Plant Succession Following Nuèes Ardentes of Mt. Merapi Volcano

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Plant Succession Following Nuèes Ardentes of Mt. Merapi Volcano
Plant Succession Following Nuèes Ardentes of Mt. Merapi
Volcano, Java Indonesia
Sutomo
Sarjana in Forestry Science (B.Sc. Hons.)
Gadjah Mada University, Indonesia
This Thesis is presented for the Degree of Master of Science
University of Western Australia
SCHOOL OF PLANT BIOLOGY
FACULTY OF NATURAL AND AGRICULTURAL SCIENCES
2010
STATEMENT OF ORIGINAL CONTRIBUTION
The research presented in this thesis is an original contribution to the field of plant ecology.
The ideas presented and discussed here unless otherwise referenced or acknowledged, are
my own original thoughts. This thesis has been completed throughout my enrolment at the
University of Western Australia and has not been used previously for a degree at any other
institution.
Sutomo
ii
ABSTRACT
Tropical volcanic ecosystems, particularly in Asia, have been relatively
understudied. Mt. Merapi volcano, in Java Indonesia provides an opportunity to study
succession following nuées ardentes events which essentially reset the successional clock
of the montane forest. Information regarding the development of the vegetation on Mt.
Merapi during early succession is scarce. Using a chronosequence approach, my study
documented the patterns of vegetation succession following nuées ardentes and wildfires in
tropical montane forests on Mt. Merapi. I examined primary succession in areas directly
affected by nuées ardentes and secondary succession in areas affected by fire associated
with nuées ardentes.
There was a rapid colonization by vascular plants in both primary and secondary
succession as the sites aged.
Imperata cylindrica, Eupatorium riparium, Anaphalis
javanica, Athyrium macrocarpum, Brachiaria paspaloides, Dichantium caricosum,
Selaginella doederleinii, Eleusine indica, Cyperus flavidus, Calliandra callothyrsus and
Acacia decurrens were the species mainly responsible in explaining the differences
between sites. In primary succession, the species richness and diversity reach their peak 14
years after disturbance. In secondary succession, the species richness was similar to the
reference site in only a little after two years - however, the peak of species diversity was 14
years after the nuées ardentes disturbance.
Native and exotic invasive species varied in abundance among sites. The native
invasive I. cylindrica dominated the early succession but then disappeared under the shade
of the emerging tree species. In contrast, invasive exotic species such as Eupatorium spp
and Brachiaria spp remained in the system long after the nuées ardentes had occurred and
forest structure had developed.
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In the secondary succession the forest structure developed throughout the
succession with the older sites, regaining nearly complete stratified forest vegetation after
14 years. In terms of species interspecific associations, positive associations were greater
than negative associations as time progressed in the primary succession. The nitrogenfixing species Calliandra callothyrsus had the highest number of positive interspecific
associations compared to the other species, which may suggest that this legume species has
a prominent role in facilitation.
An examination of soil nutrient status showed improvement in soil condition as
time progressed. There was a clear pattern of increase in N, C, and exchangeable cation
concentration with age of site, whereas the P concentration decreased with time. There was
also a significant relationship between species composition and the measured soil nutrients
(P, N, Ca++, Na+ Mg++ and K+).
Results from this study have shown that the ecosystem is resilient to volcanic
disturbance as shown by the significant increase in species richness and diversity, increase
in positive species association, and improvement in soil nutrients within 14 years of
disturbance. However, this study had also raised some concern regarding invasive alien
species in the succession. Hence, ecological intervention in the form of weed management
in Mt. Merapi succession should be considered.
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ACKNOWLEDGEMENT
Praise is to “Allah Subhanahu wataàla” for His blessings and gift of health among
many others blessings for me to be able to finish this thesis. I have been fortunate and
grateful to have received an Australian Development Scholarship, a prestigious scholarship
awarded by the Australian Agency for International Development (AusAID). Gratitude also
goes to its staff, Ms. Megan Mayne and AusAID scholarship officer at the University of
Western Australia Ms. Deborah Pyatt and AusAID representative staff in Indonesia, Mr.
Riza Reinanto and Ms. Trilia Dianti for their kind support throughout the end of my study.
To my supervisors, Professor Richard Hobbs and Dr. Viki Cramer go my sincere
gratitude. It has been a great honour for me that they were willing to accept me as their
student and gave me the opportunity to work with them. Richard’s brilliant ideas in ecology
have always been influential to me and I enjoyed every minute of his presentations. I thank
Viki, for her patience for getting into the details of my research, field site visits to
Yogyakarta, Indonesia and her constructive criticisms and ideas throughout my research. I
would also thank to all the crew in the Ecosystem Restoration Laboratory, my office room
mates and fellow research students, Christina, Lauren, and Juan for the beautiful friendship.
My appreciation also goes to the head of school of plant biology Hans Lambers and all the
staff at the school, Dr. Krys Haq for her invaluable feedback on my writing and also to
UWA visiting professor, Pierre Legendrè for his lecture and tutorial on spatial analysis of
multivariate environmental data.
My special thanks to scientists, ecologists, botanists and staff at Murdoch
University. My gratitude goes to Dr. Fiona Valessini, for teaching me the basis of
multivariate ecology and the use of PRIMER software which was widely used in this study.
Dr. Philip Ladd, Dr. Joe Fontaine for the insightful discussions, Professor Ann Hamblin for
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her kind support and first comments, discussions and personal communication for Chapter
3 of the thesis. Thanks also to Dr. Helen Allison for the friendship and discussion on
resilience thinking.
To brilliant scientists Professor Pat Kennedy from Oregon State University, for
insightful discussions, her kind support, encouragement and friendship, Professor Franck
Lavigne, from the Universite´ Blaise Pascal, France, for his kind discussion by
correspondence on Mt. Merapi with his surprisingly fluent “Bahasa” and “Javanese”
languages, and Professor Bruce Clarkson from the Waikato University New Zealand, for
visiting my presentation and for a short discussion on volcanoes in the SERI world
conference on Ecological Restoration in Perth, 2009.
I would also like to extend my gratitude to the Indonesian government, Sri Sultan
Hamengkubuwono X, the King and the Governor of the Mataram and Yogyakarta Special
Province, Professor Endang Sukara and Dr. Irawati from the Life Sciences Department and
Bogor Botanical Garden, the Indonesian Institute of Sciences-LIPI, the heads of my
establishment in Indonesia, who permitted me to pursue my study in Australia, my advisor
Ir. Soewarno Hasanbahri MS. from the Faculty of Forestry, Gadjah Mada University for his
invaluable advice, encouragement and support while I was in Yogyakarta to do the
fieldwork, Ir. Arman Mallolongan M.M. from the Directorate of Forest Protection and
Nature Conservation (PHKA) the Indonesian Ministry of Forestry, and Ir. Tri Prasetyo, the
head of the Merapi National Park (BTNGM) for permission to enter the national park and
conduct the field data collections, Mbah Maridjan, the caretaker and gatekeeper of the
Merapi Mountain, Ir. Taufik Tri Hermawan M.Si from the Faculty of Forestry, Gadjah
Mada University for the invitation to host one of his lectures in ecology for undergraduate
students.
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I would like to thank also to the diligent volunteers from Gadjah Mada University
for their invaluable help, Gunawan, Ali, Iqbal, Indri, Kuntala, and Titi. We endured the
long and ascendant paths of the mountain, we defeated the biting cold of the mountain
climate, we shared the happiness at lunch time and we made it to cross the steep ravines
and survived being lost without GPS signals and safely made it to the base camp, and those
were memories I treasured.
Finally, I wish to thank my families, my wife Laily and our son Muhammad Hafidz
Imtiyaz Sutomo for their love, patience and company during the difficult periods of my
candidature, my parents Drs. Nana Suhana M.Si. and Susilowati, my sisters Kristina and
Tika Permata Sari for their continuous support and prayers, and lastly my aunt Dyah
Trimurti in Yogyakarta for her kind support ever since I started my journey as an
undergraduate student at Gadjah Mada University in 1999.
Perth, March, 2010
Sutomo
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TABLE OF CONTENTS
Abstract
iii
Acknowledgements
v
Table of Contents
viii
List of Figures
xi
List of Tables
xiii
List of Appendices
xiv
Chapter 1 General Introduction
1
1.1. Mt. Merapi Volcano
1
1.2. Nuèes Ardentes and Their Effects on Vegetation
2
1.3. Succession in Plant Communities
5
1.4. The Chronosequence Approach
8
1.5 Research Needs and Significance
10
1.6 Research Aims and Thesis Outline
12
Chapter 2 Mt. Merapi National Park and Study Sites Descriptions
14
2.1. Geographical Position
14
2.2. Management History
14
2.3. Climate
15
2.4. Geomorphology
15
2.5. Soils
16
2.6. Vegetation
17
2.7. Study Sites
18
2.7.1. Sites Selection
18
2.7.2. 2006 Site
21
2.7.3. 1998 Site
22
2.7.4. 1997 Site
24
2.7.5. 1994 Site
24
2.7.6. Reference Site
26
viii
Chapter 3 Establishment of Plant Communities on Nuèes Ardentes Volcanic
Deposits on Mt. Merapi: Composition, Diversity and Interspecific
Associations along a Chronosequence
27
3.1. Introduction
27
3.2. Method
31
3.2.1. Study Sites
31
3.2.2. Sampling
32
3.2.3. Statistical Analysis
32
3.3. Results
34
3.3.1. Species Diversity
34
3.3.2. Variation in Community Composition
35
3.3.3. Interspecific Association (co-occurrence)
40
3.4. Discussion
42
Chapter 4 Community Structure and Composition along a Chronosequence in
Forests Burnt by Nuèes Ardentes-Induced Fire in Mt. Merapi National
Park
49
4.1. Introduction
49
4.2. Method
53
4.2.1. Study Sites
53
4.2.2. Vegetation Sampling
53
4.2.3. Soil Sampling and Analysis
54
4.2.4. Data Analysis
54
4.2.4.1. Species Diversity and Turnover
55
4.2.4.2. Forest Structure
55
4.2.4.3. Variation in Community Composition
57
4.2.4.4. Variation in Soil Nutrients and Relationship with Floristic
58
Composition
4.3. Results
59
4.3.1. Species Diversity and Turnover
59
4.3.2. Forest Structure and Composition
60
4.3.3. Variation in Community Composition
63
ix
4.3.4. Variation in Soil Nutrients
66
4.3.5. Relationship between Floristic Composition and Soil Nutrients
66
4.4. Discussion
70
Chapter 5 General Discussion
76
5.1. Introduction
76
5.2. Plant Succession on Mt. Merapi
76
5.2.1. Species Re-colonization and Diversity
76
5.2.2. Species Composition and Structure
77
5.2.3. Exotic Invasives Species
78
5.2.4. Interspecific Association
80
5.2.5. Soil Nutrients
80
5.3. Similarities to other Volcanoes in Asia
82
5.4. Management Implications
85
5.5. Conclusions and Recommendations
87
Appendices
89
References
96
x
LIST OF FIGURES
Chapter 1
Figure 1.1 A Mt. Merapi nuèe ardente, known by locals as wedhus gembel, moves
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down the volcano’s slope to Kaliadem area in Yogyakarta Province
Figure 2.1 The 2006 Mt. Merapi nuées ardentes resulted from collapse of the
4
gegerboyo lava dome at the southwest flank of the volcano that had
destroyed vegetation and houses in the village surrounding the Gendol
River in Kaliadem area
Chapter 2
Figure 2.1 Sketch of morphological unit of the Merapi strato-volcano
16
Figure 2.2 Map of sampling sites in Mt. Merapi National Park
20
Figure 2.3 Collapsed lava dome (the gegerboyo) at the southwest flank of Mount
Merapi in June 2006
21
Figure 2.4 The 2006 site condition showing the deposits of the nuées ardentes
(primary succession site) and the seared trees of Pinus merkusii (secondary
succession site)
22
Figure 2.5 Primary succession on the 1998 nuées ardentes deposits
23
Figure 2.6 The 1998 secondary succession site’s condition
23
Figure 2.7 The 1997 secondary succession site’s condition showing Imperata
cylindrica, few Melastoma sp and one tree fern of Cyathea sp.
24
Figure 2.8 The 1994 secondary succession site’s condition
25
Figure 2.9 The primary succession on the 1994 nuées ardentes deposit
25
Figure 2.10 An old secondary forest in Kaliurang as a reference site
26
Chapter 3
Figure 3.1. Deposit 2006-Steep and deep abrupt valleys, formed in the Kaliadem area
following the 2006 Mt. Merapi devastating nuées ardentes
Figure 3.2. NMDS of sites based on vegetation composition and abundance
31
35
Figure 3.3. Total number of interspecific association of species in each deposit of
primary succession sites of Mt. Merapi
xi
40
Figure 3.4. Seedling of Pinus merkusii on the 2006 deposit. This is the type of
situation at Mt Merapi in which recruitment from seed was taking place
45
Chapter 4
Figure 4.1. NMDS of sites based on vegetation composition and abundance
63
Figure 4.2. LINKTREE diagram
69
Chapter 5
Figure 5.1. Resilience of Merapi volcanic ecosystem
xii
84
LIST OF TABLES
Chapter 2
Table 2.1. Site location, nuées ardentes history and environmental information in each
study site at Mt. Merapi National Park
19
Chapter 3
Table 3.1. Differences in species richness and diversity between sites on four nuées
ardentes deposits
34
Table 3.2. ANOSIM pairwise test of vegetation plots within sites
36
Table 3.3. SIMPER result 1. Percentage contribution of species to average Bray–
Curtis dissimilarities in all pairs of sites
38
Table 3.4. SIMPER result 2. Average abundance percentage in each site of the
selected species from table 2 (SIMPER 1)
39
Table 3.5. Results of the association tests using the chi-squared test statistic (χ2)
between the differentiating species and their co-occurring species
41
Chapter 4
Table 4.1. Differences in species diversity between the burnt sites and reference site
in Mt Merapi National Park
59
Table 4.2. Species turnover rates (D) between pairs of sites in the chronosequence on
Mt Merapi
60
Table 4.3. Number of individuals in each stratum for each site of secondary
succession at Mt. Merapi
60
Table 4.4. Importance Value Index (IVI), and shade tolerance for the most important
species in each stratum at each of the study sites
Table 4.5. ANOSIM pairwise test of NMDS vegetation plots ordination
62
64
Table 4.6. SIMPER result. Percentage contribution of species to average Bray–Curtis
dissimilarities in all pairs of sites
65
Table 4.7. Global test BEST. Combinations of edaphic factors that best constrain the
community composition within the Merapi National Park study sites
66
Table 4.8. Summary of differences in concentrations of soil nutrients in secondary
forest and the reference site on Mt. Merapi
xiii
67
LIST OF APPENDICES
Appendix 1. Mean abundance of species per family in the primary succession study
sites
89
Appendix 2. Mean abundance of species per family in the secondary succession
study sites
91
Appendix 3. Sketch of Eupatorium odoratum
94
Appendix 4. Sketch of Eupatorium riparium
95
xiv
Chapter 1
General Introduction
1.1. MT. MERAPI VOLCANO
Volcanic activity is a major natural disturbance that can catastrophically change an
ecosystem over a short time (Scheffer et al. 2001). Indonesia has 130 active volcanoes
(Weill 2004), and the island of Java is the most volcanically active island in the world, with
about 20 of its volcanoes having been active in the historical period (Whitten et al. 1996).
This study examines the patterns of vegetation succession on Mt. Merapi after volcanic
disturbance.
Mt. Merapi is one of the major active volcanoes on the Island of Java, located
approximately 30 kilometres north of Yogyakarta Province, in Central Java (Figure 2.2).
The name ‘Merapi’ is derived from the Javanese words ‘Meru’ (mountain) and ‘Api’ (fire).
The first eruption of Mount Merapi was noted in 1006. The eruption, described in an
ancient manuscript of Mataram Kingdom, the “Maha Pralaya Mataram”, covered all parts
of Central Java with ashes, and buried the Borobudur temple (Bemmelen 1970; Whitten et
al. 1996).
Mount Merapi has been reported as one of the most active volcanoes in the world
with 83 recorded eruptions (Gomez et al. 2009; Newhall et al. 2000). Hence, Mt. Merapi is
Indonesia’s most frequently erupting volcano (Simkin and Siebert 1994), with small and
frequent nuées ardentes occurring between large infrequent eruptions (Bardintzeff 1984).
Mount Merapi has varying intervals of eruptions. Small eruptions occur at an interval of 25 years, medium scale eruptions occur every 7-10 years (MVO 2006) and large explosive
eruptions usually occur once in a 100 years (Newhall et al. 2000). Mt. Merapi is become
1
one of the most monitored Indonesian volcanoes (MVO 2006; Thouret and Lavigne 2005;
Voight et al. 2000).
1.2. NUÉES ARDENTES AND THEIR EFFECTS ON VEGETATION
Nuées ardentes are known amongst the local people of Mt. Merapi in Yogyakarta as
‘Wedhus Gembel’, which literally means ‘a sheep ran down from the mountain’ (Figure
1.1). A nuèe ardente (French for “glowing cloud”) is the “rapid movement of extremely hot
(often more than 700° C) turbulent gases and fragmental material across a land surface from a
volcanic vent. The denser, basal part of a pyroclastic flow hugs the ground and follows topography,
moving with great force and speed (up to 200 km/h)” (Dale et al. 2005b p. 249). Merapi has
different characteristic of nuées ardentes compared with other volcanoes, and it has become
the reference volcano for its nuées ardentes, known as the ‘Merapi type’. The nuèes
ardentes of Merapi are the product of a collapsed lava dome at the summit (Takahashi and
Tsujimoto 2000). In the Merapi-type nuées ardentes, gravity plays a significant role
(Bardintzeff 1984). In other volcanoes, such as Mt. St. Helens in the USA, the nuées
ardentes do not originate from a collapsed lava dome, but instead are blasted out from side
vents of the mountain due to the escalating volcanic activity (Bardintzeff 1984; Belousov et
al. 2007).
