Ecosystem-Based Management of Coastal Eutrophication

FACULTY OF SCIENCE
UNIVERSITY OF COPENHAGEN
Ph.D. thesis
Jesper H. Andersen
Ecosystem-Based Management of
Coastal Eutrophication
Connecting Science, Policy and Society
Submitted: 19/03/2012
FACULTY OF SCIENCE
UNIVERSITY OF COPENHAGEN
Ph.D. thesis
Jesper H. Andersen
Ecosystem-Based Management of
Coastal Eutrophication
Connecting Science, Policy and Society
Submitted: 19/03/2012
1
Ph.D. thesis
Name of department:
Department of Biology
Section:
Section for Aquatic Biology
Author:
Jesper H. Andersen
Title / Subtitle:
Ecosystem-Based Management of Coastal Eutrophication.
Connecting Science, Policy and Society.
Topic:
This thesis focuses on Ecosystem-Based Management (EBM) of
coastal eutrophication. Special attention is put on connections between science and decision-making in regard to development, implementation and revision of evidence-based nutrient management
strategies. Two strategies are presented and analysed: the Danish
Action Plans on the Aquatic Environment and the eutrophication
segment of the Baltic Sea Action Plan. Similarities and differences
are discussed and elements required for making nutrient management strategies successful are suggested.
Key words:
Eutrophication, marine, Danish waters, Baltic Sea, action plans, nutrient management strategies, adaptive management, ecosystembased management, monitoring, assessment.
Number of pages:
54 + annexes
Submitted:
19 March 2012
2
List of Contents
1:
Preface
4
Danish Summary (Resumé)
6
Summary
8
Abbreviations
12
The Danish Action Plans on the Aquatic Environment and the Baltic Sea
13
Action Plan: Two Successful Nutrient Management Strategies?
2:
Similarities and Differences Between the Danish Action Plans on the
31
Aquatic Environment and the Baltic Sea Action Plan
3:
Beyond Action Plans and Directives: Perspectives for the Future
42
4:
Conclusions: What Makes a Nutrient Management Strategy Successful?
45
5:
References
48
Annex 1: Abstract
55
Annex 2: Nutrient Discharges and Losses in Denmark 1989-2008
56
Annex 3: Curriculum Vitae for Jesper H. Andersen
60
Annex 4: Manuscripts
74
3
Preface
This thesis addresses cultural eutrophication and the management of its causes.
The objectives of the thesis are:
1. To present and evaluate data from two apparently successful evidence-based nutrient management strategies: the Danish Action Plans on the Aquatic Environment and the eutrophication segment of the HELCOM Baltic Sea Action Plan.
2. To analyse and discuss suitable management strategies based on the two case studies.
Eutrophication itself cannot be managed; only the human activities leading to nutrient enrichment and eutrophication are within reach of control. Sometimes we incorrectly speak about management of eutrophication, when we mean development and implementation of strategies to
change human behaviour with an ultimate aim of reducing direct discharges, diffuse losses
and/or emissions (to the atmosphere) of nutrients to the aquatic environment. Special focus is put
on those links between science and decision-making processes that make nutrient management
strategies effective.
The thesis is structured in the following way:


An interdisciplinary and cross-cutting synthesis with focus on the long-term implementation
and development of two nutrient management strategies, and
An annex including the peer-reviewed papers on which this cross-cutting synthesis is based.
The thesis is founded on the following publications:
1. Conley, D.J., S. Markager, J.H. Andersen, T. Ellermann & L.M. Svendsen, 2002: Coastal
Eutrophication and the Danish National Aquatic Monitoring and Assessment Program. Estuaries 25:848-861.
2. Andersen, J.H., D.J. Conley & S. Hedal, 2004: Palaeo-ecology, reference conditions and
classification of ecological status: the EU Water Framework Directive in practice. Marine
Pollution Bulletin 49:283-290.
3. Andersen, J.H., L. Schlüter & G. Ærtebjerg, 2006: Coastal eutrophication: Recent developments in definitions and implication for monitoring strategies. Journal of Plankton Research
28(7):621-628.
4. Carstensen, J., D.J. Conley, J.H. Andersen & G. Ærtebjerg, 2006: Coastal eutrophication
and trend reversal: A Danish case study. Limnology & Oceanography 51(1-2):398-408.
5. Andersen, J.H., & D.J. Conley, 2009: Eutrophication in coastal marine ecosystems: towards
better understanding and management strategies. Hydrobiologia 621(1):1-4.
6. Andersen, J.H., C. Murray, H. Kaartokallio, P. Axe & J. Molvær, 2010: A simple method
for confidence rating of eutrophication status classifications. Marine Pollution Bulletin
60:919-924.
4
7. Andersen, J.H., 2010: Eutrophication. Baltic Sea Environmental Proceedings 122:16-17. In:
HELCOM, 2010: Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment
2003-2007. Baltic Sea Environment Proceedings 122. 63 pp.
8. Andersen, J.H., P. Axe, H. Backer, J. Carstensen, U. Claussen, V. Fleming-Lehtinen, M.
Järvinen, H. Kaartokallio, S. Knuuttila, S. Korpinen, A. Kubiliute, M. Laamanen, E. LysiakPastuszak, G. Martin, F. Møhlenberg, C. Murray, G. Nausch, A. Norkko & A. Villnäs, 2011:
Getting the measure of eutrophication in the Baltic Sea: towards improved assessment principles and methods. Biogeochemistry 106:137-156.
9. Korpinen, S., L. Meski, J.H. Andersen & M. Laamanen, 2012: Human pressures and their
potential impact on the Baltic Sea ecosystem. Ecological Indicators 15:105-114.
10. Laamanen, M., S. Korpinen, U.-L. Zweifel & J.H. Andersen, in review: Ecosystem health.
Textbook chapter in “Biological Oceanography of the Baltic Sea” (Eds: P. Snoeijs, H. Schubert & T. Radziejewska).
The thesis also draws on information from the peer-reviewed reviewed book:
11. Ærtebjerg, G., J.H. Andersen & O.S. Hansen, 2003: Nutrient and Eutrophication in Danish
Marine Waters. A Challenge to Science and Management. National Environmental Research
Institute. Roskilde. 126 p.
The production of this thesis has been financially supported by DHI (RK 2006-2009). This thesis
would not have been possible without the support of the DHI-NTU Water & Environmental Research Centre in Singapore.
Special thanks are due to Daniel J. Conley for discussion of an earlier version of the thesis and
linguistic corrections as well as to Gunni Ærtebjerg, Jørn Kirkegaard and Mette Olesen.
Thanks are extended to J. Borum, J. Brøgger Jensen, J. Carstensen, U. Claussen, B.W. Hansen,
J.W. Hansen, H. Karup, S. Korpinen, M. Laamanen, J.E. Larsen, J.-M. Leppänen, O. Mark, C.
Murray, F. Møhlenberg, S. Pedersen, J.D. Petersen, J.B. Reker, B. Riemann and A. Stock.
Jesper H. Andersen
19 March 2012
5
Danish Summary (Resumé)
Denne afhandling handler om ’eutrofiering’ og forvaltning af årsagerne til eutrofiering.
Udledning og tab af næringsstoffer til vandmiljøet fører til eutrofiering, et ord som har sine rødder i to græske ord: ‘eu’ som betyder ‘godt’ og ‘trope’ betyder ’næret’. Den moderne brug af
ordet ‘eutrofiering’ er relateret til tilførsler og effekter af næringsstoffer i økosystemer, specielt
næringsstofberigelse af akvatiske økosystemer. Det er med de danske kystvande som med mennesker, der spiser for godt: for megen mad fører til en dårlig sundhedstilstand. Velkendte symptomer på, at ’patienten’ ikke har det godt er bl.a. algeopblomstring og iltsvind.
Formålet med afhandlingen er udover at præsentere og diskutere principperne for ’adaptiv management’ og ’ecosystem-based management’, at (1) analysere og diskutere to videnbaserede
forvaltningsstrategier, at (2) identificere hvad der som minimum skal til, for at strategiske miljøhandleplaner kan gennemføres med succes.
Strategiske miljøhandleplaner og egentlige forvaltningsplaner med fokus på næringsstoffer og
næringsstofforurening i vandmiljøet, skal som udgangspunkt inkludere samtlige menneskelige
aktiviteter og kilder, hvorfra der bliver udledt eller ’tabt’ næringsstoffer. Disse indbefatter landbrug, industri, husholdninger og energiproduktion og -forbrug.
To forvaltningsplaner, som vurderes at være bedste praksis i forhold til udvikling og gennemførelse af strategi- og forvaltningsplaner til nedbringelse af tilførslerne af næringsstoffer til vandmiljøet og til begrænsning af eutrofieringseffekterne i hhv. danske farvande og Østersøen, analyseres. Den danske forvaltningsplan er vandmiljøplanerne fra 1987, 1998 og 2004, som altovervejende er et eksempel på ’adaptiv management’. Den anden forvaltningsplan er Østersøhandlingsplanen fra 2007, der er baseret på ‘ecosystem approach to management of human activities’, der i
princippet er identisk med ’ecosystem-based management’.
Afhandlingen beskriver de faktorer, der er kritiske, hvis en strategi eller forvaltningsplan på sigt
skal blive en succes og føre til væsentlige reduktioner af udledninger og tab af næringsstoffer til
vandmiljøet. Nøglefaktorer er progressive reduktioner af udledningerne af næringsstoffer, generelt både kvælstof (N) og fosfor (P), landsdækkende overvågning af tilførsler til og effekter af
næringsstoffer i vandmiljøet samt politisk vilje til at fastholde oprindeligt fastsatte mål.
Politisk fokus og vilje til handling på området er afgørende. Om reduktionsmålene er tilstrækkelige, er tilsyneladende ikke afgørende for, om en strategi eller forvaltningsplan bliver en succes –
i hvert fald ikke så længe både analyse og rapportering af den gennemførte overvågning og regelmæssig evaluering af udviklingen i tilførslerne af næringsstoffer har en central plads i strategien. Overvågningen skal føre til årlig opgørelse af tilførslerne og årlige vurderinger af miljøtilstanden. Vurderingerne skal i sagens natur udarbejdes af faglige institutioner og evalueres af
beslutningstagere eller den statslige administration, der traditionel understøtter politikerne i fagligt komplicerede spørgsmål.
Der er flere forskellige måder, hvorpå økosystem-baserede forvaltningsstrategier og -planer kan
udarbejdes. Den enkelte plan skal tage udgangspunkt i det pågældende områdes karakteristika.
Forvaltningsstrategier og -planer vil desuden variere afhængigt af den anvendte lovgivning og
hvilke myndigheder, der er involveret. De to forvaltningsplaner har mange lighedspunkter, men
er på en række centrale punkter væsentligt forskellige, bl.a. med hensyn til geografisk dækning
6
og politisk ophæng: Den ene er national og gennemføres med lovgivning. Den anden er international og gennemført med en politisk aftale.
Afhandlingen konkluderer at følgende elementer er afgørende for at forvaltningsstrategier og
-planer kan blive gennemført med succes:




Miljø- og reduktionsmålene skal være klare for alle parter.
Effekter og udledninger skal løbende overvåges.
Strategien eller planen skal indeholde: 1) målfastsættelse, 2) virkemidler, 3) overvågning og
4) evalueringsfase.
Den politiske vilje skal være til stede.
Et forslag til hvordan disse elementer kan kobles, er illustreret nedenfor:
Konceptuel model for tilrettelæggelse og gennemførelse af en økosystem-baseret forvaltningsstrategi.
Bemærk at relevante eutrofieringseffekter er indeholdt, både forhøjede koncentrationer af næringsstoffer,
direkte effekter (eksempelvis primærproduktion og klorofyl-a-koncentrationer) og indirekte effekter (for
eksempel iltsvind og ændring i mængde og udbredelse af bundlevende dyr og planter).
Fremtidige forvaltningsstrategier og -planer vil ikke skulle opbygges fra grunden af, men bygge
på det eksisterende. Første trin vil være at tage det bedste fra ’adaptive management’ og kombinere det med det bedste fra ’ecosystem-based management’. Konkret vil det væsentligst omfatte
processen fra førstnævnte og den økosystem-baserede tilgang fra sidstnævnte. Den måske mest
kritiske faktor med hensyn til, om fremtidige forvaltningsplaner kan føre til en god miljøtilstand,
er, at de ikke bliver udarbejdet og gennemført for snævert, både hvad angår fokus og personkreds. Fremtidige forvaltningsplaner bør derfor være videnbaserede, involvere alle relevante
interessenter og nyde opbakning af en vedvarende politisk vilje til at begrænse tilførslerne af
næringsstoffer til vandmiljøet.
7
Summary
This thesis concerns eutrophication and the management of human activities resulting in nutrient
enrichment and the biological effects on aquatic ecosystems.
The term ‘eutrophication’ (noun) has its root in two Greek words: ‘eu’ which means ‘well’ and
‘trope’ which means ’nourishment’. The modern use of the word eutrophication is related to high
inputs and effects of inorganic nutrients in ecosystems, especially over-enrichment of aquatic
ecosystems.
Management is basically the process of getting people together to accomplish desired goals and
objectives. The verb ‘manage’ comes from the Italian ‘maneggiare’ (to handle), which originally
derives from the Latin ‘manus’ (hand). In the context of eutrophication, management is about
setting up a strategy for control of human activities resulting in discharges (direct sources), losses (diffuse sources, e.g. from agriculture) and emissions (to the atmosphere) of nitrogen, phosphorus and organic matter to the aquatic environment.
An adaptive nutrient management strategy (NMS) should include the following elements: Problem identification and four phases focusing on planning, acting, checking and evaluation. The
papers on which this thesis is based upon addresses all of these five elements. (See Table 1).
Paper no.
Context
Planning
Acting
Checking Evaluation
X
1. Conley et al. (2002)
(x)
(x)
X
2. Andersen et al. (2004)
(x)
X
3. Andersen et al. (2006)
(x)
X
4. Carstensen et al. (2006)
(x)
(x)
(x)
X
5. Andersen & Conley (2009)
(x)
X
6. Andersen et al. (2010)
(x)
(x)
X
7. Andersen (2010)
(x)
(x)
X
8. Andersen et al. (2011)
(x)
X
9. Korpinen et al. (2012)
(x)
X
10. Laamanen et al. (submitted)
(x)
(x)
‘X’ = the paper has a direct focus upon this NMS phase; ‘(x)’ = the paper has an indirect focus.
Understanding the context of eutrophication is important both from a scientific point of view,
since both definitions and conceptual understanding are constantly developing, and from an
implementation of nutrient management strategies. If decision-makers are not informed or do not
understand the concept of eutrophication then management is a difficult task. Despite a widespread common European understanding of causes and effects of eutrophication, there is no
mutually agreed definition of coastal eutrophication. However, within the European Union (EU)
there has been a sound tradition of focusing the measures on the sources causing eutrophication
(Elliot et al. 1999, Elliot & de Jonge 2002). Consequently, eutrophication has been defined in
relation to sources and/or sectors. For example, in the EU Urban Waste Water Treatment
Directive, eutrophication has been defined as “the enrichment of water by nutrients, especially
nitrogen and/or phosphorus, causing an accelerated growth of algae and higher forms of plant
life to produce an undesirable disturbance to the balance of organisms present in the water and to
the quality of water concerned” (Anon. 1991a). The EU Nitrates Directive has an almost
identical definition specifically emphasising losses of nitrates from agriculture (Anon. 1991b).
8
Nixon (1995) defined eutrophication as “an increase in the rate of supply of organic matter to an
ecosystem”. This definition is short and emphasizes that eutrophication is a process, not a trophic
state. Nixon also noted that various factors may increase the supply of organic matter to coastal
systems, but the most common is clearly nutrient enrichment. The supply of organic matter to an
ecosystem is not restricted to pelagic primary production, even though such an interpretation
makes the definition operational. The supply of organic matter to a system includes primary
production of higher plants and benthic microalgae as well as inputs of organic matter from adjacent waters or from land via rivers or point sources. Having such a broad interpretation of the
term ‘supply’ makes the definition difficult to use in a monitoring and management context.
Eutrophication and definition(s) of eutrophication are widely discussed (Jørgensen & Richardson
1996). The most common use of the term is related to inputs of mineral nutrients, in particular
nitrogen and phosphorus, to specific waters. Consequently, eutrophication deals with both the
process and the associated effects of nutrient enrichment and natural versus cultural
eutrophication. Despite the definitions in existing European directives, the implementation of the
EU Water Framework Directive (WFD) revealed a need for a common understanding and
definition of eutrophication as well as stronger co-ordination between directives dealing directly
or indirectly with eutrophication. Hence, the European Commission convened a process aiming
for a development of a pan-European conceptual framework for eutrophication assessment in the
context of all European waters and policies (Anon. 2009a). This process did not lead to a
common European definition of eutrophication, but it revealed that if ‘undesirable disturbance’ is
understood as ‘unacceptable deviation from reference conditions’, the pan-European definition
will be coherent with the normative definitions sensu the WFD (Andersen et al. 2006).
Accepting this, a pan-European definition of eutrophication, would be:
“the enrichment of water by nutrients, especially nitrogen and/or phosphorus, and organic matter, causing an increased growth of algae and/or higher forms of plant life to produce an unacceptable deviation in structure, function and stability of organisms present
in the water and to the quality of water concerned, compared to reference conditions”.
The suggested definition includes causative factors (nutrient enrichment), primary effects
(increased growth) and secondary effects (sometime referred to as ‘undesirable disturbance’).
However, it also is a matter of interpretation, in particular in regard to what an ‘acceptable deviation’ is.
In addition, the definition enables classification of ‘eutrophication status’. Using the definition as
a basic assessment principle, an eutrophication quality objective or target (EutroQO) is defined
as an indicator with an acceptable deviation (AcDev) from the reference condition (RefCon),
EutroQO = RefCon ± AcDev (Andersen et al. 2004, Andersen et al. 2011). As an additional
feature, the definition also acknowledges that eutrophication has both quantitative and qualitative
perspectives, an aspect not included in Nixon’s definition.
The setting of science-based eutrophication quality objectives (or targets) is a prerequisite for
ecosystem-based management. These target setting principles used in Europe are commonly
based of information on RefCon and setting of an AcDec from RefCon. The concept originates
from the Water Framework Directives and is described and demonstrated by Andersen et al.
(2004). The strength of the concept is that it is operational and that the data used derived by
science-based process. The weakness is that is allows for expert judgement, e.g. in regard to the
9
setting of AcDev, and thus potentially a weakening of the scientific basis. A specific problem is
related to natural variability and it potential influence of the assessment of eutrophication status
(Andersen et al. 2004).
Before implementing a nutrient management strategy, having a complete overview of the human
activities and pressures is necessary to focus on the activities resulting in impaired conditions.
An example can be found in Korpinen et al. (2012), where cumulative pressures and impacts in
the Baltic Sea region have been estimated. The estimate is based on the mapping of human
activities, maps of key ecosystem components and expert judgement of the impact of a specific
pressure upon a specific ecosystem component. A matrix is established and from it, the dominant
pressures in the Baltic Sea were estimated to be: (1) nutrient enrichment, (2) fishing activity, (3)
input of contaminants, and (4) physical modification. This study is the first ever assessment of
cumulative pressures and impacts for a regional sea, and is a useful tool for documenting the
causes of impaired conditions as well as targeting of measures, regionally and sub-regionally.
The targets of nutrient management can in principle be established in two ways: (1) the traditional way where load reduction targets are agreed upon, and (2) a more modern way where Eutrophication Quality Objectives (EutroQO’s) are established and the critical loads matching the
EutroQO’s are calculated. Two different case studies are analysed in this thesis. The Danish Action Plans on the Aquatic Environment have a strong focus on load reduction targets for agricultural discharges and losses as well as discharges from urban water treatment plants and industries with separate discharge (Conley et al. 2003, Carstensen et al. 2009, Andersen & Conley
2009). The HELCOM Baltic Sea Action Plan, which is based on an ecological target and subsequent calculation of a critical load, represent a more evidence-based way to estimate the load
reductions (Andersen et al. 2011).
A key step in any nutrient management strategy is monitoring for expected improvements in
ecological quality, specifically eutrophication status, in the marine environment. The Danish
Action Plans on the Aquatic Environment included a well-designed monitoring programme for
Danish marine waters (Conley et al. 2002, Ærtebjerg et al. 2003). The data and information originating from monitoring activities have not only resulted in annually national reports, which have
been used for regular evaluations of the effectiveness of the Danish Action Plan, but also in papers of eutrophication trends (Carstensen et al. 2006). Both the reductions in loads and the effects of the loads reductions in Danish coastal waters are well documented: (1) inputs have decreased significantly, both for nitrogen and phosphorus, (2) nutrient concentrations have decreased significantly, (3) primary productivity and phytoplankton biomass have decreased as
well, and (4) benthic communities have in some areas improved their ecological status.
The work on assessing eutrophication in Danish marine waters and the Baltic Sea has lead to
important advances in our understanding. For decades, eutrophication assessments have focused
on state for a given indicator supplemented by temporal trend assessment for individual
indicators. Recently, multi-metric indicator-based assessment tools are emerging (Andersen
2010, Andersen et al. 2011). With the development of the HELCOM Eutrophication Assessment Tool (HEAT) (Andersen et al. 2011), status assessments can now be supplemented with a
simple estimate of confidence (Andersen et al. 2010). The approach to solve a statistical
challenge in a non-statistical way is based on expert judgement of the confidence of information
in regard to RefCon, AcDec and observations of the state. This information is combined for each
indicator and integrated into an overall estimate of confidence. Information in regard to confi-
10
dence estimates is useful for setting up evidence-based nutrient management strategies, but also
essential when redesigning monitoring programmes.
Based on the lessons from the Danish Action Plans on the Aquatic Environment and the
HELCOM Baltic Sea Action Plan, the following DO’s and DON’T’s of evidence-based nutrient
management strategies can be made:


