective withdrawal events
or water transfers from/to other reservoirs)
phytoplankton vertical distribution is the result
of the interaction between physical agents
(Zmix:Zeu) and the buoyancy capacities of each
algal functional group.
As suggested by Kullemberg (1978) and
demonstrated for El Gergal reservoir study
cases, algae vertical patchiness coupled with the
A)
Figure 6. Velocity field (horizontal component) corresponding to the epilimnion of El Gergal reservoir as measured using
ADCP. Positive values mean water moving from North to the
South. Negative values mean water moving from South to the
North. Campo de velocidades (componente horizontal)
correspondiente al epilimnion del embalse de El Gergal determinado mediante ADCP. Valores positivos indican desplazamientos del agua de Norte a Sur. Valores negativos indican
desplazamientos del agua de Sur a Norte.
B)
Figure 7. The horizontal distribution of Cyanobacteria (A)
and Diatoms (B) in El Gergal reservoir. Algal biomass values
expressed in µg l-1 Distribución vertical de Cianobacterias
(A) y Diatomeas (B) en el embalse de El Gergal. Valores de
biomasa algal expresados en µg l-1.
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Figure 8. Total Chl-a vertical-horizontal coupling in El Gergal
reservoir (from Moreno-Ostos, 2004). Acoplamiento entre la
distribución vertical y horizontal de la Chl-a total en el
embalse de El Gergal (tomado de Moreno-Ostos, 2004).
wind-induced water movement at different
depths resulted in the generation of heterogeneus horizontal distribution of phytoplankton.
In addition, the different algae functional
groups constitute differential and dynamical
patches, each one having their own size, location and transport routes (Moreno-Ostos, 2004).
Previous studies on the spatial distribution
of plankton in lakes and reservoirs (George
& Edwards, 1976; Moreno-Ostos, 2004) have
revealed the existence of an intense verticalhorizontal coupling in microalgae patchiness.
Figure 8 represents the correlation between vertical and horizontal phytoplankton patchiness in
El Gergal reservoir during a two years field
study (Moreno-Ostos, 2004). In this figure,
algal patchiness was expressed as the coefficient
between mean crowding and total Chl-a average
values, following George & Edwards (1976).
Based on similar observations, Reynolds (1984)
points out that phytoplankton horizontal patchiness can be estimated from the vertical distribution of algae in a downwind sampling station
and the appropriate morphometrical and meterological data. According to that, reservoir managers can approximate the degree of horizontal
patchiness from a reduced number of vertical
profiles, a detailed bathimetry of the reservoir
and meteorological records, thus making possible to integrate horizontal heterogeneity in their
monitoring procedures with minimal increases
in time and money costs.
213
In the same context, our results demonstrate
that surface water temperature horizontal distribution can be considered as an adequate
surrogate of reservoir surficial hydrodynamics.
Thus, extensive water temperature surveys
along the horizontal plane of the system could
be used by reservoir managers as an appropriate and non-expensive methodology in
order to estimate advective processes involved
in the formation and transport of phytoplankton horizontal patchiness.
The precise knowledge of the phytoplankton
vertical and horizontal distribution patterns
under different biological, physical and meteorological conditions represent a valuable tool in
hands of reservoir managers. It is essential for
the adequate design of selective withdrawal
strategies and water transfer operations and for
the optimal selection of sampling stations and
depths in water quality monitoring surveys
(Moreno-Ostos et al., 2004b).
Additionally, the integration of the spatial
heterogeneity of phytoplankton in dynamic reservoir modelling would improve the quality and
precission of simulations and predictions, thus
advancing in the application of this kind of tools
for a scientifically-based reservoir management.
To adequately achieve these objectives it is
essential to incorporate modern high-resolution
technologies both for phytoplankton and hydrodynamical monitoring. The combination of in
vivo and in situ spectrofluorimetry and ADCP
represents a powerful and efficient device for the
analysis of physical-biological coupling.
Fee (1976) alerts on the necessity of taking
into consideration the spatial distribution of
Chl-a in plankton dynamic studies and points
out that estimations of total Chl-a derived from
just one vertical profile in the deepest area of
the lake often presents errors of 25 %-40 %. In
this context we encourage reservoir managers to
incorporate in their routinary water quality
monitoring procedures the development of
extensive samplings on the vertical and horizontal planes of the system, especially during the
thermal stratification period, when phytoplankton vertical and horizontal patchiness becomes
more intense (Moreno-Ostos, 2004).
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ACKNOWLEDGEMENTS
This study has been partly funded by the projects UE-LIFE 98 ENV/UK/000607; CICYT
HI99/0836 and CICYT REN 2003-03038.
Special thanks to the University of Granada
Research Council for funding E Moreno-Ostos
Postdoctoral position in the University of
Barcelona. Authors would also like to thank the
staff of EMASESA for their assistance in the
field sampling campaigns in El Gergal reservoir and EMASAGRA for their help in Quentar
reservoir. Dr. C. Escot and Dr. F.J. Rueda contributed with valuable comments and suggestions along all the study.
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WIENS, J. A. 1989. Spatial scaling in ecology.
Funct. Ecol., 3: 385-397.
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Limnetica, 25(1-2): 217-252 (2006)
High mountain lakes of the Central Range (Iberian Peninsula):
Regional limnology & environmental changes
Manuel Toro1,2, Ignacio Granados1,3, Santiago Robles1,4, Carlos Montes1
1
Departamento Interuniversitario de Ecología. Universidad Autónoma de Madrid. Campus de Cantoblanco.
28049 Madrid. Spain.
2 Centro de Estudios Hidrográficos (CEDEX). Paseo Bajo Virgen del Puerto, 3. 28005 Madrid. Spain.
3 Parque Natural de Peñalara. Centro de Gestión Puente del Perdón. Cta. M-604, Km. 27,6. 28740
Rascafría. Spain.
4 CIMERA Estudios Aplicados, S.L. Parque Científico de Madrid. Pol. Indust. Zona Oeste. 28760 Tres
Cantos. Spain.
ABSTRACT
High mountain lake ecosystems in the Iberian Peninsula, being more than 1700 water bodies, are represented mainly by small
or medium size lakes (75 % with a surface less than 0.5 Ha.). The knowledge of their regional limnology in Spain is yet uneven
and insufficient, as well as their ecological status and sensitivity to human activity impacts. This work describes the major limnological characteristics and functioning of high mountain lakes in the Spanish Central Range, and their relationships with
regional environmental variables and existing human pressures. Some hydrological processes (turnover rate), thermal properties (ice-cover dynamics) or hydrochemical parameters (conductivity) are discussed in more detail in those lakes with long
term monitoring data. The composition of planktonic and benthic communities responds to both human pressures and biogeographical or environmental aspects. The effects produced by tourism, cattle, lake damming, wastewater inflow, watershed erosion, introduction of the brook trout, or environmental warming, are studied in some lakes. Implemented management and restoration measures to reduce environmental impacts are described and evaluated.
Key words: High mountain lakes, environmental change, human impacts, paleolimnology, lake restoration, regional limnology, Spanish Central Range.
RESUMEN
Los ecosistemas acuáticos leníticos de alta montaña de la Península Ibérica, con un número superior a 1700 masas de agua,
se hallan representados en su mayor parte por lagos de pequeño o mediano tamaño (el 75 % presenta una superficie inferior
a 0.5 Ha.). El conocimiento limnológico regional de estos sistemas es aún muy limitado e irregular en España, así como su
estado de conservación y su sensibilidad ante los posibles impactos producidos por las actividades humanas. Este trabajo describe los principales aspectos limnológicos y el funcionamiento de las lagunas de alta montaña del Sistema Central en territorio español, en relación con las variables ambientales regionales y con las principales presiones humanas a las que se hallan
sometidas. Algunos procesos hidrológicos (tasa de renovación), térmicos (cubierta de hielo) o hidroquímicos (mineralización)
son abordados con mayor detalle en aquellas lagunas con un seguimiento limnológico más continuo. La composición de las
principales comunidades planctónicas y bentónicas responde tanto a motivos biogeográficos o ambientales, como a la presión
humana. Se analiza el impacto producido en algunas lagunas por el turismo y la ganadería, el represamiento, los vertidos de
un refugio, la erosión en la cuenca, la introducción de un salmónido o por un posible calentamiento climático. Las medidas
adoptadas para mitigar algunos de estos impactos o restaurar sus condiciones originales son descritas y evaluadas.
Palabras clave: Lagunas de alta montaña, cambios ambientales, impactos humanos, paleolimnología, restauración de lagunas, limnología regional, Sistema Central Español.
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INTRODUCTION
The study of natural processes that guide the
functioning of ecosystems is a difficult task for
researchers in the 21st century, due to the direct
or indirect interference of human activities. The
location of semi-pristine ecosystems in which to
study the Earth’s processes at different scales
and its response to global environmental changes is a goal of recent interdisciplinary studies
in the field of ecology (N.R.C., 2005). The
aquatic ecosystems are a key to global change
studies, due to the sensitivity and vulnerability
of their biological communities and their ecological processes (Poff et al., 2002). In order to
detect trends in the climate or in ecological processes, both natural and resulting from human
activity, researchers have turned to high altitude
aquatic ecosystems, as sites not altered directly
(Marchetto & Rogora, 2004; Livingstone,
2005). Their value as sensors of environmental
change come from a series of characteristics
that set them apart from other aquatic systems:
low mineralization and buffering capacity, low
nutrient concentration (N compounds), accumulation of trace metals and organic compounds in
the food chain, and the predicted higher reaction
to global warming in alpine areas (MoralesBaquero et al., 2001). These aquatic ecosystems, present in most continents, have been studied as sensors of distinct environmental
processes: acidification (Battarbee & Renberg,
1990; Camarero et al., 1995a, b; Marchetto, et
al., 1995; Wögrath & Psenner, 1995; Tait &
Thaler, 2000), climate change (Hauer et al.,
1997; Lami et al., 1998; Battarbee et al., 2002;
Abbott et al., 2003), dispersion of atmospheric
contaminants (Fernández et al., 2002; Carrera et
al., 2002; Curtis et al., 2005), land use changes
(Hausmann et al., 2001) and erosion processes
(Toro & Granados, 2002).
The management and control of the factors
that induce environmental changes, as well as
informing society about the magnitude of such
changes and the effectiveness of the adopted
measures, require extensive, continuous and
precise monitoring, which helps detect, monitor and link these variations to the parameters
that quantify the environmental conditions
(Parr et al., 2003). The researcher should make
use of three basic tools to reach this goal:
1) reliable historical information, 2) long term
monitoring networks, and 3) paleoecological
studies. The first two tools are scarce in most
of our alpine systems, which, due to their
remote character and harsh environmental conditions, have been seldom visited in the past,
with few rigorous research done until the
1980s (Pascual et al., 2000). In order to make
up for the scarcity of long-term environmental
studies, scientists have developed complex
mathematical models, both predictive as
reconstructive, in different spatial and temporal scales, for distinct geographical zones
(Castro et al., 1995; Arpe & Roeckner, 1999).
Nevertheless, it is essential to be able to use
historical and current information, organized
in databases, as calibration for the models
(Richardson & Berish, 2003) or in the paleoecological reconstructions (Veski et al., 2005).
In the alpine regions of the Iberian Peninsula,
there are few long-term studies with sufficient
time scale to detect natural ecosystems responses
to regional environmental changes. The meteorological station located in Puerto de Navacerrada
(1890 m a.s.l., Guadarrama Mountains, Central
Range) (Martinez-Molina et al., 1984) has one of
the longest databases, with more than 50 years of
measurements. The paleoecological studies performed in the Iberian Peninsula are contributing
valuable alternative data to the task of reconstructing our past, and though not homogenously
distributed, have been established in numerous
locations. A large part of these studies have been
developed in alpine regions, contributing to the
knowledge of past climate change and its effects
in the ecosystems in the Iberian Peninsula (Toro et
al., 1993; Ruíz-Zapata et al., 1997; Monserrat,
1992; Peñalba et al., 1997, Catalan et al., 2002),
land use changes on the regional scale (Luque &
Julia, 2002), long distance contaminant dispersion
(Camarero et al., 1995b), or human activity
effects on a local scale (Toro & Granados, 2002).
The present work compiles the main studies
performed during the past 15 years by a research team in the Ecology Department of the
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High mountain lakes of Central Spain
Universidad Autónoma de Madrid, on the
Central Range (Guadarrama and Gredos
Mountains) high mountain lakes and wetland
ecosystems, in the Iberian Peninsula. The focus
of the research done has been the local and
regional characterization of these systems from
a limnologic standpoint, as well as their utilization as environmental change sensors in the
local and regional scales for natural and human
induced processes. The results obtained have
served as support for different measures adopted in the management and restoration of the
natural systems studied.
The historic scientific heritage.
Ramon Margalef’s contribution
Even though these alpine ecosystems are considered today pristine sites, free from impacts or
considerable human activity, it was in the period
of the first explorers (19th century and the first
219
half of the 20th century) that these ecosystems
were in their most natural condition. The origins
of scientific research in high mountain lakes
and wetlands in Spain can be traced to this
period, motivated by an excursionist illusion and
seeking similarities with the studies done in the
Alps lakes (Casado, 2000).
Some of the first scientific or naturalistic
explorations of the Central Range mountain
lakes cited in the literature date back to the 19th
century (Aznar, 1839; Pictet, 1865) and the first
half of the 20th century (Azpeitia, 1911;
Obermaier & Carandell, 1917; Arevalo, 1921,
1931, Pardo, 1932, 1948). The first aquatic species inventories are found in the works of
González-Guerrero (1927, 1929a, b, 1965) and
Caballero (1944, 1950) on phytoplankton.
The Peñalara lakes were also visited by the
best known Spanish limnologist, Ramón
Margalef. On April 29, 1949, with his friend
Emilio Fernández Galiano, Margalef collected
Figure 1. Reproduction of Ramon Margalef ’s draws of aquatic organisms found in several lakes, streams and mires in Peñalara
Massif in 1949. (Source: R.Margalef. 1949. Datos para la hidrobiologia de la Sierra de Guadarrama. Publ. Inst. Biol. Apl., Tomo
VI: 5-21). Reproducción de los dibujos realizados por Ramón Margalef de los organismos acuáticos encontrados en varias lagunas, arroyos y turberas del Macizo de Peñalara en 1949. (Fuente: R.Margalef. 1949. Datos para la hidrobiologia de la Sierra de
Guadarrama. Publ. Inst. Biol. Apl., Tomo VI: 5-21).
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Toro et al.
Figure 2. Location of high mountain lakes in Spanish Central Range (Guadarrama and Gredos Mountains). Localización de las
lagunas de alta montaña del Sistema Central, España (Sierras de Guadarrama y Gredos).
water samples from several lakes, streams and
mires, in order to supply new biogeographic
data on Iberian fresh water organisms. The
results of this excursion, with an extensive list
of organisms identified, were published in that
same year in a paper entitled “Datos para la
hidrobiologia de la Sierra de Guadarrama”
[Data for the hydrobiology of Guadarrama
Range] (Margalef, 1949). The historical data
provided by Margalef in this work have served
as valuable reference for recent studies, such as
the effects of the introduction of the brook trout
in the lake’s community, with the description of
the species present previous to the introduction
of this salmonid. Margalef mentioned the absence of phanerogams and mosses in the lake, as
well as the absence of fish. His biological descriptions encompass phytoplankton, zooplankton, phytobenthos (herpon and pecton) and
macroinvertebrate species, including detailed
morphological and taxonomic notes on some
species, as well as some drawings (Fig. 1).
STUDY AREA
The Central Range, with a ENE-WSW orientation, divides the two Iberian Peninsula plateaus
corresponding to the basins of the Duero River
to the north, and Tajo River, to the south (Fig. 2).
This system is formed by the following mountains, from the northeast: Ayllón (1691 m),
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Somosierra (2250 m), Guadarrama (2430 m),
Gredos (2592 m), Béjar (2401 m), Peña de
Francia (1723 m), Gata (1519 m) and, in
Portugal, the Estrela and Lousa (1991 m). It is
constituted primarily by Paleozoic Precambrian
igneous rocks (granite and gneiss) and metamorphic rock slates, with very low solubility. Lakes
are only found in the Guadarrama, Gredos and
Estrela Mountains, where the landscape forming
role of the glaciers was more intense. The results
obtained in projects in the Guadarrama and
Gredos ranges are presented in this work.
Information on the Estrela mountain range lakes
can be found in the work of Boavida (2000).
Most of the glaciers of the Gredos Mountains
were located on the northern slope, where all of
the alpine lakes are situated. Their watersheds
flow into the Duero River basin, in the Northern
Iberian Plateau (Fig. 2). The glacial impacts
were much less in the neighbouring Guadarrama
Mountains. The high snow accumulation by the
wind in the leeward slopes facing southeast originated the development of these small glaciers
of ice masses on south-facing slopes (SanzHerraiz, 1977). Sanz-Herraiz (1988, 1999) and
Pedraza et al. (2004) describe and interpret glacier’s geomorphology in Peñalara Massif.
Besides a considerable number of ponds, there
are three main lakes which represent the circus
and moraine types, whose watersheds flow into
the Tajo river basin in the South Iberian Plateau
(Fig. 2). The precipitation reaches maximum
values above 2000 mm in some places in the
Central Range. The mean annual precipitation in
the Peñalara Massif (Guadarrama Mountains)
is 1350 mm, ranging between 780 mm and
2380 mm, and maximum and minimum mean
values of temperature are 20 ºC and -5 ºC respectively (data from the period of 1946-2004:
Meteorological Station of Puerto de Navacerrada, 1890 m.a.s.l. National Institute of
Meteorology). There are no meteorological stations in the high mountain zone of the Gredos
Mountains that could collect local data, but bulk
precipitation is estimated to be slightly higher
and temperature values slightly lower that in
the Guadarrama Mountains. The timberline is
situated at an altitude of 1900-2100 m.a.s.l.
221
(Martínez, 1999), with the Pinus sylvestris pine
as the forest community that reaches the highest
altitude in all of the Central Range (Luceño y
Vargas, 1991). The high mountain zone, the
focus of this study, has two vegetation zones: the
oromediterranean, typical of “piorno serrano”
(Cytisus oromediterraneus) and the cryoromediterranean, represented by the psychroxerophilous grassland. Most of the lake basin has little
vegetation or developed soils.
The two groups of lakes studied in the
Central Range are under the protection of two
regional parks: the Peñalara Natural Park,
reclassified in 1990 from the former National
Interest Natural Site of 1930, and the Gredos
Mountains Regional Park, created in 1996. The
Peñalara Massif wetlands are in the process of
approval for inclusion in the RAMSAR list of
international importance.
METHODS
The limnologic samples have been collected
from 1991 to the present with variable intervals,
according to the lakes sampled. The methodology used was based on the following references:
morphometric parameters according to
Häkanson (1981), water chemistry following
standard methods and protocols (APHA, 1992;
Catalan & Camarero, 1988; Catalan Lafuente,
1990; Krol et al., 1997; Wathne & Hansen,
1997), plankton analyses (Sournia, 1978; De
Hoyos & Negro, 2001), pigments (Jeffrey &
Humphrey, 1975), macroinvertebrates (Toro &
Granados, 1998), sediment sampling, dating and
analyses (Battarbee, 1986; Appleby et al., 1986;
Walker, 1987; Glew, 1988, 2001), and sediment
traps (Wathne & Hansen, 1997). Besides that, the
methodology used in the various specific environmental change studies is described in detail in
the following works: effects of the brook trout
introduction (Granados & Toro, 2000b), study of
erosion in a lake basin (Toro & Granados, 2002),
effects of wastewater refuse (Robles et al., 2000),
and the study of subfossil diatoms and chironomids as sensors of recent changes (Toro et al.,
1993; Granados & Toro, 2000a).
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Toro et al.
REGIONAL LAKE LIMNOLOGY IN THE
SPANISH CENTRAL RANGE
Morphometrics and genesis
The Central Range morphology is a result of
several morphoclimatic processes, with a predominance of fluvial and glacial processes.
The end of glaciers took place around 10.000
years ago (De Pedraza & López, 1980). The
lakes were formed by the melting of ice and
snow in the depressions, and by water retention
by the lateral or frontal moraines, that acted as
dams. Toro & Granados (2001), based on a previous classification done by De Pedraza &
López (1980), developed a typology of the lake
basins based on their genesis (glacial geomorphology) and morphometrics, that conditions
the ecological functioning of each of the lake
Figure 3. Distribution and total surface area (km2) of high
mountain lakes (lake surface area > 0.5 ha) in the mountain ranges of the Iberian Peninsula (adapted from Pascual et al., 2000).
Distribución y superficie total (km2) de los lagos de alta montaña (superficie mayor de 0.5 ha) en los sistemas montañosos de
la Península Ibérica (adaptado de Pascual et al., 2000).
systems (Table 1): circus (4 lakes), glacial
valley bottom (10), “hoyas” (4), fluvial (2) and
Table 1. Morphometric characteristics of high mountain lakes in Gredos and Guadarrama mountains (Central Range, Iberian Peninsula).
Características morfométricas de las lagunas de alta montaña de las Sierras de Gredos y Guadarrama (Sistema Central, Península Ibérica).
Lake
Coord.