A dome is built continuously, and the topographic position of this ‘new’ dome
determines whether, where and when it will collapse. The topographic position also
determines the frequency of eruptions (Chiu et al. 2001). Even though nuées ardentes
usually descend along the land surface, they are also capable of moving uphill, leaping over
ridges as they move away from the volcano’s crater. The 1994 nuèe ardente on Mt Merapi
moved 6.5 kilometres south-southwest and deposited approximately 2.5–3 million cubic
metres of material (Abdurachman et al. 2000). The most recent nuées ardentes occurred in
2
2006. On 4th June 2006, the “Geger boyo” flank in Kaliadem (Sleman District, Yogyakarta
Province) collapsed and nuées ardentes occurred until 14th June. The flows moved down
the slope through Gendol River (Kaliadem area), and destroyed all vegetation and buildings
in its path (Figure 1.2).
Figure 1.1. A Mt. Merapi nuèe ardente, known by locals as wedhus gembel, moves down the
volcano’s slope to Kaliadem area in Yogyakarta Province. Photo courtesy of Merapi Volcanology
Observatory/BPPTK (2006).
Volcanic eruptions have an important impact on the environment (Marti and Ernst
2005). Every year, approximately 60 volcanoes erupt on Earth, and even though almost
80% of them occurring under the oceans, terrestrial volcanic eruptions commonly cause
great impact on vegetation over large areas (Dale et al. 2005b). In East Java, Mt. BromoTengger’s violent eruptions created an un-vegetated area of ash near one of the best known
caldera called the “Sea of sand”. Similarly the 1963 violent eruption of Mt. Batur in Bali
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created hard lava stream areas cleared of vegetation in which regeneration could take more
than one hundred years (Whitten et al. 1996).
Figure 1.2. The 2006 Mt. Merapi nuées ardentes resulted from a collapsed gegerboyo lava dome
at the southwest flank of the volcano that had destroyed vegetation and houses in the village
surrounding the Gendol River in Kaliadem area. Photo by author.
Six types of volcanic activity can have complex effects upon vegetation: lava
formation, nuées ardentes, debris avalanches, lahar, tephra and blowdown (Dale et al.
2005b). Rough and hot heterogeneous mass material ejected from the nuées ardentes would
likely have an impact on any vegetation in its path due to poisonous gases, extreme heat or
by burial (Dale et al. 2005b; Kelfoun et al. 2000). Standing dead trees, damaged or buried
plants, and fine ash coverage of plant leaves are some of the impacts of nuées ardentes on
vegetation. In one of the few studies that has examined the effect of nuées ardentes on
vegetation, Kelfoun (2000) found that there were four types of effects to the tree vegetation
following the 1994 Merapi nuées ardentes: singed trees; broken trees; blown down trees;
and buried trees. Fine ash, stuck to tree foliage, can have a variety of effects. Antos and
4
Zobel (2005) which worked on tephra falls approximately 30 km from the cone of Mt. St.
Helens found that in the long term, fine ash coverage on tree foliage could reduce tree
growth due to the inhibition of the photosynthesis. The initial impact on understorey plants
is much more devastating, as they are buried under any material deposited with the flow.
Furthermore, Antos and Zobel (2005) also reported that the majority of herbaceous species
could penetrate upward through ≤ 4.5 centimetre deposits, but a 15 centimetre deep deposit
almost eliminated the herb layer. The heat released from nuées ardentes can ignite
wildfires. Intense fire will kill or damage some of the plants, but others are more resistant
and depend on fire for their regeneration (Bond and Wilgen 1996). Adaptations to fire
usually involve a plant’s capability to endure fire and to rejuvenate after the fire (Bond and
Wilgen 1996). According to Uhl (1990) species with thick bark and other protective
features exhibit greater survival. Following the devastating eruption of Mt. Semeru in East
Java in 1918, the singed Casuarina junghuhniana trees remained standing and re-sprouted
vigorously (Whitten et al. 1996). Albizia lopantha and Pinus merkusii are other examples
of tree species in Java where fire enforces their regeneration, breaks the hard seed case and
allows seeds to germinate (Whitten et al. 1996).
1.3. SUCCESSION IN PLANT COMMUNITIES
Succession is a central concept in ecology and the phenomenon of changes in
species composition over time has captured the interest of ecologists and naturalists for
centuries (McLean 1919; Smith 1914; Walker et al. 2007). Plant species establishment and
composition changes in a newly formed substrate with no biological legacy are known as
primary succession, while secondary succession refers to the species changes in an habitat
with relatively intact soil following a disturbance (Walker and del Moral 2003). Succession
5
model is a conceptual framework to elucidate the past successional trajectory and predict
the future trajectory by taking into account various mechanisms and the various stages of
the successional pathway (Walker and del Moral 2003).
The classical model proposed by Clements (1916) suggests that community
development occurs in distinct successional stages towards a “climax” community based on
processes of nudation (creation of a barren substrate following a disturbance); migration
(plant arrival on the already decaying site); ecesis (plant establishment, either generative or
vegetative); competition (plant interactions that eliminate some species) and, reaction (site
modification by plants). Later, Eggler (1954) proposed another model which emphasized
the importance of differential growth rates and longevities of species. In this model, early
successional species may be present in the late seral stages but probably will be less
abundant or slowed in their growth. Another model was proposed by Connell and Slatyer
(1977) which derived from the literature. The model was based on three concepts that affect
succession: facilitation, tolerance and inhibition. In the facilitation model, the establishment
of later species requires earlier species to ameliorate the site and thus these species become
the nurse plants for the later species. In the tolerance model, the later species that survive
will be the species that could tolerate the lower level of nutrients and other resources due to
competition with other species. Lastly, the inhibition model hypothesizes that the first
colonizer will continue to occupy the space and monopolize the resource, thus inhibiting
the establishment of the subsequently arriving species. This condition persists until the first
occupant species dies out and releases the resources that allow later species to establish and
reach maturity. More recent emphasis has shifted towards the idea of non-equilibrium
dynamics where it is recognized that disturbance can occur at any of those stages and at a
rate that makes a "climax" community unattainable (Suding and Hobbs 2009).
6
Li et al. (1999) stated that many succession theories were based on intensive work
in temperate forests. However, plant succession has also been described in the tropics
(Brearley et al. 2004; Finnegan 1996; 1981; Pena 2003). These studies have resulted in
various patterns of succession in terms of species richness, diversity, plant biomass and also
soil changes. However some general patterns may also appear in succession (Quesada et al.
2009; Radosevich et al. 2007). Increases in plant diversity, cover and biomass with time
have been observed in succession (Aplet et al. 1998; Dale et al. 2005c; del Moral 2000;
Finnegan 1996; Pena 2003; Tsuyuzaki and Hase 2005; Zhu et al. 2009). Common patterns
are also observed in the differences of the characteristics of plants which dominate each
stage of succession (Wills 2002). Early seral stages are usually dominated by plants with
rapid growth, small seeds, and a short life span (e.g. ruderal species). Late seral stages are
dominated by species with slow growth, often with large seeds and longer life spans
(Radosevich et al. 2007).
The way that species composition changes through time is described as the
successional trajectory. Walker and del Moral (2003) summarize some of the prominent
types of trajectory that may be found in a succession:
1) Convergence theory argues that random patterns of species composition at the
start of succession will be narrowed down and become more similar as time
progresses (Baer et al. 2005; del Moral 2007; Inouye and Tilman 1988).
2) A divergence pattern is identified when vegetation composition becomes more
dissimilar as the succession proceeds (Bush et al. 1992; Lepš and Rejmanek
1991).
3) Several trajectories can combine to form a network. Network trajectories occur
when there are several different invasive pioneer species that dominate and each of
the species can initiate different trajectories (del Moral et al. 2010).
7
With increasingly fragmented landscapes, convergence patterns are rare but can still be
found (although there are some limiting factors) in some primary succession sites such as
on Mt. St. Helens. Their chance of occurring increases where there are strong biotic
interactions (del Moral 2007; Walker and del Moral 2003). There are several methods
which can be used to examine a succession trajectory. A commonly used method is the
chronosequence approach (space for time substitution) and by the approach of direct
observation in the permanent plots (Walker and del Moral 2003).
1.4. THE CHRONOSEQUENCE APPROACH
There are various methods for studying succession, such as permanent plots,
chronosequence, air photo interpretation and palynology (Walker and del Moral 2003).
Permanent plots, and a chronosequence approach, are the most common methods used in
succession studies. Direct observation in permanent plots, combined with experimental and
manipulative study, is the most appropriate method to apply (when possible), in order to
understand the mechanism of succession. However, permanent plot studies which last
longer than 10 years are rare (Foster and Tilman 2000). When timing (i.e. the succession
measured in scales of decades or even centuries) and logistics are the main concerns and no
aerial photos are available, an indirect approach using a chronosequence method is
preferred. The idea is essentially to choose a series of plots showing various stages of
disturbance or abandonment and thus presume that they characterize different stages of an
analogous successional development (Foster and Tilman 2000; Gomez-Pompa and
Vazquez-Yanes 1981; Walker and del Moral 2003).
8
In order for the chronosequence approach to be useful in inferring patterns of the
succession over time, there are prominent assumptions that need to be considered. Firstly,
the environmental conditions such as the climate, substrates and topographical positions of
the sites under study need to be relatively similar (Durán et al. 2009; Johnson and
Miyanishi 2008; Pickett 1989). Secondly, these sites should also have the same disturbance
history and have similar pre-disturbance conditions. This means that the past disturbance
history should be relatively similar or at least randomised if it is unknown. When these
assumptions are met then the chronosequence approach can be a powerful tool to use in the
study of succession (Durán et al. 2009; Johnson and Miyanishi 2008; Wills 2002; Zhu et al.
2009).
There are potential disadvantages of the chronosequence approach (Johnson and
Miyanishi 2008; Milberg 1995; Pickett 1989). Pickett (1989) stated that the
chronosequence approach is disadvantageous because of its averaging property which could
hide the short term dynamics of the examined system. Meanwhile, Foster and Tillman
(2000) argued that it is this particular property which yielded broad successional trends
makes the chronosequence approach advantageous. Myster and Malahy (2008) conclude
that the decision whether to use the chronosequence approach or not will depend largely on
the parameters measured and that species richness, diversity and total plant abundance
perform best with the chronosequence approach. Foster and Tilman (2000) support the use
of a chronosequence approach to infer basic patterns in succession based by re-sampling a
chronosequence plots over an interval of time because the rate of the succession can be
measured.
9
A chronosequence approach has been applied in many types of ecosystems in both
primary succession and secondary succession. The method has been used on primary
successional seres, especially on volcanic terrain, by Vitousek et al. (1993) in Hawaii.
Meanwhile the method had also been widely applied in a wide range of sites following fire
(Gonzalez-Tagle et al. 2008; Watson and Wardel-Johnson 2004) and land abandonment
(Baniya et al. 2009; Myster and Pickett 1992a; b). The chronosequence approach has been
used to serve a specific purpose to examine the vegetation dynamics (Wills 2002; Zhu et
al. 2009), soil nutrient status and biomass (Bautista-Cruz and del Castillo 2005; Bormann et
al. 1995; Durán et al. 2009; Wang et al. 2009) and forest structure (Spencer et al. 2001).
The results of these studies suggest that using a chronosequence approach remains a robust
method to gain the fundamental patterns of changes in succession.
1.5. RESEARCH NEEDS AND SIGNIFICANCE
Montane forests such as those on Mt. Merapi have become increasingly important
for conserving Indonesia’s biodiversity (Whitten et al. 1996). The importance of the
tropical montane forest for Indonesia is understandable, considering that much of its
tropical lowland forest has been cleared for more than 30 years (Whitten et al. 1996).
Paddy fields and plantations are now common in lowland landscapes in Java, replacing the
primary lowland forest that previously dominated these landscapes. In the northern part of
Java no single forest remnant remains. In the southern part only a few patches of lowland
natural forest remain such as those in Ujung Kulon and Pangandaran National Parks in west
Java. In the southern parts of central and east Java all lowland forests were converted to
plantations. Consequently, pristine forests in Java can only be found on mountains (van
Steenis 1972; Whitten et al. 1996). Forests in Merapi are also subjected to destruction due
10
to increasing volcanic activity of Merapi as well as prolonged drought and human
disturbances (Anonym 1999; 2004; Heru 2002; MVO 2006; Newhall et al. 2000).
Increasing temperatures during the El Nino Southern Oscillation (ENSO) in 1992 caused
wildfires in this area and the escalating volcanic activity led to an eruption in 1994 which
caused catalytic destruction to the forest over 650.30 hectares (Dinas Kehutanan DIY
1999).
Rehabilitation and management of a degraded ecosystem, including a volcanic
ecosystem, will require specific knowledge of the process of succession (Walker et al.
2007; Walker and del Moral 2003). However, the scientific knowledge in this field is
limited, particularly in volcanic tropical montane forests, where scientific studies about the
effects and the ecological consequences of disturbances at high-elevation are still scant
(Horn et al. 2001; Tsuyuzaki and Hase 2005; Whittaker et al. 1999). In contrast, there is
abundant of evidence regarding the needs of ecological intervention to accelerate the rate of
succession in lowland tropics (Brearley et al. 2004; Lindig-Cisneros et al. 2006; Zahawi
and Augspurger 1999). Understanding the dynamics of forest ecosystems during succession
and how they interact with disturbances, will be an important component of forest
ecosystem conservation and restoration management practices (Hobbs et al. 2007; Swamy
et al. 2000). However, these key elements are often neglected in Indonesia and elsewhere.
In addition, although there is a large body of literature available on the subject of
the geology and vulcanology of Mt. Merapi (Berthommier and Camus 1991; Camus et al.
2000; Gomez et al. 2009; Lavigne 1999; Newhall et al. 2000), ecological research on Mt.
Merapi has received scant attention. To date there has not been any comprehensive research
on the dynamics of the vegetation (native and exotic species) following volcanic
disturbance (but see Hardiwinoto et al. 1998). Therefore, the present study is significant in
11
terms of generating useful baseline data for the management of volcanic ecosystems in the
region and elsewhere in Indonesia and Asia.
In summary, this study will:
1) Be the first project to carry out a study on plant community succession on Mt.
Merapi in both primary and secondary succession.
2) Increase the knowledge and theoretical background in succession and dynamics of
a tropical montane forest following volcanic disturbance.
3) Contribute to the efforts of linking ecological succession knowledge with
restoration/rehabilitation practices.
1.6. RESEARCH AIMS AND THESIS OUTLINE
This study uses a chronosequence approach to examine the vegetation succession on
Mt. Merapi after volcanic disturbance. A range of sites that have been affected by nuées
ardentes at different times are available on Mt Merapi. While acknowledging the
limitations imposed by site variables other than time since disturbance, the sites used still
provide a sequence of different ages which can form the basis for further studies
Overall, the aims of the present study are to:
1) Evaluate how the vegetation changes over time in terms of the structure, species
composition and diversity in primary and secondary succession sites.
2) Examine whether early interaction patterns of the species in the primary
succession can be identified by examining their interspecific association.
3) Examine whether there were significant differences in major soil nutrients
between sites of varying ages and whether these edaphic factors are correlated
with the species assemblages in the secondary succession sites.
12
4) Evaluate these findings and discuss the implications for the management plan of
the Mt. Merapi National Park.
This thesis consists of five chapters. In chapter 2, the regional context of the Merapi
National Park and the study sites is briefly described with a particular emphasis on
management history, conservation, vegetation formation and environmental condition such
as climate, soil and geology. In this chapter I also describe the criteria and rationale for the
choice of sites. The next two chapters are the topics that are covered by this thesis. Chapter
3 focuses on the establishment and succession of pioneer plant communities on the nuèe
ardente deposits (primary succession), which also includes detection of early interaction
patterns of the species in the succession by examining their interspecific association.
Chapter 4 examines changes in floristic composition in the secondary succession which
also takes into account the identification of a subset of edaphic factors that best correlates
with the species assemblages in the course of the secondary succession. Lastly, Chapter 5
summarizes the knowledge gained from the previous topic chapters and discusses the
findings in the framework of management planning of the national park. It also proposes
potential ecological restoration (intervention) activities and outlines further research.
13
Chapter 2
Mt. Merapi National Park and Study Sites Descriptions
2.1. GEOGRAPHICAL POSITION
Mt. Merapi (7º 35’ S and 110º 24’ E) is administratively located in two provinces,
Central Java (Magelang, Boyolali and Klaten Districts) and Yogyakarta (Sleman District).
In Yogyakarta Province, Mt. Merapi is located approximately 30 kilometres north of
Yogyakarta. Mt. Merapi is representative of the landforms, soils and vegetation on a
volcanic mountain that typify a large portion of montane ecosystems in Java (Whitten et al.
1996).
2.2. MANAGEMENT HISTORY
In 1931, the Dutch colonial government established the forest area located in
Central Java Province (6,962.1 ha) and Yogyakarta Province (1,510 ha) on the slope of
Merapi as protected forest (Act no. G. B. No 4197/B, 4 May 1931). The appointed
protected forest was meant to protect the local natural resources and the surrounding area.
In 1975, the Indonesian Ministry of Agriculture, the department that managed Indonesian
forests at that time, issued Act no. 347/Kpts/Um/8/1975, which assigned some parts of the
protected forest as nature reserve (198.5 ha) and tourism forest (30 ha) named CA/TWA
Plawangan-Turgo. Since 1975, villagers were prohibited from living in the protected forest
in the kumpulrejo and patuk areas because of the escalating threat from the volcano. At this
time, management was taken over by the joint Forestry and Agriculture Provincial office of
Yogyakarta. The Merapi Forest area located in Yogyakarta Province is managed by
Yogyakarta Forestry Provincial Office based on Governor Act no. 336/1974. On 4 May
14
2004, The Minister of Forestry issued decree no. 134/MENHUT-II/2004 regarding the
amendment of forest area functions of protected forests, nature reserves, and tourism forests
to become the Merapi National Park, which covers an area of 6,410 ha (Anonym 2004).