DO understand that ecosystem-based management is adaptive and science-based.
DON’T assume that decisions can not be taken because of incomplete knowledge and
uncertainty.


DO evidence-based target setting and exhaustive planning, the latter involving decisionmakers, authorities and all stakeholders.
DON’T wait for perfection and all-inclusive ecosystem understanding.


DO a full execution of the plan.
DON’T rely on voluntary agreements or guidelines.


DO monitoring with ecologically relevant resolution in time and space.
DON’T underestimate resources needed for sampling, quality assurance, analysing data and
reporting.


DO regular evaluations in regard to the progress of the nutrient management strategy.
DON’T disregard the advantages of a dual monitoring strategy focusing on both nutrient
inputs as well as ecological responses to lowered nutrient inputs.
An important lesson learned from the Danish Action Plans on the Aquatic Environment and the
HELCOM Baltic Sea Action Plan is that decisions are often made in short windows of
opportunity. It is critical to prepare for those brief moments where decisions and actions can be
taken. Preparation of decision support systems and determining the best possible scientific basis
for decision-making can provide the scientific basis for actions to be implemented.
Perhaps the most important lesson is that time is needed before the effects of changes in human
behaviour can be observed in nutrient inputs and, eventually, in the ecological quality of the
marine environment. It would, therefore, be prudent to ask if we, within a decade or two, can
expect to have a marine environment not affected by eutrophication as required by national and
international processes, e.g. the Danish Action Plan on the Aquatic Environment, the EU Water
Framework Directive, the EU Marine Strategy Directive and the HELCOM Baltic Sea Action
Plan. There are a number of factors underlying the slow response of ecosystems, e.g. delays in
nutrient inputs from fields to streams and rivers caused by retention in the soil. There is a growing recognition that the recovery trajectories differ from the well known degradation trajectories (Duarte et al. 2009; Laamanen et al. submitted). Another challenge is a shifting baseline
caused by increasing temperatures, resulting in a situation where loads of nutrients probably
have to be reduced more than estimated in a situation with stable temperatures (Laamanen et al.
submitted). Apparently, we face two counteracting process, one where nutrient loads are progressively reduced, and one where sea temperatures are rising. The prospects in regard to the
long-lasting eutrophication crisis are not good due to the lack of political will to act, a fact being
sustained by the current financial crisis.
11
Abbreviations
AcDev
Acceptable deviation
AM
Adaptive Management
APAE
Action Plan on the Aquatic Environment
BSAP
HELCOM Baltic Sea Action Plan
COMBINE
Cooperative Monitoring in the Baltic Marine Environment
DAMP
Danish Aquatic Monitoring Programme
EBM
Ecosystem-Based Management
EC
European Community
EEZ
Exclusive Economic Zone
EU
European Union
EutroQO
Eutrophication Quality Objective
HEAT
The HELCOM Eutrophication Assessment Tool
HELCOM
Helsinki Commission
NMS
Nutrient Management Strategy
MSFD
Marine Strategy Framework Directive
N
Nitrogen
ND
Nitrates Directive
NGO
Non-Governmental Organisation
NOVA
Nationalt program for overvågning af vandmiljøet (DAMP 1998-2003)
NOVANA
Nationalt overvågningsprogram for vandmiljøet og naturen (DAMP 2004-2015)
NPo
Nitrogen, phosphorus and organic matter
OSPAR
Oslo and Paris Commissions
P
Phosphorus
PACE
Plan, act, check and evaluate
RBMP
River Basin Management Plan
RefCon
Reference conditions
RT
Reduction target
TN
Total nitrogen
TL
Target load
TP
Total phosphorus
UWWTD
Urban Waste Water Treatment Directive
UWWTP
Urban waste water treatment plant
WFD
Water Framework Directive
12
1: The Danish Action Plans on the Aquatic
Environment and the Baltic Sea Action Plan:
Two Successful Nutrient Management Strategies?
Cultural eutrophication of coastal waters has been recognised as a growing global problem for
more than three decades. Many resources have been allocated for eutrophication research and
eutrophication mitigation around the world, especially in Europe and North America. The causes, processes, and effects of eutrophication are well documented (e.g. Cloern 2001, Kononen &
Bonsdorff 2001, Rabalais & Nixon 2002, Bachmann et al. 2006, Diaz & Rosenberg 2008, Andersen & Conley 2009). However, very few examples of successful nation-wide or regional
nutrient management strategies (NMS) are published (e.g. Carstensen et al. 2006, Kronvang et
al. 2008). This raises a series of questions: (1) Do we have a common conceptual understanding
of what eutrophication and nutrient management strategies are about? (2) For decades, NMS’s
have been planned and implemented, but why have plans in general not resulted in significant
improvements? (3) Do we have to wait for the effects of already implemented actions or is it
possible that we have a structural defect preventing our plans from succeeding?
A hypothesis in this thesis is that any successful NMS or action plan is characterised by the confluence of the following four steps: (1) a politically agreed plan including objectives and targets,
(2) implementation of measures, (3) monitoring activities including publication of assessments,
and (4) appropriate feedback loops from monitoring and assessment to the political level (back to
step 1). Direct testing of this working hypothesis is not possible. Instead this thesis analyses and
discusses two apparently successful nutrient management strategies, (1) a national action plan
based on Adaptive Management (AM), and (2) a trans-national action plan, aiming to be based
on the principles of ecosystem-based management (EBM).
The two action plans differ substantially. The one based on the AM approach has a long history,
while the one based on the EBM approach has had a long prologue, but is in the early phases of
its implementation. Both plans are believed to be representative in regard to AM and EBM, respectively, and they are analysed and discussed with the aim of highlighting which factors an
evidence-based NMS might include in order to be successful. The first example of an apparently
successful nutrient management strategy are the Danish Action Plans on the Aquatic Environ-
13
ment (APAE), where the first of, to date, three consecutive plans was adopted in 1987. The information about APAE originates from a combination of governmental publications and White
Papers as well as peer reviewed papers, e.g. Iversen et al. (1999), Grant et al. (2006), Carstensen
et al. (2006) and Kronvang et al. (2008). It should be pointed out that APAE 1 does not have any
reference since the first APAE 1 is a combination of a proposal from the Danish Government
(Miljøministeriet 1987) and changes to it decided by a majority of the Danish Parliament (Folketinget 1986-1987). The second nutrient management strategy is the HELCOM Baltic Sea Action
Plan (BSAP) which covers an entire regional sea. The information about the HELCOM BSAP
originates from the BSAP itself (HELCOM 2007) and a suite of peer reviewed papers, e.g. Savchuk & Wulff (2007), Wulff et al. (2007), Backer (2008), Backer & Leppänen (2008) and Backer at al. (2009). In addition, information has been extracted from HELCOM (2009). The Danish
marine waters as well as the neighbouring regional seas, the wider North Sea and the Baltic Sea
are shown in Figure 1. It should be noted that the Kattegat and the Danish Straits, being the transition zone between the North Sea/Skagerrak and the Baltic Sea, as well as the marine waters
around the island of Bornholm, are included in both plans.
Figure 1: Map of the Baltic Sea and North Sea including the transition zone consisting of the Skagerrak
between Denmark, Norway and Sweden, the Kattegat and the Danish Straits (between the main Danish
islands west of southern Sweden). EEZ = Exclusive Economic Zone.
14
Other nutrient management plans were considered in the analysis. Based on the available scientific literature and combination of criteria (plans should focus on both point and diffuse sources,
not operate on a local scale, and have been enacted), only the above introduced nutrient management strategies were selected cf. Table 1.
An additional reason for focusing on the Danish APAE’s and the HELCOM BSAP was that it is
well documented that nutrient enrichment is the key pressure followed by fisheries, inputs of
heavy metal and inputs of persistent organic pollutants (HELCOM 2010, Korpinen et al. 2012).
Similar rankings of cumulative pressures are not available for the other areas where nutrient enrichment may be an issue1 though it can not be excluded that other pressures are more important.
Table 1: An overview of potential successful nutrient management strategies including criteria for final
selection. P = point sources; D = diffuse sources; NAT = national plan or strategy; REG = regional plan or
strategy; LOC = local plan or strategy; UW = union-wide; and UWWT = urban waste water treatment.
CBA = Chesapeake Bay Agreement; GHAP = Gulf Hypoxia Action Plan; and VLSL = Venice Lagoon
Special Law.
Plan/Strategy
Adopted
Predecessor
P/D
Scale
Enacted Reference
Danish APAE
1987
NPo Action Plan
P+D
NAT
+
ATV 1990
1
HELCOM BSAP
2007
HELCOM 50%
P+D
REG
Indirectly
HELCOM 2007
HELCOM 50%
1988
None
P+D
REG
÷
Laäne et al. 2002
OSPAR 50%
1988
None
P+D
REG
÷
OSPAR 2008
Chesapeake Bay
2000
CBA 1983, 1987
P+D
REG
÷
Bosch et al. 2001
Gulf of Mexico
2008
2011 GHAP
P+D
REG
÷
Anon. 2008
Venice Lagoon
1992
VLSL 1973
P+D
LOC
+
Suman et al. 2005
EC Nitrates Dir.
1991
None
D
UW
+
Anon. 2010
EC UWWT Dir.
1991
None
P
UW
+
Anon. 1991b
1: The HELCOM Baltic Sea Action Plan is indirectly enacted via the EC Urban Waste Water Treatment
Directive and the EC Nitrates Directive as eight out of nine coastal states are EU Members States.
2: ‘Eutrophication non-problem area’ cf. OSPAR and not designated as a sensitive sensu the UWWTD.
3: The upstream catchment is designated as a ‘nitrogen vulnerable zone’ sensu the Nitrates Directive.
The Danish APAE’s and in particular the HELCOM BSAP are likely to represent best practices
in regard to European nutrient management strategies (Foden et al. 2008). No other successful
national or regional plans have been identified, although the upcoming WFD River Basin Managements Plans (RBMPs) include many features of evidence-based and adaptive nutrient management strategies (Foden et al. 2008).
1
The Chesapeake Bay, United States may potentially be considered a successful nutrient management strategy
(based on information in Bosch et al. (2001), Kemp et al. (2005), and Bosch (2006)), but detailed information on
cumulative pressures and impacts sensu Halpern et al. (2008) is not available.
15
Having an understanding of the meaning of the terms ‘Adaptive Management’ (AM), and ‘Ecosystem-Based Management’ (EBM) is important for two reasons: first, and from a societal point
of view, because policy drivers and management frameworks are being continuously developed
and updated, and second, because scientists should have knowledge regarding legislative and
political processes to understand that decision-making has to balance recommendations from
scientists with societal needs.
‘Adaptive Management’ (noun) is a
structured, iterative process of best possible decision making in the face of uncertainty, seeking to reduce uncertainty
over time via system monitoring. AM is
often characterized as "learning by doing” and depends upon an open management process which seeks to include past,
present and future stakeholders, for example those sectors discharging, losing
or emitting nutrients to the environment.
Hence, AM is characterised as being both
a social and a scientific process. In its
basic form, AM includes four phases: (1)
a planning phase, (2) an action phase, (3)
a checking phase, and (4) an evaluation
phase. This sequence is on occasion
named PACE (Figure 2).
AM is linked to the ‘Ecosystem Approach to management of human activities’ (EA). EA (noun) is defined as ”the
Figure 2: Conceptual model of the classical Adaptive
Management cycle including five phases: Identification
(Do we have a problem?) and the Plan, Action, Check
and Evaluate loop, the four last phases sometimes
grouped under a PACE heading. Feedback loops exist
from evaluation to implementation and planning, the
latter including adoption of additional measures.
.
16
comprehensive integrated management
of human activities based on the best
available scientific knowledge about the
ecosystem and its dynamics, in order to
identify and take action on influences which are critical to the health of marine ecosystems,
thereby achieving sustainable use of ecosystem goods and services and maintenance of ecosystem integrity” (HELCOM & OSPAR 2003). Hence, the Ecosystem Approach can be seen as a
fore-runner of ‘ecosystem-based management’ as indicated by Backer et al. (2009).
‘Ecosystem-Based Management’ (noun) is an integrated approach to management that considers
the entire ecosystem, including humans with the goal of maintaining an ecosystem in a healthy,
productive and resilient condition so that it can provide the services humans want and need
(McLeod et al. 2005). An important element in regard to Ecosystem-Based Management (EBM)
is the term ‘ecosystem’ (noun), which is “a dynamic complex of plant, animal and microorganism communities and their non-living environment interacting as a functional unit”, cf. the
UN Convention on Biological Diversity.
EBM differs from current approaches that focus on a single species, sector, activity or concern; it
considers the cumulative impacts of different sectors. Specifically, EBM: (1) emphasizes the
protection of ecosystem structure, functioning, and key processes; (2) focuses on a specific ecosystem and the range of activities affecting it; (3) explicitly accounts for the interconnectedness
within systems, recognizing the importance of interactions between many target species or key
services and other non-target species; (4) acknowledges interconnectivity among systems, such
as between air, land and sea; and (5) integrates ecological, social, economic, and institutional
perspectives, recognising their strong interdependences (Christensen et al. 1996, McLeod et al.
2005). At its core, EBM is about acknowledging linkages between ecosystems and human societies, economies and institutional systems (McLeod & Leslie 2009).
EBM and AM share a lot of common ground, especially in regard to the Action and Checking
phases. The most prominent differences between EBM and AM are related to the Planning
phase, which in regard to EBM is principally evidence-based since it is system-oriented and
based on the best available knowledge, and to the Evaluation and feedback phase which is a key
phase in the AM process, while it is not being given specific emphasis or is an integrated part of
the current EBM concept. Another key differentiation is that AM generally focuses on sectors
and their pressures, whilst EBM aims on setting ecologically relevant targets.
17
1.1: The Danish Action Plans on the Aquatic Environment
In early autumn of 1986, large parts of the Danish Straits and estuaries were depleted of oxygen.
Danish fishermen showed dead Norwegian lobsters on national television and thereby demonstrated to the public and politicians that the environmental status of the marine waters of Denmark was severely impaired (ATV 1990, Andersen & Carstensen 2011).
The public communication of the dead lobsters is generally considered to be the catalyst of the
Danish Action Plans on the Aquatic Environment (APAE). However, in reality the issue of eutrophication emerged gradually during the 1970s, with the Belt Project as one of the activities
launched to document the extent and severity of the effects of nutrient enrichment on marine
waters in Denmark (Ærtebjerg Nielsen et al. 1981).
The Belt Project, taking place 1975-1978, was the first Danish National Marine Research Programme focusing on nutrient enrichment and its associated effects. It showed that nutrient concentrations, primary production and phytoplankton biomass were increasing and that oxygen
concentrations were decreasing, but also that there were no general problems related to nutrient
enrichment in the open waters (Ærtebjerg Nielsen et al. 1981). This conclusion was questioned,
since a number of incidents occurred in the summer and early autumn of 1981, where fish and
benthic invertebrates were killed by oxygen depletion (Miljøstyrelsen 1984a). When the Belt
Project ended in 1978, parts of it continued as an ongoing activity named the National Marine
Pollution Monitoring Programme. Focus was on monitoring of nutrients, phytoplankton and oxygen in the open parts of the Inner Danish Waters (the Kattegat, the Danish Straits and the south
western part of the Baltic Sea).
The report “Oxygen depletion and fish kills in 1981” (Miljøstyrelsen 1984a), primarily being
based on local monitoring activities, and the NPo White Paper (Miljøstyrelsen 1984b) put focus
on an emerging and extensive problem related to nutrient enrichment and derived consequences
in Danish Marine waters. As a consequence, the Ministry of Environment developed the NPo
Action Plan from 1985 focusing on discharges of nitrogen (N), phosphrous (P) and organic matter (o) from point sources, including the setting of discharge limits in effluents from urban waste
water treatment plants (Folketinget 1984-1985). Discharges and losses of nutrients from agriculture were not taken into consideration, because of uncertainties in regard to different sources and
pathways. Linked to the plan was a national research programme, the NPo Research Programme,
18
focusing on discharges to groundwater and fresh and marine surface waters, the effects of these
discharges, and possible alleviating activities.
In 1986, a few days after the lobsters were killed by hypoxia, the Danish Society for Nature Conservation (DN), the largest nature conservation and environmental NGO in Denmark, held its
annual assembly. A resolution was adopted urging the Minister of the Environment, counties and
municipalities to substantially reduce loads from waste water treatments plants immediately and
to lessen losses from other activities, e.g. agriculture (ATV 1990). Almost in parallel, the Minister of the Environment developed and launched an “Action Plan for the Marine Environment”
(APME) for consideration and eventual adoption by the Parliament (Folketinget 1990). This
APME should be seen as a proposal from a Government without a majority in the Parliament —
hence, the plan was subject to political negotiations and eventual adoption.
On 18 November 1986, a majority in the Parliament forced the minority government, by adopting an official Parliamentary Agenda, to: (1) guarantee that all illegal discharges from municipal
waste water treatment plants, industries and agriculture would be brought to an end before 1 May
1987, and (2) to issue a Governmental Action Plan, including a plan for investments, aiming at a
reduction of N and P discharges with 50% and 80%, respectively, to be presented before 1 February 1987 (Folketinget 1986-1987).
The Governmental Action Plan, published as “Action Plan Against Pollution of the Danish
Aquatic Environment with Nutrients” (Miljøministeriet 1987), was based on the Action Plan on
the Marine Environment taking the November agenda into account and was amended by the
Danish Parliament. Hence, no publication exists. The closest we come to a first Danish Action
Plan on the Aquatic Environment (APAE 1) reference is the combination of the Governmental
Action Plan (Miljøministeriet 1987) and the summary of changes adopted by the Parliament
(Folketinget 1986-1987). On the basis of a sequence of political agreements adopted by the Parliament’s Environmental and Planning Committee in April 1987, the Government pushed forward a legislative process leading to adoption of a series of laws, including funding of monitoring (1988-1991 and onwards) and research. In addition a suite of derived statutory orders implementing specific elements of the APAE 1 were published. An overview can be found in ATV
(1990) and Miljøstyrelsen (1990).
19
The publication of the first nation-wide assessment of the state of the aquatic environment in
1990 (Miljøstyrelsen 1990) marks the end of the first cycle of the APAE. Hence, APAE 1 includes all four phases of AM: (1) Plan, (2) Act, (3) Check, and (4) Evaluate, collectively setting
the PACE not only for APAE 1, but also for the subsequent action plans.
The planning phase of APAE 1 focused strongly on the overarching aim of the APAE 1: 50%
reduction compared to the levels in the mid-80’s of nitrogen discharges and losses from agriculture, urban waste water treatment plants (UWWTP), and industries with separate discharge. For
phosphorus the aim was to reduce discharges by 80%. Diffuse losses of phosphorus from agriculture were not included in APAE 1 owing to inaccurate data on this source. The Danish Parliament agreed on reduction targets (RT), reduction percentages (%) and target loads (TL), cf.
Table 2. It should remembered that the Action Plan from 1987, despite well-known uncertainties, was based on an estimated annual loss of nitrogen from agriculture in the order of 260,000
tonnes TN. The reduction target (RT) was set to 127,000 tonnes corresponding to a reduction
percentage of 49%. Consequently, the resulting target load (TL) was set to be 133,000 tonnes.
Table 2: Danish nutrient reduction targets sensu the Action Plans for the Aquatic Environment 1 and 2.
Baseline is 1987; reductions and targets were agreed by the Danish Parliament in 1987 and subsequently
adjusted in 1990 (for UWWTP’s) and1999 (for industries). Units = tonnes per year. See Ærtebjerg et al.
(2003) and Carstensen et al. (2006) for details.
Sector
Total nitrogen loads (TN)
Total phosphorus loads (TP)
1987 ÷
RT
%
= TL
1987 ÷
RT
%
= TL
Agriculture* 260,000 ÷ 127,000
49 = 133,000
4,400 ÷ 4,000
91 = 400
UWWTPs
18,000 ÷ 11,400
63 = 6,600
4,470 ÷ 3,250
73 = 1,220
Industries
5,000 ÷
3,000
60 = 2,000
1,250 ÷ 1,050
84 = 200
Total
283,000 ÷ 141,400
50 = 141,600
10,120 ÷ 8,300
82 = 1,820
UWWTPs: Urban wastewater treatment plant effluents. RT: Reduction target. TL: Target loads. * Agricultural loads of phosphorus only concerns direct discharges from farms; diffuse losses of phosphorous
were not included.
The act phase of APAE 1 focused on implementation of (1) measures to reduce nutrient discharges, losses and emissions, (2) a nation-wide aquatic monitoring programme as well as (3)
two research programmes. It is beyond the scope of this thesis to go into details of the full package of laws and statutory orders enacted under APAE1. However, one piece of legislation – the
so-called ‘aktstykke’2 - is of particular interest since it laid down the economic basis of the Danish Aquatic Monitoring Programme (DAMP). DAMP was one of the most comprehensive na2
’Aktstykke’ translates to a legal document, agreed by a majority in the Parliament, which secures sustained funding until the Parliament decides to terminate the funding.
20
tional aquatic monitoring programmes designed and carried out, with an action phase focusing
on (1) inclusion of all relevant sources (point and diffuse, the latter including atmospheric deposition) and compartments (groundwater, lakes, streams and rivers, as well as marine water), (2)