UTM
Lake
Volume
(m3)
Altitude
(m)
Typology
(genesis)*
Watershed
Area
(ha)
Surface
Area
(m2)
Guadarrama
Peñalara
30TVL190215
Claveles
30TVL201231
Pájaros
30TVL200239
2019
2119
2170
Circus
Moraine
Hoya
44.2
12.6
4.8
5779
6263
4943
11563
11560
1021
Gredos
Duque
Cura
Barco
Cervunal
Majalaescoba
Lagunilla
Grande Gredos
Nava
Caballeros
Trampal 3
Negra
Cuadrada
Gargantón
Bajera
Brincalobitos
Trampal 2
Trampal 1
Mediana
Galana
Cimera
Trochagosta
Gutre
1595
1750
1785
1815
1830
1915
1935
1945
2025
2025
2070
2085
2085
2100
2100
2115
2125
2130
2135
2140
2210
2300
Bottom valley
Moraine
Bottom valley
Moraine
Fluvial
Hoya
Bottom valley
Circus
Circus
Bottom valley
Hoya
Hoya
Fluvial
Bottom valley
Bottom valley
Bottom valley
Bottom valley
Bottom valley
Bottom valley
Circus
-Hoya
760.0
4.8
374.4
6.3
403.1
62.5
325.0
62.4
58.0
188.0
2.3
42.6
99.0
120.0
106.3
70.0
54.2
97.9
92.3
75.6
19.5
3.1
203295
2000
74781
4200
2615
5437
63076
92268
14027
62042
1450
7773
441
9599
981
15352
6063
3240
15251
44900
-960
1644000
-357702
---145837
425500
40840
599830
-10525
-14974
1391
37715
8780
4693
54353
216890
---
30TTK716650
30TUK159614
30TTK785567
30TUK054630
30TUK042617
30TUK035621
30TUK064585
30TTK810565
30TTK798553
30TTK697658
30TTK697643
30TTK787557
30TUK054594
30TUK044605
30TUK042600
30TTK692657
30TTK689658
30TUK043602
30TUK045604
30TUK040596
30TTK691652
30TUK045594
* Hoya: local term for small lakes or ponds located at glacial valleys bottom or depressions (Toro & Granados, 2001).
Lake
Perimeter
(m)
Max.
Depth
(m)
Residence
Time
(days)
364
440
447
4.7
2.6
0.5
9.0
-5.2
2440
150
1250
235
231
362
2423
1100
450
1140
180
380
250
505
150
650
400
295
680
1275
-175
21.5
0.4
14.8
1.0
1.2
1.7
6.5
11.0
5.2
20.8
2.0
2.5
1.5
4.2
3.3
4.25
2.9
2.4
7.6
9.4
-2.0
46.4
-20.5
---9.6
146.4
15.1
68.5
-5.3
-2.7
0.3
11.6
3.5
1.0
12.6
61.6
---
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Figure 4. Bathymetric map of Grande de Gredos Lake with
location of sampling sites and mountain refuge. Mapa batimétrico de la laguna Grande de Gredos con la ubicación de los
puntos de muestreo y del refugio de montaña.
moraine (3). The lakes of the Central Range
have smaller dimensions when compared to
other mountain systems in the Iberian
Peninsula, such as the Pyrenees or the Cantabric Range, having much fewer water bodies
(Fig. 3). This reduced size determines a greater
influence of the external environmental conditions on the water bodies. The lakes with greater surface areas are the ones that were dammed: Duque (20.3 ha.), Nava (9.22 ha.) and
Barco (7.47 ha.), with a considerable increase
from their original surface areas. The larger
natural lakes are the Grande de Gredos
223
(6.3 ha.) and the Cimera (4.49 ha.), both in
the Gredos Mountains. In the Guadarrama
Mountains, the largest permanent lake is the
Grande de Peñalara (0.57 ha.), although
Claveles, a temporary lake, is slightly larger
(0.62). Table 1 contains the main morphometric parameters of Central Range lakes. Their
altitudes range between 1600 m and 2300 m,
with a mean altitude of 1994 m, and 60 % of
the lakes situated above 2000 m.
The Grande de Gredos Lake has the most
peculiar morphometrics of all the Central
Range lakes, in which the erosion of the ice on
a long bottom glacier circus generated two
depressions, originating two lake basins connected by a narrow strait (Fig. 4). The rocky
nature of the basin generated a very sinuous
perimeter, uncommon in these alpine lakes,
giving it the longest perimeter of all the Central
Range lakes (> 2400 m) and the largest surface
area of the lakes that are not dammed (6.3 ha).
The rest of the lakes have a single basin. Most
of the lake catchment have steep slopes, with a
predominance of partially fractured bedrock
and moraines and talus zones with smaller
sized materials, as well as small areas occupied
by psychroxerophilous alpine fields, or
small depressions occupied by ponds, wetlands
or peatlands. There are few catchments with
considerable alpine shrub vegetation on the
less steep slopes (Duque, Barco, Caballeros,
Cervunal, Majalaescoba and Peñalara). As a
representative example, the Peñalara Lake
watershed has 63 % of its surface occupied by
bedrock and talus, 29 % by mountain scrubs
(Cytisus oromeditarraneus, Juniperus communis ssp. alpina), some 6 % by psychroxerophilous grassland (mainly Festuca curvifolia) and
2 % by Nardus stricta wet meadows.
The bottom of most of the lakes is mainly
silty, with a relatively low organic matter content (13 % Peñalara, 16 % Cimera, 19.5 %
Grande de Gredos) with a higher percentage
of sandy material at the shores and at stream
inlets, with the presence of blocks or scattered stones, usually at the foot of the hills,
talus or slide rocks. One exception is the
Claveles Lake (Guadarrama Mountains), loca-
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Toro et al.
Figure 5. Lake level, volume and turnover rate responses to the water inflow from snow melting and liquid precipitation in
Peñalara Lake for 1999 and 2000. Respuesta del nivel, volumen y tasa de renovación de la laguna de Peñalara a la entrada de agua
procedente de la fusión de la nieve y de la precipitación líquida durante 1999 y 2000.
ted on a moraine arch, which gives it a temporary characteristic, due to the substrate permeability. This lake’s bottom is lined with
medium sized rocks (0.1-1 m in diameter)
with almost no silt or fine material, probably
due to the small catchment slope of rocky
nature and the absence of water in the summer
months, when the biological productivity is
greatest in these lakes (Toro et al., 2000).
Hydrology
All the Central Range lakes, except for the
Cervunal and Negra lakes, have a superficial
water outlet and one or two main surface water
inlets and several intermittent streams, with
greater flow during the thaw period or heavy
storms. With the exception of the large lakes
(Barco, Duque, Nava, Trampal 3 and Cimera),
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the mean annual turnover rate is less than 15
days (Table 1), reaching minimums of less than
1 day during the ice cover melting (the precipitation as snow can exceed 60 % of the total
annual precipitation; Puerto de Navacerrada
Meteorological Station data, INM). The interannual and seasonal variation in turnover rate
is apparent in the data collected in the Peñalara
Lake basin during a period of several years.
Figure 5 reflects the lake level, volume and
turnover rate responses to the phenomena of
snow melt and liquid precipitation during the
studied period, picked after the thawing and the
autumn rains, and with minimum values during
summer. Isolated precipitation events during
the summer can cause an increase of between 4
and 15 % of the lake water volume in only one
day (a precipitation of 32 mm caused a rise in
the water level of 15.5 cm in July 1998), with
theoretical lake turnover rates of up to 183 %
for October 19, 2001 (almost twice its volume).
It is relatively common to have periods of one
or more days with a turnover rate of over 50 %
of the water retained in the lake, in response to
intense precipitation.
The water discharge coming from the snowpack responds rapidly to the increased spring air
temperatures, generating a hydrogram with
strong daily fluctuations during the days of maximum melting, due to the day/night temperature
difference. Figure 6a shows the Peñalara Lake
outlet stream hydrogram during a typical thawing
month, with a diminishing fluctuation range with
temperature or by the snowpack disappearance.
On the other hand, the typical hydrogram during
the summer period is less fluctuating, though isolated sharp water level increments can be observed, resulting from heavy storms (Fig. 6b), returning to the normal flow 2 or 3 days later. It has
been assumed, for the hydrological balance and
turnover rate estimations, that the precipitation
falling on a watershed, as well as the snowpack
melting water, run almost entirely on the surface
of the soil. Nevertheless, recent studies with markers have shown the important role of the “soil
and talus reservoirs” in high mountain watersheds, where up to 50 % of the snow melt water is
infiltrated in the terrain and later discharged in
225
Figure 6. Peñalara Lake outlet stream hydrogram during: a) a
typical thaw month; b) a summer month with heavy storms.
Hidrograma del arroyo de salida de la laguna de Peñalara
durante: a) un mes típico de deshielo; b) un mes estival con
fuertes tormentas.
the watershed streams. This promotes the chemical interaction of the water and the substrate
during the subsoil storage, noticeably affecting
the water chemical composition (Williams et al.,
2005). The storage capacity of the studied Central
Range watersheds is not known, but it probably
varies according to each watershed’s size and
lithological characteristics. Nevertheless, base
flow for Peñalara Lake watershed in summer is
less than 5 L s-1, and the lake outlet uses to become dry after a 30-45 days period without rainfall,
therefore underground water storage period or
storage capacity would not be so large.
During periods of intense turnover, the smaller lakes’ limnological characteristics are closer
to a lotic ecosystem (e.g. lakes Majalaescoba,
Brincalobitos, Gargantón), with predominance
of watershed material transport by erosion and
of aquatic organisms drift, with a total instability of the water column (turbulent heat and flux
distribution) and a grater ion and nutrient dilution. The change magnitude of the water turn-
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Figure 7. Isopleth diagram of temperature at Peñalara Lake
during the period 1997-1999. Diagrama de isopletas de temperatura en la laguna de Peñalara durante el periodo 1997-1999.
over rates depends mainly on two meteorological variables: temperature and precipitation, just
as shown in the figure 6 hydrograms. Therefore,
changes in precipitation and temperature regime
in these high mountain watersheds (e.g. as a
consequence of a possible climate change)
would noticeably affect the water flow regime,
and ultimately, the water availability as a resource throughout the year (Williams et al., 2005).
Water temperature
The physical stability of the water column, as well
as it’s thermal inertia as related to the air temperature in the lakes studied are smaller than in other
deeper alpine lakes, due to their smaller size.
Mixed water column periods predominate over
stratified ones (Fig. 7). The water thermal response to air temperature changes is fast: in summer,
with maximum air temperatures above 25 ºC, the
superficial water layer of some lakes reaches temperatures of 18-20 ºC, although there is not a stable and lasting actual thermocline in most lakes.
In water depths of less than 10-15 m, stable stratification only occurs under ice cover (Fig. 7).
Figure 8 illustrates the summer water temperature
profiles in the Central Range’s deepest lakes. It
can be observed that the three dammed lakes
(Duque, Trampal 3 and Barco) have a greater surface/bottom thermal gradient, since there’s no
natural water column turnover during the summer.
Duque, the deepest lake, stands out, presenting a
clear thermal stratification with an evident thermocline. During the autumn, temperatures progressively drop, and the water column is mixed by
the rain water input, up to the formation of an ice
cover. Lakes freeze over in November or
December, coinciding with a period of at least 3
or 4 days with maximum air temperatures below
0 ºC, beginning the winter inverse stratification
(Fig. 7). During the winter period, if the ice cover
thickness is not thick enough and minimum air
temperature increase over 0 ºC for a few days, it
can melt completely until air temperature decreases again (Fig. 7). The ice cover duration in the
lakes varies spatially and temporally. Besides precipitation, temperature, winds and radiation factors during the winter months, the watershed’s
relief can prolong the duration of the ice cover, by
obstructing direct radiation incidence on the lake
surface. The Cimera Lake (Gredos Mountains) is
a clear example of this. There is a distinctive
escarpment in the south side, which exerts a noticeable shading effect on the lake. Figure 9 shows
the percentage of reduction of the potential incident radiation (direct radiation) on the lake by the
effect of the surrounding relief. As opposed to the
north shore, the south shore does not receive
direct solar radiation from October 7th to March
7th (there’s only diffuse solar radiation incidence)
and receives less than 10 % of the potential radiation during five months of the year. On the other
hand, the potential radiation only drops noticeably
for one month a year at the north shore, with a
relief-related reduction of only 40 %. This incident radiation asymmetry causes a delay in the
south shore ice cover melting of 1-2 months compared to the north shore, and the lake water temperature has a marked north-south gradient during
the melting, with temperatures of up to 12 ºC in
the north side while the south shore still has a
considerable ice cover. The possible ecological
implications in the benthic shore habitat are evident: 1) water temperature differences between 8
and 12 ºC; 2) ice cover shading effect on the benthic habitat, reducing primary production; and 3)
this situation can last up to two months. The relief
effect is less important in the Peñalara lakes, due
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to their south-southeast orientation, than in the
Gredos lakes. Due to this, the mean ice cover
duration is much more prolonged in the Gredos
lakes (e.g.: 115 days in the Peñalara Lake,
185 days in the Grande de Gredos Lake, and
220 days in the Cimera Lake). The formation and
melting of the ice cover are also quite different
between the lakes of the two massifs: they form in
November-December in Gredos, completely melt
in June-July (August 1996 in Cimera); the ice
cover period in Peñalara goes from December-
227
January to March-April (Fig. 7). Ice cover thicknesses of 142 cm (March 1996) and 165 cm
(March 1991) have been recorded for Peñalara
Lake, with the lowest maximum being approximately 30 cm (winter of 2001-2002). In the
Cimera Lake (Gredos Mountains), the maximum
ice cover thicknesses registered were >280 cm in
the winter of 1996-1997 and 187 cm in the winter
of 1997-1998. In the former, the ice cover was
composed of up to 10 layers with different thickness, texture and water content.
Figure 8. Profiles of water temperature and percentage of dissolved-oxygen saturation in the Central Range’s deepest lakes in summer months. Perfiles de temperatura y porcentaje de oxígeno disuelto en las lagunas más profundas del Sistema Central en los
meses de verano.
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Dissolved oxygen
Under natural conditions, the water column oxygen concentration in the high mountain lakes is
relatively high during the whole annual cycle, and
is not a limiting factor for the biota’s development.
Dissolved oxygen saturation between 90 and
110 % is common. There is an annual periodical
fluctuation in the lake’s surface layer, with maximum values in the winter (e.g. 10-12 mg L-1 O2)
and minimum values in the summer (e.g. 78 mg L-1 O2). Nevertheless, during the ice cover
period, there can be a progressive depletion of the
oxygen in the water layer closer to the sediment
(up to 0.1 mg L-1), as compared to the upper layers
under the ice cover (up to 13 mg L-1). This process
is directly proportional to the ice cover duration. If
the trend for increasing mean air temperatures is
happened and continued, the processes related to
the ice cover dynamics could be affected, such as
the oxygen depletion in the bottom. The intensity
of this process could have been reduced since the
beginning of the 1980s, when the recent increase
in air temperatures was first detected (Granados
and Toro, 2000a). The melting of the ice cover produces a massive input of water in the lakes, mixing
and renewing the whole volume in few days, reaching a maximum dissolved oxygen concentration
in the deeper layers (10-12 mg L-1).
In some of the dammed lakes (Trampal 3, Barco
and Duque), the general pattern in dissolved oxygen concentration during the summer is different
than the rest of the lakes, diminishing in depth as
the summer passes, reaching anoxic conditions in
the deepest layers (Fig. 8). The oxygen depletion
in the Trampal 3 Lake caused the formation of
methane bubbles in the sediment, as a result of the
organic matter degradation in anaerobic conditions. These bubbles are then release to the water
column from structures in the sediment that
resemble small volcanoes, with several centimetres of height. This inadequate condition for a
high mountain oligotrophic lake can be caused by
the following reasons: 1) the decrease in the natural turnover rate and its annual variation, increasing the quantity of organic matter that reaches
the sediment (Whiteside, 1983); 2) the damming
produces a higher fluctuation in the lake’s water
level eliminating the littoral aquatic vegetation
and favouring the organic matter decomposition
in the shore line with the alternation of dry/flooded periods, increasing the nutrient availability.
The annual fluctuation in the water level can
reach several meters in the dammed lakes, while
in most natural lakes usually is less than 50 cm.
Hydrochemistry
The chemical composition of the high mountain
lakes is determined by several factors: weathering
and biogeochemical processes, atmospheric
inputs, biological processes in the lakes, and
finally, hydrological variables such as the evaporation and turnover rates dynamics. The chemical
composition values of the Central Range lakes
(displayed data from end-summer samplings) are
shown in Table 2. The mean water column conductivity of all the lakes studied is very low, fluctuating between 4 µS cm-1 25 ºC in the lakes of
the Cinco Lagunas circus (Gredos Mountains)
and 22 µS cm-1 25 ºC in the Peñalara Lake (Guadarrama Mountains) in the summer months.
Nevertheless, there are remarkable fluctuations in
the annual and interannual scales, depending
mainly on the turnover rate (Fig. 10). The mean
water pH is slightly acid (ranging between 6.2 and
6.8), and the mean alkalinity is very low (ranging
Figure 9. Percentage of reduction of the potential incident
radiation (direct radiation) on Cimera Lake (Gredos
Mountains) and mean duration of the ice cover over the different areas of lake. Porcentaje de reducción de la radiación
potencial incidente (radiación directa) en la laguna Cimera
(Sierra de Gredos) y duración media de la cubierta de hielo en
las diferentes zonas de la laguna.
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229
between 40 and 60 µeq L-1), contributing to a very
low buffering capacity in lake waters. The pH
fluctuation range observed in some lakes (e.g.
Peñalara) is very broad, with minimum values of
4.76 during the ice cover meltdown to maximum
values of 7.70 during the summer months (Fig.
10). The maximum values for conductivity, pH
and alkalinity during the year occur in the middle
of the summer, under low turnover rates and high
evaporation, with maximum primary productivity,
increasing the concentration of dissolved ions in
the water (Fig. 10). Even though pH over 7.5 is
common in the Gredos Mountains lakes, a peak of
8.70 was registered in the Peñalara Lake, in the
year 2000. With the autumn rains, the conducti-
vity decreases temporarily until the ice cover formation, when the turnover rate falls to almost
zero. A progressive gradient in conductivity, alkalinity and pH is formed then, with higher values
in the deeper layers of the lake.
A significant correlation (r2 = 0.42; p<0.05)
has been found between the ice cover duration
and the increased conductivity under the ice
cover for a 10 year time series in the Peñalara
Lake. In the ice cover period in the years 19951996 and 2004-2005, when the ice cover was
very thick, stable and long-lasting, there was a
large increase in conductivity, reaching summerlike values (>17 µS cm-1 25 ºC). On the other
hand, during the ice cover period in 1997-1998 y
Table 2. End-summer chemical characteristics of the high mountain lakes in Central Range (Iberian Peninsula) (Gredos: 1995/1997;
Guadarrama: 1999). Características químicas de los lagos de alta montaña del Sistema Central (Península Ibérica) a finales del periodo estival (Gredos: 1995/1997; Guadarrama: 1999).
Lake
pH
Cond.
µS/cm
25°C
Ca2+
µeq/l
Mg2+
µeq/l
Na+
µeq/l
K+
µeq/l
Alk
µeq /l
SO42µeq/l
Clµeq/l
NO3µgN/l
NH4+
µgN/l
TN
µgN/l
TP
µg/l
TN/TP
Guadarrama
Mountains
Peñalara
Claveles
Pájaros
6.87
6.85
6.80
16
7.0
23.1
53
23
27
23
16
13
39
52
44
4
17
8
106
44
120
33
15
37
4
7
17
27
18
11
22
45
321
164
403
529
7
26
35
15
15
15
Gredos
Mountains
Duque
Cura
Barco
Cervunal
Majalaescoba
Lagunilla
Grande Gredos
Nava
Caballeros
Trampal 3
Negra
Cuadrada
Gargantón
Bajera
Brincalobitos
Trampal 2
Trampal 1
Mediana
Galana
Cimera
Trochagosta
Gutre
6.50
5.89
6.49
5.52
7.00
6.23
6.72
6.60
6.62
6.24
6.50
6.55
6.18
7.24
6.90
7.03
6.40
7.15
7.47
6.80
6.73
6.52
8.0
18.4
10.2
14.4
7.8
9.3
7.0
6.6
9.9
9.5
13.0
6.6
8.1
5.3
5.5
12.9
17.7
5.2
5.1
5.3
7.7
9.3
49
56
26
33
29
33
35
13
25
51
56
13
23
20
18
69
114
21
20
20
15
24
20
21
23
22
16
9
15
6
18
23
30
7
9
13
13
23
62
11
11
14
4
13
47
80
25
47
24
18
15
20
19
44
24
14
15
13
11
54
62
12
11
15
31
20
3
21
3
15
2
2
3
2
2
5
2
2
2
3
2
3
4
1
2
2
2
3
94
120
46
79
42
66
47
34
45
103
101
36
33
36
29
129
167
30
39
42
63
33
12
14
13
2
14
16
19
10
16
20
25
10
11
10
11
28
34
11
11
3
13
13
10
33
4
18
6
7
6
5
5
10
6
6
5
5
5
11
12
6
5
2
7
8
7
145
52
245
125
12
100
9
9
2
14
26
0
8
62
14
34
14
13
36
35
159
0
84
53
284
29
105
64
118
31
0
125
122
101
75
29
0
0
1
103
0
290
56
92
848
225
1559
256
355
245
--98
303
532
313
227
204
176
514
167
228
288
208
252
6
13
7
39
5
<0.1
17
8
12
3
8
22
9
24
23
7
10
9
10
9
2
14
15
65
32
40
51
-14
--33
38
24
35
9
9
25
51
19
23
32
104
18
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Figure 10. Seasonal variation of alkalinity, pH and conductivity
in Peñalara Lake during a four years period (2001-2004).
Variación estacional de la alcalinidad. pH y conductividad en
la laguna de Peñalara durante un periodo de cuatro años
(2001-2004).