2.3. CLIMATE
Based on Schmidt and Fergusson’s climate classification (1951), the Merapi area is
classified as type B - tropical monsoon area - which is characterized by a high intensity of
rainfall in the wet season (November-March) with a dry season that can often be very dry
without any rainfall (April-October). Annual precipitation varies from 2,500–3,500
millimetres (Anonym 2004). The variation of rainfall on Mt Merapi’s slope is influenced by
orographic precipitation. As in many other tropical monsoon areas, there are minor
temperature and humidity variations during the year. Relative humidity on Mt Merapi
varies from 70%-90%, with daily average temperatures from 19° to 30° C (Dinas
Kehutanan DIY 1999).
2.4. GEOMOPRHOLOGY
Mount Merapi is a basaltic-andesite volcano whose formation is known through five
periods: the pre-Merapi period (>400,000 years ago), old-Merapi period (between 400,000
to 6,700 years ago), middle-period Merapi (6,700 – 2,200 years ago), young-Merapi
(2,200-600 years ago) and the Merapi we know today which began 600 years ago (Gertisser
and Keller 2003; Newhall et al. 2000). Based on its geomorphology, Mt. Merapi can be
divided into five units: volcanic cone, volcanic slope, volcanic foot plain and fluviovolcanic plain (Pannekoek 1949) (Figure 2.1).
Mt. Merapi’s strato-volcano summit
continually changes due to dome growth and collapse (typical of the Merapi-type nuées
15
ardentes; see chapter 1) and in 2006, the Merapi summit was 2,911 metres (MVO 2006).
The topography of the Mt Merapi varies from undulating to mountainous. There are two
bedrock materials that comprise Mt. Merapi (Anonym 2004):
1. Bedrock of the new Merapi, consisting of tufa, lahar, endeictic and basaltic lava. This
deposit is widely spread on Mt Merapi.
2. Old Merapi volcanic bedrock, which is found locally in some places, especially in hill
areas such as on Turgo, Plawangan, Gono and Maron hills.
Figure 2.1. Sketch of morphological unit of the Merapi strato-volcano (Pannekoek 1949)
A spring belt occurs when groundwater moves from recharge area to the discharge
area at the slope break (Irawan and Puradimaja 2006). Figure 2.1.showed that there were
three break of slope-spring belt on Mt. Merapi which is the source of excellent quality
water resource for the lower areas (Irawan and Puradimaja 2006; Pannekoek 1949).
16
2.5. SOILS
Soils of the study area are mainly of young volcanic-ash origin (regosol) with
shallow and/or deep, low to medium fertility solums with a profile not yet developed
(Anonym 2004; Darmawijaya 1990). The soil textures are granulated, whereas the
structures are crumbly (Anonym 2004; Dinas Kehutanan DIY 1999).
2.6. VEGETATION
Based on his extensive work in Java’s volcanic mountain forests, van Steenis (1972)
described plant formation patterns that generally occupied the montane forests. Primary
forests occur both below and above 2,000 metres. High primary forest occurs at 1,0002,000 metre elevation. This forest contains tall and emergent trees species up to 25-35
metres such as Altingia excelsa, Schima wallichii and Vernonea arborea. Primary forest
that occurs at an elevation greater than 2,000 metres is named elfin or low primary forest.
This type of forest is structured by small trees and shrubs up to 10 metres in height such as
Pittosporum spp., Rhododendron spp., and Vaccinium spp.
When these primary forests experience a disturbance that damages or kills
vegetation (fire, volcanic activity), then secondary forests developed. Secondary forest that
occurs at elevation below 2,000 metres usually includes invasive grass and shrub species
such as Imperata cylindrica, Eupatorium odoratum, Lantana camara, and Saccharum
spontaneum and the tree species Homalanthus populneus, Macaranga javanica, Vernonea
arborea, Villebrunea rubescens, Casuarina junghuhniana and Pinus merkusii. Disturbed
elfin forests generally include species such as Albizia spp., Anaphalis spp. (some Anaphalis
are also can be found in lower elevation secondary forest), Engelhardia spp., Dodonaea
spp., and Lespedeza junghuhniana.
17
2.7. STUDY SITES
On Mt. Merapi, areas which have been completely buried by the nuées ardentes
deposits undergo primary succession. These areas usually occur along the streams, channels
or valleys created by the solid material flow paths of nuées ardentes. The secondary
succession areas were located adjacent to the primary succession areas. These areas are the
adjacent forest on either side of the valley, or deposit channel which escapes burial and
mainly scorched by the extreme temperature of the nuées ardentes.
2.7.1. Site Selection
Sites or areas of different ages (years since last nuée ardente) were selected to
obtain a chronosequence. Identification of site age was conducted by studying aerial
photographs, topography maps, and nuée ardente history maps (obtained from the Merapi
Volcanology Observatory) to date sites affected by recent nuées ardentes. Identification
was also conducted by reconnaissance study, interviewing long - term residents of the
surrounding villages, personal communication with the national park’s ranger and
managers, and also field site visits. Sites also had to show no obvious signs of human
disturbance and be at least 50 metres from any human activities or structures. Based on
these, I chose four sites that were affected by nuées ardentes at different times (2006, 1998,
1997, and 1994) and one forest area that was mostly undisturbed and had not been affected
by nuées ardentes for at least 50 years as a reference site (Figure 2.2). The five sites were
located in a lower montane zone and were located at a range of altitudes from 1,000 to
1,600 metres. Chronosequence assumptions were met within these sites as they had similar
environmental conditions such as climate, substrate, topography and geomorphology. Four
sites were low-slope (6-13°) while one site was of intermediate-slope (28-29°). The fieldwork
18
was conducted from March to August 2008. Summary of the average environmental
conditions in each site is described in table 1.
Table 2.1. Site location, nuées ardentes history and environmental information in each study sites at
Mt. Merapi National Park.
Average
Year of nuée
Site age
Elevation
Slope
Location
Soil type
(years)
(m)
ardente
(°)
Kaliadem
2006
2
Regosol
1,220
12.2
Kalilamat
1998
10
Regosol
1,579
28.26
Kalibedog
1997
11
Regosol
1,207
6.53
Kalikuning
1994
14
Regosol
1,180
6.23
Kaliurang
-
-
Regosol
1,000
13
19
Figure 2.2. Map of sampling sites in Mt. Merapi National Park. Ellipses show the location of the sampling sites in the nuées
ardentes affected areas. The rectangle shows the location of the reference site. The green spot in the insert of the map of Java Island
is the approximate location of Mt. Merapi on the island
20
2.7.2. 2006 Site
The 2006 site was located in Kaliadem, at an altitude of 1,220 metres. This site is
the youngest site (2 years old) in terms of time since last nuèe ardente. On 4 June 2006, the
Geger Boyo Flank in Kaliadem (Sleman District, Yogyakarta Province) collapsed (Figure
2.3) and nuées ardentes occurred until 14 June 2006. The flows moved down the slope
through the Gendol River, with areas dominated by Pinus merkusii heavily affected. The
area that was affected by the nuèe ardente flows formed steep and deep abrupt valleys.
Deposits from the nuées ardentes were estimated to reach 5.6 million cubic metres
(PVMBG 2006). In this site, the damage caused by the flows was still clearly visible seared and standing or buried dead trees of P. merkusii located around the centre, the area
that was near to the Gendol River (Figure 2.4).
Figure 2.3. Collapsed lava dome (the gegerboyo) at the southwest flank of Mount Merapi in
June 2006. Arrow sign showing the remains of the track created by the collapsed dome
(MVO 2006)
21
Figure 2.4. The 2006 site condition showing the deposits of the nuées ardentes (primary
succession site) and the seared trees of Pinus merkusii (secondary succession site)
2.7.3. 1998 Site
The 1998 Merapi eruption on 11 and 19 July 1998 produced 8.8 million cubic
metres of nuèe ardente deposits, in places up to 8 metres deep. Nuèes ardentes flow down
towards rivers such as Sat, Lamat and Senowo which are located west side of Mt Merapi, 6
kilometres from the crater. Located at an altitude of around 1,500–1,600 metres, the 1998
site was the most difficult to reach. This site is situated in an area of one of the most
dangerous tracks of volcanic flows - the Lamat River which belongs to the Central Java
section of the national park. The distance from the crater was around 2 kilometres and very
remote with a steep slope of 28°. Figure 2.5 and 2.6 shows sites that have undergone
primary and secondary succession.
22
Figure 2.5. Primary succession on the 1998 nuées ardentes deposits
Figure 2.6. The 1998 secondary succession site’s condition
23
2.7.4. 1997 Site
The nuées ardentes that occurred in 1997 flooded the Krasak, Bebeng, Boyong and
Bedog Rivers, which left deep and abrupt marks of valleys in the south-west flank and 2
million cubic metres of material deposits. The sampling sites for primary and secondary
succession were located in the areas surrounding the Bedog River at an altitude of 1,207
metres in the Yogyakarta Province. Figure 2.7 shows the secondary succession.
Figure 2.7. The 1997 secondary succession site’s condition showing Imperata cylindrica, few
Melastoma sp and one tree fern of Cyathea sp.
2.7.5. 1994 Site
In 1994, the collapsed dome initiated surges of nuées ardentes down to the Boyong
River which reached the middle of Turgo and Plawangan Hill and Kuning River,
Yogyakarta Province. Turgo and Plawangan are hills that are believed to be the remains of
what vulcanologists call old-Merapi, or the beginning of the Merapi that we see today
(Newhall et al. 2000). The estimated volume of the slide was 2.6 million cubic metres and
24
the deposits were almost 6 metres deep. The 1994 sampling site is located in an area
surrounding the Kuning River, at an altitude of ± 1,180 m. Figure 2.8 and 2.9 shows areas
that have undergone secondary and primary succession.
Figure 2.8. The 1994 secondary succession site’s condition
Figure 2.9. The primary succession on the 1994 nuées ardentes deposit
25
2.7.6. Reference Site
To compare the development of plant communities with a site that had not been
disturbed by the scorching and fire associated with nuées ardentes, I chose a reference site
named Kaliurang (Figure 2.10), which is located on the south flank of Mt Merapi, near the
Plawangan and Turgo Hills, approximately 4-5 kilometres from the crater in Yogyakarta
Province. Kaliurang is located at 7º35’ S and 110º24’ E, at an altitude approximately 1000
metres. Generally the forest condition around this site is intact and is characterized by the
presence of large trees (diameter between 40 and 100 centimetres) of primary species such
as Altingia excelsa. Some gaps due to natural tree felling were noted. In some locations, the
ground was covered by ferns, Araceae, and Palmae. However, this reference site is actually
an old growth secondary forest and it is not a “pristine” forest. This forest has been likely
been damaged or cleared (whether by volcanic and or anthropogenic disturbance) and has
re-grown at some time in the past.
Figure 2.10. An old secondary forest in Kaliurang as a reference site
26
Chapter 3
Establishment of Plant Communities on Nuèes Ardentes Volcanic Deposits
on Mt. Merapi: Composition, Diversity and Interspecific Association
along a Chronosequence
3.1. INTRODUCTION
Primary succession is the establishment of plant species and subsequent change in
community composition over time on a substrate with little or no biological legacy (Walker
and del Moral 2003). Volcanic eruptions pose a significant challenge to the study of
primary succession compared with other disturbances (i.e. erosion, landslides, floodplains
and glaciers) because of the absence of a biological legacy following the eruption (Franklin
et al. 1985). One type of volcanic disturbance is nuées ardentes or pyroclastic flows. Nuèes
ardentes are hot turbulent gas and fragmented material resulting from a collapsed lava
dome that rapidly moves down the volcanic slope (Dale et al. 2005b). The accumulation of
this material is called a nuées ardentes deposit and it may be up to ten metres thick
(Franklin et al. 1985). Such pure mineral deposits preserve no “memory” of previous
vegetation owing to the absence of a seed bank (Thornton 2007). Hence, colonization must
occur from other undisturbed places.
The phenomenon that creates the disturbance to a vegetation community and the
process of recovery that occurs plays a significant role in the dynamics of the community
and its species diversity (Connell and Slatyer 1977; Crain et al. 2008). The establishment of
vegetation on volcanic deposits has been documented in many parts of the world such as in
USA, Italy and Japan, with rates of establishment being shown to vary (Aplet et al. 1998;
Dale et al. 2005c; Eggler 1959; Tsuyuzaki 1991). For example, plant establishment and
spread on the debris-avalanche deposit of Mt. St. Helens were slow during the first years
27
after eruption (Dale et al. 2005c). In contrast, Taylor (1957) reported that six years after
Mt. Lamington in West Papua erupted, vegetation regeneration was very rapid, and
included species such as Saccharum spontaneum, Imperata cylindrica, Pennisetum
macrostachyum, Vitaceae and several ferns. Mt. Krakatau had at least 64 vascular plant
species (dominated by grass species such as Saccharum spontaneum and Imperata
cylindrica) that colonized the island three years after the eruption (Thornton 2007).
The pace of primary succession on volcanoes is usually thought to be slow,
especially in the early stage, but as abiotic processes develop on barren substrates, the rate
of succession increases (del Moral and Wood 1993). The initial conditions following
disturbance will determine the trajectory of plant establishment. Survival and recovery of
plants following volcanic disturbance will be severely affected by the type of volcanic
activity, nutrients transported by the volcanic disturbance, distance from seed sources and
the types of propagules from nearby undisturbed areas (Dale et al. 2005a; del Moral and
Wood 1993; Walker and del Moral 2003).
The study of succession may assist in recognizing the possible effects of species
interactions (i.e. facilitation or inhibition) (Connell and Slatyer 1977; Walker et al. 2007).
Facilitation promotes establishment and in the context of succession, facilitation can be
defined as any role plants play in influencing changes in species composition to the next
stage (Walker and del Moral 2003). In contrast, inhibition basically refers to the negative
effect that one species has on another species (Walker and del Moral 2003). A species
inhibits the growth or establishment of the co-occurring species through resource
competition or by directly excluding subsequent species by allelopathy effect (Collins and
Jose 2009; Soerjani et al. 1983; Walker and del Moral 2003). Previous studies have shown
that facilitation is more prominent in severely disturbed habitats, whereas competition tends
to be more significant in more productive and established habitats (Callaway and Walker
28
1997; Walker et al. 2007). The barren landscapes created by volcanic disturbance provide
excellent opportunities to examine the role of pioneer species in facilitating or inhibiting
the later species in the succession (Morris and Wood 1989; Walker and del Moral 2003).
However, much remains unknown about the initial interactions during succession that drive
the subsequent community composition (Bellingham et al. 2001; Connell and Slatyer
1977).
Observations of species co-occurrence may be seen as the first attempt to identify
species interactions and niche processes that structure the community (Walker and del
Moral 2003; Widyatmoko and Burgman 2006). By determining which species are strongly
associated with of co-occurring species (through field observation), experimental studies of
species interactions may become more effective, and may help in understanding the role of
species interaction throughout succession (Myster and Pickett 1992b).
A commonly used method in studying primary succession is the chronosequence
approach or space for time substitution (Walker and del Moral 2003). An alternative
approach is to establish permanent plots to carry out such research (Cramer et al. 2008; del
Moral 2007; del Moral and Wood 1993; Simbolon et al. 2003). Although there has been
some criticism of the chronosequence approach (see chapter 1) (Herben 1996; Johnson and
Miyanishi 2008), it is still a useful method, especially when timing and logistics are a
problem (Aplet et al. 1998; Myster and Malahy 2008). Thus, a chronosequence approach
can provide an initial approach to understanding the fundamental patterns of changes in
succession (Foster and Tilman 2000). The nuées ardentes deposits found on Mt. Merapi
present an opportunity to study succession using a chronosequence approach. The deposits
at these sites are relatively young, with the last known eruptions occurring between 1994
and 2006. The sites are not strictly analogous in terms of their slope and landscape position,
and hence care must be taken in interpreting the results (del Moral and Ellis 2004).
29
My goal was to understand the patterns of plant re-establishment in the early
successional stage on the nuées ardentes deposits on Mt. Merapi, in order to increase the
understanding of plant successional patterns caused by these unique, though ecologically
understudied Merapi-type nuées ardentes deposit materials. My research questions were:
1) How does species composition and diversity vary across deposits of different ages?
2) Which species contributed the most to the differences in composition between different
aged deposits?
3) Could early patterns of species interactions be identified by examining their
interspecific association?
30
3.2. METHOD
3.2.1. Study sites
My research sites were located in the south-west flank of Mt. Merapi, in the zone of
the Merapi National Park. These sites are more often affected by nuées ardentes, which
tend to flow down the hills in this direction. I chose four areas that were affected by nuées
ardentes between 1994 and 2006. The four deposit sites were located in a lower montane
zone at a range of altitudes from 1000 to 1500 metres (Montagnini and Jordan 2005) (for
detailed description and the map of the deposit sites, see Chapter 2). The condition of the
2006 deposit, two years after the nuées ardentes disturbance can be seen in figure 3.1.
Figure 3.1. Deposit 2006-Steep and deep abrupt valleys, formed in the Kaliadem area following
the 2006 Mt. Merapi devastating nuées ardentes. Photo was taken in April 2008.
31
3.2.2. Sampling
In 2008, I sampled vegetation on the four nuées ardentes deposits. I sampled ten
circular plots (diameter range approximately 10 metres) in each of the four deposit sites,
assigned at random to grid cells on a map (Dale et al. 2005a; Zimmerman et al. 2008). The
position and altitude of each site were recorded using a GPS (Garmin E-Trex legend) and
slope was measured using a clinometer (Suunto PM-5/360 PC Finland, clinometer).
I measured plant abundance as density, a count of the numbers of individuals of a
species within the quadrat (Endo et al. 2008; Kent and Coker 1992). Local plant names and
Latin names, when known, were noted. Whenever there was any doubt about a species
name, a herbarium sample was made. Drying and sample identification were carried out in
Dendrology Laboratory, Faculty of Forestry, Universitas Gadjah Mada. Vascular plant
nomenclature was based on “Flora of Java” (Backer and van den Brink 1963), “Mountain
Flora of Java” (van Steenis 1972), and International Plant Names Index Databases (IPNI
2008). .