coordination and documentation of all strategies and methods, (3) regional reporting, nationwide thematic assessments, and nation-wide integrated assessment, (4) regular evaluations and
revisions and (5) securing of funding (Folketinget 1986-1987, Indenrigsministeriet 1988,
Miljøstyrelsen 1989, Kronvang et al. 1993, Conley et al. 2002).
Two research programmes were initiated as part of the act phase, one on urban waste water
1987-1992 and another on marine eutrophication 1990-1994 (for details, see PH-Consult ApS
(1993), Jørgensen & Richardsson (1996), Christensen et al. (1998)). Aside from the underestimated feature of sustained funding, the DAMP is unique because the periodic cycle of design,
sampling, evaluation and revision of monitoring activities. The monitoring programme that came
out of the planning and actions of APAE 1 (Miljøstyrelsen 1989 and Kronvang 1993) has resulted in four follow-up programmes: (1) DAMP 1993-1997, NOVA 1998-2003, NOVANA 20042009, the latter including a so-called “half-way tuning” effected from 2007, and NOVANA
2010-2105 (see Miljøstyrelsen 1993, Miljøstyrelsen 2000 and Svendsen et al. 2004 for details).
The collaboration between the local partners (the counties) and the national partners (e.g. the
National Environmental Research Institute) has resulted in an accumulation of knowledge leading to many scientific publications, e.g. Conley et al. (2002), Kronvang et al. (2005), Carstensen
et al. (2006) and Kronvang et al. (2008). In addition to national assessment reports, the programme has provided data for regional marine conventions such as HELCOM (e.g. HELCOM
2009) and OSPAR (e.g. Ærtebjerg et al. 2003 and OSPAR 2008).
The direct link from monitoring and assessment activities to evaluation of the action plan(s) was
already incorporated in APAE 1. The first evaluation of APAE 1 revealed that the reduction targets for discharges of nitrogen and phosphorus from urban waste water treatment plants and industries were likely to be met in the mid-1990s. The targets for discharges of both nitrogen and
phosphorus from urban waste water treatment plants were met in 1996 and 1995, respectively,
and are today below the targets of APAE 1 (Figure 3). For industries with separate discharge,
the targets for both nitrogen and phosphorus were met in 1996 and 1995, respectively, and are
today far below the targets (Figure 4).
21
A
B
2007
2005
2003
2001
1999
1997
1995
1993
1989
Mid 80's
2007
2005
0
2003
0
2001
2
1999
5
1997
3
1995
10
1993
5
1991
15
1989
6
Mid 80's
20
1991
Phosphorus, 1000 tonnes TP
Nitrogen, 1000 tonnes TN
Figure 3: Panel A: Temporal trend from the mid 1980’s to 2008 in nitrogen discharges from Danish urban
waste water treatment plants to surface waters. Panel B: Temporal trend from the mid 1980’s to 2008 in
phosphorus discharges from Danish urban waste water treatment plants to surface waters. Data courtesy
of the Danish Nature Agency – see Nordemann Jensen et al. (2010) and Annex A2.1 for details.
B
7
1600
6
1400
Phosphorus, tonnes TP
5
4
3
2
1
1200
1000
800
600
400
200
0
2008
2007
2005
2004
2003
2002
2001
2000
1999
1998
1997
1996
1995
1994
1993
1992
1991
1989
2008
2007
2005
2004
2003
2002
2001
2000
1999
1998
1997
1996
1995
1994
1993
1992
1991
1990
1989
0
1990
Nitrogen, 1000 tonnes TN
A
Figure 4: Panel A: Temporal trends 1989-2008 in nitrogen discharges (tonnes of TN) from Danish industries with separate discharge to surface waters. Panel B: Temporal trends 1989-2008 in phosphorus discharges (tonnes of TP) from Danish industries with separate discharge to surface waters. Data courtesy
of the Danish Nature Agency – see Nordemann Jensen et al. (2010) and Annex A2.2 for details.
The discharges of nitrogen and phosphorus from urban waste water treatments plants and industries with separate discharge are estimated by a specific sub-programme of DAMP (see page 1819). All industries are monitored in the same way and the estimated discharges are considered
accurate. However, it is noteworthy that data is reported with uncertainty. This might be a negligable deficiency, since the contribution from industries with separate discharge is small. For urban waste water treatment plans, the loads are estimated by a complex combination of measure-
22
ments and statistical models (depending on size and type of UWWTPs). Hence, uncertainties
could be considerable and it would seem well-justified to analyse potential uncertainties.
The recognition of success of the plan in relation to point sources has lead to a very strong and
well-justified focus on the dominant source of nitrogen, the diffuse losses from agriculture. In
total, there have been three follow-ups on APAE 1:
1. The 1991 Action Plan of Sustainable Agricultural Development (APAE 1½).
2. The 1998 APAE 2, as a combined consequence of feedback from the DAMP activities, implementation of the Nitrates Directive (Anon. 1991b) and a momentous oxygen depletion
event in Mariager Fjord (Fallesen et al. 2000).
3. The 2004 APAE 3.
Estimating losses from agriculture is more difficult to determine then it is for the discharges from
point sources. However, the nitrogen losses from agriculture can be calculated from the overall
nitrogen balance from agriculture, especially the estimate of the nitrogen surplus (Figure 5).
Nitrogen, 1000 tonnes
800
600
400
200
N fertilizer
Animal fertiliser
Sludge
N-fixation
Deposition
20
07
20
05
20
03
20
01
19
99
19
97
19
95
19
93
19
91
19
89
19
87
19
85
0
Harvest
Figure 5: Nitrogen input and export (harvest) in Danish agriculture 1985-2008. For details in regard to
the nitrogen surplus, please see Nordemann Jensen et al. (2010) Annex A2.3.
Despite the clear trend in nitrogen surplus, it should be remembered that the APAE nitrogen target for agriculture was related to losses from fields. These losses from the root zone have been
monitored and assessed by a specific sub-programme of DAMP. Here, the strategy has been to
intensively monitor six small-scale agricultural catchments. The trends in nitrogen losses from
the root zone in monitored agricultural catchments are outlined in Figure 6.
23
180
160
N loss (kg N/ha)
140
LOOP6
120
LOOP2
100
LOOP3
80
LOOP4
LOOP7
60
LOOP1
40
20
0
1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008
Figure 6: Estimated losses of nitrogen (kg/ha) in six Danish agricultural catchments (LOOP areas) 19912008. From Nordemann Jensen et al. (2010). The weighted reduction is presented in Annex A2.4.
The selection of the agricultural catchments (LOOP areas) and the monitoring activities are welljustified. However, the sub-programme cannot be fully representative given the variation in soil
types and agricultural activities in Denmark. It is difficult to grasp why a programme taking spatial and temporal variations into account has not been developed. An answer, which is speculative, is that the sub-programme is unpopular with stakeholders living within the catchments and
politically sensitive. Further, the sub-programme is faced with another challenge. The up-scaling
from relatively few catchments using simple statistical models, to a nation-wide estimate leaves
considerable room for improvement. No assessments of uncertainties in the annually estimated
losses have been published throughout the APAE period. Considering that these estimates are the
basis for evaluation of whether Danish agriculture has met the targets for reduction of nitrogen
discharges according to the APAE’s, it is beyond understanding that obvious ways to reduce
uncertainties, e.g. by setting up a spatially representative sampling programme and by using
complex models, have apparently been ignored.
It would be prudent to mention that uncertainties in regard to the nitrogen losses were dealt with
in connection with the evaluation of APAE 2 and setting up of APAE 3. It was documented that
the original estimated losses of nitrogen from cultivated fields were underestimated (Grant &
Waagepetersen 2003). Hence, the loss was corrected to 311.000 tonnes of TN, or 20% higher
compared to APAE1. Table 3 is based on Table 2 but updated according to APAE 3.
24
Table 3: Danish nutrient reduction targets sensu the Action Plans for the Aquatic Environment 1, 2 and
adjusted sensu Action Plan 3. From Andersen & Conley (2009), based on Carstensen et al. (2006).
Sector
Agriculture*
UWWTPs
Industries
Total
Total nitrogen loads (TN)
1987 ÷
RT
%
= TL
311,000 ÷ 152,400
49 = 158,600
18,000 ÷ 11,400
63 = 6,600
5,000 ÷
3,000
60 = 2,000
334,000 ÷ 166,800
50 = 167,200
Total phosphorus loads (TP)
1987 ÷
RT
%
= TL
4,400 ÷ 4,000
91 = 400
4,470 ÷ 3,250
73 = 1,220
1,250 ÷ 1,050
84 = 200
10,120 ÷ 8,300
82 = 1,820
Hence, a correction factor of 1.2 can be used for ‘normalising’ APAE 1, the Action Plan for Sustainable Agricultural Development and APAE 2 with APAE 3 – such correction enabled direct
comparisons between APAE 3 and its predecessors and can be seen in Table 4. In the columns
labelled “old”, the figures are based on APAE 1, while the columns labelled “new” includes figures corrected by a factor of 1.2. The revised values are 127.000 or 152.400 tonnes, cf. Tables 2
and 3. Please note that the evaluation of APAE 1 is shown under APAE 1½ as “old”, the evaluation of APAE 1½ is shown under APAE 2 as “old” and the evaluation of APAE 2 is shown under APAE 3 as “old”.
Table 4: Summary of planned and implemented reduction targets in regard to nitrogen discharges and
losses from agriculture according to the Danish Action Plan 1 from 1987, the subsequent Action Plan for
Sustainable Agricultural Development from 1991 (here named APAE 1½) and the follow up Action Plans
2 and 3 from 1998 and 2004. Unit = tonnes TN per year. See Ærtebjerg et al. (2003) for evaluation results (red numbers) as well as details about the specific measures under the Action Plans 1, 1½ and 2.
Action Plan 1
Action Plan 1½
Action Plan 2
Action Plan 3
Grand total
APAE 1
Old
New
127,000 152,400
—
—
—
—
—
—
127,000 152,400
APAE 1½
Old
New
50,000
60,000
77,000
92,400
—
—
—
—
127,000 152,400
APAE 2
Old
New
—
—
89,900 107,880
37,100
44,520
—
—
127,000 152,400
APAE 3
New
—
—
—
—
113,460
38,940
152,400
The first evaluation phase of the APAE 1 as well as the evaluation of subsequent action plans
provides excellent examples of adaptive management. This is evident in the way information
from the monitoring programme was analysed, compared to the goals of APAE 1, synthesised
and eventually directed to the Minister of the Environment for political considerations and subsequent tightening of the measures needed. The last outcome of this evidence-based AM loop
included in the thesis is the 2008 mid-term evaluation of APAE 3 (Waagepetersen et al. 2008).
As word of caution it should be noted that APAE 2 introduced a measure related to reconstruction of wetlands. Consequently, the evaluation criteria not only include losses from the root zone,
25
but also retention in wetlands and streams. Whilst the estimation of losses from the root zone is
made by DAMP, the estimation of the retention in wetlands and streams has a more uncertain
origin. No nation-wide technical guidance is available, instead a variety of estimates ranging
from the use of empirical models to complex hydrological and biogeochemical models are used.
The 50% reduction target for nitrogen was estimated to be met via the full implementation of
APAE 3 (Waagepetersen et al. 2008). The legitimacy of this estimate is yet to be determined as it
is based on the monitoring of agricultural watersheds, which apparently is a non-representative
data set. However, the estimate is indirectly confirmed by sub-program of DAMP dealing with
inputs to marine waters (Nordemann Jensen et al. 2010, Carstensen et al. 2006).
1.2: The Baltic Sea Action Plan
The countries surrounding the Baltic Sea have since 1974 joined forces in order to safeguard the
Baltic Sea environment and to coordinate mitigatory efforts. The framework for this work is the
‘Convention on the Protection of the Marine Environment of the Baltic Sea Area’ – known as the
Helsinki Convention. The governing body is the Helsinki Commission, which is responsible for
the coordination of activities and day-to-day work.
Nutrient enrichment and eutrophication were dealt with for the first time at a high political level
by the Ministers of the Environment of the Baltic Sea States at a Ministerial Meeting in 1988.
The first Danish Action Plan on the Aquatic Environment played a key role in regard to this Baltic Sea-wide adoption of the 50% reduction target, which were stated in the 1988 Ministerial
Declaration:‘…efforts on reduction of the load of pollutants should aim at a substantive reduction of the substances most harmful to the ecosystem of the Baltic Sea, especially of heavy metals
and toxic and persistent organic substances, and nutrients for example in the order of 50 percent
of the total discharges of each of them, as soon as possible, but not later than 1995’ (HELCOM
1988).
The Helsinki Convention was revised in 1992 in order to embrace the changed geopolitical situation. The new convention also became more explicit in regard to eutrophication, e.g. by including an annex with a specific focus on the needs for reducing water-borne nutrient inputs from
point and diffuses sources. In order to support the implementation of the reduction targets agreed
upon at the 1988 Ministerial Meeting, the Baltic Sea Joint Comprehensive Environmental Action
Programme (JCP) was established in 1992. Identification and elimination of pollution Hot Spots
26
was an important part of this work, and initially 132 Hot Spots were identified including both
municipal wastewater treatment plants and agricultural ‘sites’. In 2002, an evaluation of
achievements revealed that the 50% reduction target for the time period from 1987 to 1995 had
been achieved for phosphorus discharges from point sources by almost all countries, while most
countries had not reached the targets for nitrogen (Lääne et al. 2002). Agricultural loading levels
showed smaller decreases than point-source loading despite the fact that almost all countries in
transition3 had achieved the 50% target for phosphorus. However, accurate estimates of changes
in agricultural loading were hampered by a lack of monitoring data. Further estimation of
achievements between 1985 and 2000 showed that as a result of improved treatment of industrial
and municipal wastewaters, nutrient discharges from point sources had greatly decreased.
The reduction targets for diffuse sources such as agriculture were not fulfilled (HELCOM 2009).
Hence, it remained clear that eutrophication was still of concern. The HELCOM Bremen Ministerial Meeting Declaration of 2003 demanded further actions, in particular in the agricultural
sector, to reduce diffuse nutrient loads. In addition, HELCOM was tasked to implement an ecosystem approach to the management of human activities and the idea of developing ecological
objectives with indicators was put forward.
In 2006 HELCOM adopted a system of ecological objectives with the specific strategic goal for
eutrophication of a ‘Baltic Sea unaffected by eutrophication’ defined by five specific ecological
objectives: (1) ‘concentrations of nutrients close to natural levels’, (2) ‘clear water’, (3) ‘natural
levels of algal blooms’, (4) ‘natural distribution and occurrence of plants and animals’, and (5)
‘natural oxygen levels’. To make these ecological objectives operational, indicators with initial
target values were agreed upon reflecting a good ecological and environmental status of the Baltic marine environment. Thus, the target values, when achieved, are intended to represent good
ecological or environmental status. It has subsequently been agreed that the ecological objectives
for eutrophication will be measured by the following indicators: (1) winter surface concentrations of nutrients, reflecting the ecological objective ‘concentrations of nutrients close to natural
levels’; (2) Chlorophyll-a concentrations, reflecting the ecological objective ‘natural level of
algal blooms’, (3) Secchi depth, reflecting the ecological objective ‘clear water’, (4) depth range
of submerged aquatic vegetation, reflecting the ecological objective ‘natural distribution and
occurrence of plants and animals’, (5) abundance and structure of benthic invertebrate communi3
Estonia, Latvia, Lithuania and Poland.
27
ties, reflecting the ecological objective ‘natural distribution and occurrence of plants and animals’, and (6) area and length of seasonal oxygen depletion, reflecting the ecological objective
‘natural oxygen levels’. More information in regard to the operationalization of the above indicators can be found in Andersen et al. (2004, 2006), HELCOM (2006), and HELCOM (2009).
To have a more targeted approach to address the symptoms of eutrophication, it was considered
necessary to have nutrient reduction targets taking into account both ecosystem functioning and
sub-regional differences. A model-based approach employing sub-regional targets related to selected ecosystem features such as water transparency was established (Wulff et al. 2007). Following the principle of adaptive management and in order to implement the ecosystem approach
to the management of human activities, HELCOM coordinated the development of the Baltic
Sea Action Plan (BSAP).
Initial estimates of nutrient reductions needed to reach the target levels for eutrophication were
produced by the MARE program (Wulff et al. 2007). In addition, scenarios were considered to
examine how far the full implementation of existing HELCOM Recommendations, as well as
EU legislation and programmes, would bring the Baltic Sea towards the agreed ecological objectives for eutrophication, using the target ‘clear water’ as a basis. These results produced by
MARE were used to develop specific reduction targets and actions related to reducing nutrient
loading to the BSAP. Hence, the BSAP defines maximum nutrient loads that will allow
achievement of eutrophication targets for the whole Baltic Sea and each of its sub-basins. The
required reductions in nutrient loads were estimated based on the objective ‘clear water’, modelling of maximum allowable nutrient loads matching the objective, and average nutrient load levels from 1997 to 2003. It was acknowledged that the maximum allowable nutrient loads and the
country-wise allocations of the BSAP were based on the best knowledge at the time and that
review and revision of the figures should start as soon as the BSAP was adopted. By using an
evidence-based target, the BSAP is partly, but not completely based on the Ecosystem Approach.
The BSAP contains measures estimated to be sufficient to reduce eutrophication to a target level
that would correspond to good ecological/environmental status by the year 2021 (HELCOM
2007). It was estimated that nutrient load reductions of 135,000 t of nitrogen and 15,250 t of
phosphorus from average annual nutrient loads (based on loads during the period 1997–2003)
would be needed. Quantitative reduction requirements were applied to each of the sub-basins and
28
provisional allocations of nutrient reduction requirements to each HELCOM country and to
transboundary loads were included in the BSAP. The main bulk of reductions were to be made in
the Baltic Proper, while the Gulf of Bothnia was at that time considered to have good ecological/environmental status and thus not in need of reductions. It was estimated that the reductions
would result in achieving the eutrophication-related targets on water transparency, primary production and nutrient concentrations (Wulff et al. 2007). The inputs to and outputs from the
MARE/NEST calculations on maximum allowable loads are summarised in Table 5. Time delays in achieving good ecological status were presumed to be significant, on the order of decades
due to long residence times, even in the case that all nutrient reductions were made immediately
(Savchuck & Wulff 2007).
Table 5: Maximum allowable annual loads of phosphorus and nitrogen to achieve ‘good environmental
status’ (calculated for water transparency) and corresponding minimum load reductions (in tonnes) calculated per sub-basin (based on HELCOM 2007).
Basin
Bothnian Bay
Bothnian Sea
Gulf of Finland
Baltic Proper
Gulf of Riga
Danish Straits
Kattegat
Sum
Maximum allowable
nutrient loads (tonnes)
TP
TN
2,580
51,440
2,460
56,790
4,860
106,680
6,750
233,250
1,430
78,400
1,410
30,890
1,570
44,260
21,060
601,710
Inputs in 1997–2003
(normalized)
TP
TN
2,580
51,440
2,460
56,790
6,860
112,680
19,250
327,260
2,180
78,400
1,410
45,890
1,570
64,260
36,310
736,720
Table 6: Country-wise nutrient load reduction allocations, in
tonnes (from HELCOM 2007).
Country
Denmark
Estonia
Finland
Germany
Latvia
Lithuania
Poland
Russia
Sweden
Transboundary pool
Total
Reductions
TP
16
220
150
240
300
880
8,760
2,500
290
1,660
15,016
TN
17,210
900
1,200
5,620
2,560
11,750
62,400
6,970
20,780
3,780
133,170
Needed reductions
(interim allocation)
TP
TN
0
0
0
0
2,000
6,000
12,500
94,000
750
0
0
15,000
0
20,000
15,250
135,000
The country-wise allocation of reductions is summarized in Table 6.
In order to reduce nutrient inputs to
the Baltic Sea to the maximum allowable level the countries have also
agreed to take actions not later than
2016 to reduce the nutrient load
from waterborne and airborne inputs
aiming at reaching good ecological
and environmental status by 2021.
29
Meeting the 2007 BSAP targets by 2021 will be a difficult and would be a significant achievement. However, it should be recognised that the Baltic Sea States have already reduced input of
nutrients significantly, especially for phosphorus (Figure 7).
The BSAP does not include a well
described checking phase (Figure 8).
However, this is not an issue since
monitoring is already dealt with via the
HELCOM monitoring and assessment
strategy (HELCOM 2005) and the
HELCOM COMBINE programme
(HELCOM 2008). Assessments include annually updated HELCOM
Indicator Fact Sheets and the production of thematic assessment reports on
Figure 7: Nitrogen and phosphorus loads to the Baltic Sea
1990-2006. The target loads of the BSAP are indicated as
‘2021’. From Andersen et al. (2010).
eutrophication (e.g. HELCOM, 2009),
which covers the period 2001-2006
and sets a baseline for the BSAP. Further, the BSAP does not include an unambiguous evaluation phase. This may not be a significant issue, since the countries have committed themselves
politically to (1) implement AM for the restoration of good ecological/environmental status of
the Baltic Sea, and (2) revisiting the nutrient reduction targets and measures, in particular the
country-wise allocation. The upcoming 2013 HELCOM Ministerial Meeting will evaluate the
effectiveness of the national programmes and review the progress towards the ecological objectives describing a Baltic Sea in good status.
Figure 8: Illustration of the core of the eutrophication segment in the BSAP. The first step is political, the
second step is in principle scientific (setting target is the basis of the politically agreed visions and objectives) while the third step is a combination of science (scenario modelling) and policy (agreeing on the
reduction targets). Based on HELCOM (2007) and Backer (2008).
30
2: Similarities and Differences Between the Danish
Action Plans on the Aquatic Environment and
the Baltic Sea Action Plan
The origin of the first Danish Action Plan on the Aquatic Environment (APAE 1) and its evolution are summarized in Figure 9. Many direct and indirect connectors are identified, the most
important ones are highlighted in the following sections.