2000-2001, when the ice cover was thinner and
lasted less time, the increase in conductivity was
small (< 9 µS cm-1 25 ºC). The clear pH peaks in
these deep layers is also only observed in the
years with a long-lasting ice cover. When the ice
is formed, there’s ionic exclusion, as observed in
other high mountain or alpine shallow lakes
(Barica, 1977; Baron, 1992; Welch & Bergman,
1985). The magnitude of this process can be
approximately quantified with the example of the
Peñalara Lake: with an ice cover of more than 1m
thick (1995-96 or 2003-04), it is assumed that
more than 40 % of the lake water volume becomes ice. There’s also a release of basic cations
from the sediment (Psenner, 1988) and a sulphate
and nitrate reduction in the anoxic condition that
can be reached in the few millimetres of the
water-sediment interface (Psenner & Catalan,
1994). The final result is a remarkable increase in
the lake deeper layers conductivity, as well as an
increment in the alkalinity –and therefore, of pH–
of the water layer closer to the sediment. Under
the ice, though, these parameters remain low. The
more potent, stable and durable the ice cover, the
more intense these changes in the lake water chemistry are. During the ice cover melting, there is
a water mineralization increment due to the early
fusion of the ion richer layers; nevertheless, the
most important result is the input of a large quantity of water in the lake, which has a dilution
effect in the medium, reaching the annual minimum conductivity, with extremely low values.
The pH and alkalinity also decrease, with the
input of acid ions retained in the snowpack and
ice cover (Fig. 10). In the Cimera Lake, there has
been a conductivity measurement of 1.34 µS cm-1
25 ºC (7/7/98), which is assumed to be the lowest
registered value in these high mountain lakes.
Therefore, the conductivity seems to be a good
indicator of the intensity of the processes present
in a lake as a consequence of the ice cover formation, such as the ion and compound fluxes from
the sediment, the oxygen depletion in the bottom
or the generation of alkalinity.
The wide pH range observed in some lakes
(Peñalara) (Fig. 10) is due to the scarce alkaline
reserves (bicarbonate) which characterizes their
hydrochemistry, in essence, the low buffering
capacity of the water. Thus, their sensitivity to the
atmospheric input of acid anions is very high, and
acid rain is one of the potential risks for this type
of lakes. There have been records of dry or wet
deposition of dust from the Sahara (e.g. the orange
snow episode in the Peñalara Lake in the winter of
1992), which contribute for the neutralization of
the possible acid compound inputs from atmospheric deposition. This dust from the Sahara is an
important source of alkalinity and nutrients for the
alpine lakes in southern Europe (Psenner 1999),
contributing with a high percentage of the total
annual calcium (70 %) and bicarbonate (100 %)
input in the lake watersheds (Rogora et al., 2004).
Nevertheless, in the absence of this dust deposition
from the Sahara, the snow is usually highly acid
(pH between 4.19 and 5.43), with an extremely
variable ionic composition, influencing in the lake
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hydrochemistry, especially during the ice cover
melting period. Even though the bicarbonate concentration is low (< 100 µeq L-1), they represent a
high percentage of the ionic composition in these
lakes, reaching values of close to 90 % of the total
anions in lakes like the Cimera, one of the least
mineralized in all of the Central Range. The predominance of sulphates over chlorides reflects the
distance from the ocean –and thus, from the
influence of marine aerosols. The following relations are found in the Central Range lakes: for
cations [Ca2+] ⭓ [Na+] > [Mg2+] >> [K+], and
for anions [HCO3-] ⭓ [SO4=] > [Cl¯]. These relations are common in high mountain lakes on silica
basins in other continental mountainous systems in
the world (Baron, 1992; MOLAR, 1999). Summarizing the chemical composition described above,
231
the water in most of the Central Range lakes can be
classified as a mix between bicarbonate-calcic and
bicarbonate-sodic, a result of the crystalline lithology of the basins (gneisses and granites).
Nitrate concentration is low in most of studied lakes (<60 µg L-1 N-NO3), with the exception of Cura, Cervunal, Majalaescoba, Grande
de Gredos or Peñalara lakes, where NO3 reached higher values. Ammonium concentration is
usually under 90 µg L -1 N-NH4, and mean
annual values are about 20 µg L-1. Total phosphorus concentration do not exceed 25µg L-1,
except for Cervunal Lake, where a maximum of
39 µg L-1 P was recorded. TN/TP ratio founded
in summer months in lakes ranges from 9 to
104. Kopacek et al. (1996) established a TN/TP
ratio of 7.2 for phytoplankton in high mountain
Figure 11. Annual succession pattern of the main phytoplankton groups in Peñalara Lake (Source: Toro & Montes, 1993). Patrón
de sucesión anual de los principales grupos del fitoplancton en la laguna de Peñalara (Fuente: Toro & Montes, 1993).
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Toro et al.
lakes; therefore, summer productivity should be
P-limited in most of lakes. In Peñalara Lake,
several years of monthly monitoring show a Plimited productivity during most all-year period,
but TN/TP trend to reach lower values is detected in summer periods, changing to a N-limited
phytoplankton productivity. This trend has been
also observed by Morales-Baquero et al. (1999)
in a survey of 31 high mountain lakes in Sierra
Nevada (Southern Spain), where the importance
of atmospheric N and P inputs together with the
size of the lake catchment is highlighted.
Phytoplancton
The phytoplankton communities in the studied
lakes are composed of widely distributed, nonendemic species, most of them typical of acidic
oligotrophic lakes. Figure 11 reflects the annual
succession pattern of the main algae groups in
the Peñalara Lake (Toro & Montes, 1993). In the
summer, with water column stability and high
nutrient content and temperatures, the main
algae groups reach their maximum density and
biovolume: chlorophytes (up to 90 % of the biomass), with an alternation of predominance of
Zygnematales (Arthrodesmus sp.) with Chlorococcales (Chlorella sp., Dydimocystis sp.,
Scenedesmus sp.), cianobacteria (summer bloom
characteristic of a filamentous species, without
heterocysts; Aphanothece sp., Pseudanabaena
sp.) and, at the end of the summer, dinoflagellates and diatoms (Aulacoseira sp., Fragilaria sp.)
favoured by the first autumn rains. The importance of planktonic diatoms is very low (<1 %)
in the lakes studied, in comparison to their
importance to the benthonic communities. The
density of flagellate algae, favoured by their
mobility in the stratification bellow the winter
cover, increases in the beginning of the winter:
dinoflagellates (Amphidinium sp., up to 50 % of
the biomass), chrysophytes (Dynobrion sp.,
Ochromonas sp.) and some flagellate chlorophytes (Pedinomonas sp.). The ice cover melting
causes a low phytoplankton density, even though
small flagellate groups persist. This community
succession annual pattern happens in the group
level, since each period’s dominant species
usually vary each year (Toro y Granados, 1997).
There is a strong correlation between the different phytoplankton communities and the lake’s
degree of mineralization, indicating a low organic contamination, or by nutrients in general,
since this is the main factor that differentiates
the algae communities (De Hoyos & Negro,
2001). The smaller and shallower lakes
have been found to be the richest in number
of species (e.g.: >200 species in the Peñalara
Lake) (Toro & Montes, 1993; De Hoyos &
Negro, 2001), in which the proportion of benthonic or shore habitats exceeds the pelagic
zone, not corresponding to typically planktonic
communities (>30 % of the algae species are
diatoms, and most are benthonic).
Macrophytes
The composition of the macrophyte communities is determined by the degree of temporality
or water level fluctuation in the lakes, and there
are two main communities: permanent lakes
with stable water level or shallow lakes or ponds
with a fluctuating water level (Aldasoro & Toro,
2001). The typical alpine aquatic systems species are in the first group: Subularia acuatica,
Isoetes velatum, Callitriche palustris or Sparganium angustifolium are the most representative. Subularia acuática, a boreoalpine species,
has in the Trampal 1 and 2 lakes their only
known distribution in the Central Range
(Aldasoro & Toro, 2001). The surface area
covered and the richness depend on the morphology of the shore and on the substrate type.
Therefore, the Grande de Gredos Lake, with a
large shore development and shallow zones, has
one of the largest macrophyte richness and density in the Central Range. The most characteristic species of the second group are Ranunculus
peltatus, Antinoria agrostidea, Potamogeton
natans or Juncus bulbosus. Bryophyte species
of the Fontinalis and Sphagnum genera are also
common in the lakes. In the Guadarrama
Mountains, the Grande de Peñalara Lake does
not have any aquatic vegetation, although a couple of Callitriche plants were observed by the
lake shore in 2002-2003 summer periods.
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Nevertheless, there is in the area a total of 27
aquatic plant species (angiosperm phanerogams,
charophytes and bryophytes), with the presence
of Nitella flexilis, a charophyte typical of oligotrophic waters (Granados & Toro, 2000b).
Zooplankton
The zooplankton species richness in the lakes
studied is smaller than in other alpine systems in
the northern Iberian Peninsula (Miracle, 1978;
Vega et al., 1991), being more like other lakes in
the southern Peninsula (Cruz-Pizarro et al.,
1981). There have been found 21 species in the
Gredos Mountains lakes (Robles & Aldasoro,
2001), similar to the 19 species found in the
Guadarrama Mountains lakes (Toro & Granados,
1998). Both mountain systems present a very
similar community composition, with low specialized and cosmopolitan, copepods and cladoceran species, such as Tropocyclops prasinus,
Ceriodaphnia quadrangula, Daphnia longispina
or Chidorus sphaericus. Nevertheless, there are
species with a more boreoalpine character, such
as Alonella nana. The main difference between
these mountain systems is the absence of the
diaptomid Diaptomus castaneti in the Peñalara
lakes, with presence in the Gredos lakes and in
other mountain systems of the northern Iberian
Peninsula (Aldasoro et al., 1984). The most frequent diaptomid in the Pyrenees lakes, Diaptomus cyaneus (Miracle, 1978), is not present in
the Peñalara lakes, but has been cited in some
high mountain ponds in an area very close to the
Peñalara Massif (Baltanas, 1985). Within the rotifer group, species such as Poliarthra remata and
Asplachna priodonta dominate in the summer
months, while Keratella quadrata dominate in
the winter, in the absence of predatory pressure
from Asplachna sp. (Toro & Granados, 1997).
Macroinvertebrates
The richness of species in the benthonic macroinvertebrate communities is high in the Central
Range lakes. More than 90 species have been identified in the Gredos Mountains, with 47 belonging
to the chironomid group, with 33 different genera
233
(Toro & Granados, 2001). A total of 59 species
have been identified in the Guadarrama Mountains
lakes, with 29 chironomids (Granados & Toro,
2000b). This is one of the groups that contribute
the most to the total benthonic invertebrate biomass in these lakes with mean values of 847 individuals/m2 for the Peñalara Lake, and peak concentrations of almost 4000 individuals/m2. The
species Micropsectra spp., Heterotrissocladius
marcidus or Parakieffieriella bathphyla, frequent
in alpine systems (Langton, 1991; Rossaro, 1982;
Soriano, 1995), stand out for their abundance. The
oligochaetes are another group that represent a
large proportion of the benthonic biomass in several of the lakes. The two most characteristic species
are Stylodrilus heringianus, typical of oligotrophic
environments, and Nais alpina, characteristic of
high mountains (Brinkhurst, 1971). The mean
abundance of individuals in the Peñalara Lake is
317 individuals m-2, with maximum abundance of
up to 1800 individuos m-2. Besides the chironomids and oligochaetes, species like Sialis lutaria, a
megalopteran predator, and Pisidium casertanum,
the only represented bivalve mollusc, are common
in the silty substrate of the lake bottom, with densities between 38 and 150 individuals/m2 in the
Peñalara Lake. In the lake shores, where there’s
greater substrate diversity, there are also other gastropod (Ancylus fluviatilis), hirudinea (Helobdella
stagnalis), ephemeroptera (Baetis fuscatus,
Siphlonurus lacustris, Habrophlebia fusca), plecoptera (Protonemura meyeri), tricoptera
(Plectrocnemia conspersa, Athripsodes cinereus),
odonata (Selysiothemis nigra), coleoptera
(Nebrioporus fabressei, Oulimnius tuberculatus)
and heteroptera (Notonecta obliqua, Arctocorisa
carinata, Sigara sp.) (Granados & Toro, 2000b).
The distribution of some species can be a response
to certain impacts suffered by some lakes.
Helobdella stagnalis, a species that tolerates anaerobic conditions or temporary dry periods, indicating an organic matter enrichment (Elliot & Mann,
1979), is present only in lakes that have been altered by damming (Barco, Trampal 3), with abundant aquatic vegetation (Grande de Gredos,
Trampal 1) or with incipient eutrophication processes from a recent past (Peñalara and Grande de
Gredos). The trichoptera are absent in the dammed
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lakes, and from the lakes with a superior trophic
level (Cura, Negra and Cervunal). The formation
of a dry section in the shores of the dammed
lakes, due to the water level oscillation, can hinder the colonization of the shore area by some
macroinvertebrate groups. There are Chironomus
sp. in some well conserved lakes, characteristic of
situations with intense organic matter decomposition in usually eutrophic environments. This is
explained by its adaptation to oxygen depletion in
the bottom during the formation of the ice cover
in the winter. Larger sized species from the odonata, coleoptera or heteroptera families are only
present in the Central Range lakes without fish or
with abundant aquatic vegetation.
2001). Finally, among the mammals, the otter
(Lutra lutra) and the American mink (Mustela
vison) are occasional visitors to the lakes searching mainly for fish, and occasionally for amphibious and other preys. The mink is an introduced
species, naturalized in the Central Range by escaping the farms where it is raised for its fur. The
Iberian desman (Galemys pyrenaica), although a
fluvial habitat species, is occasionally present at
some Gredos lakes (Lizana & Morales, 2001).
Vertebrates
Incipient eutrophication of the Peñalara Lake
caused by tourism and cattle
There are several fish, amphibious, reptile and
mammal species in the Central Range lakes.
Among the fish, the only endemic Iberian species is the brown trout (Salmo trutta), present in
only 6 lakes, although it hasn’t been confirmed
if this presence is natural or has been introduced. Besides that, the brook trout (Salvelinus
fontinalis) has been introduced in the Cinco
Lagunas complex (Cimera, Galana, Mediana,
Brincalobitos and Bajera) and in the Peñalara
Lake, and the Iberian chub (Leuciscus carolitertii) and the Iberian nase (Chondrostoma polylepis) have been introduced in the Duque Lake
(Lizana & Morales, 2001).
In the Gredos and Guadarrama Mountains
lakes and wetlands, there are 10 amphibious species: the common salamander (Salamandra salamandra), the alpine newt (Triturus alpestris)
(introduced in Peñalara in the beginning of the
1980s), the marbled newt (T. marmoratus), the
midwife toad (Alytes obstetricans), the natterjack
toad (Bufo calamita), the common toad (B. bufo),
the Portugal painted frog (Discoglossus galganoi),
the San Antonio frog (Hyla arborea), the Iberian
frog (Rana iberica) and the green frog (R. perezi)
(Lizana & Morales, 2001; Martínez-Solano et al.,
2002) . Among the reptiles, there are only two species related to the aquatic ecosystems of these
mountains: the two snakes of the genus Natrix
(N. natrix and N. maura) (Lizana & Morales,
ENVIRONMENTAL CHANGE AND THE
IMPACT IN THE CENTRAL RANGE
HIGH MOUNTAIN LAKES
In the end of the 1980s, the Peñalara Lake presented symptoms during the summer period that indi-
Figure 12. Chlorophyll a (µg L-1) and TP (µg P-PO4 L-1) evolution in Peñalara Lake during a 15 years period. Recovery of
natural levels of both variables, after restoration measures were
adopted to reduce the input of nutrients (1991-1993), is observed. Evolución de la clorofila a (µg L-1) y TP (µg P-PO4 L-1) en
la laguna de Peñalara durante un periodo de 15 años. Se observa la recuperación de los niveles naturales de ambas variables
despues de que se adoptasen medidas de restauración para
reducir la entrada de nutrientes (1991-1993).
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cated an increase in the phytoplankton and benthic
primary production. Chl-a concentration over
50 µg L-1 was recorded in the August 1990, in
waters of an intense green and Secchi disk transparency under 50 cm. The maximum phosphorous
concentration during that month was 427 µg PPO4 L-1, with a mean annual value of 242 µg PPO4 L-1. Several management practices were
adopted in 1991 in order to reduce the nutrient
input from human activity in the lake watershed:
bathing and camping were prohibited, and cattle
access to the lake was limited. The chlorophyll
concentration decreased in a short period, approaching oligotrophic levels (Fig. 12), with summer
maximums of 12.4 µg L-1 in 1992 and 4.5 µg L-1 in
1993, and phosphorus values of 137 µg P-PO4 L-1
in 1992 and 62 in 1993. From 1995 on, there was a
clear annual pattern of this variable, with a Chl-a
summer peak concentration of between 3-8 µg L-1,
and phosphorus peaks of 10-40 µg P-PO4 L-1.
Chl-a reached 7.6 µg L-1, in the summer of 2002,
coinciding with a previous winter with very low
rain and snow precipitation - which caused the lake
turnover rate to be smaller and the water level to be
the lowest. Chlorophyll concentration is minimum
in the ice cover period, with mean minimum values
under 1 µg L-1 Chl-a, although higher maximums
under the ice cover have been recorded in some
years. In the winter of 2004-2005, for example,
there was a remarkable chlorophyll increment in
235
the deeper layers close to the sediment, reaching
values of 7.7 µg L-1, and 0.4 µg L-1 in the layers
adjacent to ice cover. The sensitivity of the phytoplankton primary production in the high mountain
lakes to environmental variables such as total
annual or winter precipitation, temperature or
nutrient input, is evident in the response obtained
throughout the management practices adopted in
the Peñalara Lake to control nutrient input related
to visitors and cattle. The turnover rate of some
lakes (mean annual values of under 15 days, with
minimum values of under 1 day during the melting
of the ice cover) favour the fast recover of the oligotrophic levels, by producing a decrease in the
nutrient input, reaching natural levels in 2 or 3
years of control. The high sensitivity of these ecosystems with oligotrophic vocation is demonstrated
by the increments in summer maximum concentration of chlorophyll related to the minimum values
of winter precipitation, which generate a lower
melting volume, and thus, a lower turnover rate.
Wastewater inflow from a refuge in the
Grande de Gredos Lake
In 1971 a mountain refuge was built close to the
inlet stream of the Grande de Gredos Lake, near
its south basin (Fig. 4). Until 1995, the residual
waters were depurated in a waste treatment lagoon in a natural pond located between the refuge
Figure 13. Dry weight (%), organic matter (% loss on ignition) and number of Clostridium sp. (CFU/g: colony-forming units per
gram of dry weight of sediment) profiles of sediment cores collected at Grande de Gredos Lake in 1991 and 1997. 210Pb dating
was carried out for 1991 sediment core (adapted from Toro et al., 1993 and Robles et al., 2000). Perfiles de peso seco (%), materia
orgánica (% pérdida por combustión) y número de Clostridium sp. (CFU/g: unidades de colonias formadas por gramo de peso
seco de sedimento) en dos testigos de sedimento obtenidos en la laguna Grande de Gredos en 1991 y 1997. El perfil de 1991 fue
datado con 210Pb (adaptado de Toro et al., 1993 y Robles et al., 2000).
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Figure 14. Formation process of the gas chambers and anoxic water layer under macrophyte roots in the bottom of south basin at
Grande de Gredos Lake as a consequence of a waste water inflow from a refuge (Summer 1995). Proceso de formación de cámaras
de gas y capas de agua anóxica bajo las raices de las macrófitas en el fondo de la cubeta sur de la laguna Grande de Gredos como
consecuencia del vertido de un refugio (verano de 1995).
and the lake. In that year, a more complex depuration system was constructed. It consisted basically of an initial grid, a 1st decant-anaerobic
digestion tank, and a 2nd aerobic digestion tank
with an electric air diffuser. The resultant
effluent or wastewater circulated through some
drainage pipes surrounded by gravel and sand in
a ditch towards the Grande de Gredos Lake.
During the following winter, one depuration
system tank broke, due to the ice and the weight
of the snow mantle over the installation, generating a continuous flow of wastewater into the
lake. At the same time, the faulty operation has
not produced the desirable purification levels. In
this way, a second source of scarcely purified
wastewater with diffuse character, passed
through the drainage pipes, because the depuration system did not work due to low winter temperatures and to the lack of power. Tank repairs
during the next summer originated a wastewater
spill of about 6000 litres of black waters without
depuration treatment over the surrounding area,
close to the lake. This fact, together with diffuse
pollution caused by the high number of visitors
and campers in the lake watershed, originated
serious changes in the lake’s trophic level.
The paleolimnological surveys carried out in
1991 and 1997 reflected a big increase in
the organic matter content accumulated in the
period between both studies, as well as an increase in the sedimentation rate. A general trend
to increase the organic matter content (measured as LOI) since the beginning of the 1960’s,
according to the 210Pb dating, is observed along
the sediment record (Fig. 13). Before 1991, the
sedimentation rate was around 2.2 mm/year,
increasing after that year to reach values of
approximately 5 mm/year. This increase clearly
reflects an extraordinary input of organic matter
(measured by LOI) in the past few years. Since
the lake has two different basins (Fig. 4), comparisons between north and south basins sediment profiles were made to look for differences
in the sedimentation rates and organic matter
content. The input of soluble and particulate
materials (of natural origin from the watershed
drainage or from the refuge and campground
area) is higher in the south basin, which acts as
a sediment and nutrient trap prior to the north
basin. The LOI concentration in the surface
sediment of the south basin is much higher than
in the north basin, although there is an acceptable correlation between both curves’ trends.