3.2.3. Statistical analysis
Data were analysed using multivariate and univariate statistics. Multivariate
analyses such as Non Metric Multidimensional Scaling (NMDS), Analysis of Similarity
(ANOSIM), Similarity Percentage (SIMPER), were conducted using PRIMER V.6 (Clarke
and Gorley 2005). Univariate analyses such as ANOVA and post hoc tests were conducted
using SPSS package V.11.5.
Species richness and Shannon-Wiener species diversity on each deposit were
calculated. Changes in these values across the deposits were tested for significance using
one-way ANOVA test. I tested differences in community composition between deposits
using data on species abundance (density) per plot. The data was square root transformed
32
prior to constructing a resemblance matrix based on Bray-Curtis similarity (Valessini
2009). An NMDS ordination diagram was then generated based on the resemblance matrix.
Ordination results were then overlaid with the results of a cluster analysis, using 20%
similarity to show grouping of plots in each deposit and separation between different
deposits. Variation in community composition between deposits was subsequently tested
for significance using one-way ANOSIM (analysis of similarity). ANOSIM is basically
analogous to standard univariate ANOVA, and tests a priori defined groups against random
groups in ordinate space. The RANOSIM statistic values, generated by ANOSIM, are a
relative measure of separation of the a priori defined groups. A zero (0) indicates that there
is no difference among groups, while one (1) indicates that all samples within groups are
more similar to one another than any samples from different groups (Clarke 1993). The
SIMPER routine was then used to explore the relative contribution of individual species to
dissimilarity among deposits.
Interspecific association between species was measured using the chi-square (χ2)
test of the species presence/absence data on a 2 x 2 contingency table (Kent and Coker
1992; Ludwig and Reynolds 1988). The significance of the chi-square test statistic is
determined by comparing it to the theoretical chi-square distribution (P = 0.05, df = 1).
Based on this test, there are two types of association. A positive association occurs if χ2 test
> χ2
theoretical,
that is, the pair of species occurred together more often than expected by
chance. Negative association occurs if χ2 test < χ2 theoretical, that is, the pair of species occurred
together less often than expected by chance. The strength of the association was measured
using the Ochiai index which gives a value of 0 at ‘no association’ and 1 at
‘complete/maximum association’.
33
3.3. RESULTS
3.3.1. Species diversity
Fifty two species belonging to 23 families were recorded in the four sites (Appendix
1). The highest number of species belonged to the Asteraceae (14), then Poaceae (11),
followed by Fabaceae (4). Species richness was significantly different between deposits
(ANOVA P = 0.05, table 3.1). Species richness was lowest in the 2006 deposit and was
highest in the oldest deposit (1994). Species diversity was lowest in the 1998 deposit and
highest in the 1994 deposit (Table 3.1). The two youngest deposits (2006 and 1998) had
significantly lower species diversity than the two older deposits (1997 and 1994) (ANOVA
P = 0.05; table 3.1). Despite a difference in deposit age of only one year, the species
diversity was significantly lower in the 1998 than the 1997 deposit.
Table 3.1. Differences in species richness and diversity between sites on four nuées ardentes
deposits. Superscript letters (a-b) after mean values (±SD) indicate significant differences between
sites as assessed with Tukey’s HSD test.
Time of eruption/age of deposit Species richness
Species diversity
at sampling
2006 site/2 years
4.0 (±1.49)a
1.03 (± 0.3)a
1998 site/10 years
6.7 (±1.41)b
0.95 (± 0.34)a
1997 site/11 years
7.4 (±1.07)bc
1.38 (± 0.16)b
1994 site/14 years
8.4 (±2.11)c
1.61 (± 0.22)b
34
3.3.2. Variation in community composition
Figure 3.2 showed NMDS ordination that clustered the sites into three groups . The
right cluster consists of the plots from the youngest deposit (2006), and the left cluster of
plots from the 1998 deposit. The middle cluster contains plots from the 1997 and 1994
deposits. This cluster is slightly intersected with the left cluster. An analysis of similarity
(ANOSIM) test confirmed that there were significant differences in Bray-Curtis species
similarities between deposits on Mt. Merapi (Global RANOSIM = 0.91, P < 0.001). Pairwise
comparison tests between sites generally showed R values ranging from 0.9 to 1.0 (Table
3.2) except for the comparison between the 1997 deposit and the 1994 deposits, which had
an R-value of 0.42.
Figure 3.2. NMDS of sites based on vegetation composition and abundance of the 2006 deposit
(triangles), 1998 deposit (inverse triangles), 1997 deposit (squares), and 1994 deposit (diamonds).
Ellipses indicate groups resulting from cluster analysis.
35
Table 3.2. ANOSIM pairwise test of vegetation plots within sites. Sample statistic (Global R): 0.91,
significance level of sample statistic P < 0.001, number of permutation: 999. A zero (0) indicates
that there is no difference among groups, while one (1) indicates the groups are very different with
each other.
Groups
R Statistic
2006-1998
1
2006-1997
0.97
2006-1994
0.99
1998-1997
0.94
1998-1994
0.96
1997-1994
0.42
Twelve species (Anaphalis javanica, Athyrium macrocarpum, Calliandra
callothyrsus, Cyperus flavidus, Eupatorium riparium, Imperata cylindrica, Panicum
reptans, Paspalum conjugatum, Pinus merkusii, Polygala paniculata, Polyosma ilicifolia,
and Polytrias amaura) were most responsible for the dissimilarity between the sites (Tables
3.3 and 3.4). The species that contributed to the dissimilarities among almost all pairs of
deposits were A. javanica, E. riparium, and Polytrias amaura, while I. cylindrica was the
species that was responsible for dissimilarities among all pair of deposit sites (Table 3.3).
The average of the Bray-Curtis dissimilarities between deposit 2006 and 1998 is
98% and this is made up of five species. Athyrium macrocarpum contribution is 10.27% of
the total 98%; E. riparium gives 6.85% of this total. I. cylindrica and Panicum reptans give
15.75% and 13.33% (Table 3.3). Therefore the highest dissimilarity contribution was made
by Panicum reptans as its abundance was 3.98% in the 1998 deposit and was absent in the
2006 deposit (Table 3.4). For the 2006 and 1997 pair of deposit sites, the dissimilarity is
93.13% (Table 3.3) characterized mainly by three species, namely A. javanica, Athyrium
macrocarpum and Polytrias amaura, with A. javanica providing the highest contribution
(16.35%) of the dissimilarities. A. javanica was absent in the 2006 and in the 1997 its
abundance was 4.19% (Table 3.4).
36
Five species were mainly responsible in explaining the difference between deposits
1998-1997 with C. flavidus having the highest contribution (19.51%) compared to the
other species (Table 3.3) as its abundance was 6.92% in the 1998 deposit and was absent in
the 1997 (Table 3.4). Differences between the 2006 and 1994 deposit were mainly due to
the contribution of three species, namely A. javanica, Athyrium macrocarpum and Polytrias
amaura with the highest contribution made by Polytrias amaura (16.04% Table 3.3) where
it was absent in the 2006 deposit. Its abundance in the 1994 was 3.78% (Table 4).
Deposits 1998 and 1994 were characterized by four grass species, namely C.
flavidus, I. cylindrica, Panicum reptans and Polytrias amaura, with the highest
contribution made by C. flavidus (19.31%). C. flavidus was abundant in 1998 (6.92%) and
less abundant in the 1994 deposit (0.20%). Lastly, in the 1997 and 1994 deposits the
highest contribution was shown by A. javanica (13.05%) where it was more abundant
(4.19%) in 1997 compared to the 1994 deposit (2.06%).
37
Table 3.3. SIMPER result 1. Percentage contribution of species to average Bray–Curtis
dissimilarities in all pairs of sites. Only those species with a contribution to average dissimilarity of
>5% are included. The average dissimilarity value (%) is also shown for each pair of the sites.
Asterisks indicates exotic species.
2006-
2006-
1998-
2006-
1998-
1997-
1998
1997
1997
1994
1994
1994
-
16.35
9.99
8.08
5.11
13.05
10.27
9.86
-
12.25
-
-
Fabaceae
-
-
-
-
-
5.00
Cyperus flavidus*
Poaceae
23.89
-
19.51
-
19.31
-
Eupatorium riparium*
Asteraceae
6.85
7.07
5.35
-
5.31
7.55
Imperata cylindrica
Poaceae
15.75
8.21
13.06
6.72
13.07
8.16
Panicum reptans
Poaceae
13.33
-
10.99
-
11.21
-
Paspalum conjugatum*
Poaceae
-
-
-
6.48
-
-
Pinus merkusii
Pinaceae
-
7.11
-
7.83
-
-
Polygala paniculata*
Polygalaceae
-
-
-
5.87
-
5.80
Polyosma ilicifolia
Poaceae
-
7.75
5.30
-
-
7.26
Polytrias amaura
Poaceae
-
13.98
10.41
16.04
11.12
7.66
98.00
93.13
82.82
94.48
85.74
68.17
Species
Families
Anaphalis javanica
Asteraceae
Athyrium macrocarpum
Polypodiaceae
Calliandra callothyrsus
Average dissimilarity (%)
38
Table 3.4. SIMPER result 2. Average abundance percentage in each site of the selected species
from table 2 (SIMPER 1). Asterisks indicate exotic species.
Species
2006 deposit
1998 deposit
1997 deposit
1994 deposit
-
1.32
4.19
2.06
Athyrium macrocarpum
2.69
0.24
1.10
0.38
Calliandra callothyrsus
0.14
0.24
1.23
0.24
-
6.92
-
0.20
0.10
2.18
2.77
1.28
Imperata cylindrica
-
5.31
2.04
1.80
Panicum reptans
-
3.98
-
-
Paspalum conjugatum*
2.75
-
-
0.56
Pinus merkusii
1.67
-
-
-
Polygala paniculata*
0.65
-
-
1.70
Polyosma ilicifolia
-
0.32
1.89
0.24
Polytrias amaura
-
-
3.52
3.78
Anaphalis javanica
Cyperus flavidus*
Eupatorium riparium*
Different species peak in abundance at different stages of early succession (Table
3.4). E. riparium, Athyrium macrocarpum and Calliandra callothyrsus were present at all
four sites. E. riparium abundance was lowest in the 2006 deposit, increased in abundance in
the 1998 and 1997 deposits, and then decreased in the 1994 deposit. E. riparium peaked at
the intermediate stage (1997 deposit). Athyrium macrocarpum abundance was highest in
the 2006 deposit whereas Calliandra callothyrsus abundance peaked in the 1997 deposit.
A. javanica, I. cylindrica and Polyosma ilicifolia were present in all but the
youngest (2006) deposit. A. javanica and Polyosma ilicifolia reached their peak of
abundance in the 1997 deposit whereas I. cylindrica reached its peak in the more or less
middle stages of the successional time range (1998 deposit). Paspalum conjugatum and
Polygala paniculata were only present in the youngest (2006) and the oldest (1994)
deposits. Paspalum conjugatum abundance was greatest in the 2006 deposit whereas
Polygala paniculata abundance was greatest in the 1994 deposit.
39
3.3.3. Interspecific association (co-occurrence)
The number of positive associations was greater in all sites than the number of
negative associations (Figure 3.3). Positive associations were highest in 1994, lowest in
1998 and at an intermediate value in the 2006 and 1997 deposits. Negative associations
were highest in the 1997 deposit, lowest in 1998 and similar in the 2006 and the 1994
deposits.
Num ber of as s oc iation
35
31
30
25
20
16
15
10
Positive
15
Negative
11
7
5
8
5
4
0
2006
1998
1997
1994
Deposits
Figure 3.3. Total number of interspecific association of species in each deposit of primary
succession sites of Mt. Merapi
A. javanica had the largest number of negative interspecific associations compared
with other species pairs, and in fact all of its associations were negative (Table 3.5). In
contrast, Calliandra callothyrsus had the highest number of positive interspecific
associations. I. cylindrica did not exhibit any association with other species at any of the
sites. Athyrium macrocarpum co-occurred with Polygala paniculata as shown by their
strong positive association (Table 3.5). E. riparium always co-occurred with Melastoma
affine whereas Calliandra callothyrsus was present together with Cyperus rotundus
40
Table 3.5. Results of the association tests using the chi-squared test statistic (χ2) between the
differentiating species and their co-occurring species (as derived from SIMPER analysis; table 3).
Association is significant at P values 0.05. Values of the Ochiai Index (strength of association) are
equal to 0 at ‘no association’ and to 1 at ‘complete/maximum association’. Asterixis indicates the
species pair that is also included in the most differentiating species. NA stands for no associations.
Type of
Ochiai
Species
Paired species
association
Index
0.40
Debregeasia longifolia
Anaphalis javanica
0.40
Humata repens
0.40
Rubus flaxinifolius
Athyrium macrocarpum
Polygala paniculata*
Polygonum chinense
+
-
1.00
0.40
Calliandra callothyrsus
Crassocephalum crepidioides
Cyperus rotundus
Polygala paniculata*
Panicum reptans*
Eleusine indica
Polytoca bracteata
Polytrias amaura*
+
+
+
+
+
-
0.84
1.00
0.84
0.40
0.77
0.70
0.40
Eupatorium riparium
Gnaphalium japonicum
Stachytarpheta jamaicensis
Melastoma affine
+
+
0.84
0.40
1.00
Imperata cylindrica
NA
NA
NA
Pinus merkusii
Polygala paniculata*
Shuteria vestita
-
0.28
0.40
41
3.4. DISCUSSION
In the first decade of primary succession, plant re-colonization on the nuées
ardentes deposits at Mt Merapi was rapid, with 52 species belonging to 23 families
recorded. Species abundance and composition were significantly different between the
younger and the older sites. Younger deposits were dominated by species such as Athyrium
macrocarpum, Polygonum chinense, Paspalum conjugatum and Cyperus flavidus. These
species were either present at very low abundance or were not present in the older deposits.
The older deposits were dominated by species such as Anaphalis javanica, Imperata
cylindrica, Polytrias amaura and Eupatorium riparium. These species were either present
at very low abundance or were not present in the younger deposits, which indicated that the
dominance of these species is related to the development of the habitat over a period of
several years at the sites.
E. riparium, Athyrium macrocarpum, Calliandra callothyrsus, I. cylindrica,
Polyosma ilicifolia and A. javanica were the most common species, although their
abundance varied throughout the succession. Native invasive species such as A. javanica
and I. cylindrica had lower abundance in the oldest deposit. A comparable phenomenon
was also found at Mt. St Helens, where the cover of the early pioneer pearly everlasting
(Anaphalis margaritaceae) increased after the eruption but then declined in the older phase
(Dale et al. 2005c). The decrease in I. cylindrica abundance may be due to the absence of
subsequent disturbance and fire that inhibited the stimulation of flowering (Collins and Jose
2009) and perhaps competition with other dominant species such as Polyosma ilicifolia and
E. riparium in the older sites. Invasive alien species like E. riparium were recorded in all
deposits, with relatively high abundance in the older deposits. According to Heyne (1987) ,
E. riparium is a fast growing species, usually found in steep slope in a wide range of soil
42
conditions. Thus, E. riparium may have indirectly facilitated the co-occurring species such
as Gnaphalium japonicum and Melastoma affine, perhaps by assisting in stabilizing and
preventing erosion on the deposit site. However, over-domination by this invasive species
could be a problem itself. The Mistflower or Eupatorium is native to South America, and
this unpalatable and highly competitive species has become a problem elsewhere such as in
Nepal (Kunwar 2003).
Stability in species richness and diversity after a decade of succession was
observed in Mt. Usu Japan (Tsuyuzaki 1991), while in Mt. St. Helens, the stabilization rate
was different in different locations or type of volcanic disturbances (del Moral 2007). At
the Mt. Merapi sites, species richness and diversity was still increasing after a decade of
succession. Species diversity was significantly different between the younger and the older
deposits, with the oldest deposit having the highest species diversity. The differences in the
rate of change in species diversity between the Mt Merapi sites and those of Mt Usu and Mt
St Helens was perhaps due to faster plant regeneration in the warm wet tropics than in the
cool or cold high mountain chains (A. Hamblin, personal communication, 2009).
Establishment of plant from seeds on volcanic deposits depends on a range of
factors. Plant establishment on Mt. St. Helens was influenced by factors such as distance
from seed sources, species-specific dispersal capabilities, germination and growth
characteristics of colonizing species and the substrate condition (Dale et al. 2005a; Dale et
al. 2005c). Distance from the safe sites that act as a seed source also plays an important role
in influencing the rate of plant establishment in primary succession (del Moral and Wood
1993; McClanahan 1986; Titus and del Moral 1998; Walker and del Moral 2003). From the
study of primary succession in a post-mining area, McClanahan (1986) found that distance
to a seed source was best in predicting the establishment of the subsequent species. Early
primary succession on Mt. St. Helens showed that although sites located near the
43
undisturbed vegetation still exhibited low plant cover, they retained higher species richness
comparable to the species richness in the undisturbed area (del Moral and Wood 1993). The
youngest (2006) deposit on Mt. Merapi provides a good example of the initial plant
establishment conditions. This site is located at 1,220 metres above sea level on a moderate
slope, 4 kilometres from the volcano’s crater. Seed source areas are located at a close
distance and so seeds are likely to disperse easily into the site - potentially by winds or
animals (mostly bird). The two major seed source areas for the 2006 deposit were a recently
harvested area of pine forest that was burnt in 2006 by fire associated with the nuées
ardentes, and an intact forest near the deposit. Rapid re-establishment in the 2006 deposit
was perhaps due to high abundance of perennial herbs in the understorey of the pine forest
which rapidly colonized the site after the forest was burnt. These perennial herbs species
were then perhaps dispersed by winds into the deposit site. The second seed source for the
2006 site was an intact forest on the west and east side of the valley that was minimally
affected by the eruption. Hence it was most likely that some tree seedlings found in the
2006 deposit site such as Pinus merkusii (Figure 3.4) and Calliandra callothyrsus may have
come from seed dispersal from the burnt and intact forests. Safe sites that act as seed
sources for the recovery of the vegetation in the affected areas were also observed in Mt.