The Belt Project and especially its successor, the National Pollution Monitoring Programme
identified a large-scale eutrophication problem in Danish marine waters. This, in combination
with results originating from regional monitoring (including ground and freshwaters), led to a
good understanding and wide acceptance of the cause-effect relations leading to oxygen depletion in the inner Danish marine waters.
The interactions between research and monitoring and subsequent evaluations of the APAE’s
have proved to be working as intended by APAE 1. Of particular importance are the links from
the Danish Marine Environmental Research Programme, initiated via APAE 1, to the revisions
of national monitoring programs leading to the Danish Aquatic Monitoring Program (DAMP)
1993-1997 and NOVA-2003. Three elements of the monitoring programs worth highlighting are:
(1) weekly sampling at open water stations in dynamic areas, (2) coastal areas with extended
sampling programmes including mass balances for nutrients, and (3) the inclusion of a 3D Marine Modelling Complex covering open Danish marine waters, including the North Sea, Skagerrak, Inner Danish Waters, and the western Baltic Sea.
Input from the nation-wide monitoring programmes (DAMPs, NOVA, NOVANA) have been
directly linked to revisions of the APAEs: (1) The Action Plan for Sustainable Agricultural Development (from 1991, sometimes called APAE 1½ ) was very much influenced by DAMP
1989-1992, (2) APAE 2 was directly influenced by DAMP 1993-1997 as well as implementation
of the Nitrates Directive, the latter being amplified by the incident in Mariager Fjord where the
whole water column became anoxic, and (3) APAE 3 was influenced by NOVA-2003, in particular the sub-programmes for catchment monitoring and riverine loads. Hence, the monitoring
programs have been and still are a backbone, providing data for assessments and evaluations.
31
Figure 9: Key interactions between the Danish Action Plans on the Aquatic Environment, monitoring of
the aquatic environment and marine research. Colours refer to a pre-phase (white), the Action Plans as
such and derived activities (grey) and indirectly related activities (light grey).
An overlooked and to some extent refreshing feature in regard to the Danish APAE 1 is that it is
evidence-based, both in regard to its roots (being the Belt Project and its successor, the National
Pollution Monitoring Programme (Ærtebjerg Nielsen et al. 1981, ATV 1990)) and partly also in
regard to its reduction targets (Miljøstyrelsen 1987 &1990 and Jens Brøgger, pers. comm.). The
APAE 1 has from time to time been criticized for being based on uncertain estimates of the existing loads, especially the losses from agriculture, and the required reductions (ATV 1990). Parts
of the criticism are understandable, but perhaps not entirely justified. The Government’s Action
Plan for the Marine Environment (APME), developed and published by Miljøstyrelsen (1987),
was evidently not based on a nation-wide mass balance or comprehensive scenario modelling,
32
but still it was based upon the best available information (e.g. the NPo Action Plan (Folketinget
1984-85)) as well as justified estimates for load reduction prepared by the counties (e.g.
Hovedstadsrådet 1983). Given the media pressure at that time, the criticism seems to underestimate an inevitable political need for action.
There is no single element in the Danish APAE’s that makes them successful; it is a combination
of activities. Four elements are likely to be of particular significance. First of all, it is the consistent, direct and sustained implementation of measures, especially in regard to discharges of
phosphorous from point sources and losses of nitrogen from diffuse sources. Secondly, it is the
monitoring. Thirdly, it is both the regional and national assessments of environmental status.
Fourthly, it is a solid political will, at least in the first decade of the APAE’s, to follow up on the
results of the evaluations, especially when it comes to APAE 1 and 2.
A key characteristic of the Danish APAEs is their focus on all four phases of AM, e.g. the PACE
principles. APAE 1 in particular had a strong focus on Planning (in particular the overall reduction targets, well designed monitoring and assessment systems, the addition of a pre-planned
evaluation), on Actions (in particular the focus on relevant sources and reduction of both N and
P), on checking (in particular establishment of DAMP 1989-1992, annual reporting at three levels, sustained funding) and on a pre-planned Evaluation (in particular the first evaluation and the
agreement on the Action Plan on Sustainable Agricultural Development (APSAD being nicknamed APAE 1½), which in reality primed the basis for all subsequent evaluations and follow up
plans).
Another characteristic is the role of science: the Belt Project and the NPo Research Programme
initially fed results into the APs and monitoring programmes. Especially the Danish Marine Environmental Research Programme (Jørgensen & Richardson (1996), Christensen et al. (1998))
played a crucial role, not only in regard to re-design of the marine monitoring programmes but
also in regard to building a joint and widely accepted understanding of causes and effects of eutrophication as well as capacity building locally (counties/environmental centres) and nationally
(in particularly at the National Environmental Research Institute). In later years, another important feature has been the vast number of scientific publications based on the long-term monitoring of inputs and effects in all compartments of the aquatic environment (e.g. Conley et al.
33
2000, Nielsen et al 2002a, Nielsen et al 2002b, Ærtebjerg et al. 2003, Josefson & Hansen 2004,
Carstensen et al. 2006, Conley et al. 2007, and Carstensen et al. 2011).
When scrutinising what has happened during the evaluation of the Danish Action Plans, not everything is clear-cut. When the 1987 reference losses for root zone nitrogen from agriculture were
corrected in 2003, as preparation for the 2004 AP III, they were assumed to be 20% higher than
earlier estimated (260.000 to 311.000 t cf. Table 3) (Grant & Waagepetersen 2003). However,
the percentage used for the correction of the 1987 reference loss could be a negotiated figure and
there was no scientific, public or political debate of the corrected figure and in particular of its
implications.
Some points worth reiterating are that the APAE 1 overruled the different agendas adopted by
the Parliament and that the APAE had three ‘targets’: (1) a reduction target (RT), a reduction
percentage (%) as well as a target load (TL). While APAE 1 and 2 focused on the numbers in the
original AP (See Table 3), the APAE 2 evaluation, because of the correction of the reference
load for nitrogen from agriculture (see Table 4), created an opportunity for APAE 3 to put emphasis on either the reduction target or the reduction percentage or the target load, cf. the scenarios presented in Table 7. An interpretation of the three scenarios presented in Table 7 could be
that the course set by APAE 3 is likely to be the least stringent. The scenarios being more stringent in regard to alleviation of eutrophication were simply disregarded by APAE 3. A likely explanation would be that the agricultural sector had regained a political influence as strong as before the adoption of APAE 1.
Table 7: Summary of estimated differences between the 1987 load targets in Action Plan I. Scenarios are
based on the corrected reference loads and setting of a fixed reduction target (scenario I), a fixed reduction percentage (scenario II), and a fixed load target (scenario III). Numbers in red originates from APAE
1. The difference is calculated as the 1987 load target minus the revised load target. Units = tonnes TN.
AP 1
Scenario I
Scenario II
Scenario III
260,000
Reference loads from agriculture
311,000
311,000
311,000
127,000
127,000
Reduction target
152,400
178,000
49%
49%
Reduction percentage (%)
41%
57%
133,000
133,000
Load target
184,000
158,600
Difference
÷51,000
÷25,600
0
Another remarkable point is that the APAE 1 seems to be inspiration for other European 50%
reduction plans, e.g. those by HELCOM and OSPAR, which were agreed in February 1988 and
in June 1988, respectively (OSPAR 1988, HELCOM 1988). Denmark promoted key principles
34
originating from the APAE 1, but it is unclear if the countries around the Baltic Sea and the
North Sea simply acknowledged the Danish adoption of the APAE 1 or agreed on the necessity
of almost identical reduction targets.
The overture leading to the adoption of the HELCOM Baltic Sea Action Plan (BSAP) can be
described as protracted. However, it reflects the scientific understanding of eutrophication in the
Baltic Sea (Elmgren 2001) as well as the shift in geopolitical conditions. The BSAP can be said
to have had a long prelude, starting with the 1974 Convention and the 50% reduction target in
the 1988 Ministerial Declaration. Reductions in nutrient loadings have been achieved by most
Baltic Sea countries; the long-term results are remarkable while the recent short-term developments (2004–2006) are not as encouraging. The reductions have not yet resulted in a Baltic Sea
unaffected by eutrophication. Hence, the good environmental status in terms of eutrophication as
defined by the HELCOM BSAP had not been reached via the predecessors to the BSAP. The
links between the 1974 Convention and the onset of regular Baltic Sea-wide assessments including assessments of eutrophication status are illustrated in Figure 10.
The drivers that will result in a decrease in loads are proper implementation of national action
plans and HELCOM recommendations as well as a number of legally binding international
agreements and legislation such as the European directives addressing eutrophication. The most
recent additions to the list of drivers are the BSAP and the Marine Strategy Framework Directive
(MSFD). Implementation of the Urban Waste Water Framework Directive (UWWTD), Nitrate
Directive (ND), Water Framework Directive (WFD) and MSFD is essential, because tangible
and durable improvements in the eutrophication status of the Baltic Sea rely on the load reductions provided via these directives and without their proper implementation, progress, if any, will
be very slow and difficult to document. Moreover, the implementation of these directives has already been taken into account when establishing the eutrophication segment and load reduction
allocations of the BSAP.
The eutrophication segment of the BSAP envisages provisional national load reductions tentatively set up on the basis of: (1) overall objectives and a set of targets for water transparency, (2)
model calculations of maximum allowable loads and country-wise reduction targets, and (3) reduction scenarios and cost-efficiency. The approach employed is well-justified and welldocumented and should be seen as an appropriate first step. The BSAP thus acknowledges that
35
the figures related to targets and maximum allowable nutrient loads should be periodically reviewed and revised using a harmonized approach based on the most recent information and data.
Figure 10: Key interactions between political agreement under the Helsinki Conventions (Conventions,
Ministerial Declarations and Action Plans), Baltic Sea-wide assessment of the state of the environment
and other eutrophication related policy drivers. BIO = integrated thematic assessment of biodiversity and
nature conservation; BSAP = Baltic Sea Action Plan; EUT = integrated thematic assessment of eutrophication; HAZ = integrated thematic assessment of hazardous substances; HOLAS = Holistic Assessment
2003-2007; JCP = Baltic Sea Joint Comprehensive Environmental Action programme; MSFD = Marine
Strategy Framework Directive; WFD = Water Framework Directive; UWWTD = Urban Waste Water
Treatment Directive; ND = Nitrates Directive.
36
Hence, the BSAP includes an element of Adaptive Management, but no progress reporting or
evaluations as such. Further technical development of the modelling approach should be carried
out by including a broader range of indicators, such as nutrient concentrations, chlorophyll-a
concentrations and oxygen in addition to the currently employed water transparency. In addition,
greater coherence is needed between the modelling approach and the practical use of modelling
results, and most likely also future eutrophication assessments. Coherence could be enhanced by
increasing the temporal resolution of the model to the level which is employed, inter alia, in this
status assessment enabling a distinction between the different seasons instead of data averaged
over the annual cycle. This would not only improve the reliability of the approach and load allocations, but also lead to greater credibility among the public, which has not yet been achieved by
any regional marine convention.
The total acceptable loads sensu the 2007 HELCOM BSAP and the 50% reduction targets sensu
the 1988 HELCOM Ministerial Declaration cannot directly be compared because there are differences in the approaches used. An indirect comparison indicates that the 2007 BSAP is stricter
in terms of phosphorus than the 1988 Ministerial Declaration. In terms of nitrogen, however, it
could appear that the 1988 Ministerial Declaration might be stricter. Nonetheless, this may not
be significant for the following reasons: (1) the BSAP, addressing eutrophication using a holistic
ecosystem approach, specifies a number of indicators with associated targets which are comparable with what would have been the ultimate effect of implementing the 50% reduction target,
(2) the BSAP does not (yet) take a consistent implementation of the WFD into account in terms
of expected load reductions, and (3) the BSAP will pursue declining loads and allow the Baltic
Sea to recover from its present status.
A characteristic of the BSAP eutrophication segment is that it builds on a suite of principles
which in combination makes it ecosystem based. The principles include, cf. Fig. 5: (1) setting of
visions, objective and selection of indicators (Backer and Leppänen (2008), (2) operational targets (EutroQOs) based on RefCon and AcDev (Andersen et al. (2004, 2006, 2010), HELCOM
(2009), Andersen (2010)), (3) linking targets and loads leading to estimation of Total Allowable
Loads and scenarios for cost-effective country-wise reduction targets (Wulff et al. 2007) and
subsequent, but not yet implemented (4) actions according to the Water Framework Directive,
which shall be fully implemented by 2016. However, when scrutinising the basis of the BSAP
eutrophication segment, it may not look as good as claimed. The BSAP is currently based on a
37
single indicator/target i.e. water transparency (Secchi depth). This does, at least in principle, base
the BSAP on Ecosystem Based Management (EBM). However, it might appear that the BSAP
has been rushed since it is not based on any other targets/indicators such as causative factors
(e.g. nutrient concentrations), primary effects (e.g. primary production, chlorophyll-a), or secondary effects (e.g. changes in benthic communities, oxygen concentrations). To be ecosystembased beyond doubt would require inclusion of more targets/indicators, cf. Figure 11.
Revision of the load calculations would also be problematic if not based on: (1) updated load
figures including atmospheric deposition, (2) an update of the targets and (3) inclusion of more
targets/indicators. HELCOM (2009) and Andersen et al. (2010, 2011) provide valuable information in regard to target setting for all major basins of the Baltic Sea. Further development of
the model is in progress (Maria Laamanen, pers. comm.), and taking this into account when updating the allowable loads and the load allocation would turn the eutrophication segment of the
BSAP into a state-of-the-art ecosystem-based nutrient management strategy.
Figure 11: Suggested framework for implementation of the eutrophication segment of the Baltic Sea Action Plan. Please note that this framework is ecosystem-based, taking relevant eutrophication effects into
account, e.g. elevated nutrient concentrations, primary effects (e.g. Chlorophyll-a concentrations) and
secondary effects (e.g. benthic communities and oxygen depletion). Based on HELCOM (2009).
38
An important similarity between the Danish Action Plans and the Baltic Sea Action Plan is that
both are closely related to the implementation of a suite of eutrophication-related EU directives,
e.g. the EC Urban Waste Water Treatment Directive (Anon 1991a), the EC Nitrates Directive
(Anon 1991b), the EU Water Framework Directive (Anon 2000) as well as the EU Marine Strategy Framework Directive (Anon 2008). The interactions between these eutrophication related
directives are many and complex (HELCOM 2009), but in no way conflicting. The directives are
despite differences in focus and terminology all striving toward a better ecological status of surface waters and are focused on cuts in nutrient loads.
Both the Danish APAE’s and the BSAP are based on a political agreement but the Danish
APAE’s are enacted by national law and statutory order while the BSAP is implemented via ‘soft
law’ such as the Helsinki Convention and HELCOM Recommendations and not legally binding.
The Danish APAE can be viewed as “dark green” while the BSAP is “light green” in the sense
that the APAE are enacted whilst the BSAP is incentive based sensu Ernst (2010), who have
characterized the spectrum of “Mainstream Environmental Thought” as “dark green” (left wing
with a preference for mandatory agreements), “light green” (moderate with a preference for voluntary agreements) and “conservative” (right wing with a preference for market-oriented solutions).
Another significant difference is the strong focus of the Danish APAE’s on progress reporting
based on monitoring and assessment and evaluations where the progress is compared to the goals
of the APAE 1. This is in line with the concept of Adaptive Management, where an Evaluation
phase is an integral and re-occurring element. Although the BSAP is in its early phases of implementation, it could appear that there is a risk of losing momentum, e.g. in regard to: (1) the
recalculation of the load allocation, (2) progress reporting, and (3) evaluations where the progress is compared to the already agreed load reductions. The advantage of being based on the
concept of Ecosystem-Based Management might not be as great as assumed, and even outweighed by the weakness of not having included a well-planned evaluation phase with ‘rules of
procedure’ if loads are not reduced as agreed. Finally, another difference is that that Danish
APAE’s are focused on discharges and losses to freshwater and marine waters, while the BSAP
is focused only on inputs to the sea.
39
Based on the Danish Action Plans on the Aquatic Environment and the HELCOM Baltic Sea
Action Plan, four key characteristics of the plans have been identified:
1. A good nutrient management strategy (NMS) must include four well defined phases: (1)
Planning, (2) Action, (3) Checking, and (4) Evaluation. No weak links are allowed in this
PACE sequence. A perfect plan and accurately estimated reduction needs are of no value
without monitoring and regular evaluation because improvements in water treatment, agricultural practises and ultimately eutrophication status span over timescales longer than political
election periods. High-quality monitoring without progressive reductions of loads will not result in improvement of environmental quality. The Danish APAE’s, being strong in all phases, are currently the only NMS that successfully have proven to reduce inputs from both
point sources and diffuse sources (Carstensen et al. 2006, Kronvang et al. 2008).
2. A good NMS must have overarching goals and reduce loads of relevant nutrients, in particular nitrogen and phosphorus. Whether the goals are reduction goals as in the Danish APAE’s,
eutrophication targets (EutroQOs) or a combination of EutroOQs and reduction targets as in
the BSAP, is of minor importance. The key thing is that goals are being set and are easy to
understand. The reduction targets of the Danish APAE were clear and unambiguous, and the
BSAP objectives of total allowable loads and country-wise load allocation are in principle
not open to interpretation. And, as illustrated by the Danish APAE, some degree of uncertainty in regard to the target setting does not affect the likelihood of fulfilling the plan as long
as loads are progressively reduced.
3. The effects of a good NMS must be documented by high-quality monitoring and assessment
activities. Here, the monitoring activities under the Danish APAE’s are examples for others
to learn from.
4. A good NMS relies upon political will to carry out evaluation as planned and, above all, to
conclude and follow up. Neither the best available scientific advice nor good planning can
substitute a sustained political will to alleviate eutrophication symptoms. Both the Danish
APAE’s and the BSAP are based on a political will to lessen eutrophication effects, and the
long-term sustained political will in Denmark, especially in the period 1986-2001, is perhaps
exceptional.
In addition to the key characteristics above, a number of specific requirements in regard to successful implementation of the PACE sequence are identified:
1. Planning, meaning both agreeing on overall objectives and targets and developing a strategic
plan on how to achieve these should, on the basis of the analyses of the two case studies, include:
a) A document including not only political objectives and targets but also description of and
strategies for the three other phases (Actions, Checking and Evaluation), and
b) A plan for multi-year funding of actions (measures) and monitoring.
2. Actions, understood as agreeing on measures and subsequent implementation of these,
should on the basis of the two case studies analysed include:
a) Reduction targets for both nitrogen and phosphorus – and covering all contributing sectors (e.g. agriculture, industries, households, energy production, and transport), and
b) Legal instruments such as laws, statutory orders (as in the Danish APAE’s) or alternatively recommendations (as in the case of the BSAP).
40
3. Checking, understood as monitoring and assessment, should on the basis of the case studies
include:
a) Effects and inputs, a marine programme alone is not enough, information about inputs
and activities in upstream catchments is essential, cf. the Danish Aquatic Monitoring
Programme (DAMP), and
b) Adequate spatial and temporal coverage of primary and secondary effects, nutrients, inputs, sectors, human activities.
c) A detailed and unambiguous programme manual including description of methods, data
flow, and quality assurance/quality control procedures, and
d) Publication of assessments reports, in principle covering all relevant compartment of the
aquatic environment (point sources, diffuse sources, groundwater, lakes, rivers and marine waters) – in order to maintain awareness.
e) Evidently, sustained funding is an immense advantage, alternatively multi-year funding
e.g. for a 4-6 year period can be used as a guiding principle.
4. Evaluation, especially in the understanding of a strategic check as to whether the objectives
and target are being fulfilled, should include evaluation criteria as well as actions to be
agreed during the planning phase.
The Danish APAEs and the HELCOM BSAP share many principles including the PACE sequence. However, there are three major differences:



First, the BSAP is still in its first round, while the Danish APAE’s are approaching a fourth
round. Therefore, the interim judgement of the BSAP’s checking and evaluation phases
might turn out to work better than predicted.
Second, the Danish APAEs are adopted by a majority in the Danish Parliament and implemented in national law. The HELCOM BSAP is not legally binding, it is merely a political
expression of interest. Since it is closely linked to a suite of EU directives, in particular the
Urban Waste Water Treatment Directive (UWWTD) and Marine Strategy Framework Directive (MSFD), this might compensate for the non-legally binding character of the BSAP.
Third, whilst the APAEs are focusing reduction of discharges and losses from three sectors,
the BSAP is focusing on an ecological target being ‘clear water’ on basis of which load reductions are estimated.
Although the APAE 2 was partly adopted in response to the Nitrates Directive (ND), the followup plan, APAE 3, might offer an example of changing the 1987 goals during the APAE 3 evaluation in 2008, cf. Table 3 and 7. The outcome could be seen by some as a less stringent course in
regard to losses of nitrogen from diffuse sources. Further, it seems unclear how this change fits
with the implementation of the Nitrates Directive, but also the Water Framework Directive.
41
3: Beyond Action Plans and Directives:
Perspectives for the Future
Although the Danish Action Plans and the HELCOM Baltic Sea Action Plan (BSAP) might be
considered as best practices and examples from which to learn, both plans are far from being
perfect, even though they are evidence-based.
Six issues are of interest to consider when developing a next generation of evidence-based nutrient management strategies: (1) the N/P controversy, (2) so-called ‘Technical Solutions’ to abate
eutrophication, (3) target setting, (4) thresholds, (5) socio-economical aspects of eutrophication,
and (6) climate change manifested as shifting baselines.
The N/P controversy is a discussion about which nutrient input should be reduced in order to
combat eutrophication. For coastal marine waters, nitrogen has historically been considered the
limiting nutrient. However, anthropogenic phenomena affecting both sides of the N:P ratio have
combined to increase that ratio in coastal waters: Human activities have contributed to an overabundance of nitrogen in coastal waters, while upstream nutrient controls focusing mainly on removing P have also increased downstream N:P ratios. Schindler et al. (2008) suggested that effective eutrophication control can be achieved in both freshwater and coastal ecosystems by controlling P only, based on research done in an experimental lake. Schindler et al. (2008) conclude
that fixation of atmospheric nitrogen can respond to meet ecosystem N requirements in a regime
of P enrichment, P ultimately controls eutrophication and there is no pressing need for N input
controls. Both Conley et al. (2009a) and Paerl (2009) question this finding. Evidence is presented
that both N and P must be reduced to battle eutrophication in coastal waters. Paerl (2009) points
out that nutrient dynamics in coastal and estuarine waters are quite different from those in freshwater systems. Coastal N2 fixation generally does not satisfy ecosystem-level N demands, causing these waters to remain N-limited and hence sensitive to N over-enrichment. HELCOM
(2009), in line with Conley et al. (2009a) and Paerl (2009), emphasize that despite regional variations (the Gulf of Bothnia is P limited) control of both N and P is needed for long-term management of eutrophication in the Baltic Sea region.
The usefulness of so-called ‘technical solutions’ as an alternative to nutrient reductions in the
Baltic Sea is debatable (see Conley et al. 2009b for a summary). Proposals for technical fixes
42
include (1) artificial aeration/oxygenation, (2) large-scale manipulation of the circulation, (3)
chemical removal of phosphorus, and (4) bio-manipulation. Virtually all engineering methods
proposed to date for the Baltic Sea’s pelagic waters seem unrealistic. At best they can only speed
up recovery while nutrient reductions begin to have an effect. Conley et al. (2009b) conclude that
these large-scale attempts at remediation are unlikely to substantially improve the short-term
conditions in the Baltic Sea and several pose substantial risks for the environment. It should be
mentioned, as a precautionary note, that engineering solutions have been evaluated and rejected
(Conley et al. 2009b). However, reconstruction of stone reefs, a protected habitat type under the
EC Habitats Directive, in shallow coastal waters is likely to have a positive impact on both eutrophication and biodiversity (Møhlenberg et al. 2009).
A third area of concern where research is urgently required, is in regard to the setting of targets,
in particular improvement of the current understanding of and values for reference conditions
(RefCon, being the anchor of target setting) and acceptable deviation (AcDev, the acceptable
deviation from RefCon). Establishment of reference conditions in aquatic systems can be made
in a number of different ways. The methods currently used are (1) spatially based reference conditions (including historical data and paleo-ecological studies), (2) modelling (empirical or dynamic), (3) combinations of (1) and (2), and (4) expert judgement (Andersen et al. 2004, 2010,
2011). The setting of AcDev is very critical. Experiences from the implementation of the WFD
suggest that the ‘scientific’ interpretation of the politically agreed ‘normative definitions’ (of
what an ‘acceptable’ deviation is about) has been carried out by a process not able to adequately
balance scientific advice and policy.
It is important to consider ecosystem thresholds and regime shifts as well as shifting baselines
when developing and implementing evidence-based nutrient management plans, e.g. Duarte
(2009), Duarte et al. (2009), Kemp et al. (2009), and Carstensen et al. (2011). These newly published results are important since management of eutrophication is not only about reducing nutrient losses from human activities, in particular agriculture, but also about keeping the ecological
impact of human activities at a sustainable level leading to realization of politically agreed ecological objectives, e.g. a Baltic Sea ‘unaffected by eutrophication’ or ‘good ecological status in
coastal waters’.
43
Some potential implications of ecological threshold and shifting baselines in regard to eutrophication, especially in regard to reduction of loads are illustrated in Figure 12.
Panel A shows a straight forward linear recovery,
where loads have to be reduced to the loads equivalent to fulfilment of the target (EutroQO = RefCon
÷ AcDev). Panel B is slightly different since it includes a threshold, meaning that the reduction required to fulfil the target is slightly larger than in a
situation without a threshold. Identification of
threshold is imperative when trying to achieve ecological targets, e.g. as those in the WFD and BSAP.
Climate change will also affect the eutrophication
status of temperate coastal water in the future – and
introduce a shift in the baseline as indicated in panel C – development and execution of evidencebased nutrient management strategies ought to take
such shifts into account as soon as possible (Duarte
2009).
Combining a shifting baseline and a threshold (panel D), which might be appropriate for many shallow
coastal waters, may indicate that the load reductions
required to achieve the already agreed targets might
be larger than currently acknowledged (or in extreme cases impossible to achieve).
Figure 12: Conceptual models of the consequences of shifting baselines, regime shifts
(thresholds) as well as the combination of
shifting baseline and regime shifts for nutrient
management strategies. Dashed red line indicated target, while the red arrow indicates
required reduction of human pressures, e.g.
reductions of loads. Based on Laamanen et
al. (submitted).
Finally, inclusion of socio-economic considerations
and cost-benefits are becoming more and more usual, e.g. Wulff et al. (2007) and Anon. (2008). Sweden has been at the forefront of this issue, e.g.
Turner et al. (1999), Garpe (2008), and Gren &
Elofsson (2008), the latter updating the estimate of
44
the net benefits of alleviating eutrophication in the Baltic Sea. Depending on the choice of target
for nutrient reductions and choices of discount rate the overall annual net benefit ranges between 0.2 and 7.4 billion Euro per year. Such findings are very interesting, in particular since
politicians often, if not always, focus on the costs of actions. If the Swedish results are correct,
and it should be stressed that there are no reasons for questioning their methods and models, then
solving the eutrophication problem would be a good bargain. So why hasn’t it been done yet?
Probably because the Danish agriculture is better at lobbying decision-makers than environmentalists are. It is likely that this situation will change once national authorities and politicians become informed in regard to the socio-economic consequences of continued eutrophication. The
costs of non-action are often far greater than the costs of action.
Getting more science on board when revising and developing existing nutrient management
strategies, both in terms of improved ecological understanding and better socio-economic models, will ultimately lead to more and better ecosystem-based strategies. Here, we are talking
about the fine art of balancing knowledge and economy without forgetting that policy is about
avoiding making unpopular decisions and regulations.
4: Conclusions: What Makes a Nutrient
Management Strategy Successful?
Cultural eutrophication will continue to be a significant issue for decades. However, we have
now reached a level in regard to our conceptual understanding of coastal eutrophication and its
causes where nutrient management strategies should be both informed and ecosystem-based.
With current policy drivers (WFD, MSFD, and BSAP), existing nutrient management strategies
will be updated and include more outreach and stakeholder involvement. Combining this with
the upcoming development and implementation of maritime spatial planning as well as the experience from successful nutrient management strategies, it is clear that any so-called value-based
policies aiming to hamper endeavours to reduce coastal eutrophication will be clearly exposed.
Any attempt to redefine what the problems are about should be rejected.
45
But how do we improve existing planning and make them even more evidence-based and consequently ecosystem-based? Answering this question is not simple, but based on the experiences
from the two action plans discussed, an answer can be split in four ‘areas of concern’.
The first area of concern is related to political will. Generating and maintaining political will is
required for solving the problems of eutrophication in coastal waters, in particular:


‘Education’ of politicians – especially when it comes to: (1) the basic concepts and ecological and economic consequences of eutrophication, (2) a political acceptance of the uncertainties related to the estimation of load reductions, and (3) an acceptance of using the sequence
of i) Planning, ii) Action, iii) Control, and iv) Evaluation, e.g. the PACE sequence.
Goals should not to be subject to revision – the only themes to be revised by politicians are:
(1) the measures to be implemented to fulfil the goals, (2) the timing when the measures are
to be fully implemented, and (3) the overall strategy to monitor progress.
The second area of concern is related to research. If future ecosystem-based nutrient management strategies are to be ecosystem-based, then focus ought to be put on the following themes:



Better targets – reference conditions, acceptable deviations from reference conditions, and
functional relations used for target setting should be evidence-based.
Socio-economy – a better understanding of ecosystem services and the socio-economic benefits of marine waters not affected by eutrophication is required.
Habitats and species – a better understanding of the links between eutrophication, biodiversity and fisheries is required.
The third area of concern is related to long-term monitoring and assessment, which is a prerequisite for documenting the effect of measures and for improving the evaluation phase:


Monitoring – all ecologically relevant indicators should be monitored, spatial and temporal
coverage should be decided according to ecosystem structures and functioning, not by the
availability of funds.
Funding – should be long term, at least for 4-6 years.
The fourth area of concern is related to strategy development. Existing nutrient management
strategies should, whether they are based on the concept of either AM or EBM, be further developed and optimized. Future nutrient management strategies ought to be both adaptive and ecosystem-based. The first step should be to develop better evaluation phases of existing ecosystembased nutrient management strategies, a second step should be to make use of evidence-based
decision support systems linking targets, loads and costs while a third step should be to factor in
climate change.
46
It should be clear that there are many different ways leading to ecosystem-based management
taking into account differences in geographic scales, each with its own unique historical, ecological, and social context. The future nutrient management strategies (NMS) will also vary depending on the types of legislative and managerial framework already in place. Notably, we no longer
start from scratch but need to refine the existing NMS. Perhaps most critically, the development
and implementation of any NMS cannot act in isolation. NMS’s need to be evidence-based, to
include all relevant stakeholders and to be supported by a sustained political will to alleviate eutrophication, in particular to reduce nutrient loads.
The DO’s and DON’T’s of evidence-based nutrient management strategies are:


DO understand that ecosystem-based management is adaptive and science-based.
DON’T assume that decisions can not be taken because of incomplete knowledge and
uncertainty.


DO evidence-based target setting and exhaustive planning, the latter involving decisionmakers, authorities and all stakeholders.
DON’T wait for perfection and all-inclusive ecosystem understanding.


DO a full execution of the plan.
DON’T rely on voluntary agreements or guidelines.


DO monitoring with ecologically relevant resolution in time and space.
DON’T underestimate resources needed for sampling, quality assurance, analysing data and
reporting.


DO regular evaluations in regard to the progress of the nutrient management strategy.
DON’T disregard the advantages of a dual monitoring strategy focusing on both nutrient
inputs as well as ecological responses to lowered nutrient inputs.
It is critical to be patient as emphasised by Fulweiler et al. (2010) since time is needed before the
effects of changes in human behaviour can be seen in inputs and eventually in the ecological
quality of the marine environment. It is also critical to prepare for those moments where decisions and actions can be taken by building up the best possible scientific basis for decisionmaking as well as decision support systems. An important lesson learned from the Danish Action
Plans on the Aquatic Environment and the HELCOM Baltic Sea Action Plan is that things come
about in windows of opportunity.
47
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Kronvang, B., H.E. Andersen, C. Børgesen, T. Dalgaard, S.E. Larsen, J. Bøgestrand & G. Blicher-Mathiasen, 2008: Effects of policy measures implemented in Denmark on nitrogen
pollution of the aquatic environment. Environmental Science & Policy 11(2):144-152.
Laamanen, M., S. Korpinen, U.-L. Zweifel & J.H. Andersen, submitted: Ecosystem health. Textbook chapter in “Biological Oceanography of the Baltic Sea” (Eds: P. Snoeijs, H. Schubert
& T. Radziejewska).
Läine, A., H. Pitkänen, B. Arheimer, H. Behrendt, W. Jarosinski, S. Lucane, K. Pachel, A.
Räike, A. Shekhovtsov, L.M. Svendsen & S. Valatka, 2002: Evaluation of the implementation of the 1988 Ministerial Declaration regarding nutrient load reductions in the Baltic Sea
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Miljøministeriet, 1987: Handlingsplan mod forurening af det danske vandmiljø med næringssalte. 13 pp. (in Danish)
Miljøstyrelsen, 1984a: Iltsvind og fiskedød i 1981. Omfang og årsager. Miljøstyrelsen. 247 pp
(In Danish)
Miljøstyrelsen, 1984b: NPo redegørelsen. 218 pp. (In Danish)
Miljøstyrelsen, 1989: Vandmiljøplanens overvågningsprogram. Miljøprojekt nr. 115.
Miljøstyrelsen. 64 pp (In Danish)
Miljøstyrelsen, 1990: Vandmiljø-90. Redegørelse fra Miljøstyrelsen nr. 1, 1990. 204 pp (In Danish)
Miljøstyrelsen, 1993: Vandmiljøplanens overvågningsprogram 1993-1997. Redegørelse fra Miljøstyrelsen nr. 2, 1993. 172 pp. (In Danish)
Miljøstyrelsen, 2000: NOVA-2003. Programbeskrivelse for det nationale program for overvågning af vandmiljøet i Danmark, 1998-2003. Redegørelse fra Miljøstyrelsen nr. 1, 2000. 198
pp. (In Danish)
Møhlenberg, F., J.H. Andersen (eds.), C. Murray, P.B. Christensen, T. Dalsgaard, H. Fossing &
D. Krause-Jensen (2008): Stenrev i Limfjorden: Fra naturgenopretning til supplerende virkemiddel. DHI Teknisk Rapport til By- og Landskabsstyrelsen. 41 p + bilag. (In Danish)
Nielsen, S.L., K. Sand-Jensen, J. Borum & O. Geertz-Hansen, 2002a: Depth colonisation of eelgrass (Zostera marina) and macroalgae as determined by water transparency in Danish
coastal waters. Estuaries 25:1025-1032.
Nielsen, S.L., K. Sand-Jensen, J. Borum & O. Geertz-Hansen, 2002b: Phytoplankton, nutrients
and transparency in Danish coastal waters. Estuaries 25:930-937.
Nixon, S.W., 1995: Coastal marine eutrophication: a definition, social causes, and future concerns. Ophelia 41:199-219.
Nordemann Jensen et al. (2010): Vandmiljø og nature 2008. NOVANA. Tilstand og udvikling –
faglig sammenfatning. Faglig rapport fra DMU nr. 767. 106 pp (In Danish)
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area. Eutrophication Series. OSPAR Commission. 107 pp
Paerl, H.W., 2009. Controlling eutrophication along the freshwater-marine continuum: Dual nutrient (N and P) reductions are essential. Estuaries and Coasts 32:593-601. DOI:
10.1007/s12237-009-9158-8.
PH-Consult ApS, 1993: Spildevandsforskning 1987-1992. Spildevandsforskning fra Miljøstyrelsen nr. 53. 337 pp. (In Danish)
Rabalais, N.N. & S.W. Nixon (eds.), 2002: Nutrient Over-enrichment in Coastal Waters: Global
Patterns of Cause and Effect. Dedicated Issue. Estuaries 25:639-900.
Rabalais, N.N., R.E. Turner & D. Scavia (2002): Beyond science into policy: Gulf of Mexico
hypoxia and the Mississippi River. BioScience 52(2):129-142.
Savchuk, O.P., & F. Wulff, 2007: Modelling the Baltic Sea eutrophication in a decision support
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Schindler, D.W., R.E. Hecky, D.L. Findlay, M.P. Stainton, B.R. Parker, M. Paterson, K.G.
Beaty, M. Lyng, & S.E.M. Kasian, 2008: Eutrophication of lakes cannot be controlled by
53
reducing nitrogen input: Results of a 37 year whole ecosystem experiment. Proceedings of
the National Academy of Science 105:11254-11258. DOI: 10.1073/pnas.0805108105.
Suman, D., S. Guerzoni & E. Molinaroli, 2005: Integrated coastal management in the Venicelagoon and its watershed. Hydrobiologia 550(1):251-269.
Svendsen L.M., L. van der Bijl, S. Boutrup & B. Norup (eds.), 2005: NOVANA: Nationwide
Monitoring and Assessment Programme for the Aquatic and Terrestrial Environments.
Programme Description - Part 2. National Environmental Research Institute, Denmark. NERI Technical Report No. 537. 137 pp.
Turner, R.K., S. Georgiou, I.M. Gren, F. Wulff, S. Barrett, T. Söderqvist, I.J. Bateman, C. Folke,
S. Langaas, T. Zylick, K.G. Mäler & A. Markowska, 1999: Managing nutrient fluxes and
pollution in the Baltic: an interdisciplinary simulation study. Ecological Economics
30:333-352.
Waagepetersen, J., R. Grant, C.D. Børgesen & T.M. Iversen, 2008: Midtvejsevaluering af
Vandmiljøplan III. Det Jordbrugsvidenskabelige Fakultet, Århus Universitet og Danmarks
Miljøundersøgelser, Århus Universitet. 36 pp. (In Danish)
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the possible future of the Baltic. Ambio 36:243-249.
Sources of unpublished information

Kristoffer Colding, pers. comm., the Danish Nature Agency, Ministry for the Environment,
Haraldsgade 53, 2100 Copenhagen Ø, Denmark

Jens Brøgger Jensen, pers. comm., the Danish Nature Agency, Ministry for the Environment,
Haraldsgade 53, 2100 Copenhagen Ø, Denmark