One of the main impacts and extreme consequences in the lake ecosystem because the high
organic matter input from the refuge was the for-
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mation of gas chambers, under macrophyte roots
in the south basin. These gases originate as a
consequence of intense organic matter degradation in anaerobic conditions. In another lake of
the Central Range, the Trampal 3, a similar process has been verified as a consequence of the
modification of the hydrological cycle due to
damming lake (Toro & Granados, 2001). The
formation of bacterial mats in the upper layers of
the sediment, together with the dense net of
macrophyte roots, hinders the efflux of the gas to
the atmosphere. Because of this accumulation,
the macrophytes and their roots are lifted up
towards the water surface. In this way, gas chambers and a layer of anoxic water with relatively
high concentrations of nutrients ([P-PO4=] = 43
µg L-1, [NH4+] = 1.025 mg L-1), reach a height of
2 m and a surface area of several square metres.
The formation process of the gas chambers and
anoxic water layer is schematised in figure 14.
Robles et al. (2000) studied the effects of outdoor activities in the Grande de Gredos Lake
watershed by the analysis of sulphite-reducing
clostridia in the lake sediment from the south
basin. Sulphite-reducing clostridia are good
indicators of past human pollution because of
their longevity in natural habitats, and they cannot multiply at temperatures below 20°C, or in
the presence of O2. There was a great increase in
the numbers of clostridia (expressed as colonyforming units per gram (CFU g-1) of dry weight
of sediment) in the sediment record of this lake
since the 1970s, showing the rise of human pressure caused by the practice of outdoor activities.
Clostridia CFU g-1 increased dramatically after
the breakdown of the refuge’s depuration system
in 1995 (Fig. 13). Concentration levels of clostridia in the north basin of this lake were similar to
those found in the surface sediment of the
Cimera Lake, located in a more remote area
without this high tourist pressure: both were two
orders of magnitude lower than in the Grande de
Gredos Lake (Robles et al., 2000).
Additional evidences of a recent nutrient and
organic content enrichment in the lake has been
provided by Toro et al. (1993) by means of the
subfossil diatoms analysis from the sediment
core taken in the north lake basin. It reflected a
237
remarkable trend towards an increase in lake
trophic conditions since the end of 60’s. Some
species, such as Navicula radiosa var. tenella,
Pinnularia microstauron, Fragilaria pinnata or
Aulacoseira sp., showed a response to a possible
nutrient enrichment increasing their densities
in more recent sediment layers, whereas other
species associated to oligotrophic conditions,
such as Achnanthes austriaca or Cymbella perpusilla, decreased their densities towards surface layers (Toro et al., 1993).
The progressive recovery that has been
observed in the Grande de Gredos Lake after
the repair and control of the wastewater inflow
is another proof of the great self-recovery
capacity which seems to be characteristic of
these alpine systems. Nevertheless, the studies
being carried on the follow up of these lakes
biological communities (phytoplankton, zooplankton, macroinvertebrate and macrophytes)
will provide knowledge in the future about the
long term effects of these impacts.
Erosion processes in the Peñalara Lake
watershed
The intense tourism and cattle raising activities
in the proximity of the Peñalara Lake caused an
advanced erosion process at the shores of the
lake basin, in the beginning of the 1990s. The
results obtained in the paleolimnological sediment study (Fig. 15) show clearly some of the
effects of this pressure (Toro & Granados, 2002).
Firstly, the dispersed and irregular radiometric
profile obtained from sediment dating shows a
possible recent alteration in the surface sediment
layers (0-20 cm), probably due to its re-suspension by the numerous bathers who visited the
lake during the summer, as well as by livestock.
A period of increased sediment accumulation is
observed from the 1970’s onwards, associated
with denser sediment at a depth of 8.5 cm in the
core. Furthermore, this period is the inflexion
point of the relationship between the sedimentation rate (measured as g m-2 y-1 or cm y-1) and
annual precipitation. Figure 15b shows that prior
to 1970, the sedimentation rate is inversely
correlated to annual precipitation, but after that
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Figure 15. Relationship between sedimentation rate evolution and annual precipitation in Peñalara Lake during second half of 20th century. Correlation changes between both variables prior and post 1970 is emphasized (Toro & Granados, 2002). Relación entre la evolución de la tasa de sedimentación y la precipitación anual en la laguna de Peñalara durante la segunda mitad del siglo XX. Se señala el
cambio producido en la correlación entre ambas variables antes y después de 1970 (Toro & Granados, 2002).
year, a significant positive correlation is initiated
between precipitation and sediment accumulation rate. Thus, before 1970, the sedimentation
process was driven mainly by the opposed
influence between the sedimentation and flushing rates: as water turnover time is reduced (i.e.
more precipitation), the sedimentation rate is
also reduced (Fig. 16). On the other hand, after
1970, a notable rise in the erosion rate in the
watershed and at the shore of the lake starts as a
consequence of the loss of vegetation cover
(mainly mountain pasture) caused by visitors’
trampling. Because of these large areas of bare
soil, erosion increases with increased precipitation. In addition, at the shoreline and in shallow
parts of the lake, bathing visitors cause a process
of re-suspension of the sediment, which accumulates in the deeper area. Prior to the increase in
organic content, at a depth of 8-10 cm there is a
sharp drop in these values, coinciding with an
increase in the percentage of dry weight and wet
density, and less 137Cs and 210Pb activity (Toro
& Montes, 1993). This may represent a strong
erosion process in the watershed or at the shore
of the lake, as has been previously discussed. In
some areas of critical erosion level, soil losses
up to almost 1 m of thickness were recorded between 1985 and 1995 (Toro and Granados, 1999).
For some zones of the frontal moraine that
encloses the lake, this loss meant almost half of
the height of the same above the water level
of the lake. After the prohibition of access to the
lake, the vegetation cover was recovered in
the least degraded zones, but the erosive process
continued in the critical zones. In 1997, it was
necessary to interfere, artificially seeding the
soil with various species of grasses and protecting it with vegetable fibre matting.
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239
Figure 16. Annual model of the erosion-sedimentation processes in the watershed of a small high mountain lake in natural conditions (prior 1970) and with high soil erosion problems (post 1970): Peñalara lake (Granados et al., 2002). Modelo anual del proceso de erosión-sedimentación en la cuenca de un pequeño lago de alta montaña en condiciones naturales (antes de 1970) y con problemas de alta erosión del suelo (después de 1970): laguna de Peñalara (Granados et al., 2002).
The sedimentation rate in a lake basin may be
used as an indirect measurement of sediment
influx, i.e. the rate of erosion in the watershed
(Dearing, 1986). The natural sedimentation processes in the studied lake watersheds follow a
general annual pattern according to the erosion
process degree (Fig. 16) (Granados et al.,
2002). The sedimentation under the winter ice
cover is minimal, without watershed material
inflow and with a negligible photosynthetic primary productivity. During the melting period,
the sedimentation rate is still low, due to the
high lake water turnover rates, in spite of the
very high available energy for particle transportation form the watershed (erosion) and the
maximum soil erodibility due to its higher
water content and the winter cryofraction phenomenon. During the summer, sedimentation
reaches the annual maximum values, because of
the higher water column stability, more visitors
(watershed erosion) and annual maximum primary productivity (organic matter generation
through photosynthesis), besides the highly erosive summer storms. The persistent autumn
rains cause an increase in the turnover rates, a
decrease in the primary productivity and, therefore, a decrease in the sedimentation rates.
To monitoring the response of erosion-sedimentation processes to the restoration measures
adopted in the Peñalara Lake watershed, two sediment traps were located in the deepest area of the
lake (1.3 and 2.6 m above lake bottom) in 1997.
The annual evolution of the rate of organic/inorganic material caught in the sediment traps from
1997 to 2005 validates this model, since the largest percentage of inorganic material is observed
in the summer (Fig. 17). In this period, the mineral material input by watershed erosion is higher,
increasing the percentage of organic matter
during the rest of the year (Granados et al., 2002).
Besides this clear annual pattern, the sedimentation rates have varied in the last few years, with a
trend of decreased sedimentation, due to the
watershed erosion control measures adopted and
the re-vegetation of the most affected zones (Toro
& Granados, 2002). The sedimentation rates in
the summer of 1997 were between 24.3 and
29.2 g m-2 d-1 (1.9 kg m-2 y-1), reached a maximum of only 10.9 g m-2 d-1 (0.83 kg m-2 y-1) in
1998, and have since reached annual maximum
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and 2003 (between 0.22 and 0.56 kg m-2) would
be comparable to the mean values of the years
preceding the impact of tourism.
Climate change detection through the study
of sediment subfossil chironomid
Figure 17. Interannual evolution of the sedimentation rate in
Peñalara Lake according to the material collected in the sediment traps during the period 1997-2005. Evolución interanual
de la tasa de sedimentación en la laguna de Peñalara según
el material recogido en las trampas de sedimento durante el
periodo 1997-2005.
values of between 2.6 and 8 g m-2 d-1 (0.22 –
0.56 m-2 y-1) (Fig. 17). According to the data
obtained in the Peñalara Lake paleolimnological
study (Toro and Montes, 1993; Granados et al.,
2002), there was an annual maximum accumulation of 1.4 kg m-2 y-1 between 1918 and 1970,
with a mean of 1.06 kg m-2 y-1 for this period.
From 1970 on, with the beginning of tourism related erosion processes, the maximum accumulation was 2.5 kg m-2 y-1, with 1970-1991 mean of
1.59 kg m-2 y-1. Therefore, it is shown that both
the paleolimnological and the sediment trap techniques provide results of the same order of magnitude. However, when comparing the results of
both techniques, it is necessary to consider some
aspects related to how each method represents the
watershed erosion processes. The sediment traps
produce values of percentage of organic matter
sedimentation, since part of this material could
have been degraded in the consolidated lake sediment. On the other hand, the traps underestimate
the quantity of total sediment, since they do not
collect the sediment that reaches the bottom
through lateral “focusing” movement bellow the
height of the trap’s entrance (Håkanson, 1977;
Crusius and Anderson, 1995). Assuming certain
compensation between both factors, the value
obtained in 1997 with sediment traps (1.9 kg m-2)
would be comparable to the average of the years
when the tourism related erosion was more intense. Therefore, the values obtained between 1999
A paleolimnological study was carried out in
Cimera Lake sediment, where some changes
were observed in the diversity and abundance of
the chironomid head capsules in the more recent
sediment layers. It wasn’t possible to detect the
possible sources of direct impacts on the lake
that could be responsible for the biological
changes during this period (Granados & Toro,
2000a). To explain these changes a hypothesis
about a possible influence of local climate change (temperature) on chironomids communities
was tested by temperature reconstruction.
Several authors have used the subfossil chironomid to develop temperature reconstruction
models (Walker et al., 1991; Lotter et al., 1997:
Olander et al., 1997), though none of them is
specific for the Iberian Peninsula. Granados and
Toro (2000a) used the model developed by
Lotter et al. (1997) in the Alps to estimate the
summer mean temperature (June, July and
August), since it is the mountainous region biogeographically most similar to the Iberian
Peninsula, with a similar species composition.
The Cimera Lake species taxonomy was previously harmonized with the one use in the
model, to avoid incoherence in its application.
Figure 18 shows the mean summer temperature
reconstruction by way of subfossil chironomid
transfer functions, as well as by the use of long
term climate series (Agusti-Panareda &
Thompson, 2002). When applying the transfer
functions developed to reconstruct summer past
temperatures in the Alps to fossil chironomids
of the Cimera Lake, it is also well correlated
with reconstructed air temperatures (n = 20,
r = 0.45, p <0.01), especially when only the
most accurate dating levels (top of the core, ca
75 years) are taken into account (n = 13,
r = 0.75, p <0.01). However, 1) the linear
regressions of both models show significantly
different slopes, and 2) chironomid reconstruc-
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241
Figure 18. Upper graph: Reconstructed annual mean summer air temperatures using long climate series (Agusti-Panareda &
Thompson, 2002). Lower graph: Reconstructed mean summer air temperatures in Lake Cimera by means of climate model and chironomids calibration model (Granados & Toro, 2000). Gráfica superior: Reconstrucción de las temperaturas del aire medias anuales del verano mediante largas series climáticas (Agusti-Panareda & Thompson, 2002). Gráfica inferior: Reconstrucción de las
temperaturas del aire medias del verano en la laguna Cimera mediante un modelo climático y un modelo de calibración basado en
el análisis de quironómidos (Granados & Toro, 2000).
tion underestimates air reconstruction in ca.
3 ºC. The later is probably because the fossil
chironomid model has been developed for a different geographical region. Nevertheless, the
results of the two different and independent
models suggest the existence of an environmental warming of over 1.5 ºC in the summer mean
temperature since the 1980s in the Central
Range mountains. Our data also supports the
use of chironomid head capsules as an effective
tool for past temperatures inference.
The introduction of the brook trout
The effects of the introduction of the brook trout
(Salvelinus fontinalis), a salmonid from the
northeast region of North America, have been
studied in one of the Central Range lakes
(Peñalara Lake). Although this species was introduced in Spain at the end of the 19th century
(Gómez Caruana & Díaz Luna, 1991), it was in
the 1970s that mountain rivers and lakes were
populated for sport fishing purposes. The brook
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trout eats mainly aquatic invertebrates, though it
preys upon a large range of organisms throughout
its life cycle (zooplankton, juvenile fish, amphibious larvae) (Massabuau, 1997; Mullen, 1958;
Newman & Dubois, 1996; Scott & Crossman,
1973), causing, besides, nutrient cycling in the
lake water (mainly phosphorus), stimulating primary production (Slusarczyk,1997) and causing
an important alteration in the lake ecosystem.
Before the introduction of the brook trout, there
were no fish species in the Peñalara Lake, as
reported by researchers or naturalists who visited
the lake in that period (Margalef, 1949; Arévalo,
1921, 1931). Besides, the existence of a small
cascade in the lake outlet would be a barrier for
the fish to swim upstream to colonize the lake.
During 1997 a field experiment in situ was
carried out in the lake to test the effects of fish
predation on aquatic invertebrates by using two
types of limnological enclosures, pelagic and littoral, to prevent the access of fishes. The results
obtained with the shore limnological enclosures
do not reflect differences in the species composition within and outside these enclosures. A year
is probably not a long enough period to observe
the re-colonization of aquatic organisms.
Nevertheless, in some larger macroinvertebrate
taxa, some differences were observed in the densities of the communities inside and outside the
limnological enclosures, probably as a response
to the presence/absence of the brook trout (Fig.
19). About the macroinvertebrates communities
(littoral enclosures), the tricoptera and megaloptera, the larger sized taxa in the lake, have a larger population density inside the enclosures (Fig.
19a). There are no significant differences in the
coleoptera present, small in size. Those taxa
which live on top of the substrate (e.g. tricoptera)
are more susceptible to the brook trout predation
than the species which live buried in the sediment. In the presence of brook trout, no swimming or pelagic taxa individuals have been detected. The diptera of the chironomid and
ceratopogonid families, of small size, were not
different in both environments. However, the
tabanids, larger sized, had a slightly higher density inside the limnological enclosures. The oligochaetes have a higher population density outside the enclosures, while the only bivalve species
found in the lake tends to have a higher density
inside the enclosures. The Peñalara Lake diversity has been much lower during the years with
brook trout presence (9-13 species during the
introduction, compared to 23 after its eradica-
Table 3. Zoooplanktonic crustacea recorded in Peñalara Lake. In 1995 and 1997 columns, the percentage of samples with each taxa is indicated.
Maximum length is related to partenogenetic female about cladocera (according to Alonso, 1996), but in copepoda it is related to both sexes
(according to Dussart, 1969). Crustáceos zooplanctónicos citados en la laguna de Peñalara. Para los años 1995 y 1997 se cita entre paréntesis
el porcentaje de muestras en que se ha encontrado cada especie. La longitud máxima se refiere a la de la hembra partenogenética en los cladóceros, según Alonso (1996), mientras que en los copépodos se presenta la longitud máxima para ambos sexos, según Dussart (1969).
Cladocera
Margalef, 1949
(1 sampling)
Toro & Montes, 1995
(13 samplings)
Toro & Granados, 1997
(17 samplings)
Maximum lenght
Daphnia pulex
(=pulicaria)
—— —— ——
—— —— ——
—— —— ——
—— —— ——
—— —— ——
2.5 mm
—— —— ——
—— —— ——
Copepoda
—— —— ——
Daphnia longispina (18 %)
—— —— ——
Ceriodaphnia reticulata
Ceriodaphnia
Ceriodaphnia quadrangula (82 %)
quadrangula (100 %)
Alona quadrangularis (23 %)
Alona quadrangularis (47 %)
Chydorus sphaericus (92 %)
Chydorus sphaericus (76 %)
Eucyclopslilljerborgi
Eucyclops serrulatus (46 %)
(=serrulatus)
—— —— ——
Tropocyclops prasinus (100 %)
—— —— ——
—— —— ——
Eucyclops serrulatus (12 %)
Tropocyclops prasinus (94 %)
Cyclops strenuus (6 %)
2.3 mm
1.3 mm
0.8 mm
0.8 mm
0.5 mm
씸: 0.55-0.60 mm
씹: 0.67-0.95 mm
씸: 1.20-1.76 mm
씹: 1.46-2.30 mm
씸: 0.68-0.80 mm
씹: 0.80-1.45 mm
Limnetica 25(1-2)02
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243
Figure 19. Differences observed in the densities of the aquatic invertebrates communities inside (fish absence) and outside (fish
presence) the limnological enclosures in Peñalara lake during Summer 1997. a) Macroinvertebrates in littoral enclosures.
b) Zooplankton in pelagic enclosure. Diferencias observadas en las densidades de las comunidades de invertebrados acuáticos
dentro (con ausencia de peces) y fuera (con presencia de peces) de los limnocorrales instalados en la laguna de Peñalara durante
el verano de 1997. a) Macroinvertebrados en los limnocorrales litorales. b) Zooplankton en el limnocorral pelágico.
tion) when compared to the macroinvertebrate
communities found in other Peñalara Natural
Park ponds or lakes (Toro y Granados, 1998),
where there are no fish species. In the case of the
pelagic limnological enclosure, the densities of
organisms (zooplankton) both inside and outside
were relatively small, though presenting significant differences, and there were no differences in
species composition in the two environments
(Fig. 19b). The cladocera Ceriodaphnia reticulata, not cited before in this lake, much larger
(maximum length: 1.3 mm) than the Ceriodaphnia species previously cited, quickly appeared in the enclosure, reaching higher densities in
the absence of the brook trout pressure. The
second cladocera species of larger size, Alona
quadrangularis, had a lower density outside the
enclosure. The third cladocera species found,
Chydorus sphaericus, smaller in size and with a
low and irregular density in both environments,
had slight larger numbers inside the enclosure.
Regarding the copepods, the two species found
had a similar size (Tropocyclops prasinus and
Eucyclops serrulatus), and had similar population densities, though with a trend for higher densities within the enclosure. Lastly, the two rotifer
species identified (Asplancha priodonta and
Keratella quadrata) presented the opposite trend,
with a higher population density outside the
enclosure, in the presence of the brook trout.
Comparing the results of this and recent studies (Toro and Montes, 1993; Toro and Granados,
1997) after the fish introduction to the species
composition data in Margalef’s historic research,
previous to the introduction of the brook trout
(Table 3), there are some interesting changes,
which are probably related to the introduction of
this salmonid. Margalef (1949) only cites 2 species of plankton crustaceous in the Peñalara
Lake: Daphnia pulex and Eucyclops lilljerborgi,
Limnetica 25(1-2)02
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Toro et al.
which possibly would be correspond with
D. pulicaria and E. serrulatus according to upto-date or later taxonomy (Alonso, 1996;
Dussart, 1969, Margalef, 1953). In order to give
a comparison of relative sizes of each species,
there is a column with the species maximum
length, since this is an essential factor in the
probability of being predated by a brook trout.
The difference in the number of species found is
probably due to the fact that Margalef cited species found in one single sampling day (April
29th, 1949), while the other authors integrated
the species found in 30 lake samplings, varying
between 2 and 4 species per sample. After the
brook trout introduction, two crustaceans clearly dominated: Ceriodaphnia quadrangula and
Tropocyclops prasinus, representing at least
50 % of the individuals present, being frequently found with Chydorus sphaericus and, in
a smaller proportion, with Alona quadrangularis and Eucyclops serrulatus, all species of
small or medium size. On the other hand, the
larger species, the cladocera Daphnia longispina and the copepod Cyclops strenuus, are only
found sporadically and always represent less
than 5 % of the total of individuals in a sample.
Margalef ’s citation in 1949 of the presence of
Daphnia pulex (= D. pulicaria), a macrofilterfeeder of a large relative size, without other
smaller sized cladocera, is possibly evidence
that the presence of this predator has clearly
favored the smaller species in relation to the
macrofilters. Besides, Margalef (1949) himself
comments that “the absence of planktonic rotifers is remarkable”, while ten species of this
group have been found in later samplings. The
changes in the rotifer community structure caused by different fish population densities have
been observed in experiments in other lakes
(Stenson, 1982). In summary, the larger species
tend to dominate in the zooplankton not submitted to the brook trout pressure. The copepod and
cladocera population densities are higher inside
the limnological enclosure. In contrast, the density of rotifers is higher outside the limnological
enclosure. The brook trout juvenile stages have
predominantly plantivore habits, although their
incidence is directly proportional to the zoo-
plankton species size. The rotifers, microscopic
organisms, would be, thus, benefited by the
brook trout presence, being able to use the trophic resources that would be otherwise consumed by the larger species in the zooplankton. As
additional information, Bosch et al. (2000)
observe that the common salamander (Salamandra salmandra) and the midwife toad (Alytes
obstetricans) reproduced in the lake before the
introduction of the brook trout. There are no
citations of salamander larvae after the introduction of the brook trout (Bosch et al., 2002).