Guntur volcano in West Java (van der Pijl 1939; van Steenis 1972).
The early establishment patterns recorded in this study also showed that there was a
convergence of floristic composition in the older deposits. A convergence can be identified
when two succession sites become similar in their floristic composition (Baer et al. 2005;
del Moral 2007). Chances of convergence occurring also increases where there are strong
biotic interactions and also similarities in climatic and edaphic conditions (del Moral 2007;
Walker and del Moral 2003). Pairwise comparison tests between deposits generally showed
an R value of 0.9 to 1.0, indicating large differences between the communities. Yet the
44
comparison between the 1997 deposit and the 1994 deposit was R = 0.42, indicating a
moderate degree of similarity between the sites. This result indicates that there was a
similar floristic composition in these two deposits. However, it is interesting to note that the
deposits that are only one year different in age - the 1997 and 1998 deposits - were not
more similar to each other. This finding indicates that factors other than age, such as sites
characteristics, are at play in influencing the species composition. The 1998 deposit was
located at the highest elevation (approximately 1,500 metres) compared to the other deposit
sites and hence the vegetation distribution and abundance in this area would also likely be
influenced by site characteristics such as elevation and the substrate’s nutrient status (van
Steenis 1972; Whitten et al. 1996).
Figure 3.4. Seedling of Pinus merkusii on the 2006 deposit. This is the type of situation at Mt
Merapi in which recruitment from seed was taking place. Kaliadem, May 2008.
45
Plant establishment in primary succession is largely influenced by the development
of the site’s physical environment (Walker and del Moral 2003). Generally, nitrogen and
phosphorus are the most limiting essential macro nutrients in new volcanic soil materials
(Lambers et al. 2007). By means of physical weathering of new substrates over time,
phosphorus will become available for plants (Walker and Syers 1976). Nitrogen is
introduced to the ecosystem by N2 fixing organisms such as the Leguminoceae. Similar to
the primary succession on the debris-avalanche deposit at Mt. St. Helens (Dale et al.
2005c), there was also an increase in the abundance of nitrogen fixing species on the nuées
ardentes deposits at Mt Merapi. While Lupinus lepidus was the most abundant nitrogenfixing pioneer in early volcanic substrates in Mt. St. Helens (del Moral 2007), Mt. Merapi
had Calliandra callothyrsus as its nitrogen-fixing pioneer species during the early primary
succession. Calliandra callothyrsus is able to grow on a wide range of soil types, including
the moderately acidic volcanic origin soils that are a common feature in Southeast Asia
(Palmer et al. 1994). At Mt. Merapi primary succession, Calliandra callothyrsus was
present in all deposits of different ages. Its abundance was lowest in the youngest (2006)
deposit and was highest in the 1997 deposit but not in the oldest deposit (1994). The
decrease in Calliandra callothyrsus abundance was perhaps due to the dominance of grass
species Polytrias amaura in the oldest (1994) deposit. Given that nitrogen is generally one
of the most limiting essential macro nutrients in new soil materials (Lambers et al. 2007),
N2-fixing species are often regarded to have a facilitative function in succession
(Bellingham et al. 2001; Walker et al. 2003) because they can directly influence in assisting
establishment of subsequent species in the succession by providing nitrogen (Walker et al.
2003). Furthermore, the addition of organic matter by other pioneer species aids the
retention of water and nutrients to support the growth of co-occurring species (Hodkinson
et al. 2002).
46
Changes in species composition are not only driven by the changes in the physical
environment, but can also be driven by the effects of species on each other. This makes
species interaction an important indicator factor in succession and ecosystem development
(Muller 2005; Walker and del Moral 2003). Species co-occurrence observations may be
seen as the first attempt to detect species interactions, such as facilitation and inhibition,
and niche processes that structure the community (Walker and del Moral 2003;
Widyatmoko and Burgman 2006). At the Mt. Merapi sites, there was an increase in the
number of positive species associations over time. This observation might support the idea
that in the severely disturbed habitat in which primary succession occurs, the role of
facilitation will have more value for species change than competition (Callaway and
Walker 1997; Connell and Slatyer 1977; Walker and del Moral 2003).
At Mt. Merapi primary succession, Anaphalis javanica had the largest number of
negative interspecific associations with a low association index (0.4). This result may
suggest that Anaphalis javanica and Debregeasia longifolia, Humata repens and Rubus
flaxinifolius favour different environmental conditions to establish at a particular site. In
contrast, the nitrogen fixing legume, Calliandra callothyrsus, had the highest number of
positive associations (association index ranging from 0.7 to 1) with other species, such as
Cyperus rotundus and Eleusine indica grasses. This result may indicate that not only that C.
callothyrsus and its co-occurring species have similar requirements in terms of
environmental conditions, but it may also indicate the facilitative function of the legume
species (Walker et al. 2003). It is also interesting to note the absence of interspecific
association with Imperata cylindrica, which might indicate the aggressiveness of this
species in utilizing the resources, and possibly that, I. cylindrica, with its allelophatic
capability, may have inhibited other species from co-occurring together with it at a
particular site (MacDonald 2009).
47
Plant interspecific associations have also been found in other volcanic sites across
the globe. The association of Honkenya peploides, a low-growing, sand-binding pioneer,
with lyme grass, Elymus arenarius, and the lungwort, Mertensia maritima, has contributed
to the development of a relatively unstable ecosystem on Surtsey, a volcanic island in
Iceland (Fridriksson and Magnusson 1992; Thornton 2007). On the volcanic island of
Krakatau in Indonesia, the beach-creepers Ipomoea pes-caprae and Canavalia rosea, and
the grasses I. cylindrica (alang alang) or Saccharum spontaneum (glagah), have been
found to form associations in
slow-growing sand dune communities on the island
(Thornton 2007). Furthermore, on a volcanic desert of Mt. Fuji, Japan, a dwarf pioneer
shrub, Salix reinii, was clumped together and positively associated with successional tree
seedling Larix kaempferi, demonstrating its role as nurse-plant in the primary succession
(Endo et al. 2008).
As a first attempt to detect species interaction (i.e. facilitation and inhibition) the
study of interspecific association is insufficient to determine the processes and mechanisms
of species interaction during succession. However, studies of the role of species interaction
during the succession can be done more effectively by targeting those species that are
strongly associated (Myster and Pickett 1992b). If the process and mechanism of recovery
and establishment in primary succession is to be investigated, a study of long-term
vegetation dynamics with manipulative controlled research is needed.
48
Chapter 4
Community Structure and Composition along a Chronosequence in
Forests Burnt by Nuées Ardentes-Induced Fire in Mt. Merapi
National Park
4.1. INTRODUCTION
In active volcanoes, volcanic activity remains the most significant threat to forest
vegetation (Lavigne and Gunnell 2006; Whitten et al. 1996). Fire is an integral part of
volcanic disturbance and has shaped community composition in montane forests of Java
(van Steenis 1972; Whitten et al. 1996). On Mt. Merapi, the intense heat (often more than
700° C) released from nuées ardentes ignites wildfires (Bardintzeff 1984).
Intense fire most likely kills or damages some plants, but others are more persistent
and even depend on fire for their regeneration (Bond and Wilgen 1996). Some species that
inhabit mountainous area of Java and Bali exhibit this phenomenon. In 1918, Mt. Semeru in
East Java erupted violently. Falls of hot ash stripped the branches of Casuarina
junghuhniana trees, but many trunks remained erect and re-sprouted. This species is a longlived pioneer, nevertheless, as with many other pioneers, it will be replaced by other
species when there are no subsequent fire disturbances (Whitten et al. 1996). Other
examples are Albizia lopantha and Pinus merkusii. Regeneration of these species is reliant
on fire, which breaks the hard seed case and allows seeds to germinate (Whitten et al.
1996).
The montane forests of Java and Bali are not resistant to fire (Marrinan et al. 2005).
The forests are easily ignited under conditions of prolonged drought, such as when
lightning strikes oil-rich species such as Vaccinium spp. On Mt Merapi, nuées ardentes are
the primary cause of forest fire (Simon 1998; Whitten et al. 1996). Recovery of the
49
montane forest following fire is usually slow (Horn et al. 2001). Fire destroys the
aboveground part of shrubs and some surviving species may be covered with ash, which
could slow the rate of the secondary succession (Antos and Zobel 2005; Whitten et al.
1996). Severely burned areas on mountains in Java and Bali are usually characterized by
the increase in abundance of invasive species, such as alang-alang grass (Imperata
cylindrica), and also white-leaved ‘edelweis’ (Anaphalis longifolia) and bracken fern
(Pteridium aquilinum) (Whitten et al. 1996). Homalanthus giganteus is also a common
pioneer tree species that occurs during secondary succession in these areas (van Steenis
1972).
Diversity measures are important indicators of successional processes in a plant
community (Hobbs and Huenneke 1992; Hobbs and Norton 1996; Magurran 1988; van der
Putten et al. 2000; Zhu et al. 2009). There are three contrasting hypotheses concerning
species diversity in a succession. In the first hypothesis, species diversity increases
continuously as succession progresses and the ecosystem becomes more complex (Odum
1969). The second hypothesis states that, due to plant density, species diversity is highest at
the beginning of the succession and decreases gradually as the succession proceeds
(Hubbell et al. 1999). The third hypothesis divides the succession into three stages: early,
mid and late. Species diversity, according to this hypothesis, will increase during the early
succession stage, reach a maximum in the mid succession stage and decrease in the late
succession stage (Aubert et al. 2003).
One of the main objectives of community ecology is to resolve the hypotheses
concerning the relationship between species assemblages and factors that may have
influenced the composition of the community (Pan et al. 1998). Temporal and spatial
heterogeneity of the abiotic environment is strongly correlated with the variation and
heterogeneity of the floristic assembly (Ruprecht et al. 2007; Walker et al. 2003). In the
50
tropical forests of Indonesia, the importance of edaphic factors that affect the floristic
composition of the forest ecosystem has been widely studied. However, these studies were
conducted on the lowland forests of Sumatra or Kalimantan Island (Brearley et al. 2004;
Herrera and Finegan 1997; Widyatmoko and Burgman 2006). Highland and/or mountain
rain forests, especially on Java Island, remain less studied (but see van Steenis 1972). A
significant proportion of the surface of the Indonesian archipelago is covered by highlands
or (volcanic) mountain regions. In Java, they are estimated to cover 21,950 square
kilometers, or 17% of the entire area of the island (Tan 2008). Soil changes are probably
the long-term impact of volcanic disturbances and are likely to have an effect on
vegetation.
Li et al. (1999) stated that many succession theories were based on intensive work
in temperate forests. Gomez-Pompa and Vasquez-Yanes (1981) and Chazdon et al. (2007)
studied secondary succession that occurs in the tropics, however their findings were based
on work on old fields or in lowland tropical forests. Other forest types such as volcanic
tropical montane forest have received little attention (Tsuyuzaki and Hase 2005; Whittaker
et al. 1999). Furthermore, we are now acknowledging that ‘one model fits all’ is not
appropriate for all communities and ecosystems due to the complexity of each system
(Hobbs et al. 2007).
51
The aim of this chapter was to describe plant species composition and diversity
along a chronosequence of sites that had been burnt by fires caused by nuées ardentes in
the tropical montane forest of the Merapi National Park. I then compared the successional
patterns to environmental variability. The research questions were:
1) Are there any differences in species diversity, turnover, and community structure and
composition across sites of different ages?
2) Which species contributed the most to the differences in composition between sites of
different ages?
3) Are there significant differences in major soil nutrients between sites of different ages?
4) Which subset of the soil nutrients best correlates with the species assemblages in the
course of secondary succession?
52
4.2. METHOD
4.2.1. Study sites
I chose four areas that were affected by nuées ardentes fire between 1994 and 2006
and one unaffected forest area in Kaliurang as a reference site. These study sites had acidic
soil and were located at an elevation from 1000-1500 metres above sea level. Note also that
although the sites had undergone nuées ardentes fire, their soils were of volcanic origin.
Detailed descriptions of the study sites are given in Chapter 2.
4.2.2. Vegetation sampling
In April 2008, vegetation was sampled in each of the four areas/sites burnt by fire
generated by nuées ardentes in 1994, 1997, 1998 and 2006. One area of unburnt forest (the
reference site) was also sampled. The position and altitude of each site were recorded using
a GPS, and slope was measured using a clinometer. At each site, an area of approximately
2.5 hectares was chosen and five circular plots (diameter range approximately 60 metres)
were randomly placed in the chosen area. In each of these larger plots, three sets of circular
plots of 10, 5 and 2 metres diameter were nested within each other to measure trees (10
metre plots), groundcover (5 metre plots) and seedlings (2 metre plots) (Isango 2007;
Supriyadi and Marsono 2001). The species name, height and diameter of trees (dbh ≥ 10
cm) and young trees (dbh 2-9.9 cm, height ≥ 1.3 m) were recorded. The number of
understorey plants and seedlings was counted (Kent and Coker 1992). All plants were
identified to species level when possible. Identification was conducted at the dendrology
laboratory, Faculty of Forestry, Gadjah Mada University Yogyakarta, Indonesia.
Identification was done using flora books such as “The Flora of Java” (Backer and van den
53
Brink 1963) and “Mountain Flora of Java” and the results were confirmed by a botanist in
the Faculty.
4.2.3. Soil sampling and analyses
To assess possible interactions between edaphic factors and species composition,
soils were sampled from each of the sites. Three to five soil samples were collected from
random points within the 60 metre plots. Soil samples were taken from 0-20 centimetre
depth using an auger (5 cm diameter), bulked and sealed in a plastic bag and transferred to
the laboratory. After removing stones, pebbles and large pieces of plant material, the
samples were sieved by 2 millimetre mesh screen and used for further physicochemical
analysis.
Soil organic matter was determined by the Walkey and Black (1934) method. Total
nitrogen and phosphorus were estimated following the Kjeldahl procedure (Bremmer and
Mulvaney 1982). The pH of the soil sample was measured in a soil-water suspension (1:2.5
w/v H2O) using a digital pH meter. Availability of exchangeable base cations (Ca++, Na+
Mg++ and K+) were extracted from the soil using a neutral (pH 7) salt extractant of 1 M
NH4-acetate in a mechanical vacuum extractor (Suarez 1996). The analyses were carried
out in the Soil Science Department, Faculty of Agriculture Gadjah Mada University
Yogyakarta, Indonesia.
4.2.4. Data analysis
Data were analysed using multivariate and univariate statistics. Multivariate
analyses such as Non Metric Multidimensional Scaling (NMDS), Analysis of Similarity
(ANOSIM), Similarity Percentage (SIMPER), and Biota-Environment matching (BEST)Linkage trees analysis (LINKTREE) were done using PRIMER V.6 (Clarke and Gorley
54
2005). Univariate analyses such as ANOVA and post hoc tests were conducted using SPSS
package V.11.5.
4.2.4.1. Species Diversity and Turnover
Species diversity at each site was calculated using the Shannon-Wiener diversity
index. Differences in diversity between sites were tested for significance using a one-way
ANOVA. To examine short term species turnover (beta diversity), a modified Sorensen’s
community correspondence index or CCI was used (Barbour et al. 1980; Cook et al. 2005)
with the formula as follows:
CCI =
2c
a+b
Where a = the number of species present in the first community, b = the number of species
present in the second community, and c = the total number of species found in both
communities. I then calculated D, which is an index of how much a species list changes
across sites with the formula as follows (Cook et al. 2005):
D = 1 − CCI
This index ranges from 0 to 1, and a low value indicates little change in the species
composition between sites whereas a high value indicates the opposite.
4.2.4.2. Forest Structure
In order to examine the vertical structure, forest vegetation was divided into five
strata (A, B, C, D and E), as recognized for humid tropical forests (Simon 1996). Stratum A
consisted of emergent trees more than 35 metres tall. Stratum B was the main canopy layer
with trees 18-35 metres in height. Stratum C consisted of young trees 8-18 metres tall.
Stratum D consisted of shrubs and sapling (of trees) with height ranges from 1.5-5 metres.
55
Stratum E was the groundcover layer, including grasses, herbs, tree seedlings and fern allies
(Simon 1996). The number of trees, young trees (poles), sapling and shrubs that have the
characteristics of stratum A, B, C and D were noted, while the number of groundcover
species was noted for the E stratum in each of the study sites.
Importance Value Index (IVI) (Curtis and McIntosh 1950; Kent and Coker 1992)
was used to describe the quantitative structure of the community. This statistic represents
the contribution that a species makes to the community in terms of the number of plants
within the quadrats (density), its contribution to the community through its distribution
(frequency), and its influence on the other species through its dominance. Importance Value
Index was calculated for each species of tree and groundcover in each of the study sites.
The formula for tree IVI is as follow:
IVI = RD + RF + RDom
Where RD = relative density of a species, RF = relative frequency of a species and RDom =
relative dominance of a species.
Relative Density of species A
=
Number of individual of A species
x 100%
Total number individual of all species
Relative Frequency of species A
=
Frequency value of A species
x 100%
Total frequency value of all species
Relative Dominance of species A
=
Dominance value of A species
x 100%
Total dominance value of all species
56
Dominance values for a tree species were obtained by dividing the basal area of the tree
with the size of the plot (Simon 1996; Supriyadi and Marsono 2001). The IVI formula for
groundcover species (including seedlings) was similar to the tree layer but without the
calculation of relative dominance (Kusmana 1995), and so the formula is as follow:
IVI = RD + RF
Where RD = relative density of a species, and RF = is relative frequency of a species.
4.2.4.3. Variation in Community Composition
Species abundance data were square root transformed prior to all multivariate
analyses. A resemblance matrix based on a Bray-Curtis similarity index was generated as a
basis for the subsequent ordination and cluster analyses. Plant species composition and
abundance at each site were compared using non-metric multidimensional scaling
ordination (NMDS) (Clarke 1993). Statistically significant differences in species
composition and abundance between the sites were determined by analysis of similarity
(ANOSIM), which tests the null hypothesis that there is no difference in species
composition and abundance among groups (Clarke 1993). SIMPER, an analysis that
calculates the average Bray-Curtis dissimilarity between all samples, was used to identify
the species that differentiate sites (Clarke 1993).