Maria Laamanen, pers. comm., HELCOM Secretariat, Katajanokanlaituri 6B, FIN-00160
Helsinki, Finland
54
Annex 1: Abstract
Nutrient management strategies have to deal with all human activities resulting in discharges and
losses of nutrients, e.g. land-use, fertiliser use, industrial production, households and energy consumption.
Two Nutrient Management Strategies, one national based on Adaptive Management (AM) and
one trans-national based on an Ecosystem-Based Approach to management of human activities,
in practise being equivalent to Ecosystem-Based Management (EBM), are analysed.
The aim of this thesis is to analyze the critical factors likely to be required for a successful management strategy. Obviously, two key factors are nutrient reductions, generally of both nitrogen
and phosphorus, and monitoring of environmental status and trends.
It is likely that the prescribed quantity of measures ( e.g. nutrient reductions) is not a critical factor as long as periodic evaluation of progress and publication of assessments play a strong role in
the strategy, especially in combination with a sustained political will to follow up on the evaluations and pursue the visions and objectives of the strategy in question.
A combination of AM and EBM should be advanced to improve evidence-based nutrient management strategies.
55
Annex 2: Nutrients Discharges and Losses in Denmark
1989-2008
Table A2.1
Discharges of nitrogen and phosphorus (in tonnes) from urban waste water treatment plants
1989-2008.
Year
Nitrogen (tot N)
Phosphorus (tot P)
1989
18,000
4,470
1990
16,900
3,710
1991
15,100
2,800
1992
13,100
2,260
1993
10,800
1,760
1994
10,200
1,570
1995
8,900
1,230
1996
6,390
900
1997
4,850
670
1998
5,170
600
1999
5,130
580
2000
4,650
540
2001
4,220
470
2002
4,530
510
2003
3,610
400
2004
4,030
430
2005
3,810
410
2006
3,610
390
2007
4,360
470
2008
3,550
460
56
Table A2.2
Discharges of nitrogen and phosphorus (in tonnes) from industries with separate discharge 19892008.
Year
Nitrogen (tot N)
Phosphorus (tot P)
1989
6,500
1,410
1990
4,080
650
1991
3,770
520
1992
4,180
410
1993
2,540
240
1994
2,680
310
1995
2,440
200
1996
1,790
130
1997
1,760
130
1998
1,350
120
1999
970
70
2000
900
60
2001
820
50
2002
760
50
2003
510
30
2004
70
30
2005
440
20
20061
-
-
2007
320
20
2008
400
20
1: Discharges from 2006 have unfortunately not been estimated as a significant part of the primary data from individual industries with separate discharges has not been reported. The reason
is the Structural Reform implemented by 1 January 2007, especially the cessation of the countries in Denmark.
57
Table A2.3
Nitrogen surplus (in tonnes) in Danish agriculture 1980-2008.
Year
1980
1981
1982
1983
1984
1985
1986
1987
1988
1989
1990
1991
1992
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
Nitrogen (tot N)
434,000
394,000
381,000
453,000
400,000
420,000
408,000
423,000
376,000
374,000
379,000
396,000
431,000
360,000
356,000
320,000
303,000
290,000
292,000
278,000
267,000
260,000
247,000
224,000
229,000
215,000
188,000
206,000
209,000
58
Table A2.4
Losses of nitrogen (tot N) from the root zone of sandy soils and clay soils 1990/91 to 2007/08.
Year
Sandy soils
Clay soils
Weighted mean
Weighted reduction
(kg N/ha)
%
1990/91
154
76
107
0
1991/92
144
72
101
5,6
1992/93
139
68
96
10,3
1993/94
129
64
90
15,9
1994/95
118
66
87
18,7
1995/96
109
60
80
25,2
1996/97
102
58
76
29,0
1997/98
101
60
76
29,0
1998/99
83
53
64
40,2
1999/2000
84
51
64
40,2
2000/01
84
51
64
40,2
2001/02
82
49
62
42,1
2002/03
80
47
60
43,9
2003/04
81
46
60
43,9
2004/05
85
47
62
42,1
2005/06
83
43
59
44,9
2006/07
85
47
62
42,1
2007/08
88
46
63
41,1
The weighted mean is based on Børgesen, C.D. & R. Grant, (2003): Baggrundsnotat til Vandmiljøplan II - slutevaluering. Vandmiljøplan II - modelberegning af kvælstofudvaskning på landsplan, 1984-2002. Danmarks JordbrugsForskning og Danmarks Miljøundersøgelser. 22 pp, which
is available via:
http://www.agrsci.dk/var/agrsci/storage/original/application/phpE3.tmp.pdf.
59
Annex 3: Curriculum Vitae for Jesper H. Andersen
Name
Date of birth
Nationality
Education
Jesper Harbo Andersen
May 4, 1962
Danish
M.Sc. (Aquatic Ecology), University of Copenhagen (1990)
Courses
2006: Introduction to Microsoft Project (MaCom A/S)
2004: Introduction to Management (DIEU)
2002: Coaching (CONFEX)
2002: How to write a Technology Implementation Plan (Hyperion Ltd., Ireland)
2000: Teambuilding (DME)
1999: Communication and collaboration (DIEU)
1997: English Language Performance Training Programme (Danida)
1997: Course on Design of Water Quality Monitoring Networks (VKI and ColStat)
1996: Logical Framework Approach (COWI)
1995: Meetings and Negotiations – A Seminar in English (FLEX-SPROG)
1991: Crayfish Management and Culture (University of Kuopio, Finland)
Key
Jesper H. Andersen’s primary fields of interest are aquatic ecology, development of assessment tools (multiqualifications metric and indicator-based), design and planning of monitoring networks (and coupled courses of action regarding data flow, QA, data storage and reporting), ecosystem-based management as well as feeding scientific
information into adaptive and evidence-based management processes. He has a comprehensive knowledge of
aquatic ecology with special emphasis on impacts of human activities, environmental regulatory processes and
their administrative and economic contexts.
Jesper H. Andersen has broad experiences in relation to management of teams and large projects and is Project
Director at Institute of Bioscience, Aarhus University. He coordinates the institutes work in relation to the EU
Marine Strategy Framework Directive. Recently, he has been national expert in the MSFD Eutrophication Task
Group, the Danish Marine Strategy Expert Network as well as chair of the HELCOM HOLAS Task Force. He has
earlier on worked as a national expert in the WFD CIS COAST group and the WFD CIS Eutrophication Activity.
Another key qualification is organisation of international conferences, e.g. EUTRO 1993, the Symposium on the
North Sea QSR in 1994, EUTRO 2006, the ICES Indicator Symposium in 2007, EUTRO 2010 and the upcoming
Marine Strategy 2012 Conference during the upcoming Danish EU Presidency.
He has been heading of the Danish National Marine Monitoring Centre (M-FDC) in 2001-2005, which coordinates the national marine monitoring and assessment program (then: 18 partners and an annual budget of
~7.0 mio. €). He has been Danish HOD in the HELCOM Environmental Committee (EC, now MONAS) and the
OSPAR ASMO and worked in many other international groups. Further, he has participated in EU funded RTD
projects (CHARM, MOLTEN, DANLIM, EUSeaMap) and EEA’s ETC Water 2007-2010.
Project manager of large projects, such as: Danish EPA projects on the Water Framework Directive 1998-2004
(phase I, IIa, IIb, III and IV), NMR RETRO project 2002-2004, HELCOM EUTRO-PRO 2005-2008, BALANCE
2005-2007 (nominated for the UN ENERGY GLOBE Award 2009 and selected by the jury as national winner of
ENERGY GLOBE Award 2009 and also by WWF rated as one out of five 2008 high-lights in the Baltic Sea
area), HELCOM HOLAS 2009-2010 and HARMONY 2010-2012.
Jesper H. Andersen’s list of publications includes scientific papers, technical reports, policy papers, books, news
paper articles and includes more than 135 references.
Memberships  Chairman, Committee on Public Sector Consultancy and Applied Research, AU BioScience (2012 - )
 Expert member of the National Nature Protection and Environmental Board of Appeal 2011-2014
 Member, Advisory Board of the FORMAS project “Managing Multiple Stressors in the Baltic Sea”
 Coastal and Estuarine Research Federation (CERF – formerly ERF)
 Baltic Marine Biologists (BMB)
 Danish Society for Environmental Engineering (IDAmiljø)
Teaching
Lecturing in:
 “Systems Ecology”, Roskilde University Centre (RUC), focusing on nutrient enrichment of marine waters and
adaptive ecosystem-based management in an international and national perspective. Duration: 2003–ongoing.
 “The Baltic Sea: Yesterday, Today and Tomorrow”, Ph.D. course at Lund University. Duration: 2009–ongoing.
 “Environment and Resources”, Danish Technical University. Duration: 2006–2007.
 “Freshwater Ecology”, Copenhagen University. Duration: 2005–2006.
60
Employment record
Year
Firm
Position and responsibilities
2011 –
Department of Bioscience,
Aarhus University
Jesper H. Andersen is employed as projektchef (Project Director) at Department
of Bioscience at Aarhus University. He is currently chairman of the institute’s
Committee on Public Sector Consultancy and Applied Research.
(formerly Department of Marine Ecology,
National Environmental Research Institute Projects under negotiation and/or development:
at Aarhus University)

DEVOTES, a European MSFD FP7 research project focusing on GES
(indicators, modelling, targets), sea-based and land-based pressures, a
suite of case studies (e.g. the Kattegat) and guidance in regard to ecosystem-based management strategies. Planned budget: 9 mio. €.

SYMBIOSE, a national research and development in support of the Marine
Strategy Framework Directive. Planned budget: 4.,5 mio. dkr.

MONET, phase 2 (Development of innovative methods for monitoring and
design of monitoring networks for characterisation of Baltic Sea ecosystems), an application for BONUS. Planned budget: 4 mio. €.

HARMONY, phase 3, specific spin-off activities from the HARMONY project
Planned budget: 45.000 €.
Ongoing projects and supervision:

Project partner in HELCOM TARGREV, phase 2 (2012), focusing on
revision and publication of the HELCOM TARGREV project.
Budget: 15.000 euro.

Project partner in Baltic NEST Institute (2012), contribution to specific task
related to ecosystem-based management of eutrophication in the Baltic
Sea.
Budget: to be decided.

Project manager of HARMONY, phase 2 (2011-2012), the continuation and
finalisation of the HARMONY project initiated in 2010. Focus in on the development of tools for 1) indicator-based assessment of ‘good environmental status’ and cumulative anthropogenic pressures in the North Sea.
Budget, phase 2: 145.000 €.

MONET, phase 1 (Development of innovative methods for monitoring and
design of monitoring networks for characterisation of Baltic Sea ecosystems), an application for BONUS. Budget: 200.000 dkr.

Conference Secretary for Marine Strategy 2012, a three day conference
(14-16 May 2012) during the Danish EU Presidency focusing on research
and ecosystem-based management strategies in support of the EU Marine
Strategy Framework Directive.
Budget: > 3 mio. dkk.

Project partner in Review of Femerbelt EIA (2011-2012), a a review of the
draft Environmental Impact Assessment for the fixed link in Femerbelt.
Budget: Confidential.

Project partner in WATERS (2011-2015), a Swedish RDI project aiming to
develop indicators and assessment tools in regard to the EU Water Framework Directive.
Budget, NERI: 2,3 mio dkk.
61
Year
Firm
Position and responsibilities
Completed projects and supervision from January 2011:
2005 – 2010
DHI Water  Environment  Health

Faglige baggrundsnotater til havstrategidirektivets basisanalyse
(“Havet omkring Danmark”, phase 2), drafting of specific technical contributions to the Danish initial assessments pursuant to the EU Marine Strategy
Framework Directive.
Budget: 1 mio. dkk.

Sub-consultant in HELCOM TARGREV, phase 1 (2010-2011), a specific
NERI contribution to DHI in order to fulfil DHI’s commitments under the
HELCOM TARGREV project.
Budget for sub-contract: 29.376 euro.

Sub-consultant in BWO (2010), a specific NERI contribution to DHI in order
to fulfil DHI‘s commitments under the BWO Interreg project.
Budget for sub-contract, phase 1: 5.000 euro.
Budget for phase 2: 45.000 euro.

Sub-consultant in MARCOS (A NMR funded project focusing on “Marine
European Directives: Concepts, Overlap and Synergies”), a specific NERI
contribution to DHI in order to fulfil DHI’s commitments under the MARCOS
project.
Budget for sub-contract: 140.000 dkr.

Project manager of “MSFD synopsis” (phase 1), drafting of a synopsis for
the Danish initial assessments pursuant to the EU Marine Strategy Framework Directive.
Budget: 75.000 dkk.
Head of EU Water Policy Team, Department of Ecology & Environment: business
area manager for DHI’s activities in relation to the EU Water Framework Directive
and the EU Marine Strategy Framework Directive as well as key account manager. Number of staff implicated: 14 persons.
Projects and supervision in the period 2005-2010:

Project partner in Service contract for support to the implementation of
the Marine Strategy Framework Directive, a tender published by DG
ENV. Total budget: 450.000 € per year (2010-2013).

HARMONY, phase 1 (2010); a Danish, Norwegian and Swedish RDI project aiming to develop tools for initial assessments cf. the EU Marine Strategy Framework Directive.
Budget, phase 1: 100.000 euro

HELCOM TARGREV (2010-2011), a RDI project focusing on updating the
eutrophication segment and the country-wise load allocations of the Baltic
Sea Action Plan.
Budget: 63.000 €.

Project partner in the EEA ETC/ICM for 2011-2014 with focus on marine
and maritime tasks.

Project Partner in Ballast Water Opportunity (BWO), a North Sea INTERREG project focusing on risk management of ballast water. Duration: 20092012.
Budget, DHI: 100.000 €.
62
Year
Firm
Position and responsibilities

Project Manager of “Udvikling/tilvejebringelse af marine data til implementering af Havstrategidirektivet i Østersø- og Nordsøregionen”, a
project funded by the Danish Spatial and Environmental Planning Agency.
The project is directly related to the BWO project. Duration: 2009-2011.
Budget: 750.000 dkk.

Project Partner in EUSeaMap, a DG MARE funded project aiming to develop broad-scale habitat maps for the North Sea, Baltic Sea and western
Mediterranean Sea. Duration: 2009-2012.
Budget, DHI: ~ 72.000 €.

Ecosystem-based management of eutrophication: Linking monitoring,
assessments, and the society. Internal project aiming at a Ph.D. degree.
Co-funded by DHI via the RK contact 2006-2009 from the Danish Ministry of
Science, Technology and Innovation. Duration: 2009-2010.
Budget: 475.000 dkk.

Project Partner in an EU WFD Support Contract. Funded by the European
Commission (DG ENV). Focus is on compliance checking of European
WFD River Basin Management Plans. Duration: 2009-2012.
Total budget: 1.0 mio. €.

Contributor to several marine tasks under the EEA Topic Centre on Water
(ETC/Water). Duration: 2008-2010. In 2010, focus is on tasks related to
marine protected areas, assessment tools as well as pan-European marine
assessments (SoE 2010).
Annual budget for ETC/Water: ~ 1.150.000 €.
Annual budget for DHI in 2007-2009: ~ 100.000 €. Budget for 2010 is ~
45.000 €.

Project Manager of KARMA, a project funded by the Danish Spatial and
Environmental Planning Agency and the Swedish EPA focusing on delineation of and data availability within the Kattegat and its catchment area. Other partners are the National Environmental Research Institute and Swedish
Meteorological and Hydrological Institute. Duration: 2009-2010. Budget:
55.000 €, DHI’s share is 40%.

Conference Secretary for the 3rd International Symposium on Research
and Management of Eutrophication in Coastal Ecosystems (EUTRO
2010), which takes place 15-18 June 2010. Duration: 2009-2011. Please
see www.eutro2010.dhi.dk for details.
Budget, phase I: 300.000 dkk.
Budget, phase II: ~ 2.0 mio. dkk.

Project Partner in HELCOM HOLAS: The HELCOM Holistic Assessment of
the Baltic Sea Environment. Chair of the HELCOM HOLAS Task Force as
well as contributor. Funded by DG Environment and the Ministry of Environment in Sweden. Duration: 2009-2010.
Budget, DHI: 79.600 €.

Member of the EU Task Group on Eutrophication (MSFD ETG). Funded
by the Danish Spatial and Environmental Planning Agency. Duration: 2009.
Budget: 125.000 dkk.

Guest editor of a Special Issue of Hydrobiologia (together with Prof.
Daniel J. Conley, Lund University, Sweden). The Special Issue includes 22
papers based on presentation from the International Symposium on Research and Management of Eutrophication in Coastal Ecosystems (EUTRO
2006), which took place in June 2006.
63
Year
Firm
Position and responsibilities

Project Manager of MARCOS: A NMR funded project focusing on “Marine
European Directives: Concepts, Overlap and Synergies”. Duration: 20072009.
Budget: 725.000 dkk.

Project Supervisor of CONFIRM: A NMR funded project focusing on confidence rating of eutrophication assessments. Duration 2008-2009.
Budget: 300.000 dkk.

Work Package Leader (WP6) in HELCOM BIO, phase II: An integrated
thematic assessment of biodiversity in the Baltic Sea. The focus is on development of a tool for assessment of conservation status in the Baltic Sea.
Duration: 2007-2008. Budget is linked to EUTRO-PRO and “Udvikling af et
marint tilstandsvurderingsværktøj for Natura 2000 områder”.

Project Partner in MOPODECO: A NMR funded project focusing on modelling of habitats in the Baltic Sea. Duration: 2008-2009.
Budget: 1.2 mio. dkk.

Project Manager of OxyBas, phase I: A project aiming at a Full Application
to the Swedish EPA / FORMAS describing the project “Oxygenation of sediments in the Baltic Sea by ecological engineering” (OxyBas, phase II). Duration: 2008.
Budget: 100.000 sek.

Project Manager of BALANCE Synthesis, part 1 and 2: A follow-up project
on the BALANCE project, funded by the Danish Spatial and Environmental
Planning Agency. Duration: 2008-2009.
Budget: 300.000 dkk.