The results of the experiments with shore and
pelagic limnological enclosures, as well as the
existing historic data, reflect remarkable changes in the aquatic vertebrate community structure in the lake, due to the brook trout predatory
activity. The negative effect that the fish fauna
can have on the benthonic and planktonic communities has been demonstrated in other lakes:
Braña et al. (1996) found a significant decrease
in the abundance of amphibious larvae in high
mountain lakes in the Cantabric Range in relation to those without fish; Balvay (1978) observed the almost complete disappearance of benthos just a few years after the introduction of
salmonids in a lake in the French Alps. Johnson
et al. (1996) demonstrated a clear decrease in
the number of benthonic (triclads, mollusks,
odonate, ostracode) and planktonic (large sized
daphnid) invertebrates due to predation by fish
in mesocosmos experiments. Therefore, the eradication of the brook trout in the lake was proposed, in order to recover the aquatic population
previous conditions. In order to achieve this, gill
nets were used for 5 years, until the total absence of the brook trout was confirmed. Before the
brook trout eradication, the maximum number
of aquatic invertebrate families in the months
without the ice cover (1991-2000) was 9-13.
This taxonomic richness increased to 14 families in 2001, with a reduced brook trout population. Once the brook trout was eradicated,
17 families were found during the period
without the ice cover in 2002, and 23 families
were found in the two following years (2003 and
2004). It is evident that the brook trout eradication has brought about an increment in the taxo-
Limnetica 25(1-2)02
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Página 245
nomic richness of the macroinvertebrate fauna,
and an absolute transformation in the benthonic
macroinvertebrate community. Practically all
the species that inhabited the lake before the
eradication maintained their population, and a
large number of taxa, mainly large sized aquatic
insects, were added to the community. These
taxa have been found in other Natural Park
ponds and lakes (Toro y Granados, 1998), from
which they were able to re-colonize the lake
once their predator was eliminated.
CONCLUSIONS
The ecological characteristics of the Iberian
Peninsula Central Range high mountain lakes
make them extraordinary sensors for natural or
human induced environmental change. The chemical composition of their waters, their thermal
and hydrological dynamics, as well as their biological communities respond with high sensitivity
to changes in climate variables such as temperature or precipitation, to excess nutrients from the
watershed or the atmosphere, to erosion processes related to tourism or cattle raising activities,
to water regulation by damming, to the introduction of exotic species, or to the organic contamination of the water. However, their self-recovery
capacity once the pressure or the origin of the
impact is eliminated is also extraordinary, due,
mainly, to their high water turnover rates.
One of the main contributions of this work is
the joint utilization of paleoinformation and
modern monitoring systems as support for management and restoration of the lake ecosystems
and their watersheds. Besides, the importance of
the historic information provided by the first
researchers who dedicated their efforts to discover remote, unaltered ecosystems, leaving a writte heritage of invaluable documental and scientific importance, is emphasized.
Finally, this work contributes indirectly by stressing the importance of communication and of cooperation between scientists and managers, in order
to achieve a greater efficiency in environmental
conservation and to technically back up management measures. The need for institutional support
245
to develop long-term monitoring systems which
can provide invaluable information about local,
regional and global trends is also emphasized.
ACKNOWLEDGEMENTS
This research has been funded by following institutions and projects: The Natural Park of
Peñalara (Consejeria de Medio Ambiente,
Regional Government of Madrid), the Delegación Territorial de Ávila (Consejería de Medio
Ambiente y Ordenación del Territorio, Junta de
Castilla y León), Acciones Integrades between
UK (British Council) and Spain (M.E.C.) No.
93A (1991-1992), MOLAR project (EU contract
ENV4-CT95-0007). Collaborators from the
Department of Ecology (Universidad Autónoma
de Madrid), Natural Park of Peñalara staff and
personnel from the Mountain Refuge Elola
(Sierra de Gredos) have helped in field or laboratory works since the beginning of the projects.
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Limnetica, 25(1-2): 253-270 (2006)
Instability of shallow lakes: A matter of the complexity of factors
involved in sediment and water interaction?
I. de Vicente1, 2,*, V. Amores1 and L. Cruz-Pizarro1
1 Instituto
del Agua, Universidad de Granada, 18071 Granada, Spain
of Biology, University of Southern Denmark, 5230 Odense, Denmark
*corresponding author: [email protected]
2 Institute
ABSTRACT
Instability and extreme fluctuations in water quality characterizing shallow lakes are to a large extent due to rapid changes in
the internal supply rates of nutrients. In fact, is in these systems that sediment and water interaction plays a major role in
nutrients’ dynamics. For this study, a double-approach perspective with both field measurements and laboratory experiments
has been used in order to determine the contribution of physical, biological, and chemical mechanisms to nutrients’ dynamics
in two shallow adjacent lakes in Andalusia (Spain). Despite their geographic proximity, strong differences between the study
systems have been recognised. In the Lake Honda, the high nutrient concentrations, which ultimately support a large algal biomass, are the result of: i. Resuspension of the surface sediment favoured by its morphometry, hydrologic regime and sediment
granulometry. ii. Intense organic matter mineralization due to the labile nature of the organic settled matter (planktonic). In
Lake Nueva, by contrast, physical constrains (i.e. wind-induced resuspension) have a limited effect due to the coarse surface
sediment and to the development of macrophytes (Najas marina, Potamogeton pectinatus). In addition, the structurally more
complex organic matter of its sediment regulates the low nutrients turnover. In this lake, nutrient exchange rates across the
sediment-water interface are also controlled by chemical processes, such as P adsorption onto CaCO3, a mechanism that is
favoured by the high Ca+2 concentration in the interstitial water. In this way, the joint effect of physical, chemical, and biological mechanisms determine the fast nutrients’ benthic regeneration in Lake Honda, while in contrast, a large fraction of the
nutrients present in the sediment of the Lake Nueva is in particle form.
Keywords: instability, shallow lagoons, benthic nutrient regeneration
RESUMEN
La inestabilidad y las extremas fluctuaciones en la calidad del agua que caracterizan a los lagos someros, se deben en gran
medida, a cambios rápidos en la carga interna de nutrientes. De hecho, es en éstos sistemas, donde la interacción agua-sedimento juega un papel esencial en la dinámica de los nutrientes. Para este estudio, se ha empleado una doble aproximación,
basada tanto en datos de campo como en experimentos en laboratorio, para determinar la contribución de mecanismos físicos, químicos y biológicos a la dinámica de los nutrientes en dos lagunas costeras en Andalucía (España). A pesar de su proximidad geográfica, se han reconocido fuertes diferencias entre ambos sistemas. En la laguna Honda, las elevadas concentraciones de nutrientes, que en última instancia soportan una elevada biomasa algal, son el resultado de: i. Resuspensión del
sedimento superficial, favorecida por la morfometría, el régimen hídrico y la granulometría de su sedimento. ii. Intensa mineralización de la materia orgánica debida al carácter lábil de la materia orgánica sedimentada (origen planctónico). En la
laguna Nueva, por el contrario, el impacto de los factores físicos (p.e. resuspensión inducida por el viento) se encuentra limitado por la gruesa granulometría del sedimento superficial así como por el desarrollo de macrófitos (Najas marina,
Potamogeton pectinatus). Más aún, la naturaleza estructuralmente más compleja de la materia orgánica presente en su sedimento determina unas menores tasas de regeneración de nutrientes. En esta laguna, las tasas de intercambio de nutrientes a
través de la interfase agua-sedimento se encuentran, además, controladas por procesos químicos, tales como la precipitación
de P sobre CaCO3, mecanismo que se encuentra favorecido por las elevadas concentraciones de Ca+2 presentes en el agua
intersticial. Por tanto, es la conjunción de mecanismos físicos, químicos y biológicos los que determinan la rápida regeneración béntica de nutrientes en la laguna Honda; mientras que, por el contrario, una importante fracción de los nutrientes presentes en el sedimento de la laguna Nueva se encuentra en forma particulada.
Palabras clave: inestabilidad, lagunas someras, regeneración béntica de nutrientes
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de Vicente et al.
INTRODUCTION
Lake sediments play an outstanding role in limnological studies as they can both reflect and
affect what is occurring in the overlying water
column (Håkanson, 1984). In fact, sediments
are the product of lake life and, consequently,
they reflect the lake type. This notion was firstly
stated by Lundqvist, who made several major
investigations on the relationships between lake
type, sediment type and lake surroundings
during the thirties and forties (Lundqvist, 1938;
1942 in Håkanson, 1984).
In this sense, sediments can be regarded as a
bank of information about environmental changes occurring in both the water body and in the
catchment area (Kalff, 2002; Luque and Juliá,
2002; Schmidt et al., 2002). Accordingly, there
are many studies focused on the reconstruction
of a lake’s trophic level using decay-resistant
remains (diatom frustules and pigments) present
in the sediment (Adams and Prentki, 1986; Lo
Bianco, 1997; Musazzi, 2000; Bennion et al.,
2000; Lotter, 2001). Furthermore, and regarding
the chemical composition of the bottom sediments, many studies have stressed the relevance
for analysing the sedimentary phosphorus (P)
fractions to get some insights about the mechanisms involved in CaCO3 precipitation (i.e. de
Vicente et al., 2006 and references there in).
Besides considering lake sediments as a historical record, sediments may also affect the
water quality as a consequence of their dynamic and active character resulting from a
great variety of biogeochemical reactions
and transformations. In particular, it is in shallow lakes where this sediment-water interaction is extremely important for understanding
the whole nutrient dynamics (i.e. Ryding,
1985; Boström et al., 1988).
For the case of P, this internal loading has
been identified as an important mechanism in
delaying recovery of shallow lakes following
reduced external P loading (Marsden, 1989;
Sas, 1989; Ryding and Rast, 1992; Harper,
1992; Istvánovics and Somlyódy, 1999;
Søndergaard et al., 1999; Schauser et al.,
2003, among others). As Golterman (1995)
established, the phosphate concentrations in
the sediment and the overlying water are in a
dynamic equilibrium where the position of
this equilibrium, that controls whether input or
output dominates, is determined by the interaction of multiple factors that may change
over different time-scales.
We will synthesize some of the most relevant processes that could account for the
dynamic sediment-water interactions characterizing shallow lakes.
Firstly, physical processes, such as resuspension of unconsolidated sediment, usually play a
fundamental role in shallow lakes, where sediment often undergoes continuous wave action
(i.e. Kristensen et al., 1992; Nõges et al., 1999;
Weyhenmeyer and Blosech, 2001). Evidence
for the importance of resuspension is plentiful
(Kristensen et al., 1992; Evans, 1994; Bloesch,
1995; Weyhenmeyer et al., 1995; Golterman,
2004): Water quality in a lake is affected by
reduced light penetration, which can ultimately
promote biological changes, inducing a transition from a macrophyte-dominated community
to a plankton-dominated one. Furthermore,
nutrients recycling increases due to sedimentary
nutrients (in particulate and dissolved forms)
being brought back to the water column where,
because of previously low phosphate levels,
available P adsorbed to sediment particles could
also be released and alter the lake’s trophic status (Peters and Cattaneo, 1984). Nevertheless,
the final effect of resuspension events on P
availability depends on the particular properties
of the lake’s water (i.e. o-P concentration) and
of the sediment (i.e. FeOOH:TP), being therefore to a certain extent lake specific (Søndergaard
et al., 1992; de Vicente, 2004).
Apart from wind-induced resuspension, shallow lakes are also characterized by the fact that
most of the organic matter produced in the water
column, reach the sediment without being mineralised. Hence, it can be emphasized that sediments represent a crucial place for organic matter
decomposition in these systems (Mann, 1982).
There is controversial evidence on the relative importance of microbial activity on aerobic
P-release from the sediment. Evidence is
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Instability of shallow lakes
mounting that microbial activity has a significant role in P-release (de Montigny and Prairie,
1993; Gächter and Meyer, 1993). Although
many studies have been developed with an
inappropriate P-fractionation methodology, the
P-release from the sediment was shown to
decrease in experiments were bacterial activity
had been inactivated (Boers, 1986; Sinke and
Cappenberg, 1988). The decomposition of
organic matter in the sediment is a simple
mechanism of o-P release that cannot be ruled
out (de Montigny and Prairie, 1993). Qiu and
McComb (1995) showed that the bacterial
influence on P-release was mostly through their
influence on the breakdown of org-P.
However there also exist many studies revealing that microbial activity causes an immobilization of sedimentary P. In those cases, Prelease is lower in non-sterilised (biotic and
abiotic processes) than in sterilised samples
(abiotic process) (Kamp-Nielsen, 1974;
Doremus and Clesceri, 1982; Eckert et al.,
1997; Clavero et al., 1999; Watts, 2000). A
likely explanation for these controversial
results is the different P limitation of bacterial
activity. Then, when bacteria are P limited is
likely to occur a P retention during organic
matter mineralization, while they tend to
release P when there is no a P limitation.
Another biologically mediated process that
may account for the complexity of the sediment
and water interactions typical of shallow lakes is
the P translocation by some species of
Cyanobacteria that live part of their life cycle in
the sediment (i.e. Aphanizomenon, Gloeotrichia
echinulata) (Osgood, 1988; Pettersson et al.,
1993; Istvánovics et al., 2002). Furthermore,
daily cycles production and respiration of benthic algae may affect both directly, through uptake and release of nutrients, and indirectly, by
modifying redox potential, the nutrients exchange across the sediment-water interface (Carlton
and Wetzel, 1988; Kelderman et al., 1988).
Lastly, radicular systems of the aquatic
macrophytes add even more complexity to P
exchange across sediment-water interface. By
one hand, they can uptake dissolved nutrients
from the pore-water (Carignan and Kalff, 1980;
255
Carignan, 1982; Barko and Smart, 1980, 1981;
Carignan, 1985; Barko et al., 1991). By the
other hand, they can also release O2 from the
rhizome promoting the precipitation of P bound
to FeOOH (Christensen and Andersen, 1996).
Aquatic macrophytes may, moreover, indirectly affect sedimentary P dynamic by sediment stabilization and the subsequent resuspension reduction (i.e. James and Barko, 1990;
Horppila and Nurminen, 2003).
Regarding chemical processes occurring at
the sediment-water interface, their study has
been traditionally focused on the Fe and P compounds’ chemistry (Mortimer, 1941 in 1971).
Nevertheless, recent studies have reviewed the
redox-controlled P retention in lake sediments
concluding that, apart from O2 availability, the
Fe:P ratio plays an outstanding role for P retention during aerobic conditions (Jensen et al.,
1992; Gächter and Wehrli, 1998; Lehtoranta and
Heiskanen, 2003; Gächter and Müller, 2003).
Since the classical view about sediment and
water interactions was mainly focused on the
abiotic redox-dependent fixation and release
of P, adsorption and desorption processes of P
onto CaCO 3 have been continuously underestimated. However, it has recently been outlined the relevance of the anaerobic P release
in hard-water lakes, as a result of apatite dissolution occurring at low pH (Driscoll et al.,
1993; Gómez et al., 1999; Golterman, 2001).
The interactions between P and Ca compounds are of major interest for shallow lakes,
due to their predominant eutrophic condition
and usually high alkalinity, and also due to
their quick water heating, both factors reducing CaCO 3 solubility.
All in all, instability and extreme fluctuations in water quality characterizing shallow
lakes may be to a large extent regulated by
rapid changes in the internal rates of nutrient
supply. In this context, the present study is
based on a multi-approach perspective focused
on both field measurements and laboratory
experiments in order to determine the contribution of physical, biological and chemical
mechanisms to nutrients’ benthic dynamics in
two shallow adjacent lakes.
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de Vicente et al.
Figure 1. Location of the study lakes. Localización geográfica
de las lagunas estudiadas.
STUDY SITE
The “Albufera de Adra” is the most important
wetland located in south-eastern Spain (Fig. 1). It
is composed of two shallow and eutrophic coastal
lakes: Honda and Nueva. Because of the high biodiversity, an area of 230 ha surrounding both
lakes was converted into a Natural Reserve (1989)
and more recently (1994), Lake Honda and Lake
Nueva have been included in the list of Protected
Areas of the Ramsar Convention. Nevertheless,
since the seventies, a steady process of land reclamation for agriculture (greenhouses) has resulted
in an accelerated eutrophication process (Martinez-Vidal and Castro, 1990; Carrillo et al.,
1996). During the last few years, both lakes have
been deeply studied in the framework of a comprehensive Project (“Conservación de las Albu-
feras de Adra”), partly financed by the UE-LIFE
NATURE programme, which aims to insure that
any water quality remedial measure is proposed
on the basis of a thorough diagnosis and evaluation of the system (Cruz-Pizarro et al., 2002a).
A recent study has evidenced that although both
study lakes show a notable temporal variability, in
agreement with the unstable features of eutrophic
systems (i.e. Barica, 1980; Dokulil and Teubner,
2003), Lake Honda is characterized by large longterm, seasonal and diel fluctuations in water quality (de Vicente et al., 2004) (Table 1).
Among the triggering mechanisms for the
seasonal variability of Lake Honda, we can
firstly remark the rapid changes in the external
inputs and its high flushing rate (de Vicente et
al., 2003). Because the hydrologic regime is
dominated by surface water inputs (Benavente
and Rodríguez, 2001; Benavente et al., 2003),
Lake Honda can be considered as an epigenic
and recharge wetland (González-Bernáldez,
1992). It explains its extreme high temporal
variability through weather conditions and
human activities on the drainage basin, and also
its hypertrophic state as a result of the incoming
of large amounts of allochthonous material from
run-off (Cruz-Pizarro et al., 2002a,b; de Vicente
et al., 2003; de Vicente and Cruz-Pizarro, 2003).
By contrast, Lake Nueva can be classified as a
hypogenic and discharge wetland where the predominant entry of groundwater plays an important role in buffering most wetland characteristics and in increasing its temporal stability.
In the present study, we hypothesize that,
apart from external forcing (i.e. weather and
watershed processes), the noteworthy unpredic-
Table 1. Inter-annual variability in water transparency (ZSD), Chl-a, o-P, TP and TN concentrations (modified from de Vicente, 2004).
Mean (Min-Max). Variabilidad interanual en la transparencia del agua (ZSD) y concentraciones de Chl-a, o-P, TP y TN (modificado de
de Vicente, 2004).
Lake Honda
o-P (µg l-1)
TP (µg l-1)
TN (mg l-1)
Chl-a (µg l-1)
ZSD (cm)
Lake Nueva
1999-2000
2000-2001
1999-2000
2000-2001
40 (2-210)
255 (112-425)
2.94 (1.57-5.59)
160 (57-396)
34 (20-50)
77 (0-275)
295 (146-471)
3.86 (0.99-7.14)
129 (4-292)
64 (10-250)
7 (0-32)
79 (25-147)
1.10 (0.80-1.67)
57 (7-126)
119 (70-210)
4 (0-24)
99 (24-155)
1.68 (1.16-2.00)
54 (8-125)
75 (30-140)
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tability observed in the nutrients’ dynamic of
Lake Honda compared to Lake Nueva (de
Vicente et al., 2003; de Vicente, 2004) is the
result of the specially intense physical (resuspension), chemical (oxidation vs. reduction),
and biological (i.e. organic matter mineralization) processes, occurring at the sediment-water
interface of this hypertrophic lake.
lake-water samples were collected using a Van
Dorn sampler at three different depths in the
vertical profile (surface, mid-depth and bottom). Once at the laboratory, total phosphorus
(TP) and total nitrogen (TN) were directly measured from non-filtered water (APHA, 1995).
A sub-sample was filtered for the analysis of
inorg- and org-Cdis concentrations using a TC
Autoanalyser (Dohrman, DC-190).
Sediment Monitoring
Meteorological and morphometric data
Chemical analyses
Daily wind-speed data were gathered from a
meteorological station at the experimental station “Las Palmerillas”, located in El Ejido, the
closest town to the lake area (less than 5 km).
Hypsographic curves and morphometric
variables were based on the bathymetric map
(Cruz-Pizarro et al., 1992).
Hydrological regime is described in detail
in de Vicente, 2004).
Surface sediment samples (0-5 cm) were
collected monthly, from July 2000 to August
2001, at the deepest site of each lake, using an
Ekman dredge. Within 24 hours, the interstitial water was separated from the sediment
particles by centrifugation at 5000 r.p.m.
during 10 min. (Enell and Löfgren, 1988). The
supernatants were then f iltered through
Whatman GF/C f ilters. The wet sediments
were kept at 4ºC until they were fractionated
1-2 weeks later. No treatment (drying, freezing or sieving) was performed on the sediment
samples before fractionation.
The concentration of TN and TP were quantified in the non-filtered interstitial water
(APHA, 1995). A sub-sample was filtered for
o-P quantification as molybdate reactive phosphorus (Murphy and Riley, 1962), the inorgand org-C diss using a TC Autoanalyser
(Dohrman, DC-190) and Ammonium (NH4+)
concentration following Rodier (1989).
Sediment was analysed for TC and TN determination with a CNH Elemental Analyser. The
P-fractionation in the sediment was performed
following the EDTA method, based on a sequential extraction with chelating compounds
(Golterman, 1996), (Table 2).
Water Column Monitoring
A fortnightly monitoring of Lake Honda and
Lake Nueva was conducted from August 1999
to August 2001. From a sampling station located at the maximum depth site of each lake,
Table 2. Abbreviations used in the text. Abreviaciones usadas en el
texto.