4.2.4.4. Variation in Soil Nutrients and Relationship between Floristic Composition
and Soil Nutrients
Differences in soil nutrients between sites were tested for significance using oneway ANOVA. To match species composition and abundance patterns in each site with the
soil nutrients, the BIOENV method in BEST routine was used (Clarke and Ainsworth
1993). BEST analysis looks at abiotic variables in combination, and tries to identify a
57
subset which is sufficient to ‘explain’ all the biotic structure capable of explanation.
Correlation values (ρ) in BEST describe how strongly these best soil nutrients affected
community composition. Statistical significance testing on BEST results was done by
generating global ρ value by 999 permutations. In addition, LINKTREE analysis (Clarke
and Gorley 2005; Mitchell et al. 2008) was conducted to demonstrate which variables take
high or low values for which samples. LINKTREE take the subset of abiotic variables
identified by BEST and use them to describe how best the assemblage samples are split into
groups, by successive binary division. Each division is characterized by a threshold on
environmental variables.
58
4.3. RESULTS
4.3.1. Species Diversity and Turnover
Sixty one species belonging to 29 families were recorded in the study sites. The
highest number of species belonged to the Poaceae (10), followed by Fabaceae (9) and then
Asteraceae (6). There were significant differences in species richness between sites
(ANOVA P = 0.05, table 4.1). Species richness was lowest at the 2006 site and highest at
the 1994 site. Species richness in the reference site (undisturbed site) was much lower when
compared to the 1994 site. Species richness in the reference site was significantly lower
than in all but the 2006 site. The changes in species diversity are not as distinct as the
changes in species richness over time (ANOVA P = 0.05, table 4.1). The reference site is
significantly different to the 1994 site, but not significantly different from 2006, 1998, and
1997. The 1998 site is not significantly different from 2006 and 1997 sites and also the
1997 site is not significantly different from the 1994 site.
Table 4.1. Differences in species diversity between the burnt sites and reference site in Mt Merapi
National Park. Superscript letters (a-c) after mean values (±SD) indicates significant difference
between sites as assessed with Tukey’s HSD test. Dates are those in which the site was burnt by
fire generated by nuées ardentes.
ANOVA group/years since fire
Species richness
Species diversity
2006 site
9.20 (±1.48)a
1.91 (± 0.19)a
1998 site
14.0 (±3.39)b
2.13 (± 0.27)ab
1997 site
15.4 (±1.51)b
2.41 (± 0.19)bc
1994 site
19.4 (±2.96)c
2.7 (± 0.21)c
Reference site
10.6 (±1.67)a
2.21 (± 0.27)ab
59
Species turnover was highest (lowest species similarities) between the 2006 and the
1998 sites, and then between the 1997 and 1994 sites (Table 4.2). Species turnover between
the 1998-1997 sites was similar to the turnover between the 1994 site and the reference site.
Table 4.2. Species turnover rates (D) between pairs of sites in the chronosequence on Mt Merapi.
2006-1998
1998-1997
1997-1994
1994-Ref site
D
0.89
0.63
0.83
0.66
Sorenson Index
0.11
0.37
0.17
0.34
4.3.2. Forest Structure and Composition
In terms of vertical structure, the number of individuals found in each stratum
indicates the presence of the particular stratum in each site (Table 4.3). In the 2006 site,
stratums B, C, D, and E were recorded. The 1998, 1997 and 1994 sites also had four
stratums (B, C, D, and E) whereas in the reference site, all five stratums (A, B, C, D, and E)
were present.
Table 4.3. Number of individuals in each stratum for each site of secondary succession at Mt.
Merapi. Stratum A refers to the number of trees that are more than 35 m in height. Stratum B is
number of trees that are 18 to 35 m in height. Stratum C comprises of trees that are 8 to 18 m tall.
Stratum D is the total number of saplings and Stratum E is the total number of groundcover species.
Stratums
2006 site
1998 site
1997 site
1994 site
Ref site
A
-
-
-
-
41
B
3
5
16
25
28
C
5
45
18
17
1
D
4
10
15
6
11
E
12
20
23
25
16
In terms of quantitative structure, tree and groundcover species in the sites were
compared on the basis of the Importance Value Index, (IVI) (Table 4.4). In the 2006 site,
the tree layer was dominated by Pinus merkusii, whereas in the 1998 and 1997 site,
Homalanthus giganteus and Paraserianthes falcataria were the most important tree
60
species. In the oldest site (1994), Schima wallichii and P. merkusii were the most important
tree species whereas in the reference site, Altingia excelsa was the most important tree
species. In the groundcover layer, the 2006 site was dominated by Imperata cylindrica,
whereas Eupatorium riparium was the most important species in the 1998 and 1997 sites.
In the oldest site (1994) Brachiaria reptans was the most important species, while in the
reference site, Selaginella doederleinii was the most important species in the groundcover
layer. In the tree seedling layer, Acacia decurrens was the most dominant tree seedlings
species in the 2006 site, followed by P. merkusii. Albizia lopantha dominated the seedling
layer in the 1998 site, while in the 1997 and 1994 sites Calliandra callothyrsus was the
most important seedling. In the reference site, A. excelsa was the dominant seedling.
In addition to the IVI, Table 4.4 also shows the presence and absence of the most
important (dominant) species in each layer throughout the succession. In the tree layer, A.
excelsa and P. merkusii were present at the youngest site (2006) and then absent in the next
two older sites (1998 and 1997), and then reappeared in the oldest (1994) and the reference
site. Erythrina sp, H. giganteus, Albizia lopantha and Macaranga javanica were only
present at the 1998 site. In the groundcover layer, I. cylindrica was recorded in all four of
the burnt sites, but was more dominant in 2006, 1998 and 1997 sites than in the 1994 site,
and was absent in the reference site. In contrast, Brachiaria reptans was absent in the 2006
site and then present throughout the rest of the chronosequence and was at a very low
abundance in the reference site. Selaginella doederleinii started to appear in the 1997 and,
1994 sites and became dominant in the reference site.
61
Table 4.4. Importance Value Index (IVI), and shade tolerance for the most important species in each
stratum at each of the study sites. Asterisks indicate exotic species.
Shading
1994
Reference
Species
2006 site 1998 site 1997 site
tolerance
site
site
Trees
Acacia decurrens*
Intolerant
17.27
6.17
-
58.88
-
Albizia lopantha
Intolerant
-
33.68
-
-
-
Altingia excelsa
Tolerant
60.63
-
-
8.44
217.51
Erythrina sp
Intolerant
-
12.51
-
-
-
Homalanthus giganteus
Intolerant
-
148.65
-
-
-
Macaranga javanica
Intolerant
-
31.55
-
-
-
Intermediate
-
-
116.20
46.83
-
Intolerant
-
6.53
29.32
-
-
Pinus merkusii
Intermediate
222.09
-
-
87.85
54.05
Schima wallichii
Intermediate
-
-
104.92
89.0
21.80
Brachiaria reptans*
Intermediate
-
2.19
19.90
16.54
2.23
Eleusine indica
Intermediate
-
-
8.82
14.49
3.11
Eupatorium riparium*
Intermediate
-
55.35
57.03
-
13.43
Eupatorium odoratum*
Intermediate
-
8.69
4.70
1.95
-
Intolerant
73.77
14.18
23.51
9.55
-
Intermediate
26.83
-
2.35
7.60
-
Tolerant
-
-
0.87
0.56
61.80
Acacia decurrens*
Intolerant
99.04
13.88
-
13.57
-
Albizia lopantha
Intolerant
-
77.77
-
-
-
Altingia excelsa
Tolerant
9.52
-
11.05
7.46
33.35
Calliandra callothyrsus
Intermediate
-
-
146.56
84.04
-
Pinus merkusii
Intermediate
62.85
-
-
41.31
-
Schima wallichii
Intermediate
-
-
17.10
19.69
22.40
Paraserianthes falcataria
Parkia sp
Groundcover
Imperata cylindrica
Polygala paniculata*
Selaginella doederleinii
Seedling
62
4.3.3. Variation in Community Composition
There was clear separation between the sites as shown by the NMDS ordination
result (Figure 4.1). Plots from the youngest site (2006) were separated from plots from the
older sites (1998, 1997 and 1994), and from the undisturbed site. An analysis of similarity
(ANOSIM) test confirmed that there were significant differences in Bray-Curtis species
similarities between sites (Global RANOSIM = 0.93, P < 0.001).
Figure 4.1. NMDS of sites based on vegetation composition and abundance: 2006 site
(triangles), 1998 site (inverse triangles), 1997 site (squares), 1994 site (diamonds) and reference
site (circles).
Six pairwise comparison tests between sites (2006 and 1998, 2006 and 1997, 2006
and 1994, 2006 and reference site, 1998 and 1994, and 1994 and reference site) had an R
value of 1.0 (Table 4.5). The comparison between the 1997-1998 sites and 1997-1994 sites
had R-values of 0.72 and 0.86 and also, the comparison between 1997-undisturbed and
1998 undisturbed had R-value of 0.98 (Table 4.5).
63
Table 4.5. ANOSIM pairwise test of NMDS vegetation plots ordination. Sample statistic (Global
R): 0.93, significance level of sample statistic P < 0.001, number of permutation: 999
Groups
R Statistic
2006, 1998
1
2006, 1997
1
2006, 1994
1
2006, Reference site
1
1998, 1997
0.72
1998, 1994
1
1998, Reference site
0.98
1997, 1994
0.86
1997, Reference site
0.98
1994, Reference site
1
In Table 4.6, Eupatorium riparium contributed most to the dissimilarity between the
2006 and 1998 sites (21.27%), 2006 and 1997 sites (20.96%), 1998 and 1994 sites
(16.31%), and 1997 and 1994 sites (21.39%). Brachiaria reptans contributed most to the
dissimilarity between the 2006 and 1998 sites (9.96%) and 1998 and 1997 sites (9.89%).
Dichantium caricosum contributed most to the dissimilarity between the 2006 and 1994
sites (10.60%). Selaginella doederleinii was the most important species contributing to
dissimilarities between the reference site and the burnt sites. Imperata cylindrica was the
second most important species in the comparison between 2006 and 1998 sites and 2006
and the reference sites.
64
Table 4.6. SIMPER result. Percentage contribution of species to average Bray–Curtis dissimilarities in all pairs of sites. Only those species with a
contribution to average dissimilarity of >5% are included. The average dissimilarity value (%) is also shown for each pair of the sites. Asterixis
indicates exotic species.
Site comparison
2006
2006
1998
2006
1998
1997
2006
1998
1997
1994
and
and
and
and
and
and
and
and
and
and
1998
1997
1997
1994
1994
1994
Ref site
Ref site
Ref site
Ref site
6.21
-
6.51
-
5.07
-
-
5.98
-
-
Brachiaria reptans*
-
9.96
9.89
8.01
5.42
8.45
-
-
8.30
5.89
Calliandra callothyrsus
-
8.01
8.68
-
-
6.31
-
-
6.80
-
Dichantium caricosum*
-
-
-
10.60
7.81
7.22
-
-
-
8.08
Eleusine indica
-
-
-
8.96
6.53
-
-
-
-
5.97
Eupatorium odoratum*
-
-
-
9.84
5.29
6.42
-
-
-
7.42
Eupatorium riparium*
21.27
20.96
5.76
-
16.31
21.39
6.15
13.94
13.60
-
Imperata cylindrica
9.89
-
7.41
9.02
-
5.63
15.14
5.83
9.89
5.13
Polygala paniculata*
6.82
-
-
-
-
-
6.81
-
-
-
-
-
-
-
-
-
18.45
16.94
13.77
14.33
88.98
79.38
61.51
75.50
85.38
60.75
95.56
83.35
80.67
87.82
Species
Brachiaria paspaloides*
Selaginella doederleinii
Average dissimilarity (%)
65
4.3.4. Variation in Soil Nutrients
Soil nutrients and characteristics were significantly different (ANOVA P = 0.05)
across the system (Table 4.8). Soil in all sites had pH values around 5. Soil organic matter
increased with time since the fire, and there was also a clear pattern of increases in N
content with the age of the site. The opposite was true with P; there was a clear and
statistically significant decrease in P concentration with the age of the site. The availability
of exchangeable potassium, magnesium and sodium content dropped slightly at first and
then increased throughout the succession. Calcium also increased significantly with the age
of the site.
4.3.5. Relationship between Floristic Composition and Soil Nutrients
BIOENV analysis in BEST showed a significant correlation between species
composition and soil nutrients (ρ= 0.6, P < 0.01) at the sites. BEST results showed that the
degree to which the chosen abiotic data ‘explains’ the biotic pattern (ρ) is optimized at
0.598 for the three soil nutrients P, Na and N, and slowly decreased beyond that, as more
variables were added or reduced (Table 4.7). In short, these three soil nutrients explain
about 60% of the variation in the species data.
Table 4.7. Global test BEST. Combinations of edaphic factors that best constrain the community
composition within the Merapi National Park study sites.
∑ variables
Selections
ρ
3
0.598
P, Na and N
4
0.597
P, Na, K and N
5
0.579
P, Na, Mg, K, and N
3
0.578
P, Ca and N
2
0.571
P and N
66
Table 4.8. Summary of differences in concentrations of soil nutrients in secondary forest and the reference site on Mt. Merapi. Superscript letters after
mean values (±SD) indicates significant different assessed with Tukey’s HSD test at p = 0.05
2006 site
1998 site
1997 site
1994 site
Reference site
Soil organic matter (%)
0.75(±0.56)a
0.79(±0.09)ab
0.84(±0.22)ab
1.19(±0.19)b
1.05(±0.11)ab
Phosphorus (µg/g)
37.54(±5.17)d
23.72(±3.38)c
17.3(±2.25)bc
10.82(±3.37)b
1.05(±0.11)a
Nitrogen (%)
0.02(±0.01)a
0.05(±0.05)ab
0.11(±0.06)ab
0.15(±0.11)ab
0.21(±0.02)b
Magnesium (µg/g)
4.77(±0.53)a
2.67(±0.18)a
27.72(±1.03)b
21.29(±1.29)b
48.36(±12.56)c
Calcium (µg/g)
32.9(±2.43)a
57.78(±14.09)a
516.29(±45.49)b
567.77(±31.2)b
775.48(±77.71)c
Sodium (µg/g)
73.71(±1.92)b
28.22(±0.54)a
175.5(±11.14)c
221.12(±9.81)d
225.29(±22.7)e
Potassium (cmol(+)kg¹ֿ)
0.05(±0.02)a
0.02(±0.01)a
0.22(±0.12)a
0.14(±0.05)a
0.43(±0.2)b
5.37
5.08
5.21
5.19
5.36
pH
67
The LINKTREE diagram splits the study sites into four groups (nodes A, B, C, and
D) with abiotic inequalities. B% is an absolute measure of group differences and provides
the y-axis for the tree (Figure 4.2). LINKTREE analysis showed a large split for A
(B%=85%) that decline (split B at 69.6%) as groups got closer together. The first split (A)
in the species data was between the reference site (plots 13-15) and the burnt sites (plots 112). This had an ANOSIM of R = 0.61. It was characterized by low or high P content (P<1.71 to the left, and >0.31 to the right). Alternatively, the same split of sample is obtained
by choosing high or low Mg or Na (left to right). The next split (B) divides the 2006 site
(plots 1-3) from the other remaining burnt sites (R = 0.65), based on Ca concentrations,
which were low at the 2006 site. The next split (C) divided the 1998 sites (plots 4-6) from
the 1997 sites (plots 7-9) and the 1994 sites (plots 10-12) (R = 0.94), with five possible
explanations: Na, Mg, P, K. Split D divided the 1997 and 1994 sites (R=0.93) on the basis
of Mg, Na and P.
68
Split
A
B
C
D
B%
85
69.6
59.5
19
R
0.61
0.65
0.94
0.93
Soil factors characteristics
P<-1.71(>-0.319) or Mg>0.853(<0.691) or Na>0.806(<0.794)
Ca>-1.11(<-1.31)
Na<-1.63(>0.41) or Mg<-1.29(>0.371) or P>0.864(<0.417) or K<-0.803(>-0.401)
Mg>0.624(<0.478) or Na<0.56(>0.699) or P>0.182(<0.161)
Figure 4.2. LINKTREE diagram. 2006 site = plots 1-3; 1998 = plots 4-6 site; 1997 = plots 7-9;
1994 = plots 10-12; and the reference site = plots 13-15.
69
4.4. DISCUSSION
In the first decade after disturbance by fire there was a rapid recovery at the sites,
with 54 species belonging to 23 families recorded in the secondary forest at the study sites.
The highest number of species belonged to the Poaceae (10), followed by Fabaceae (9) and
then Asteraceae (6) (Appendix 2). Species richness and diversity increased with time since
the fire, however species richness and diversity in the reference site was not significantly
different from the youngest (2006) site. This pattern was similar to that reported in other
studies where species diversity reached its peak in older succession sites after most of the
climax species had entered the system, and then decreased with the loss of the species
present in early successional stages (Magurran 1988; Peet 1992; Zhu et al. 2009). The
results support the hypothesis of Aubert et al. (2003) that species diversity will increase
during the early succession stage, reach a maximum in the mid-succession stage and
decrease in the late succession stage. A decrease in the light availability at the forest floor
as the succession proceeds might be the cause of the decline of species diversity in the
reference site (Gomez-Pompa and Vazquez-Yanes 1981). Direct shading of overstorey
species inhibits the existence and regeneration or growth of less tolerant and intolerant
understorey species in the reference site (Lepš 1990).
There was progressive development of forest structure over time. Although all of
the burnt sites had four strata (B, C, D, E), the number of individuals in each stratum
differed. The number of individuals of stratum B (tall trees 18-35 m) was lowest in the
2006 site, greater in the older sites, but was the greatest in the reference site. The reference
site had five strata (A, B, C, D, and E) with the lowest number of individuals of stratum E
compared with the proportion of stratum E in the burnt sites. There were also differences in
the patterns of abundance of the most important species with different light requirement
70
characteristics (shade tolerant/intolerant) in the groundcover layer. The gradual decrease in
I. cylindrica (shade-intolerant species) abundance over time contrasted with the gradual
increase in the abundance of Selaginella doederleinii (shade-tolerant species), suggesting
that there was a decrease in the light availability at the forest floor as the canopy developed
and the succession proceeded.
Over the course of succession, the characteristics of species found at a site will
change (Wills 2002). In the Mt. Merapi sites, the younger sites were characterized by
shade-intolerant colonizer species with good dispersal capability. I. cylindrica is a widely
distributed invader grass that has a long record of colonising cleared lands in Indonesia
(Eussen and Soerjani 1975; Soerjani et al. 1983). I cylindrica has wide-spread rhizomes
and its seeds are wind-dispersed (Jonathan and Hariadi 1999), making it an effective
colonizer following fire (Murniati 2002). A. decurrens, however, is a nitrogen-fixing shrub
that is usually recruited after fire. At Mt. Merapi, it may have regenerated following the
nuées ardentes fire from a soil seed bank (Hardiwinoto et al. 1998; Spurr and Barnes
1980). I cylindrica and A. decurrens can also be found in other degraded areas on
volcanoes in Java, such as in Mt. Bromo-Tengger and Mt. Semeru (Anonym 2009; Whitten
et al. 1996). The species that occurred in the older sites and reference site on Mt. Merapi
were characterized as intermediate to shade-tolerant species with greater longevity. In the
older sites, A. decurrens was replaced by the leguminous tree, C. callothyrsus, which
occurred with other tree species such as Altingia excelsa. A. excelsa is a native emergent
tree species and its seedlings are shade tolerant. Older sites were also characterized by the
presence of the fern Selaginella doederleinii and the exotic invasive Eupatorium spp. in the
groundcover layer. Eupatorium spp. is a fast growing species, usually found on steep slopes
in a wide range of soil conditions and light availability (Heyne 1987).
71
Many studies have shown that generally species composition changes with time
after a fire (Clearly et al. 2006; Reilly et al. 2006; Ross et al. 2002; Spencer and Gregory
2006). The result of NMDS ordination was notable in that species composition differed
among each site, suggesting that the species composition changes with time after a fire. The
2006 and 1998 sites were different in terms of floristic composition and abundance with
highest species replacement rate when compared with the replacement rate in the other sites
(D = 0.89). Altingia excelsa, Pinus merkusii, and Polygala paniculata which were present
in the 2006 site, dropped out in the 1998 site whereas there was an increase in the number
of species from the Fabaceae family in the 1998 site with the colonization of Albizia
lopantha, Erythrina sp and Parkia sp. The increase in the number of N-fixing species from
the 2006 site to 1998 site seems to be followed by an increase in soil nitrogen. The
differences in species assemblages between the 1998 and 1997 sites (short interval) was the
lowest in all of the site pair-wise comparison, but they were still significantly different from
each other. Consistent with this, the species replacement rate was also lower (D = 0.63)
when compared with the replacement rate in the other site comparisons. Although
ANOSIM showed that the reference site and the 1994 site were significantly different, the
turnover rate between these sites was more or less the same as the rate in the 1998-1997
sites (D = 0.66). This result indicates that some of the species that characterized the
reference site, such as Altingia excelsa, Schima wallichii and Selaginella doederleinii, had
appeared earlier in the 1994 site and thus suggested convergence of floristic composition in
these sites.
In the Mt. Merapi succession, the changes in abundance of some invasive species
such as I. cylindrica, Brachiaria spp., and Eupatorium spp. are important to note. I.
cylindrica is an invasive native of south-east Asia. I. cylindrica dominated the early
succession sites, but was not recorded in older sites as it was most likely shaded out by
72
increasing canopy cover. In contrast, invasive exotic species Eupatorium spp. remained in
the system long after the fire had occurred and forest structures had developed. Eupatorium
is native to South America, and this noxious and highly competitive species has become a
problem elsewhere in Asia, such as in Nepal (Kunwar 2003). In the longer periods,
domination of invasive exotic species may limit the chance of recruitment of other native
species including seedlings of woody species, thereby reducing diversity and even changing
the successional trajectory and ecosystem functioning (Dale et al. 2005c; Hobbs and
Huenneke 1992; Raghubanshi and Tripathi 2009; Standish et al. 2009).
Changes in the physical environment, especially soil properties, can also be crucial
factors in driving the species diversity and ultimately the successional trajectory (Dzwonko
and Gawrofiski 1994; Velazquez and Gomez-Sal 2007; Whittaker 1960). Examination of
soil nutrients status generally showed improvement in soil condition as time progressed.
Nitrogen concentration was generally lower relative to P concentration in the early stage of
succession (Peet 1992; Walker and Syers 1976). At Mt. Merapi, total nitrogen content was
lowest in the 2006 site and gradually increased with time since fire. An increase in N may
also be the result of the interplay between biotic factors such as the occurrence of nitrogenfixing legumes, changes in other abiotic factors such as water content, soil pH, and soil
organic matter, or atmospheric deposition (Lambers et al. 2007; Le Brocque 1995a; Walker
and del Moral 2003). At the Mt Merapi sites, P concentration showed an opposite pattern to
N (Walker and Syers 1976), with highest concentrations in the earliest stage of succession
that declined over time. Kennards and Gholz (2001) reported that following intense fire and
extremely high temperatures (as is the case with nuées ardentes), there is a gradual increase
in extractable P. Soil organic matter showed a pattern of increase with time since fire. Fire
burns soil organic matter in the upper layer of the soil. Debano and Conrad (1978) found
significant losses in soil organic matter following intense fire. Exchangeable base cations
73
(Ca++, Na+ Mg++ and K+) generally increased with time. Capogna et al. (2009) also reported
an increase in Mg after fire in the Mediterranean basin and a gradual increase in the other
exchangeable base cations with time since fire has also been reported (Kennard and Gholz
2001). Fire burns the soil organic matter, and Mg that was previously complexed with the
organic matter is released into the soil through mineralization processes (Capogna et al.
2009).
BEST test result suggested that there was a significant correlation between species
composition and soil nutrients (ρ= 0.6, P < 0.01) with nitrogen, phosphorus and sodium
explained about 60% of the variation in the floristic composition. Furthermore, LINKTREE
analysis results indicate which soil variables take high or low values for which biotic
samples (Clarke and Gorley 2005; Mitchell et al. 2008). Phosphorus and exchangeable
cation concentrations differentiated the floristic composition in the reference site from the
burned sites. Phosphorus concentration was much lower in the reference site than in the
burned sites. The reverse was true for the exchangeable cations. It was also notable that the
increase in the number of N-fixing species such as Albizia lopantha, Erythrina sp and
Calliandra callothyrsus, Paraserianthes falcataria and Parkia sp. in the older sites appear
to be followed by an increase in soil nitrogen. Similarly, there were also increases in soil
organic matter and species richness as the burnt sites aged. In older sites, soil organic
matter is expected to be higher than in the younger sites since more accumulation or
decomposition of plant debris of the more vegetated old sites. These findings may support
the hypothesis that temporal and spatial heterogeneity of the abiotic environment is
correlated with the variation and heterogeneity of the floristic assembly or vice versa
(Ruprecht et al. 2007; Walker et al. 2003). However, it should also be noted that the
changes in community composition during succession is the result of complex abiotic and
biotic processes (Temperton et al. 2004). Hence, it would be rare for one environmental
74
variable to adequately explain floristic assemblages. Other biotic factors, such as
competition, are perhaps also at play (Le Brocque 1995b; Peet 1992; Temperton et al.
2004; Walker and del Moral 2003).
This study suggested that the Merapi forest exhibited a high resilience for site
recovery following nuées ardentes-induced wildfire with the rapid re-colonisation of
species and improvement in soil conditions in the burnt sites. However, unlike N, P
concentrations continued to decline as the sites aged. It is also important to consider the
potential problems of invasive species Eupatorium spp. as weeds, as these species remain
abundant even in the much more developed sites. These findings may have important
consequences for forest management as there is still much to learn about the capability of
alien invasive species to change soil chemical properties, which can be crucial factors in
driving the successional trajectory (Collins and Jose 2009; Dzwonko and Gawrofiski 1994;
Hughes and Denslow 2005; Velazquez and Gomez-Sal 2007).
75
Chapter 5
General Discussion
5.1. INTRODUCTION
The research presented in this study fills a gap in our understanding of an important,
though understudied, Indonesian tropical volcanic ecosystem. The aim of this study was to
describe patterns of plant community succession that occur in the tropical volcanic montane
ecosystem at Mt. Merapi, Java, Indonesia. The major findings of the research are identified
in the next sections of this chapter. The detailed results and discussions on each topic are
presented within the relevant chapters. Overall, results from this study have shown that the
succession progress is relatively rapid, as shown by the significant increase in species
richness and diversity, increase in positive species association and improvement in soil
nutrients over more than a decade of succession.
5.2. PLANT SUCCESSION ON MT. MERAPI
5.2.1. Species Re-colonization and Diversity
There was a close similarity in the number of species and families found in the
primary and secondary succession. Fifty-two species belonging to 23 families were
recorded in the four nuées ardentes deposits along the chronosequence, whereas 54 species
belonging to 23 families were recorded in the four secondary succession sites. Species
richness and diversity increase progressively from early to late successional stages in
tropical forests (Brearley et al. 2004; Pena 2003), though some studies have found that
diversity is often greatest in the mid-successional stage rather than in the late successional
stage (Aubert et al. 2003; Zhu et al. 2009). In this study, species richness and diversity
76
increased significantly with time since nuées ardentes in both primary and secondary
succession. However, species diversity decreased in the late successional stage in the
secondary succession sites, with species richness and diversity in the reference site not
significantly different from the youngest (2006) site. This result may be caused by the loss
of the species present in early successional stages due to decrease in the light availability at
the forest floor in the reference site (old secondary forest) (Pena 2003).
5.2.2. Species Composition and Structure
Shade-intolerant colonizer species with good dispersal capability have been
observed to dominate the early successional stage of succession, whereas intermediate to
shade-tolerant species with greater longevity characterize late successional stages
(Mahecha et al. 2009; Powers et al. 2009). In this study, species composition differed
significantly with site age in both the primary and secondary succession sites. In the
primary succession, younger deposits in Mt. Merapi were dominated by species such as
Athyrium macrocarpum, Polygonum chinense, Paspalum conjugatum, and Cyperus flavidus
whereas older deposits were dominated by species such as Anaphalis javanica, Imperata
cylindrica, Polytrias amaura and Eupatorium riparium. In the secondary succession sites,
I. cylindrica (alang-alang grass) dominated the sites soon after the fire and decreased in
abundance with site age. The lowest abundance of I. cylindrica was recorded 14 years after
the nuèe ardente fire. At the same time as I. cylindrica abundance declined, the invasive
shrub species E. odoratum began to dominate the sites and other tree species such as
Calliandra callothyrsus, Schima wallichii and Pinus merkusii were also recorded.
In the secondary succession, the oldest site (14 years after fire) had similar
structural characteristics (i.e. almost complete stratums and similar numbers of individuals
occupying the B stratum) as the reference site. Similarly, in 2 to 40 year-old
77
chronosequence sites in the Bolivian Amazon, stand structures also developed as the sites
aged (Pena 2003). In tropical dry forests in Costa Rica, recovery of the forest structure
comparable to those of mature forest required 4 to 5 decades of successional period
(Powers et al. 2009) therefore it is evident that the Merapi forests recover relatively rapidly.
Walker and del Moral (2009) proposed their view that stability and predictability in a
succession is maximum if there is a minimum or maximum frequent and severe
disturbance. Rapid recovery of plant communities found in this study may be due to the
ecosystems being adapted to severe and frequent volcanic disturbance.
Acacia decurrens, Altingia excelsa, Schima wallichii and Pinus merkusii were the
most important species in the tree layer. Brachiaria reptans, Eleusine indica, Eupatorium
riparium, Imperata cylindrica, Polygala paniculata and Selaginella doederleinii were the
dominant species present in the groundcover layer. The gradual decrease with time since
fire of shade-intolerant invasive species such as I. cylindrica and the opposite pattern
shown by a shade-tolerant fern species, Selaginella doederleinii, suggested that there was a
decrease in the light availability at the forest floor as the succession proceeded. This further
supports the view of a development in forest structure and canopy cover over time.
5.2.3. Exotic Invasive Species
Imperata cylindrica and Eupatorium spp. (E. odoratum and E. riparium, appendix 3
and 4) were the most important invasive species in this study. While the native invasive I.
cylindrica is not a problem at the moment because its abundance decreased as canopy cover
developed, the exotic invasive Eupatorium spp will likely continue to spread on Mt Merapi
because it tends to remain in the system even after a canopy had developed. The dominance
of Eupatorium riparium and Eupatorium odoratum in the groundcover layer of an old
secondary forest of Kaliurang (also termed as a reference site in this thesis) (Sutomo 2004),
78
the wide spread of Eupatorium riparium and Eupatorium sordidum in a closed forest of an
extinct volcanoes of Mt. Gede-Pangrango National Park in West Java (Wuragil 2009) and
the domination of Eupatorium inulifolium and Eupatorium adenophorum in Mt.
Papandayan Nature Reserve also in West Java (Setiawan 2008) support this view.
Eupatorium is native to South and Central America, and for decades, Eupatorium
spp. has gained attention as a noxious weed in Asia and Africa (Rogers and Hartemink
2000). In Indonesia, Eupatorium spp. were first introduced by the Dutch colonials in the
early 19th century and were used to reduce erosion in the large scale plantations of tea
(Camellia sinensis) and quinone (Cinchona spp.) in West Java (van Steenis 1972). Wind
dispersed seeds, an allelophatic strategy, and the capability to adapt to a wide range of soil
conditions has made this species a successful invasive weed (Grashoff and Beaman 1970;
Heyne 1987; Kunwar 2003). Eupatorium spp. also has the capability to benefit from
increased human disturbance (Kunwar 2003).
Fifty years following the eruption of the Paricutin Volcano in Mexico, Eupatorium
glabratum dominates the tephra deposits known as “arenales” and slowed the rate of the
succession (Lindig-Cisneros 2009; Lindig-Cisneros et al. 2006). Eupatorium spp. are not
exclusively a problem of high elevation only, but also in low elevation forest such as in
India and Indonesia (Swamy et al. 2000). In Indonesia, Eupatorium spp. were also found in
rubber plantations and savannas such as in Pangandaran Nature Reserve and in East Nusa
Tenggara (Prawiradiputra 2007). Invasion of Eupatorium spp. has also been a severe
problem elsewhere, such as in Nepal, where this species has been observed to have a
tendency to become over-dominant and thus decrease species diversity (Kunwar 2003;
Raghubanshi and Tripathi 2009). Hence, ecological intervention in the form of weed
management in Mt. Merapi succession should be considered.
79
5.2.4. Interspecific Association
Observation of interspecific associations in this study suggested that in a harsh
environment such as in early primary succession, plant facilitation is more prominent than
competition (Callaway and Walker 1997; Walker et al. 2003; Walker and del Moral 2003).
In the Mt. Merapi primary succession, the number of significant associations fluctuated as
succession proceeded, with positive associations being more apparent than negative
associations. Similarly, observation of interspecific association in an old field succession in
Romania resulted in no apparent trend, but instead species associations fluctuated as time
progressed (Ruprecht et al. 2007). In Mt. Merapi primary succession, the nitrogen-fixing
species Calliandra callothyrsus had the highest number of positive interspecific
associations compared with the other species, which may suggest that this N2-fixing species
has a prominent role in facilitation. In other volcanic ecosystem such as in Mt. St. Helens,
Lupinus lepidus was the most abundant nitrogen-fixing pioneer in early volcanic substrates
and was reported to have facilitated subsequent colonizers (del Moral 2007; del Moral and
Wood 1993). As a consequence of the paucity in nitrogen, leguminous species were
prominent in the succession on Mt. St. Helens.
5.2.5. Soil Nutrients
Succession involves changes in both the composition and structure of vegetation,
and changes in soil properties (Peet 1992). In a secondary succession of a tropical montane
cloud forest area in Mexico, there were significant changes with time since abandonment of
the majority of measured soil nutrients (Bautista-Cruz and del Castillo 2005). Similarly in
this study, most soil nutrient concentrations generally increased over time in the secondary
succession sites. Soil N and P showed contrasting patterns in concentration over time. In
just over a decade since fire, total N concentration increased significantly whereas the
80
reverse was true for total P concentration, which was greatest in the earliest stage and then
declined over time. The pattern of N:P found in this study was similar to what is generally
recorded during succession and therefore supports the generalization that in early years of
succession N is the limiting major soil nutrient, whereas P is limiting at the late stages of
succession (Walker and Syers 1976). Similarly, in a primary succession in Hawaii
Volcanoes National Park, N was reported to be the most important factor affecting plant
growth in the younger sites (Vitousek et al. 1993). The N concentration is expected to
increase with the invasion of nitrogen-fixing species (Walker and del Moral 2003). In this
study, invasion by nitrogen-fixing species does occur and is presumably associated with the
increase of N in the system. In contrast, in a secondary succession in a tropical montane
cloud forest in Mexico, N concentration remained low due to the absence of nitrogen-fixers
(Bautista-Cruz and del Castillo 2005). P becomes limited as the succession proceeds due to
the demanding usage by growing vegetation in each progressive stage of the succession,
and also the remaining P is also disappearing whether by erosion or becoming complexed
with other minerals (Bautista-Cruz and del Castillo 2005; Lambers et al. 2007; Walker and
Syers 1976). Lambers et al. (2007) discuss the changes in total soil P and N in a geological
time scale chronosequence whereas, in this study, the pattern of changes in total soil P and
N were observed over a very short time period.
Examination of the relationship between vegetation composition and environment,
especially soils, has been an important aspect in ecology (Pan et al. 1998; Zuo et al. 2009).