CaCO3≈P
Fe(OOH)≈P
Fe(OOH)
org-P→acid
org-P→alkali
org-Presidual
org-P
inorg-P
TP
TN
TC
o-P
org-Cdis
inorg-Cdis
C.F.U.
calcium bound phosphate
iron bound phosphate
ferric oxyhydroxides
acid soluble organic phosphate
hot NaOH soluble organic phosphate
residual organic phosphate
organic P
inorganic P
total P
total N
total C
ortho-phosphate;
dissolved organic C
dissolved inorganic C
Colony Formation Units
Physical properties
The granulometric composition of the surface
sediment layer was determined using the
method proposed by Robinson (1922), basically
based on the Stokes law.
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Biological activity
Contribution of biotic vs. abiotic processes on P
sedimentary recycling
For the continuous-flow experiment, 250 ml of
H2O were added to a set of 5 beakers (10 cm
diameter; 500 ml capacity) each containing 35 g
of fresh sediment (1cm sediment layer) collected from Lake Honda and Lake Nueva in July
and August 2001, respectively (Serrano et al.,
2005). Three beakers were autoclaved (120ºC,
20 min) following Clavero et al. (1999) and two
others were kept as controls. Continuous flow
system consisted of incubating each beaker with
a continuous input of H2O (controls) and H2O +
1 % ClCH3 (sterilised treatments), controlled by
electrical pumps. In order to maintain a constant
water volume inside the beaker (250 ml), output
flow was exactly the same as input (1 ml min-1).
The water from the outlet, collected in PET bottles, was kept for pH determination and for o-P
concentration (Murphy and Riley, 1962).
Conditions for incubation were darkness and an
average temperature of 30ºC.
Sediment respiratory activity
Electron Transport System Activity (ETSA)
was determined at the sediment top layer (0-5
cm) at the deepest site of each lake, from
February to November 2002. ETSA was measured by using the method proposed by Broberg
(1985) with slight modifications. In principle,
this method is based on the biological reduction
of the tetrazolium salts to their respective tetrazolium formazan by sediment microorganisms.
The formazan produced can then be used as a
measure of the ETSA in the sediment. For measuring ETSA in the study lakes, 2-3 grams of
wet sediment was mixed with 10 ml of homogenisation buffer (MgSO4·7H2O, PVP, Triton and
EDTA) and sonicated in an ice bath for 4 min
(0ºC). The mixture was clarified by centrifuging it at 10.000 rpm for 10 min (0ºC). Then,
we mixed 0.5 ml of the supernatant with 1 ml of
substrate solution (NADH, NADPH and Nasuccinate), 0.5 ml of INT (2-(p-iodophenil-)-3(p-nitrophenol)-5 phenil tetrazolium chloride)
and 0.5 ml of the homogenisation buffer. The
mixture was incubated at the same temperature
measured in the field (15-28ºC) for 20-30 min,
depending on the temperature. Immediately
after time incubation, the addition of Quench
(phosphoric acid and formaldehyde in 1:1 proportion) was carried out in order to stop the
reaction. Absorbance of the sample at 490 nm
was read with a spectrophotometer. In calculating ETSA, the molar adsorption coefficient of
INT-formazan of 1.42 (Kenner and Ahmed,
1975) was used. All ETSA values were determined within 24 hours of field sampling. For
optimising the method, a set of preliminary
experiments were performed in order to determine the amount of wet sediment, homogenisation conditions (method, time and homogenisation solution) and also to establish the
sufficient concentration of substrates to achieve
Vmax of the INT reduction (de Vicente, 2004).
In addition, Chl-a, Phaeopigments, Colony
Formation Units (CFU) of anaerobic and aerobic bacteria, and Organic Matter (O.M.) content were determined at the surface sediment.
All of these methods are thoroughly described
in de Vicente (2004).
Figure 2. Estimated value for the wave height (m) and estimated lake area (%) that is affected by waves (from de
Vicente, 2004). Valor estimado de la altura (m) de la onda
generada por el viento y del área (%) del lago afectada por
las olas (tomado de de Vicente, 2004).
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Sedimentation rates
A pair of sediment traps, Plexiglas cylinders
with an aspect ratio higher than 6 (Bloesch and
Burns, 1980), were set at three different layers
(50, 135 and 260 cm) at the deepest site of each
lake. The particulate matter collected in the
traps, every two weeks (February 2000-August
2001), was dried (60ºC for 24 h) and weighed,
to quantify settling fluxes (g d.w. m-2 d-1). The
settled material was also analysed for TC, TN
and TP using the method described above.
Achieving the impact of sediment resuspension
Model to predict resuspension events
Following Carper and Bachman (1984), wave
period (T, s) has been estimated as a function
of wind speed (V, m s-1) and of the effective
distance over water that the wind blows (fetch,
F, m), by the equation:
gT
2πV
冋
= 1.20 tanh 0.077
冉 V 冊册
gF
2
(1)
Table 3. Main morphometric and hydrological features of the studied
lakes (July 1999–August 2001). 1de Vicente et al. (2003). Principales
características morfométricas e hidrológicas de las lagunas estudiadas (Julio de 1999–Agosto de 2001). 1de Vicente et al. (2003).
Lake Honda Lake Nueva
Lake area (A) (m2, 103)
Maximum lenght (m)
Shoreline lenght (m, 102)
Maximum depth (Zmax) (m)
Volume (m3, 103)
Catchment area (Ac) (m2, 105)
–
Mean depth (Z ) (m)
Relative depth (m)
Shore development
Zm:Zmax
Ac:A
苴 /Z– )
Dynamic ratio (冪A
Residence time (yr)1
External Areal loading
(g P m-2 yr-1)1
80.1
586
14.66
3.19
91.52
137.2
1.14
1.00
1.46
0.36
171.30
253
0.17
1.73
260.4
759
20.66
3.80
594.70
5.0
2.28
0.66
1.14
0.60
1.92
223
2.95
0.03
cific details about these experiments are provided in de Vicente (2004) and de Vicente (2004).
Statistical analysis
Thus, wavelength (L, m) is related to its
period (T) by the equation:
gT 2
L=
2π
(2)
where g is the gravitational constant (9.8 m s-2).
Wave height is then calculated as one-half of
the wavelength. It can be assumed that windinduced waves touch the bottom when the water
depth is less than a half of its wavelength
(Carper and Bachmann, 1984).
Adsorption experiments
The flocculent layer was sampled in November
2002 at the maximum depth station in each lake,
using a horizontal Van Dorn sampler, which was
bounced off the bottom a few times to resuspend
the sediment (Doremus and Clesceri, 1982). In
the laboratory, the flocculent layer was concentrated by centrifugation (10 min, 10000 rpm)
and P adsorption experiments were carried out
using the batch-experimental technique. Spe-
Statistical analysis was performed using
StatSoft, Inc. (2001). For t-student test, unless
otherwise stated, the significance level was
established at p < 0.05.
Results and discussion
Although both study lakes show a notable temporal variability, Lake Honda is characterized
by large long-term, seasonal and diel water
quality fluctuations (de Vicente, 2004). Instability and extreme fluctuations in Lake
Honda water quality are to a large extent regulated by rapid changes in the internal rates of
nutrient supply as a result of intense biological,
physical, and chemical mechanisms at the sediment-water interface.
Physical forcing of water quality
Some of the most important morphometric features of both lakes are summarised in Table 3.
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Figure 3. Areal hypsographic curves. Curvas hipsográficas de
superficie.
Table 4. Results of the adsorption experiment with the floculent layer
(from de Vicente, 2004). Resultados del experimento de adsorción
con el sedimento resuspendido (tomado de de Vicente, 2004)
Figure 4. Granulometric composition of the surface sediment.
Composición granulométrica del sedimento superficial.
Table 5. Some of the biological variables measured at the surface
sediment. Mean (SD). (n=7). Algunas de las variables biológicas
medidas en la superficie del sedimento. Media (SD) (n=7).
Lake Honda Lake Nueva
Honda
Nueva
Pinitial (µg P-PO4-3 l-1)
Pads (µg g-1 d.w.)
37
4
-22
+4
The higher values reported for the dynamic
ratio, a measure for bottom dynamics (Håkanson and Jansson, 1983), in Lake Honda might
suggest that resuspension may play a relatively
major role in this lake compared to Lake
Nueva. Similarly, the application of the empirical model showed that the impact of sediment
resuspension is higher in the former lake than
in the later one. In particular, as figure 2 shows,
more than half of the lake area was affected by
waves in 80 % of the wind events in Lake
Honda and only in 25 % in Lake Nueva.
The present study highlights that, apart from
extrinsic factors (wind velocity), sediment resuspension depends on lake morphometry and sediment properties, factors that cause Lake Honda
to be much more affected than Lake Nueva by
resuspension events. Thus, while in Lake Honda
a great proportion of sediment is located at relatively low depth (Fig. 3), sediment of the more
recent Lake Nueva is less subjected to windinduced turbulence, and sediment resuspension is
also ultimately limited by the relevant contribution of the sand-to-mineral matrix (Fig. 4).
ETSA (µl O2 g-1 d.w. h-1)
O.M. (%)
Phaeopigments (µg g-1 d.w.)
Chl-a (µg g-1 d.w.)
Anaer. Bacteria (log C.F.U./g dw.)
Aer. Bacteria (log C.F.U./g dw.)
ETSA:OM
50.1 (15.7) 63.6 (48.6)
8.6 (1.9)
16.65 (1.4)
103.5 (3.5) 101.33 (19.9)
41.1 (8.8) 68.5 (29.8)
6.55 (0.50) 6.15 (0.42)
7.41 (0.18) 7.00 (0.25)
5.9 (1.9)
3.8 (2.7)
Finally, the impact of sediment resuspension has
also been examined simulating the effect of lake
water enrichment in resuspended material. Our
results have shown that while the flocculent
layer tends to release phosphate to the water
column in Lake Nueva, resuspension causes a
phosphate removal from the water column in
Lake Honda. Such patterns are a consequence
of the large differences in the o-P concentrations in the two study lakes (Table 4).
Additionally, we may outline the feedback relationship between resuspension and light climate
in both lakes. In Lake Honda, resuspension events
increase the total suspended solids, thereby attenuating light and limiting the development of
macrophytes and benthic algae, which would
otherwise aid sediment stabilization. By contrast,
resuspension in Lake Nueva is also limited by the
presence of cohesive agents in the sediment such
as algal mats and submerged macrophytes, their
growth being encouraged by the strong light
penetration (Cruz-Pizarro et al., 2002a).
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As de Vicente (2004) have suggested, several
facts indicate that resuspension plays a greater
role in Lake Honda than in Lake Nueva.
Firstly, the C:P ratio is similar in the surface
sediment than in the material collected by the
bottom sedimentation traps in Lake Honda,
while Lake Nueva shows major differences
between chemical composition of the two
materials. Secondly, organic-matter concentrations in the surface sediment from the hypertrophic Lake Honda are far lower than in Lake
Nueva (Table 5), presumably because of the
dynamic transport of sedimentary particulate
organic matter to the water column, which
would stimulate its overall mineralization, in
the way some recent studies have also suggested (Wainright and Hopkinson, 1997).
Biological activity at the surface sediment
Biotic vs. abiotic P release
In Lake Nueva, the sediment retained o-P and
there were no significant differences between
poisoned and non-poisoned treatments (Fig. 5).
Figure 5. Variation in time of o-P concentrations in the water
from control and sterilised treatment (modified from Serrano
et al., 2005). Variación de la concentración de o-P presente en
el agua del control y del tratamiento esterilizado (modificado
de Serrano et al., 2005).
261
Sediment from Lake Honda by contrast, showed
very high P release rates in the control compared
to the sterilised treatment. In this lake, abiotic
mechanisms may play a secondary role, while
biotic ones mainly control sedimentary phosphate mobilization. These results would hence confirm the temperature-dependence of o-P release
rates that has been previously reported by de
Vicente (2004). Actually, the impact of biotic
activity on o-P release could be both direct by
hydrolysing organic P, and indirect, by lowering
the redox potential in the surface sediment which
ultimately may induce release of Fe(OOH)≈P, a
P-fraction that presents a large pool in the sediment of Lake Honda (Table 6). The relevance of
biologically mediated aerobic o-P release, as stimulated by raised temperatures, was shown by
Jensen and Andersen (1992), who found a notable reduction in the thickness of the oxidized surface layer and a subsequent increased in P release
rates when temperature was raised.
Differences observed in aerobic P-release
from the sediment of both studied lakes are
likely to be explained by differences in organic
matter quality. De Vicente et al. (2003) showed
that although the sestonic material had a similar
concentration of TP in both lakes, the average
C:P ratio was higher in Lake Honda than in Lake
Nueva. However, the biodegradability of the top
sediment was lower in Lake Nueva as the C:P
ratio of the top sediment was significantly higher
in this lake. Then, it is likely that the top sediment of Lake Nueva had a different source of
organic matter, and hence C, that increased the
C:P ratio. In this sense, the concentration of
planktonic chlorophyll a was higher in Lake
Honda, while the lower turbidity of the water in
Lake Nueva favoured the growth of submersed
macrophytes in the littoral area during spring
and summer (Cruz-Pizarro et al., 2002b). As it is
well known phytoplankton detritus is easily
degradable, while vascular plants remains are
structurally complex and their degradation is
slower (Kristensen et al., 1995). Therefore, the
presence of vascular plants that grew on the
sediments could have accounted for the higher
C:P ratio in the top sediment of Lake Nueva.
This could cause a lower mineralization rate and
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Table 6. P-fractions at the surface sediment (µg g-1 d.w.) (n=12).
Fracciones de P en sedimento superficial (µg g-1 d.w.) (n=12).
Fe(OOH)≈P
CaCO3≈P
Org-P→acid
Org-P→alkali
Org-Presidual
ΣPsed
Inorg-P:Org-P
FeOOH (mg Fe g-1 d.w.)
FeOOH:TP (atomic ratio)
Honda
Nueva
135
313
61
174
11
694
1.82
15.6
15
18
142
53
166
14
393
0.68
5.2
5
thereby leave a higher concentration of organic
matter and org-P in the surface sediment as compared to Lake Honda. We suggest that the relative importance of planktonic (phytoplankton) and
benthic (macrophytes) primary producers can
explain the differences observed in the P-recycling and sediment-P composition of both lakes.
primary production in Lake Balaton was comprised by benthic primary production.
Furthermore, and contrarily to what was
expected, there was no relation in the sediment of
the studied lakes between the seasonality in organic matter content and ETSA within a single lake.
This lack of relation may be due to the fact that
benthic community does not respond immediately
to organic matter inputs but it requires a time lag
that ultimately depends on the organic matter
nature (Newrkla, 1982; Sommaruga, 1991).
Finally, and following Relexans et al.
(1992), we have estimated the ETSA:O.M.
ratio as a useful indicator of organic carbon
quality. The significantly higher values in Lake
Honda compared to Lake Nueva corroborate
the already mentioned biodegradable and
refractory nature of the settled material of
Lake Honda and Nueva, respectively.
Chemical monitoring at the sediment-water
interface
Sediment respiratory activity
ETSA was higher in Lake Nueva than in Lake
Honda (Table 5). Seasonal variation of ETSA was
especially important for the case of Lake Nueva,
reflected by the extremely high values of SD.
Contrary to Relexans (1996), who found that
the ETSA provided a good estimate of bacterial
activity, our results have shown a weak relation
between aerobic and anaerobic bacteria and
ETSA. One likely explanation may be that, as
Trevors (1984) suggested, the number of viable
microbial cells in the sediment is much lower
than the number of total cells. Moreover, ETSA
is the result of the activity of both intra and
extracellular enzymes, being these ones able to
remain active for several days or even weeks
(Trevors, 1984; Tóth, 1992; 1994).
In the present study, among all analysed
variables that may affect the ETSA (bacterial
and algal biomass; organic matter concentration), we have recognized a large contribution of
phytobenthos, especially in Lake Nueva. The
relevance of this community for the whole lake
metabolism was already stated by Tóth (1992),
who found that a great proportion of the total
Although settling fluxes of TC and TN were
much higher in Lake Honda, its sediment
showed lower concentrations than that of Lake
Nueva (Fig. 6). These results together with the
huge nutrient concentrations (o-P, NH4+ and
inorg- and org-Cdis) measured in the interstitial water in Lake Honda again suggest a fast
nutrient recycling in this lake. By contrast, the
majority of nutrients in Lake Nueva sediment
are in particulate forms, being slowly mobilized to the dissolved pool.
In fact, we could relate the organic matter
degradability with the ratio between the decomposition products (nutrients in the pore-water)
and their sources (nutrients in the solid phase).
Hence, the much higher values recorded for orgCdis:O.M., NH4+:TN and o-P:TP in the sediment
of Lake Honda compared to Lake Nueva, support the idea of a labile organic settled matter
(planktonic) in Lake Honda and a structurally
more complex organic matter in Lake Nueva.
Likewise, the much higher value for the ratio
between inorganic and organic phosphate in the
sediment from Lake Honda with respect to Lake
Nueva once more confirms those statements.
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Figure 6. C (a), N (b) and P (c) regeneration at the sediment and water interface. Units: liquid phase: mg (C or N) l-1 and µg P l-1;
solid phase: % (C, N or P); settling rates: g (C or N) m-2 d-1 and mg P m-2 d-1. All figures are referred to annual average concentrations of TC, TN or TP, unless otherwise stated. Regeneración de C (a), N (b) y P (c) en la interfase agua-sedimento. Unidades: fase
líquida: mg (C o N) l-1 y µg P l-1; fase sólida: % (C, N o P); tasa de sediemntación: g (C or N) m-2 d-1 y mg P m-2 d-1. Todos los
datos se refieren a la concentración media anual de TC, TN o TP, al menos que se especifique lo contrario.
Regarding to the inorganic P pools (FeOOH≈P
and CaCO3≈P), figure 7 shows that the distribution of each fraction ultimately depends on
the amount of added P, depth of the water
column, pH, Ca +2 concentration in the lake
water, and FeOOH concentration in the sediment (Golterman, 1998). This diagram and the
results obtained for Lake Honda suggest that
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the much higher concentrations of Fe(OOH)≈P
compared to Lake Nueva are the result of a
much higher availability of FeOOH in the former lake (Table 6). Moreover, and although o-P
concentration in the pore-water increases as
Fe(OOH)≈P is forming, the lower Ca+2 concentration in Lake Honda delays CaCO3≈P precipitation. In Lake Nueva, by contrast, the high
values for interstitial water Ca+2 concentrations
and the lower values for oxyhydroxides in its
sediment limit the presence of Fe(OOH)≈P.
If we finally compare the TN concentration in
both study lakes, and also consider that accordingly to Martinova (1993) and Golterman
(2004) more than 90 % of TN is present as orga-
nic N, we could assume that ammonification of
organically bound N represents a key process for
losing N from the sediment of Lake Honda. By
contrast, in Lake Nueva, macrophytes patches
and the presence of N2-fixing Cyanobacteria may
account for an additional N input to the sediment.
SYNTHESIS
Strong differences between the two studied
systems have been revealed. In Lake Honda,
high in-lake nutrient concentrations, which ultimately support a large algal biomass, are basically maintained by: i. Sediment resuspension
Figure 7. Concentrations of o-P, FeOOH≈P, CaCO3≈P (mmol m-2) in the successive sedimentary layers (lr) in a lake (2 m depth),
as a function of the cumulative P load (g m-2). pH = 8; Ca+2 = 40 mg l-1 (Golterman, 2004). Concentraciones de o-P, FeOOH≈P,
CaCO3≈P (mmol m-2) en los sucesivos estratos del sedimento (lr) en un lago de 2 m de profundidad, en función de la carga de P
acumulada (g m-2). pH = 8; Ca+2 = 40 mg l-1 (Golterman, 2004).
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that is favoured by its morphometry, hydrologic
regime and sediment granulometry. ii. Intense
organic matter mineralization due to the labile
nature of the organic settled matter (planktonic).
In Lake Nueva, physical constrains (i.e. windinduced resuspension) have a limited effect due
to the coarse surface sediment and to the development of macrophyte patches (Najas marina,
Potamogeton pectinatus). In addition, the structurally more complex organic matter of its sediment regulates the low nutrients turnover. In this
lake, nutrient exchange rates across the sediment-water interface are also controlled by the
chemical equilibrium between the solid and the
liquid phase, such as P adsorption onto CaCO3,
a mechanism that is favoured by the high Ca+2
concentration in the pore-water.
In conclusion, physical, chemical, and biological mechanisms govern the fast nutrients’
benthic regeneration in Lake Honda, while a
large pool of nutrients is buried into the sediment of Lake Nueva.
ACKNOWLEDGEMENTS
We thank Henning Jensen for his valuable suggestions and helpful discussions. This research
was supported by the projects CICYT HID 990836 and UE-LIFE B4-3200/98/458. I. de Vicente was supported by a Postdoc grant from the
Spanish Ministry of Science and Technology.
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Limnetica, 25(1-2): 271-286 (2006)
Comparison of resource and consumer dynamics in Atlantic
and Mediterranean streams
Isabel Pardo1 & Maruxa Álvarez
Área de Ecología, Universidad de Vigo, 36310 Vigo, Spain
1corresponding author: [email protected]
ABSTRACT
This study integrates a wide range of ecological data into a comprehensive framework describing the functional and structural
dynamics of streams located under different climatic conditions in Spain. Standing stocks of particulate organic matter, algal
biomass and macroinvertebrate communities are examined and compared between Atlantic streams of North-West Spain and
Mediterranean temporary streams of the island of Majorca. This data set, which link potential food sources to consumer dynamics, allows explanation for the structural and functional variability exhibited by these two fluvial systems. Results indicate
that the differences in the dynamics of community trophic structure of the two types of streams are a reflection of their hydrology and climate characteristics.