BIOENV analysis in this study showed a moderate but significant correlation between
species composition and soil nutrients, suggesting that soil organic matter, P, N, Ca++, Na+
Mg++ and K+ may be important abiotic factors in the species composition, and accounted
for approximately 60% of the species-environment relationships among the sites. In a
tropical forest in Sumatra, Indonesia, soil organic C, total N, Ca++, Mg++ and K+ were
81
reported to influence the distribution and abundance of a rare palm, Cyrtostachys renda. In
a dune succession in China, 68.1% of the species-environment relationship was correlated
with soil organic C, total N, electrical conductivity, pH, slope, very fine sand content and
soil water content (Zuo et al. 2009). However it should also be noted that the changes in
community composition during succession is the result of complex abiotic and biotic
processes and that in addition to the edaphic factors, biotic factors, such as facilitation and
competition, are also at play in shaping the community composition (Pan et al. 1998;
Temperton et al. 2004).
5.3. SIMILARITIES TO OTHER VOLCANOES IN ASIA
Knowledge gained from this study may also be relevant for other volcanoes in Asia,
as there may be similarities between the conditions on Mt. Merapi with other Asian
volcanoes in terms of the characteristics of volcanic activity, floristic composition,
environmental conditions and interaction between the biotic and abiotic factors in the
volcanic ecosystem.
Papua New Guinea has three active volcanoes, Mt. Lamington, Mt. Waiowa, and
Mt. Victory. Although nuées ardentes are not characteristic of these volcanoes, they have
an abundance of I. cylindrica and also Dysoxylum spp. in their younger deposits and
secondary succession sites (Taylor 1957). As is the case at Mt. Merapi, at Mt. Waiowa and
Mt. Lamington, there was a rapid recovery of plant communities and the communities
showed significant response to edaphic factors (Taylor 1957). In contrast, at Mt. Victory,
climate was the prominent driving factor of the succession.
82
Mt. Pinatubo in the Philippines is one of the few other examples of Asian volcanoes
which exhibit the phenomenon of nuées ardentes (Scott et al. 1996). However, information
regarding the development of its vegetation following its 1991 eruption is scarce. The
differences between Merapi and Pinatubo are that, unlike Pinatubo, Merapi is a very active
volcano with small and frequent nuées ardentes that occur between large infrequent
eruptions (Bardintzeff 1984). Once thought to be extinct, Mt. Pinatubo erupted violently in
1991 and created a local increase in carbon dioxide that rapidly boosted the regeneration of
its surviving and colonizing plants (Dale et al. 2005b).
Rapid recovery is also apparent in the Mt. Merapi ecosystem (Figure 5.1). Results
from this study have shown that the ecosystems on Mt Merapi are largely resilient to this
type of volcanic disturbance. However, as the threat of habitat and ecosystem destruction
due to the consequences of climate change and anthropogenic disturbance increases, these
volcanic-forests highlight the continuing need and importance of research on plant
community succession and restoration on a volcanic terrain in Indonesia and Asia in
general.
83
(A)
(B)
Figure 5.1. Resilience of Merapi volcanic ecosystem. (A) Kaliadem nuées ardentes deposit areas
three months after May 2006 eruption (retrieved from the internet http://omhanif.multiply.com)
(B) Kaliadem nuées ardentes deposit areas, two years since eruption (photo by author).
84
5.4. MANAGEMENT IMPLICATIONS
In line with Lavigne and Gunnel’s (2006) view that vegetation in Java is resilient to
repeated volcanic activity, the results from this study show that the Merapi ecosystem is
resilient to nuées ardentes disturbance. However, concerns regarding problematic invasive
alien species such as Eupatorium odoratum and Eupatorium riparium mean that weed
management is an important management activity in the national park. Still, there is limited
information on the control of E. odoratum and E. riparium (Norgrove et al. 2000; Yadav
and Tripathi 1985; Yadav and Tripathi 1981).
Weed management to control species such as Eupatorium spp. can be done by plant
or plant part removal, or by changing resource availability (i.e. nutrient status, water and
light) (Luken 1990). Mechanical control, primarily by mowing has been used traditionally
to suppress the spread of Eupatorium capillifolium in pasture areas. This way, the regrowth
of this species is hindered and the seed production is reduced (MacDonald et al. 1994).
Mowing and/or slashing are physically demanding and impractical on Mt Merapi so other
options are preferred, such as chemical control using herbicides. Nevertheless, cover of
Eupatorium compositifolium was reduced by only 5% one year following the application of
Dicamba + 2, 4-D (Meyer and Bovey 1991). Thus, according to MacDonald (1994) a better
result is obtained when mechanical and chemical methods (mowing and herbicide usage)
are combined, as mowing and the use of Dicamba + 2,4-D or Triclopyr + 2,4-D herbicide
successfully reduced over 94% of the regrowth of Eupatorium capillifolium.
Changing resource availability can also be used in weed management (Luken 1990)
and to be able to perform this, first we have to understand what factors limit their growth
and development. Soil moisture stress was proposed by Yadav and Tripathi (1985; 1981)
as one of the possible factors regulating the population of E. riparium. This method will not
85
be easy to implement on Mt. Merapi as the wet mountain climate retains high moisture (van
Steenis 1972). Light and competition with adult plants were considered to be the main
factors that regulate the population of E. odoratum throughout secondary succession
(Kushwaha et al. 1981; Norgrove et al. 2000; Yadav and Tripathi 1981). Hence the use of
shade, for example by planting native, fast-growing legume tree species such as Calliandra
callothyrsus or Paraserianthes falcataria, then adjusting their density once E. odoratum
has been removed, could be an important action to suppress this species and kick-start the
vegetation succession sequence (Norgrove et al. 2000). So far it has not been possible to
successfully control Eupatorium using a single method or over a short time frame – instead,
a detailed monitoring and research program to find the most suitable long-term result
management options is required (Kunwar 2003).
Anthropogenic disturbance is likely to be one of the factors that may alter a
successional trajectory (Baeza et al. 2007; Ramaharitra 2006). In other areas of Indonesia,
conversion of forest into agricultural land has increased the occurrence of I. cylindrica
(alang-alang grass) as a weed (Murniati 2002; Soerjani et al. 1983). While the results from
this study showed that I. cylindrica does not appear to be a problem in the areas affected by
the nuées ardentes, expanding highland farming activity would likely cause I. cylindrica to
become a potential problem on Mt. Merapi. Imperata can become a fire hazard as it can
easily ignite, either from extreme temperature, lightning or by fire sparked by the falling of
rocks from the upper slope of the volcano (Andrews 1983). It is also adapted to fire as the
fire stimulates flowering (MacDonald 2009). Thus I. cylindrica dominance could
potentially increase fire frequency and reduce biodiversity (Murniati 2002; Soerjani et al.
1983).
86
Therefore, the park authority should also consider these possibilities when thinking
about the future management options. Mt. Merapi may benefit from wildfire-control
management in its strategic planning, as wildfire not caused by nuées ardentes also occurs;
for example, such wildfires burnt 300 hectares of forest in 1992 and 2002 (Heru 2002).
Fuel reduction can be done by removing I. cylindrica from the infested site to reduce the
risk of fire and the construction of firebreaks to prevent the spread of wildfire. It is evident
from this study that the abundance of I. cylindrica decreased when there is a development
in the vegetation structure, therefore shading by plantation forest or agroforestry using a
fast-growing legume tree species may be more effective in suppressing I. cylindrica if this
species becomes problematic in the future (Brook 1989; MacDonald 2009; Murniati 2002;
Soerjani et al. 1983).
5.5. CONCLUSIONS AND FURTHER STUDY
Understanding succession in tropical volcanic ecosystems is important in
developing its conservation and restoration strategies (Quesada et al. 2009; Walker et al.
2007). The primary conclusion of this study is that the ecosystem on Mt. Merapi is largely
resilient to this type of volcanic disturbance, as shown by the significant increase in species
richness and diversity, increase in positive species association and improvement in soil
nutrients in just over a decade since disturbance. However, this study had also raised some
concern regarding the role of invasive alien species in the succession. Hence, intentional
ecological intervention (rehabilitation) in the form of weed management on Mt. Merapi
should be considered.
87
While this study has been based on the chronosequence approach, which assumes
space-for-time substitution, clearly there are limitations caused by non-equivalence of sites
of different ages in terms of other environmental factors. This is a frequent problem with
studies aiming to infer dynamics from spatial patterns, and it is important to consider the
possible impacts of site differences on the patterns observed. Factors such as slope and
landscape context may be important in determining the dynamics of particular sites (e.g. del
Moral and Ellis 2004). It was beyond the scope of this study to examine these other factors
in more detail, but future research should focus on elucidating their effects on the patterns
found here.
There are also a number of additional issues that should be addressed in future
research. Firstly, the role of Eupatorium and Imperata in plant community development is
important to understand, as these species are becoming more widespread in many disturbed
ecosystems and has been observed to change soil properties and cause an arrested state of
succession (Collins and Jose 2009; Lindig-Cisneros et al. 2006; Murniati 2002). Secondly,
observation of interspecific associations in this study showed that positive associations
were more apparent than negative associations - however an experimental study on species
interactions is needed in order to understand the causation behind the associations. Such an
experimental study will benefit from the available data on the observed interspecific
association because it can focus on a relatively few species that have strong associations.
Lastly, even though the results from this study showed that the ecosystem is resilient to the
nuées ardentes disturbances, escalating human activities on Mt. Merapi are likely to be one
of the factors that may alter the successional trajectory. Therefore a long-term ecosystem
dynamics study is needed to examine the successional trajectory, not only under the
influence of volcanic disturbance, but also escalating anthropogenic disturbance. Long-term
88
and experimental studies are also imperative if the process and mechanism of recovery and
establishment in succession is to be investigated.
89
APPENDICES
Appendix 1. Mean abundance of species per family in the primary succession study sites.
Asterixis indicates exotic species
Families/species
Amaranthaceae
Gomphrena celosioides*
Apiaceae
Centella asiatica
Asteraceae
Adenostema phirsutum
Anaphalis javanica
Ageratum conyzoides*
Blumea lacera
Crassocephalum crepidioides*
Dichrocephala chrysanthemi*
Emilia sonchifolia*
Erechtites valerianifolia*
Erigeron sumatrensis
Eupatorium odoratum*
Eupatorium riparium*
Galingsoga parviflora
Gnaphalium japonicum
Laggera alata
Brassicaceae
Nasturtium indicium
Commelinaceae
Aneilema nudiflorum*
Cyperaceae
Cyperus flavidus*
Cyperus rotundus*
Davalliaceae
Humata repens
Fabaceae
Calliandra callothyrsus
Dalbergia sissoo
Lespedeza junghuhniana
Uraria lagopodioides
Labiatae
Leucas lavandulaefolia
Lomariopsidaceae
Bolbitis sinuata
Life form
2006
Deposits
1998
1997
1994
-
-
-
0.1
Herb
-
-
-
1.5
Shrub
Shrub
Herb
Herb
Herb
Herb
Herb
Herb
Herb
Herb
Herb
1.6
0.3
0
0.1
0.1
0.7
-
3.1
2.3
0.1
0
0.1
1
6.1
-
0.1
22.9
0
0.1
0.5
0.4
0.3
16.4
-
9.5
0.4
0.1
0.3
0.2
1.3
5
0.1
0
0.1
Herb
-
-
0.6
-
Herb
-
-
0.2
-
Grass
Grass
0.1
57.4
-
4.5
0.4
0.5
Fern
-
-
0.1
-
Tree
Tree
Herb
Shrub
0.2
-
0.3
0.2
-
3.8
0.1
-
0.3
0.1
0.2
0.2
Herb
-
-
1.3
-
Fern
-
-
-
0.1
Herb
Shrub
Shrub
90
Families/species
Melastomataceae
Melastoma affine
Pinaceae
Pinus merkusii
Piperaceae
Piper decumanum
Plantaginaceae
Artanema longifolia
Poaceae
Brachiaria mutica*
Eleusine indica
Imperata cylindrica
Panicum reptans
Paspalum conjugatum
Paspalum longifolium
Pennisetum purpureum
Polyosma ilicifolia
Polytoca bracteata
Polytrias amaura
Setaria sp
Polygalaceae
Polygala paniculata*
Polypodiaceae
Athyrium macrocarpum
Rosaceae
Rubus flaxinifolius*
Rubiaceae
Borreria alata
Pshychotria malayana
Theaceae
Schima wallichii
Tiliaceae
Grewia sp
Urticaceae
Debregeasia longifolia
Verbenaceae
Lantana camara*
Stachytarpheta jamaicensis*
Life form
2006
Deposits
1998
1997
1994
Shrub
0.7
-
-
3.7
Tree
3.3
-
-
-
Climb
-
0.2
-
-
Herb
-
0.4
-
-
Grass
Grass
Grass
Grass
Grass
Grass
Grass
Grass
Grass
Grass
Grass
13.9
0.1
1.5
-
54.6
21.1
0.6
-
10.2
0.2
6.4
0.8
17.1
-
0.2
0.1
5.5
1.5
0.7
0.3
0.1
18.6
2.5
Herb
2.8
-
-
10.5
Fern
11.6
0.6
3.4
0.5
Shrub
-
-
0.7
-
Herb
Shrub
-
0.1
-
2.9
-
Tree
-
-
-
0.1
Shrub
1.6
-
-
-
Shrub
-
-
0.1
-
Shrub
Herb
-
0.1
0.7
-
-
91
Appendix 2. Mean abundance of species per family in the secondary succession study sites.
Asterixis indicates exotic species
Families/species
Apiaceae
Centella asiatica
Araceae
Strobilus asper
Arecaceae
Calamus sp
Asteraceae
Ageratum conyzoides*
Anaphalis javanica
Erigeron sumatrensis
Eupatorium odoratum*
Eupatorium riparium*
Eupatorium sp*
Balsaminaceae
Impatiens sp
Cyatheaceae
Cyathea contaminan
Cyperaceae
Cyperus rotundus*
Cyperus sp*
Euphorbiaceae
Homalanthus giganteus
Phyllanthus urinaria
Sauropus androgynus
Fabaceae
Acacia decurrens*
Albizia lopantha
Calliandra callothyrsus
Erythrina sp
Lespedeza junghuhniana
Leucaena glauca*
Paraserianthes falcataria
Parkia sp
Mimosa sp
Fagaceae
Lithocarpus costata
Sites
1997
1994
Life
form
2006
1998
Herb
8.6
-
1.4
7
-
-
-
-
0.6
3
Palm
-
-
-
-
0.8
Herb
Shrub
5.8
2.8
7.6
0.2
1.4
13.6
382.2
-
0.2
11.4
463
-
8.8
75.6
-
0.4
51.6
-
-
1.4
-
-
-
Tree fern
-
0.6
0.2
0.2
-
Grass
Grass
1
-
1
0.8
0
0.2
0.2
0.2
-
Tree
Shrub
Tree
0.2
14.4
0.2
0.8
2.6
0
0
Tree
Tree
Tree
Tree
Herb
Tree
Tree
Tree
Herb
20.4
0.2
6.4
1.2
4.8
1
1
0.2
-
78.6
1.2
2.2
0.4
-
2.2
8
0.2
1.6
-
0.4
-
Tree
-
1
-
2
-
Shrub
Shrub
Shrub
92
Ref. Site
Families/species
Hammamelidaceae
Altingia excelsa
Hymenophyllaceae
Trichomanes maximum*
Labiatae
Leucas lavandulaefolia
Pogostemon auricularia
Melastomataceae
Melastoma malabathricum
Melastoma sp
Meliaceae
Dysoxyllum caulostacyhum
Orchidaceae
Herminium lanceum*
Papilionaceae
Shuteria vestita
Pinaceae
Pinus merkusii
Poaceae
Andropogon citratus
Brachiaria paspaloides*
Brachiaria reptans*
Dichantium caricosum*
Digitaria longiflora*
Eleusine indica
Imperata cylindrica
Oryza granulata
Panicum reptans
Pennisetum purpureum
Polygalaceae
Polygala paniculata*
Polygonaceae
Polygonum chinense*
Polygonum paniculata*
Polypodiaceae
Athyrium dilatatum
Athyrium macrocarpum
Nephrolepsis sp
Portulacaceae
Talinum sp*
Sites
1997
1994
Life
form
2006
1998
Tree
1.6
-
0.4
0.6
17.8
Fern
-
-
-
-
4
Herb
Grass
-
-
6.8
1.4
1.6
0.4
-
Shrub
Shrub
-
2.6
7.2
-
2.6
-
-
Tree
-
-
-
-
2
Orchid
-
2.8
-
-
-
Shrub
-
3.8
-
6.8
-
Tree
4.4
-
-
6.2
3.2
Grass
Grass
Grass
Grass
Grass
Grass
Grass
Grass
Grass
Grass
181.2
0
6.8
1
84.4
3.6
72
176.8
-
146.8
44
31
126.6
37.6
17.6
-
2
100
125
19.4
68.4
35.4
9.8
-
11.2
8.6
14
67.4
-
Herb
45.8
-
2.2
14
-
Shrub
Shrub
-
5.4
0.8
-
-
-
Fern
Fern
Fern
-
-
-
4
1.6
41.6
9.6
Shrub
2.8
-
-
-
-
93
Ref. Site
Families/species
Rubiaceae
Coffea sp*
Mitracarpus villosus
Saprosma arboreum
Selaginellaceae
Selaginella doederleinii
Smilacaeae
Smilax sp*
Theaceae
Schima wallichii
Urticaceae
Laportea sinuata
Verbenaceae
Lantana camara*
Life
form
Sites
1997
1994
2006
1998
Herb
-
3
0.4
3.8
4.2
4.4
21.6
9
1
Fern
-
-
1.6
0.2
318
Tree
-
-
2.2
-
3.4
Tree
-
-
5.2
5.4
5.2
Shrub
-
-
-
-
1.6
Shrub
-
0.6
3.6
4.2
-
Shrub
94
Ref. Site
Appendix 3. Sketch of Eupatorium odoratum
95
Appendix 4. Sketch of Eupatorium riparium
96
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