Keywords: climate, disturbances, trophic structure, Mediterranean streams, Atlantic streams.
RESUMEN
En este estudio se proporciona un marco conceptual que contrasta la dinámica estructural y funcional de ríos localizados en
distintas regiones climáticas existentes en España. Con este propósito se examina un amplio rango de datos de materia orgánica particulada bentónica, de biomasa algal y de comunidades de invertebrados en ríos atlánticos del Noroeste de la Península
y en ríos temporales Mediterráneos localizados en la isla de Mallorca. Esta serie de datos ecológicos, que relacionan las fuentes de alimentación potenciales con sus consumidores, suministra una base que permite diferenciar no sólo la estructura, si no
también el funcionamiento fluvial de ambos sistemas. Los resultados indican que las variaciones en la importancia y dinámica
de la estructura trófica en ambos tipos de ríos son un reflejo de sus características hidrológicas y climáticas.
Palabras clave: clima, perturbaciones, estructura trófica, ríos Mediterráneos, ríos Atlánticos.
INTRODUCTION
Over the last few decades a number of studies
have reviewed the structural and functional differences between temperate and Mediterranean
aquatic systems (Gasith & Resh, 1999; ÁlvarezCovelas et al., 2005). These studies point out
various gaps in the scientific understanding of
Mediterranean systems, mainly in relation to the
ecosystem approach. Seasonal patterns of rainfall
and temperature across climatic regions have
important implications for water availability and
landscape structure in fluvial ecosystems. It is
well-known that flow dynamics in Atlantic and
Mediterranean regions differ strongly, both in
quantity and seasonal variation. As a consequen-
ce, the two types of stream systems are likely to
be affected by different types of disturbances.
While flooding is the main physical disturbance
occurring in Atlantic areas, both floods and dry
periods control Mediterranean streams. Indeed,
while most Atlantic streams are permanent,
depending on the availability of water and geographical setting, Mediterranean streams are subject to a permanence gradient, from permanent to
temporary flow (Álvarez-Covelas et al., 2005).
Moreover, Mediterranean temporary streams are
subject to a predictable disturbance, drying out
most of the summer, thus becoming terrestrial
systems, disrupting the biota processes typically
driven by water flow (Lake, 2000). When
water flow resumes in early autumn, high levels
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Pardo & Álvarez
of light and temperature and high nutrient concentration initiates algae colonization, thus promoting a rapid increase in biomass (Dieterich &
Anderson, 1998; Álvarez, 2004).
Climate-related parameters also influence the
composition, density and phenology of terrestrial vegetation. Differences in riparian forest
affect the dynamics of instream allochtonous
and autochthonous resources, with the biota
being adapted to availability and variability
(Cummins et al., 1989). Mediterranean streams
have different types of riparian vegetation
than temperate Atlantic streams. Mediterranean
temporary streams in mountainous areas lack
stream-dependent riparian vegetation, thus the
stream corridor is primarily composed of terrestrial species with a low interaction with the wet
stream channel. Typical Mediterranean riparian
vegetation is dominated by evergreen plants,
and the ecological significance of leaf longevity has been the subject of many hypotheses
(Terradas, 2001), suggesting an adaptation to
dry climatic conditions and drought (Cherbuy et
al., 2001), scarcity of nutrient resources (Monk,
1966), and resistance to herbivores. In contrast,
the riparian corridor of Atlantic temperate streams is dominated by deciduous species with
pulses of litterfall coupled with the vegetative
seasonal activity. Moreover, ensuring a continuous nutrient supply, some riparian species,
e.g. Alder, acquire atmospheric nitrogen from
symbiotic relationship with nitrogen fixing
organisms in root nodules (Quispel et al., 1993).
Algal communities also differ in Atlantic and
Mediterranean streams, reflecting the differences in number of sunny hours in the year, light,
temperature and nutrients dynamics that characterize the two stream types. In temporary streams, there is a seasonal sequence of biotic and
abiotic regulation of stream assemblages in response to the seasonal events of floods and dry
periods (Gasith & Resh, 1999). Undisturbed
temporary streams show a predictable pattern of
algae growth in response to the annual variation
of the magnitude of temperature and light.
However, their importance may be limited by
the consumers (Biggs, 1996). In Atlantic temperate streams, dense shading from riparian trees
prevents light from reaching the stream bed
during warmer seasons, even though springsummer climatic conditions favour algae growth
(Hynes, 1970). Therefore, the shorter growing
season due to limited solar energy of temperate
Atlantic streams compared to similar latitude
Mediterranean streams may affect the variation
in resource availability. Moreover, this seasonal
variation in resources may reflect seasonal differences of macroinvertebrate assemblages,
both in structure and function.
Study rationale and objectives
This study provides data from various streams
under different climatic conditions (Atlantic vs
Mediterranean) with different flow regimes
(permanent vs. temporary), ranging from first
order to second order streams. The study was
designed to present a useful set of data for a
general framework. Therefore, not all factors
that might explain the natural variation in
energy sources and macroinvertebrates in these
systems are tested. Data are provided for
macroinvertebrates, periphyton biomass, allochthonous inputs and benthic organic matter storage and retention. These data were recorded on
different occasions and in different streams over
the last two decades. The data available for these
streams are integrated here in an attempt to
explain the differences in underlying mechanisms and processes that characterize Atlantic
and Mediterranean streams.
METHODOLOGICAL REVIEW
Study streams
Atlantic streams: Mera, Deva, Cea and Louro
The Mera, Deva and Cea are 2nd order streams
(sensu Strahler), located in Northwestern Spain.
The Mera is subject to continental influence,
located at 545 m a.s.l. (UTM 606000, 475805);
The Deva at 624 m a.s.l. (UTM 558844, 467797)
and the Cea at 210 m a.s.l. (UTM 555018,
466396) (Figure 1). Additional information on
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Dynamics of Atlantic and Mediterranean stream ecosystems
the sites can be found in Cillero (2001). The
Louro localities, LC1 at 320 m a.s.sl. (UTM
535648 467929) and LC2 at 120 m a.s.l. (UTM
533601 467749), are 1st and 2nd order streams
respectively, located in the proximity of the
Galician coast in NW Spain. Additional information on the sites may be found in Pardo (1992).
Mediterranean streams
Most of the data provided for Majorcan streams
were obtained from the torrent Gorg Blau (GB),
a hardwater, spring-fed stream in the Sant Jordi
catchment in the mountainous Northern area of
the island (UTM 499564, 441616). The lower
spring reach, which was the study site, is located around 500 m below the springhead and
is generally dry during summer (see further
details in Álvarez & Pardo (2005), Álvarez,
2004). The study was conducted in the northern
mountainous area of the island (Fig. 1). Some
data were also recorded from other sites located
on the main axis of the torrent Ternelles (T4 and
T5), of which the torrent Gorg Blau is the main
tributary. These three sites are located ~1 km
apart (Álvarez, 2004) (Fig. 1).
Climatic data
Seasonal and annual patterns of precipitation and
temperature of the streams have been registered
at recording stations of the national Spanish
Institute of Metereology. For the Atlantic streams, data were recorded at three stations located
in Galicia, one serving as a reference for the continental stream Mera (ID = 1518, Lugo, colegio
Fingoy), and two other for the coastal streams
Cea, Deva (ID = 1723, Ponteareas) and Louro
(ID = 1495, Vigo, Peinador). For the Mediterranean streams, data were recorded on Majorca in
the Can Serra gauging station (ID = B745) and
Port de Pollensa (ID= B780).
Physico-Chemical measurements
Current velocity, water depth, and wet channel
width were recorded for the studied streams.
Current velocity was estimated from three inte-
273
grated recordings taken for one minute on the
stream bottom with a velocity-meter (Global
Water D-2466, England). Cover percentage of
substrate type was determined in an area enclosed
by a Surber sampler frame (Álvarez, 2004), or
estimated along 20 perpendicular linear transects
(Cillero, 2001). Substrate categories were classified by particle size according to the Wentworth
Scale (after Cummins, 1962; Minshall, 1984).
Allochthonous Inputs and related processes
Allochthonous input estimates are only available for the continental Atlantic stream (Mera).
Vertical litterfal was determined from 5 traps
randomly placed over the stream. Each trap consisted of a 1 m2 frame with a funnel shaped net
of 1 mm–mesh, tied with a rubber band to allow
collection of samples. The traps were placed
randomly, suspended ~ 1 m above the river surface from surrounding trees (see Cillero et al.,
1999). All samples from each trap were kept
separately. After collection, each leaf litter was
dried (80 ºC, 20 h), sorted and weighed to the
nearest 0.0001 g. Samples were collected
monthly from June 1998 to June 1999.
Benthic particulate organic matter standing
stock (BPOM) was obtained for 3 Atlantic streams (Mera, Cea, Deva) and for the Torrent Gorg
Blau. After removing all organisms, organic
detritus was separated, dried at 60 °C for 48 h
and weighed to estimate dry weight of BPOM.
Samples were taken on two occasions (December and June) within a year period.
Retention characteristics were measured for 3
Atlantic streams (Mera, Cea, Deva). Plastic
strips were used as set of artificial leaves to
measure retention capacity (following Webster
et al., 1994). At each site, 100 plastic strips
(4 x 8.5 cm) were released across the width of
the channel at the upstream end of the studied
reach. Thirty minutes after release, the entire
reach was examined for strips retained, and the
number of strips and distance travelled were
recorded. Three releases were made on each
sampling occasion. Leaf retention was adjusted
to an exponential model, following Lamberti &
Gregory (1996). This procedure allowed estima-
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Pardo & Álvarez
Figure 1. Location of the study streams. Atlantic streams are located in Galicia (NW Spain) and the Mediterranean streams in the
island of Majorca. Localización de los ríos estudiados. Los ríos atlánticos se localizan en Galicia (NO España) y los ríos mediterráneos en la isla de Mallorca.
tion of the instantaneous retention rate and average distance travelled by a specific particle in
each stream. Samples were collected monthly
from June 1998 to June 1999.
Periphyton organic matter
Data on periphyton biomass are available for 3
Atlantic (Cea, Deva and Mera) and three
Mediterranean (GB, T4, T5) streams. Periphyton
samples (examined by analysis of chl a) were
taken from individual rocks collected from each
site. From each rock, an area of 8.05 cm2 was
scraped with a chisel and a nylon toothbrush. Chl
a was extracted in 90 % acetone in darkness and
at low temperature and measured spectrophotometrically (Unicam UV/VIS UV4 Spectrometer,
Cambridge, UK). Values were determined using
the equations described in Lorenzen (1967).
Macroinvertebrate sampling
Data on annual patterns of macroinvertebrates
are available for 2 Atlantic (LC1, LC2) and one
Mediterranean (GB) stream. Quantitative samples were collected with a Surber sampler
(0.09 m2) of 500 µm in the Atlantic streams and
of 100 µm mesh size in the Mediterranean stream. Small invertebrates (ostracoda and copepo-
da), were not considered in the analysis.
Samples were preserved in the field in 10 % formaldehyde and sorted in the laboratory. All
invertebrates were picked from the fractions
> 1 mm. However, when the volume of sample
was large, smaller fractions were sub-sampled.
Macroinvertebrates were identified to genus,
family or species level under a dissecting
microscope. Each taxa was assigned to a functional feeding group (FFG) based on mouthpart
morphology and feeding behavior following
Cummins (1973) and supported by examination
of the gut contents of the dominant taxa.
STREAMS CHARACTERIZATION
Geomorphic setting: geography, geology
and climate
The Atlantic streams are located at medium
elevations in North-West Spain (Galicia)
(Fig. 1). They are relatively small (annual mean
across streams ± 1SE = 2.1 ± 0.6 m wide,
24.4 ± 4.5 cm deep) and are characterized by a
typical sequence of riffles and pools. The substrate is primarily composed of variously sized
cobbles and pebbles placed on top of gravel
and coarse sand (Pardo, 1992; Cillero 2001)
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Table 1. Selected physico-chemistry parameters at the study streams. Values represent the mean for an annual flow period (1986-87 in the
Atlantic stream Louro; 1998-1999 in the Atlantic streams Mera, Cea and Deva; 2000-2001 in the Mediterranean streams). Parámetros físicoquímicos de los ríos estudiados. Los valores representan los valores medios para un periodo hidrológico anual (1986-87 en el río atlántico
Louro; 1998-1999 en los ríos atlánticos Mera, Cea y Deva; 2000-2001 en los ríos mediterráneos).
Temperature
ATLANTIC
Conductivity Oxygen Flow Water Channel Boulder Stone- Gravel Sand-silt
depth width
Pebble
(µS cm-1) (mg L-1) (L s-1) (cm)
(cm)
(%)
(%)
(%)
(%)
STREAM
(ºC)
MERA
CEA
DEVA
13.4
14.7
13.3
6.2
6.0
5.7
41.0
24.3
18.9
9.6
9.9
9.6
113.8
127.5
310.0
13.8
6.0
28.1
9.7
16.9
18.3
13.8
7.4
7.6
7.9
763.2
1104.9
709.4
16.3
7.6
859.2
mean
MEDITERRANEAN
pH
GB
T4
T5
mean
(Table 1). The parent rock is siliceous granite,
reflected in waters of a slightly acidic character
(pH mean of streams annual means of 6.1)
and low salt content (electric conductivity
mean of streams annual means of 42.0 µS cm-1)
(Table 1) (see Pardo, 1995) for further details
on the chemistry dynamics of the Louro stream
sites). The climate is Atlantic temperate,
characterized by mild to cool winters and warm
summers, although the temperature and precipitation are influenced by a coastal –inland gradient of continentality (Table 2). Mean
air temperature for a 24-year period was
25.0
190.8
_
_
214.6
21.5
53.5
25.1
29.2
21.8
43.2
14.4
7.9
5.9
34.8
16.9
25.8
36.8
183.8
30.9
202.7
33.4
31.4
9.4
25.8
9.2
9.6
10.2
8.9
3.3
13.3
10.8
5.9
7.4
87.7
132.9
202.6
12.8
49.5
70.6
56.4
29.9
19.0
19.7
16.4
9.4
11.1
4.1
1.0
9.7
8.5
8.0
141.1
44.3
35.1
15.1
5.4
13.7 ± 1 ºC (1975-98) (mean monthly temperature average for each year across streams ±
1SE) (Table 2). Mean annual precipitation for
the same period was 1449.7 mm (± 265.0)
(Table 2). Generally, precipitation in Atlantic
streams is higher in autumn and winter months
reducing progressively through spring and
summer (Fig. 2). However, these streams receive relatively constant year round precipitation,
which, for the mentioned period, was reflected
in the high average number of days with rain
(mean across streams ± 1SE = 152 ± 4 days)
and the low percentage contribution of monthly
Figure 2. Climate diagrams for the coastal Atlantic streams (Ponteareas) and for the Mediterranean streams (Can Serra, Pollensa).
Graph scales follow Molles (1999). Diagramas climáticos de los ríos costeros atlánticos (Puenteareas) y de los ríos Mediterráneos
(Can Serra, Pollensa). Las escalas de los gráficos siguen a Molles (1999).
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Table 2. Climate characteristics of the study streams. See text for further details. Características climáticas de los ríos estudiados. Ver el texto
para más detalle.
Continental
Lugo
Precipitation
Mean annual value
SE
Inter-anual variation (CV)
SE
Range (inter-annual)
Maximun monthly contribution
SE
% of days with rain per year
Mean number of days with rainper year
SE
Temperature
Annual mean
SE
Inter-anual variation (CV)
Atlantic
Coastal
Ponteareas
Mediterranean
GB
Louro
991.5
40.3
0.78
0.05
0.45 - 1.33
4.9 %
0.3 %
41.2 %
150
4
1448.1
63.5
0.89
0.04
0.48-1.34
5.1 %
0.3 %
39.9 %
146
8
1909.5
63.5
0.85
0.04
0.50-1.41
4.5 %
0.2 %
43.5 %
159
3
895.7
54.9
0.92
0.04
0.59 -1.34
12.0 %
0.7 %
15.7 %
57
3
12.6
0.2
0.4
14.8
0.2
0.3
13.8
0.1
0.3
17.0
0.1
0.3
precipitation to the annual totals (mean across
streams ± 1SE = 4.9 ± 0.2 %) (Table 2).
Majorca is the largest (3640.16 km2) of the
Balearic Islands, located at a distance of 167 km
off the East coast of the Iberian Peninsula. The
studied Mediterranean stream reaches are located in mountain areas, at relatively low altitude
(125-150 m). Streams on the island of Majorca
are not found at higher elevation than around
300 m a.s.l. The studied Mediterranean streams
are small (annual mean across stream reaches ±
1SE = 1.4 ± 0.1 m wide, 8.0 ± 0.5 cm deep)
(Table 1). Percentages of dominant substrate differed among streams. Substrate composition
change among stream reaches, but mainly includes stones, pebbles and boulders (Table 1).
However, in some streams (see for example T5),
the stream bed is highly affected by calcite precipitation, which is reflected in an almost continuous layer of tightly consolidated substrate
(i.e., bedrock) (Álvarez, 2004). The catchment
geology is rather complex and largely consists of
sandstone, limestone and dolomite (GarcíaAvilés, 1990). As expected, the geology is reflected in the chemistry of the waters, which are
high in pH (annual mean across stream reaches ±
1SE =7.6 ± 0.1) and high in electric conductivity
(859.2± 47.4 µS cm-1) (Table 1). Moreover, these
mountain streams are fed by karstic aquifers
with high infiltration capacity, which strongly
affects the hydrology of these streams.
The island of Majorca has a typical Mediterranean climate with warm to hot temperatures
all year round (mean temperature for 1962-1991,
was 17.0 ± 0.1 ºC) and low annual rainfall (for
1975-1998, was 895.7 ± 54.9 mm) (Table 2).
Moreover, most of the precipitation over Majorca
falls as torrential rain, which, for the mentioned
24-year period, is reflected in the low average
number of days with rain (57 ± 3 days), the high
percentage contribution of monthly precipitation
to the annual total (12.0 ± 0.7 %) and in the relatively high average CV (0.92 ± 0.04) (Table 2).
Hydrology and water temperature
The high precipitation levels and high runoff
that characterize the Atlantic catchments, generates a landscape drained by hundreds of small,
fast flowing streams of a permanent character.
Annual flow patterns in Atlantic streams
usually show maximum peaks in autumn and
winter (with frequent floods), progressive
reduction towards spring and summer, and
minimum discharges occurring in September
(Fig. 3) (Pardo, 1995; 2000).
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for permanent streams (e.g., Dieterich &
Anderson, 1998). In our comparison, the wider
water temperature range corresponded to the
temporary stream, fact that can be clearly attributed to the Mediterranean climate influence.
RIPARIAN INFLUENCES ON THE
STREAM
Figure 3. Discharge for the Atlantic Mera stream and the
Mediterranean torrent Gorg Blau for an hydrologic year, from
October to July/August (O = October, S = September, N =
November, D = December, J = January, F = February, M =
March, A = April, M = May, Jn = June, Jl = July, Ag =
August). Caudales del atlántico río Mera y del torrente mediterráneo Gorg Blau durante un año hidrológico, desde octubre
a Julio/agosto (O = octubre, S = septiembre, N = noviembre,
D = diciembre, J = enero, F = febrero, M = marzo, A = abril,
M = mayo Jn = junio, Jl = julio, Ag = agosto).
The Mediterranean stream, showed a delayed
response to rainfall patterns compared to the
Atlantic streams (Fig. 2 & Fig. 3), which may be
attributed to the hydrogeologic features of the
karstic aquifers (Del Rosario & Resh, 2000).
Moreover, as occurs in many Mediterranean
environments (Vidal-Albarca et al., 1992;
Cattaneo et al., 1995), the negative hydrologic
balance during the warm dry summer interrupts
the stream flow (Fig. 3). For example, from
1999 to 2000 (considered as an average hydrological year), the flow period of the Gorg Blau
lasted for about 10 months, starting with high
discharges in autumn (end of October) and
drying out completely in late-July (Fig. 3).
As expected, discharge in the streams varied
over time and space by several orders of magnitude during the years of study (e.g., ranged from
0.4 to 31.8 L s–1 in the torrent Gorg Blau and
from 57 to 659.1 L s–1 in the Mera river) (Table
2, Fig. 3). Water temperature was also variable
among streams and time (e.g., ranged from 7.4
to 17.9 ºC in the Mera river and from 12.6 to
29.6 ºC in the torrent Gorg Blau) (Table 2).
Other studies which have compared temporary
vs permanent streams in geographically close
areas, provided greater water temperature range
In Atlantic streams, riparian vegetation is
usually dominated by Alnus glutinosa L., Betula
alba L., and Fraxinus excelsior L., mixed with
oak species (Quercus robur L., Q. pyrenaica
Willd.). However, Alder is the dominant tree
species in small, low-gradient streams. The riparian canopy in these streams is well-developed
along the channel, and usually shade the stream
completely. Alder asymmetric distribution of
tree branches creatings a greater production of
leaves towards the stream (Cillero et al., 1999).
The riparian vegetation of Majorcan streams
is composed of the typical Mediterranean
terrestrial vegetation of holm-oak (Quercus ilex
L.) and Aleppo pine (Pinus halepensis Miller),
the latter dominating on southern slopes. These
streams usually have wide active channels.
Despite larger variations in flow over the years,
the flow is mainly concentrated in the center of
the channel. This reduction in the wet area
towards the center of the channel allows the
access of direct light to the stream bottom.
Litterfall
Litter inputs to the Atlantic Mera stream were
730.0 g·m-2·yr-1, and the vegetative-seasonal
cycle of various species gives rise to a distinct
litterfall seasonal pattern of leaves and fruits
(Cillero et al. 1999), which was apparent in
these riparian forests. Leaf shed of the dominant
riparian alder showed a bimodal pattern with
maximum values in July and November (Cillero
et al., 1999). However, leaf shed from chestnut
and oak was unimodal, beginning in October,
and reaching a maximum for chestnut in
November, and for oak in December. The input
of reproductive structures was highest in
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February-March due to higher amounts of alder
flowers, while no clear seasonal patterns were
observed for wood and debris fall.
Although data are not available for the studied
Mediterranean streams, studies conducted in
other Mediterranean streams (e.g., Cherbuy et
al., 2001; Bussotti et al., 2003; Bellot et al. 1992)
show that evergreen holm-oak (Q. ilex) shed leaves throughout the year, with recorded maximum
values between April and July. This more substantial inputs occur at a time of increasing temperatures and progressively reduced water levels in
the Majorcan streams. Moreover, other studies
conducted in central Italy (Bussotti et al., 2003)
provided annual litter inputs values for holm-oak
in xeric and mesic forests of 485 and 694 g m–2
respectively, which may be comparable to the litter input to the studied Majorcan streams.
Benthic organic matter retention and storage
Small Atlantic streams are generally very retentive due to a diversity of in-stream structures such
as tree roots and dominance of coarse substrates
such as stones and blocks. In the smallest channel
of the Cea (2 m wide) and under low discharge
ranges (< 0.31 m3 s–1), released plastic strips were
retained within the first 7.1 m (1/k). However, in
wider channels of the Mera and Deva (around 3 m
wide) and under higher discharges (up to
0.72 m3 s–1), the mean distance travelled by the
strips increased to 16.5 m (Cillero, 2001).
Figure 4. Benthic organic matter (BOM) for the three Atlantic
streams and for the Mediterranean stream (GB). The graph
shows mean values (± 1SE). D = December samples; Jn= June
samples. Materia orgánica bentónica (MOB) para los 3 ríos
atlánticos y el río mediterráneo (GB). Los gráficos muestras
los valores medios (± 1EE). D = muestras de diciembre; Jn=
muestras de junio.
BPOM estimated in the 3 continental Atlantic
streams ranged from a mean of 5.2 (in the Deva
in June) to 52 g m–2 (in the Mera in December)
(Cillero, 2001), representing between 0.7 and
7 % of total annual inputs to the streams. In the
Mera stream, BPOM follows the predicted pattern of higher standing stocks in autumn
(December) (Fig. 4) before leaves have been processed completely. However, other streams
(Fig. 4) (i.e., Cea and Deva) show only a slight
difference between autumn (December) and summer (June). In Atlantic streams, spring-summer
leaf processing of different riparian and terrestrial
species occurs in less than 2 months. In these
streams, decay rate is mainly explained by greater
nutritional value and softer consistency of green
leaves entering the stream, as well as to active
processing by shredders. Autumn-winter processing is also similarly fast, influenced by higher
frequency and intensity of flow disturbance and
by physical fragmentation of leaf packs (López et
al., 1997; López et al., 2001).
In the torrent Gorg Blau, mean particulate
organic matter standing stock over the studied
flow period was 157.7 ± 33.9 g AFDM m –2
(Álvarez, 2004). On a seasonal basis, this temporary stream had from 7.3 (December) to 8.5
(June) times more BPOM than the Atlantic streams (Fig. 4). Several factors may explain these
differences. Moderate flow-related physical forces, in combination with channel characteristics
may contribute to higher retention of BPOM in
the Mediterranean streams. Therefore, in spite
of their shallowness and the near absence of
retentive structures in the channel, temporary
streams are very retentive during stable flow
conditions. Only major floods, which mainly
occur in autumn at the beginning of the flow
period seem to be able to mobilise litter packs
accumulated on the dry streambed during summer. Moreover, invertebrate leaf consumption
that in turn is influenced by litter quality may
influence stream organic matter standing stocks.
Although research on leaf decomposition has
not been conducted in Majorcan streams, other
studies have observed that the leathery and lownutrient leaves of the evergreen holm-oak
(Q. ilex) decompose slower than those of other
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species (Fano et al. 1988; Gessner & Chauvet,
1994; Schwarz & Schwoerbel, 1997). Additionally, other experimental studies on leaf
decomposition in temporary streams attribute
the lower breakdown of leaves to the lower
microbial activity occurring in these systems
due to the time lag required for the development
of the microflora after submersion (e.g. Maamri
et al., 1997). Rapp & Leonardi (1988) found
that under Mediterranean climatic conditions,
Holm-oak leaves entering the stream in summer
required a minimum conditioning time of 5
months, and the leaves completed their decomposition only during the following flow season,
thus also favouring BPOM accumulation in the
stream bed. This should explain the stable
values of BPOM found in the torrent Gorg Blau
over the studied flow period (Álvarez, 2004).
PRIMARY PRODUCERS: PERIPHYTON
BIOMASS
Overall, the Mediterranean temporary streams
had 10 times more periphyton biomass (mean chl
a over three streams ± 1SE = 49.6 ± 28.28 mg
cm-1) than the Atlantic streams (mean over three
streams ± 1SE = 4.6 ± 28.8 mg cm-1) (Fig. 5).
Similar to other forested temperate streams
around the world, values of chl a in the Atlantic
streams remained relatively constant over time
(Rosemond, 1994). However, chl a reached
higher values in spring (e.g., 5.9 mg m-2 in May
in Mera stream) and lower values in winter (e.g.,
1.5 mg m-2 in January in Mera stream), coinciding with the less favourable season for
periphyton growth in these streams (Bott et al.,
1978; Rounick & Gregory, 1981) (Fig. 5).
Moreover, higher values of benthic biomass in
some of the Atlantic streams (Deva) was inversely related to flow, with higher values corresponding to the Cea and Mera streams with lower
annual discharge amplitudes (Cillero, 2001).
These results are similar to results reported by
other studies (Sabater et al., 1998; Elosegui &
Pozo, 1998; Lohman, 1992).
In Mediterranean streams, maximum values
of periphyton chl a were reached at the begin-
Table 3. Components of taxonomic richness in two Atlantic streams
(LC1 &LC2) and one Mediterranean stream (GB). Componentes de
la riqueza taxonómica en dos ríos atlánticos (LC1 y LC2) y en un
río mediterráneo (GB).
Taxonomic Level
Species
Genus
Family
order/class
LC2
Total Richness
Annual average
± 1SE
Total Richness
Total Richness
Total Richness
108
50
3
80
57
16
LC1 GB
76
28
2
61
42
18
41
31
1
40
34
14
ning of the flow period (e.g., 43.4 mg m-2 in
Octuber in the torrent Gorg Blau) and subsequently declined towards the end of the flow
period, which has been attributed to grazing
(Álvarez, 2004) (Fig. 5). Depending on the
strength of grazers-periphyton interactions,
Mediterranean streams showed a gradient of
periphyton availability. Streams with highest
values of chl a correspond to those with lowest
grazer density (e.g., T5), and values of chl a
were relatively low in streams highly affected by
grazing (e.g., GB) (Fig. 5).
SECONDARY PRODUCERS:
MACROINVERTEBRATES
Taxonomic composition
Patterns in macroinvertebrate density and community composition, examined using nonmetric multidimensional scaling (MDS) on the
Bray Curtis similarity matrix of macroinvertebrate densities (at family level), showed a clear
separation in the ordination diagram of permanent Atlantic (2 sites in the Louro stream, LC1
and LC2) and temporary Mediterranean streams
(one site in the Gorg Blau) (Fig. 6). Macroinvertebrate density was thirty times higher in
Gorg Blau (mean annual ± 1SE = 25803 ind m–2
± 4257) than in LC1 (mean ± 1SE= 850 ind m–2
± 236), the two sites most clearly differentiated
by their taxa composition (Fig. 6, Table 3).
Invertebrate families responsible for >90 % of
dissimilarity between the Gorg Blau (GB) and
the two Atlantic stream sites together (LC1,
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Pardo & Álvarez
Figure 5. Dynamics of periphyton chl a in three Atlantic (5A) and three Mediterranean (5B) streams for an hydrological year, from
September/October to July/August. The graph shows mean values +1SE.(O = October, S= September, D = December, J =January,
A = April, M = May, Jn = June, Jl = July, Ag = August). Dinámica de la Cl a del perifiton en los tres ríos atlánticos (5A) y en los
tres mediterráneos (5B) para un año hidrológico, de septiembre/octubre a julio/agosto. Los gráficos muestras los valores medios
+1EE. (O = octubre, S= septiembre, D = Diciembre, J =enero, A = abril, M = mayo, Jn = junio, Jl = julio, Ag = agosto).
LC2), from a total of 70 families recorded
(SIMPER analysis, programme PRIMER),
were: Hydrobiidae, Caenidae, Ancylidae,
Physidae, Glossosomatidae and Gammaridae
(more abundant in GB), and Nemouridae and
Psychomyiidae (more abundant in the Atlantic
streams). These eight families contributed
32.6 % to the dissimilarly between the stream
communities (total dissimilarity = 68.73 %).
In general, permanent streams are inhabited
by a richer fauna than temporary streams (Table
3). Taxa richness of macroinvertebrates, with
similar identification level for insects, were
compared in two permanent headwater Atlantic
streams (LC2 and LC1) and the torrent Gorg
Blau for a hydrological year. On an annual
basis, total taxa richness per sampling area was
considerably lower in GB (41) than in the other
two sites (Table 3). However, such large variation was lower when comparing taxa at the
family level table 3). These differences could
be attributed to the low species-richness within
groups that characterize the Majorcan streams,
with most genera represented by a single species (Álvarez, 2004). In addition, average
macroinvertebrate taxa richness (mostly family
level) for the torrent Gorg Blau (31) was similar to that in LC1 (28), although somewhat
lower than in LC2 (50). LC1 is a headwater
Atlantic stream comparable in size to GB,
which explains the lower taxa richness in comparison to LC2, an Atlantic second order stream
with higher habitat heterogeneity and with a
more developed riparian corridor (Pardo,
1992). Moreover, although GB and LC1 show
similar average macroinvertebrate taxa richness, the total taxa richness was higher in LC1,
indicating that the Atlantic stream has a higher
replacement of taxa through time. Therefore,
rather than attributing these differences to the
insularity of GB, they may be attributed to the
hydrologic regime of the compared streams. In
fact, Vivas (2003) found that in a temporary
spring-fed stream of Southeast Spain, total taxa
richness was between 27 and 32 (overall total
45); values comparable to the totals of 27 - 34
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(overall 41) observed in the torrent Gorg Blau.
These results suggest that spring feed, temporary streams on the island of Majorca have taxa
richness comparable to similar streams on the
mainland. However, their temporality may
explain the lower species richness than similar
small permanent headwaters streams in Atlantic
climate, which in turn may allow a longer growing season for aquatic insects and the occurrence of taxa with different species traits (i.e.,
longer life cycles) (Delucchi, 1989; Williams,
1996). Even though many studies have found a
remarkable similarity between fauna in temporary stream habitats and fauna found in nearby
permanent streams (Boulton & Lake, 1992;
Feminella, 1996). Others have noted rather distinct differences between permanent and temporary forest streams (Delucchi, 1988;
Dieterich & Anderson, 2000; Muñoz, 2003).
MDS ordination indicated that differences in
community composition between sites were
greater than temporal differences (Fig. 6). However, there was a clear intra-annual variation in
community structure in the GB, (Fig. 6), which
281
has been shown to reflect its temporality, with
distinct signs of succession after the summer
dry period (Álvarez, 2004).
Functional organization of the
macroinvertebrate communities
On an annual basis, the relative importance of all
FFG was significantly different between the three
streams; Atlantic streams (LC1 and LC2) and
torrent GB (ANOVA, p<0.001). The Mediterranean spring-fed stream consisted largely of
collector-gatherers (38 %) and scrapers (45 %)
followed by filter-feeders (9.7 %), shredders (4 %) and predators (3.4 %). However, in the
permanent Atlantic streams, mean collectorgatherers and scrapers comprised (17.3 ± 2 %)
and (26 ± 2.8 %), respectively. Filter-feeders
(15 ± 2.5 %) were better represented in Atlantic
streams, and shredders (32.3 ± 4 %) and predators (9.7 ± 1 %) were more abundant.
In the permanent Atlantic streams (LC1 and
LC2) relative abundances of FFGs showed significant temporal patterns, with most FFG
Figure 6. MDS ordination plot of monthly samples of invertebrate assemblages. Analysis were based on arcsin(冪苴
x)-transformed
abundances of the dominant taxa and a Bray-Curtis similarity matrix. Lines have been drawn around samples from the same stream
(and thus, with similar taxa compositions) as an aid to visual interpretation (see text). Ordenación MDS de las muestras mensuales
de comunidades de invertebrados bentónicos. Los análisis se basan en las abundancias de los taxones dominantes transformadas
x ), e índice de similitud de Bray-Curtis. Las líneas agrupan muestras de un mismo río (y por tanto con
mediante el arcoseno(冪苴
similar composición taxonómica) para ayudar en la interpretación visual (ver texto).
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Pardo & Álvarez
Figure 7. Dynamics of scrappers (7A) and shredders (7B) in two Atlantic coastal streams (LC1 = Δ & LC2 = G ) and in one
Mediterranean stream (GB = G). Hydrologic months are 1= October until 12= September. Dinámica de raspadores (7A) y desmenuzadores (7B) en los dos ríos atlánticos costeros (LC1 = Δ & LC2 = G ) y en un río mediterráneo (GB= G). Los meses hidrológicos son 1= octubre, hasta 12= septiembre.
dynamics fitted to quadratic and cubic distributions (p<0.05) (Fig. 7). Percentage of collectorgatherers tended to be low in autumn for LC2
and in summer for LC1, whereas filter-feeders
showed an increase in representation in springsummer months. Scrapers showed a cubic S
shaped trend with a maximum value of 53 % in
winter for LC2, while no significant trend was
observed for LC1 (Fig. 7a). Predators showed a
peak of maximum representation at the end of
the spring (in June, they made up 11 % and
22 % of total FFG abundance in LC2 and LC1
respectively). Shredders showed opposite dynamics in the headwater stream (LC1) and in the
second order stream (LC2) (Fig. 7b). In LC1,
the highest shredder representation in Atlantic
streams, showed a maximum (>50 %) of shredders between August and January, when litter
inputs and BOM tend to be maximum.
However, the bell-shaped curve described by
shredders in LC2, with maximum representation in spring, may reflect the seasonal pattern
of green alder leaves entering this reach, which
has a better developed riparian alder dominated
corridor than the upstream site (LC1), which is
accompanied by sparse vegetation of Salix spp.
adjoined by grasslands (Pardo, 1992). In the
forested reach (LC2), spring-summer consumption of mostly green leaves will impulse shredders production after the winter shortage of
resources (López et al., 2001). However, at the
end of summer and in the beginning of autumn
when rainfall levels increase, the rising water
levels contribute to the physical abrasion and
transport of leaves and individuals within the
channel (López et al., 2001).
In the Mediterranean temporary stream (GB)
the dynamics followed by the different FFG
were significatively fitted to cuadratic and
cubic distributions (p<0.05). The relative abundance of collector-gatherers followed a U-shaped curve with maxima at the beginning and
the end of the flow season, where they made up
respectively 68 % and 57 % of total FFG abundance. Moreover, scrapers, which were maintained over 40 % for most of the year, showed the
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opposite temporal dynamic, represented by a
hump-shaped curve (Fig. 7a), thus reaching
their minimum representation at the beginning
and at the end of the flow period, where they
represented respectively 10 % and 34 % of total
FFG abundance. The relative abundance of filter-feeders tend to be higher at the beginning of
the flow period, and only the relative abundance
of shredders and predators increased significantly over the flow period (Fig. 7b). The disappearance of water in the temporary stream at
the beginning of the summer, caused a sudden
mortality of shredders, which otherwise in a
wetter year should be expected to continue consuming litter during autumn-winter months
(Álvarez, 2004). Moreover, as shown in other
studies conducted in temporary streams (e.g.,
Boulton & Lake, 1992), predator densities
increase with flow reduction at the end of the
study period, when discharge is lower and
invertebrate distribution becomes more concentrated (Álvarez, 2004).
The annual hump-shaped curves described
by monthly densities of shredders in the
Atlantic forested stream and by scrapers in
the Mediterranean temporary stream and the
semi-annual distribution followed by scrapers
in the Atlantic LC2 indicate that both groups
may be seasonally limited by resource availability. In the Atlantic forested streams, the
supply of allochthonous and autochthonous
food sources is discontinuous half of the year,
thus dynamics of relative abundance of functional feeding groups in these systems may
reflect the phenology of energy sources availability. However, a recent study in the Gorg
Blau showed no relation between the relative
abundance of detritus feeding invertebrates
(collectors or shredders) or grazers and their
food supplies (Álvarez, 2004), thus food may
be in continuous supply in these temporary
systems. The main constraints on the trophic
composition of the macroinvertebrate community in GB are related with the resumption or
reduction of water flow that characterize temporary streams (Álvarez, 2004) and the associated alteration in habitat structure and availability (Lake, 2003).
283
CONCLUSION: ECOSYSTEMS
GENERALIZATIONS
1. Climate promotes the availability and dynamics of food resources in the studied stream
ecosystems:
1.1. In Atlantic temperate streams, inputs of
food resources are discontinuous, due to
a clear seasonal (summer-autumn) pattern of production. In these systems,
allochtonous inputs, in the form of nutritious leaves, dominate from June to
December, thus shading by the riparian
vegetation simultaneously influences
autochtonous production. As a consequence, autochthonous resources are
relatively low over the year, slightly
increasing in spring-summer months.
1.2. Mediterranean streams receive high
inputs of solar radiation, thus increasing
the importance of authoctonous production and plant-herbivore interactions
compared to temperate Atlantic streams
with closer canopies. A continuous
supply of allochtonous resources along
the year, of otherwise very low quality,
may initially create resource bottleneck
for detritivore shredders, thus explaining
low abundances of this FFG in the
beginning of the flow period. Therefore,
allochtonous resources may be a seasonal alternative and less important energy
pathway to consumers.
2. Temporary streams have a high retentive
capacity. They accumulate large quantities of
allochthonous material despite low input.
Regardless of high retention values in
Atlantic streams, standing stocks of organic
matter are low, even during the major input
season, pointing at the influence of biotically
mediated higher processing rates coupled
with physical processes in these streams.
3. Conditioning and processing of allochthonous materials is re-initiated once flow is
resumed in temporary streams, probably
needing longer times that the mean annual
water period, before they are fully conditioned, therefore usually constituting an alter-
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Pardo & Álvarez
native food resource. Leaf processing in
fast flowing Atlantic streams is a fast process occurring in approximately less than 2
months. Green leaves input of alder from
the beginning of summer constitutes
nutrient-rich materials to which shredders
may adapt due to benign flow conditions
during these predictable food supplies. It is
to be expected that secondary production
will be promoted during spring-summer
after winter shortage of resources.
4. High densities of macroinvertebrates in the
torrent Gorg Blau are explained by the following factors, fast rates of autochtonous production, close proximity to spring refugia,
reduced predator pressure and the low competition, coupled with fast invertebrate colonization, which occurs 1-2 month after flow resumes (Álvarez, 2004; Álvarez & Pardo, 2005).
In fact, high densities and sustained feeding
by the dominant grazers were the most important factors depleting periphyton biomass in
the Mediterranean temporary stream.
5. In Atlantic streams, low invertebrate densities relate with low resources amounts and
limited availability on a seasonal basis, probably pointing at a seasonal bottom-up control of shredders and scrapers by food
resources and prolonged adverse hydrodynamic conditions. Moreover, the presence of
large predators in these small streams (primarily trout) may indicate a top-down control during benign flow periods.
FUTURE RESEARCH
It is recommended that a main line of research
in both Atlantic and Mediterranean streams
should focus on improving the understanding
of the functioning of aquatic communities in
both systems. Based on the above findings
and observations we expect to conduct experiments on the mechanisms underlying predation, competition and resource-consumer
interactions across stream types.
The analysis of benthic energy budgets or
energy flow through food webs should provide a
more precise determination of the importance of
autotrophic and heterotrophic food resources in
streams with different riparian vegetation (e.g.,
deciduous vs. ever green) and flow permanence
(temporary vs. permanent).
Temporary streams are also good environments for studying patterns of colonization and
community succession, as well as the function
and adaptation mechanisms of their fauna.
Within this line, studies will address the mechanisms allowing maintenance of temporary stream populations when facing disturbances
(reproduction patterns of oviposition behaviour,
egg hatching, diapause, mortality and emergence), as well as the determination of the importance of the hyporheic zone as refugia.
ACKNOWLEDGEMENTS
We would like to thank the editors Joan
Armengol for his kind invitation to participate in
this special volume dedicated to the Dr.
Margalef. The data base analysed in this article
reflects the help and support of many people and
institutions. Some people were technicians who
helped with sample collection and analysis (Mar,
specially); others were past fellows in the lab
who conducted data compilation and analysis and
supported some methods (Eva, Núria, Carmen),
and some others were instrumental in dealing
with language issues and presentation (thanks
Asger!). Several funds contributed to the realization of the works involved, FEDER (1DF971481-C02-02), (XUGA29106A96) from Xunta
of Galicia, and a research grant from the
University of Vigo. The work conducted on
Majorca was partially supported by a grant from
the Spanish Education Council (HID98-0323C05-02) as part of the project GUADALMED.
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