Project Manager of HELCOM EUTRO-PRO: An integrated thematic assessment of eutrophication in the Baltic Sea. Duration: 2006-2008.
Budget for 2006/2007: 27.000 €.
Budget for 2007/2008: 30.000 €.

Project Manager of BALANCE, an INTERREG IIIB project focusing on
mapping and management of marine habitats as well as development of
templates and tools for marine spatial planning in the Baltic Sea. Duration:
2005-2007.
Budget: 4.700.000 €.
Budget related to spin-out projects: 1.2 mio. dkk.

Project Supervisor of “Fagligt grundlag for genetablering af stenrev i
Limfjorden”: A pilot project funded by the Danish Forest & Nature Agency
focusing on restoration of stone reefs in Limfjorden as well as development
of ‘supplementary measures’ sensu the WFD. Duration: 2007-2008.
Budget for pilot project: 500.000 dkk.

Project Partner in IGLOO: A NOVANA funded project developing climate
change indicators. Duration: 2007-2008.
Budget for DHI: 42.000 dkk.

Project Partner in “Udvikling af et marint tilstandsvurderingsværktøj for
Natura 2000 områder”, sub-contracted by the National Environmental Research Institute: The project is funded by the Danish Spatial and Environmental Planning Agency focusing on developing a tool for assessment of
‘conservation status’ sensu the EC Habitats Directive. Duration: 2007-2008.
Budget for DHI: 52.000 dkk.
64
Year
2001 – 2004
Firm
Position and responsibilities

Marine Team Leader: European Topic Centre on Water (ETC/Water).
Duration: 2007.
Annual budget for ETC/Water: ~1.045.000 €.

Project Supervisor of OSPAR COMP-2: A project funded by the Danish
EPA focusing on assessment of eutrophication in the North Sea, Skagerrak
and Kattegat. Duration: 2007-2008. Budget: 300.000 dkk.

Project Manager of HELCOM BIO, phase I: A pilot project producing a
synopsis for the production of an integrated thematic assessment on biodiversity in the Baltic Sea. Duration: 2006. Budget: 20.000 €.

Project Manager of BSPC EUTRO: Writing a booklet on eutrophication in
the Baltic Sea focusing on effects, causes and solutions. Duration: 2006.
Budget: 100.000 dkk.

Project Manager of MST CO-EUTRO. Duration: 2005-2007. Budget:
300.000 dkk.

Conference Secretary for the 2nd International Symposium on Research
and Management of Eutrophication in Coastal Ecosystems (EUTRO
2006), which took place 20-23 June 2006. Duration: 2005-2006. Budget: 2.8
mio. dkk.

Project Manager of MST HEAT, a project supporting HELCOM EUTRO.
Duration: 2005.
Budget: 200.000 dkk.

Project Manager of HELCOM EUTRO (Development of tools for a thematic
eutrophication assessment). Duration: 2005. Budget: 30.000 €.

Project Manager of SNS/MST SYNERGY (Synergies and overlap between
the EC Habitats Directive and the EU Water Framework Directive). Duration: 2005.
Budget: 100.000 dkk.
National Environmental Research Institute, 
Dept. of Marine Ecology





Chief Consultant and Head of the Danish National Marine Monitoring Focal
Point (M-FDC), secretary of the National Steering Group on Marine Monitoring and Assessment.
Co-ordination of the revision of the NOVA programme (1998-2003) and the
design of the NOVANA programme (2004-2009).
Co-ordinator of the departments work in relation to HELCOM and OSPAR
plus participation in relevant committees and working groups working with
monitoring and assessment of environment and nature in Danish waters.
Danish head-of-delegation in the OSPAR Assessment and Monitoring
Group (ASMO). Danish marine representative in Nordic Council of Ministers’ Sea and Air Group, and Danish representative in the HELCOM/ICES
SGQAB.
Participant in the WFD CIS Coast group and the pan-European Eutrophication Activity.
Participation in EU projects under the 5th framework programme (FP5):
CHARM, DANLIM and MOLTEN.
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Year
Firm
Position and responsibilities
1999 – 2000
Danish Environmental Protection Agency, 
Waste Water and Groundwater Division.
(Until May 2000: Waste Water and Aquatic 
Monitoring Division)


1999
National Forest and Nature Agency, 
Ecological Division (ED)


Implementation of the Water Framework Directive, in particular regarding
principles for establishing ecological quality standards.
In addition, (i) political/ administrative summary report on status and perspectives for management of the Danish aquatic environment (Vandmiljø2000), (ii) co-ordination of the Agency’s viewpoints in relation to the National Monitoring Board and NOVA-2003, and (iii) National Steering Group on
Hydrological Point Sources.
Co-editor of the NOVA-2003 programme document, Vandmiljø-99, national
guidelines on reporting of the NOVA programme, guidelines on annual
evaluation of sampling, data flow and reporting.
Co-ordination of Danish participation in the HELCOM MONAS group.
Legal and technical casework in relation to the Danish Watercourse Act.
Supervision in relation to the surveillance monitoring of freshwaters carried
of by the Danish counties
ED Work Programme 1999 and RTD projects.
1998
Danish Environmental Protection Agency, 
Freshwater and Waste Water Division



Casework in relation to the Danish Watercourse Act.
Evaluation report on the 1993-1997 National Monitoring Programme.
Guidelines for 1999 reporting of the NOVA programme.
A synopsis for HELCOM’s 4th Periodic Assessment of the State of the
Marine Environment of the Baltic Sea 1994-1998 on behalf of the Marine
Division.
1994 – 1998
Danish Environmental Protection Agency, 
Marine Division

Participation in the preparation of the 4th North Sea Conference in 1995.
Marine fish farming 1994-1998 (technical casework, data management and
reporting).
Danish Head-of-Delegation HELCOM EC 1994-1998.
Professional Secretary to the National Monitoring Board (1995-1998) and
the National Revision Task Team (1996-1998).
Chairman of the National Steering Group on Marine Monitoring (19951998).
Preparation of a large number of meetings and an equal number of summary records.
Co-ordination of the ’closing down’ of the Danish Marine Research Programme (Hav90), and contribution of the report summarising the most important results of Hav90.





1992 – 1994
BioConsult (now SBH-consult)
Consultant for the Danish EPA. Task: management of the Danish Marine Research Programme (Hav90). The work included general project management,
including reimbursement, budgets, publication of project reports, secretary to the
National Marine Research Advisory Board well as planning of the International
Symposium on Nutrient Dynamics in Coastal and Estuarine Environments, 13-16
October 1993 (EUTRO 1993), the latter including work in relation to the production of Symposium Proceedings.
1990 – 1992
Nordic Council of Ministers (Danish Envi- Project secretary working for the Nordic Council of Ministers (NCM) project on
ronmental Protection Agency)
marine monitoring. The main objective of the project was to compile a framework
for an improved co-ordination of marine monitoring in the Nordic countries. Conclusions and recommendations from this work have to a large extent influenced
the Baltic Monitoring Programme/ COMBINE and the marine sub-programme of
NOVA-2003.
1988 – 1990
National Environmental Research Institute, Student assistant in M-FDC and for some periods working with research projects
Dep. of Marine Ecology
(NPo Research Programme and Hav90).
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Languages
Danish
English
German
French
Swedish
5
5
5
5
5
5
2
2
1
1
2
1
4
5
3
Speaking
Reading
Writing
(Mother tongue/excellent: 5; Average: 3-4; Poor: 1-2)
Publications in English
Currently, I have more than 80 publications in English. Most are Technical Reports, but the list also includes a growing number of peer reviewed papers as well as three reviewed books and a special issue of a
scientific journal.
As a spin off, I have reviewed manuscripts for the following scientific journals: (1) Biogeochemistry; (2)
Environmental Management; (3) Environment International; (4) Estuaries & Coasts (formerly Estuaries); (5)
Estuarine, Coastal, and Shelf Science; (6) Hydrobiologia; and (7) Marine Biology Research.
Bold
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2012
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2011
= indicates a reviewed publication.
= indicates an assessment.
Laamanen, M., S. Korpinen, U.L. Zweifel & J.H. Andersen (submitted): Ecosystem health In: Biological Oceanography of
the Baltic Sea (Eds: Pauline Snoeijs, Hendrik Schubert & Teresa Radziejewska). Springer.
Andersen J.H. (accepted): Ecosystem-based management of coastal eutrophication. Connecting science, policy and society. Ph.D. thesis. University of Copenhagen. 56 pp + annexes.
Andersen, J.H. (accepted): BEAT, the HELCOM Biodiversity Assessment Tool. In: Skov et al. (accepted): Modelling of the
potential coverage of habitat-forming species and development of tools to evaluate the conservation status of the marine Annex I habitats. Tema Nord Report. Nordic Council of Ministers.
Carstensen, J., J.H. Andersen, K. Dromph, V. Fleming-Lehtinen, S. Simis, B. Gustavsson, A. Norkko, H. Radke, D.L.J.
Petersen & T. Uhrenholdt (accepted): Approaches and methods for eutrophication target setting in the Baltic Sea region. Baltic Sea Environemt Proceesings. 133 pp.
Andersen, J,H., J.W. Hansen, M. Mannerla, S. Korpinen & J. Reker (accepted): A glossary of terms commonly used in the
Marine Strategy Framework Directive. NERI Technical Report. 31 pp.
Korpinen, S., L. Meski, J.H. Andersen & M. Laamanen (2012): Human pressures and their potential impact on the Baltic
Sea ecosystem. Ecological Indicators 15:105-114. http://dx.doi.org/10.1016/j.ecolind.2011.09.023
Andersen, J.H., P. Axe, H. Backer, J. Carstensen, U. Claussen, V. Fleming-Lehtinen, M. Järvinen, H. Kaartokallio, S.
Knuuttila, S. Korpinen, M. Laamanen, E. Lysiak-Pastuszak, G. Martin, F. Møhlenberg, C. Murray, G. Nausch, A.
Norkko, & A. Villnäs (2010): Getting the measure of eutrophication in the Baltic Sea: towards improved assessment
principles and methods. Biogeochemistry. DOI 10.1007/s10533-010-9508-4.
http://www.springerlink.com/content/x76wq76863458471/fulltext.pdf
Andersen, J.H. & J. Carstensen (2011): Reference conditions and acceptable deviation: Concepts, definitions and their
practical use. HELCOM TARGREV Working Document. 21 pp.
Ferreira, J.G., J.H. Andersen, A. Borja, S.B. Bricker, J. Camp, M. Cardoso da Silva, E. Garcés, A.-S. Heiskanen, C. Humborg, L. Ignatiades, C. Lancelot, A. Menesguen, P. Tett, N. Hoepffner & U. Claussen (2011): Indicators of human-induced eutrophication to assess the environmental status within the European Marine Strategy Framework Directive.
Estuarine, Coastal and Shelf Science. DOI: 10.1016/j.ecss.2011.03.014.4
Murray, C., J.H. Andersen, H. Kaartokallio, P. Axe, J. Molvær, K. Norling & M. Krüger-Johansen (2011): Confidence rating
of marine eutrophication assessments. Tema Nord 2011:504. 75 pp.
Andersen, J.H., S. Bricker, J. Carstensen & J.E. Larsen (2011): ICES/DHI/NOAA Third International Symposium on Research and Management of Eutrophication in Coastal Ecosystems. 6 pp + Online Supplementary Material. In: ICES
(2011): ICES Symposium Report 2010. ICES CM 2011/GEN 0x.
This paper was highlighted by the European Commission in ’Science for Environment Policy’, a news alert service from DG
Environment, please see: http://ec.europa.eu/environment/integration/research/newsalert/pdf/252na4.pdf.
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2010
ETC/W (2010): Delineation of Marine Regions in EU Directives and Regional Sea Conventions. Draft Scoping Report edited
by A. Stock, J.H. Andersen, B.M. Sharry & H.M. Jensen. 22 pp.
EUSeaMap (2010): EUSeaMap Final Report. Preparatory Action for development and assessment of a European broadscale seabed habitat map. Contribution by J. Reker & J.H. Andersen. EC contract no. MARE/2008/07. 223 pp.
Kaartokallio, H., J.H. Andersen, J.N. Jensen, A. Künitzer, N. Green & M. Peterlin (2010): Review of precursors on initial
assessment from Regional Marine Conventions. Scoping report from ETC/W to EEA. 28 pp.
Korpinen, S, L. Meski, J.H. Andersen & M. Laamanen (2010): Towards a tool for quantifying anthropogenic pressures and
potential impacts on the Baltic Sea marine environment. A background document on the method, data and testing of
the Baltic Sea Pressure and Impact indices. Baltic Sea Environmental Proceedings No. 125. 73 pp.
Christiansen, T. (Ed.), A. Meiner, B. Werner, C. Romao, E. Gelabert, R.P. Collins, R. Uhel, A. Ruus, A.-S. Heiskanen,
A. Künitzer, A. Raike, B. Bjerkeng, C. Emblow, G. Coppini, H. Sparholt, J.H. Andersen, J.-M. Leppänen, J.N.
Jensen, M. Lago, M. Peterlin, N. Pinardi, N. Holdsworth, N. Green, P. Degnbol, B. Mac Sharry, S. Condé, A.I.
Campos, J. Orr & S. van den Hove (2010): The European Environment. State and Outlook 2010. Marine and
Coastal Environment. European Environment Agency, Copenhagen. 58 pp.
J.H. Andersen & S.B. Bricker (2010): EUTRO 2010 Report-out. CERF Newsletter 36(3):24.
DHI, ICES & NOAA (2010): Third International Symposium on Research and Management of Eutrophication in Coastal Ecosystems. 15-18 June 2010, Nyborg, Denmark. Programme and Abstracts. Edited by J.H. Andersen. 44 pp.
Andersen, J.H., L. Hasselström, S. Korpinen, M. Laamanen, A. Soutukorva & U. Volpers (2010): Chapter 1: Introduction. Pages 6-13 in: HELCOM (2010): Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment
2003-2007. Baltic Sea Environmental Proceedings 122. Helsinki Commission. 63 pp.
Andersen, J.H., S. Korpinen, M. Laamanen & C. Murray (2010): 2.1 Integrated and Holistic Assessments. Pages 1415 in: HELCOM (2010): Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment 2003-2007.
Baltic Sea Environmental Proceedings 122. Helsinki Commission. 63 pp.
Andersen, J.H. (2010): 2.2 Eutrophication. Pages 16-17 in: HELCOM (2010): Ecosystem Health of the Baltic Sea.
HELCOM Initial Holistic Assessment 2003-2007. Baltic Sea Environmental Proceedings 122. Helsinki Commission. 63 pp.
S. Korpinen, J.H. Andersen, M. Laamanen & C. Murray (2010): 2.3 Hazardous substances. Pages 18-21 in: HELCOM
(2010): Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment 2003-2007. Baltic Sea Environmental Proceedings 122. Helsinki Commission. 63 pp.
Zweifel, U.L., S. Korpinen, R. Ljungberg & J.H. Andersen (2010): 2.4 Biodiversity. Pages 22-26 in: HELCOM (2010):
Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment 2003-2007. Baltic Sea Environmental
Proceedings 122. Helsinki Commission. 63 pp.
M. Laamanen, J.H. Andersen, U. Claussen, M. Durkin, S. Korpinen, J. Reker, M. Stankiewicz & U. Volpers (2010):
Chapter 4: What are the solutions? Pages 42-49 in: HELCOM (2010): Ecosystem Health of the Baltic Sea.
HELCOM Initial Holistic Assessment 2003-2007. Baltic Sea Environmental Proceedings 122. Helsinki Commission. 63 pp.
U. Volpers, Andersen, J.H., S. Korpinen & M. Laamanen (2010): Chapter 6: Conclusions and outlook. Pages 54-57
in: HELCOM (2010): Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment 2003-2007. Baltic Sea Environmental Proceedings 122. Helsinki Commission. 63 pp.
HELCOM (2010): Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment 2003-2007. Edited by
J.H. Andersen, S. Korpinen, M. Laamanen & U. Wolpers. Baltic Sea Environmental Proceedings 122. Helsinki
Commission. 63 pp. http://www.helcom.fi/stc/files/Publications/Proceedings/bsep122.pdf
HELCOM (2010): Hazardous substances in the Baltic Sea. An integrated thematic assessment of hazardous substances in the Baltic Sea. Edited by S. Korpinen & M. Laamanen with contributions from J.H. Andersen, L.
Asplund, U. Berger, A. Bignert, E. Boalt, K. Broeg, A. Brzozowska, I. Cato, M. Durkin, G. Garnaga, K. Gustavson, M. Haarich, B. Hedlund, P. Köngäs, T. Lang, M.M. Larsen, K. Lehtonen, J. Mannio, J. Mehtonen, C.
Murray, S. Nielsen, B. Nyström, K. Pazdro, P. Ringeltaube, D. Schiedek, R. Schneider, M. Stankiewicz, J.
Strand, B. Sundelin, M. Söderström, H. Vallius, P. Vanninen, M. Verta, N. Vieno, P. Vuorinen and A. Zaharov.
Baltic Sea Environmental Proceedings 120B. Helsinki Commission. 116 pp.
Andersen, J.H., C. Murray, H. Kaartokallio, P. Axe & J. Molvær (2010): A simple method for confidence rating of eutrophication status assessments. Marine Pollution Bulletin 60:919-924. doi:10.1016/j.marpolbul.2010.03.020.
http://www.sciencedirect.com/science?_ob=MImg&_imagekey=B6V6N-4YT7KN7-97&_cdi=5819&_user=684530&_pii=S0025326X10001104&_orig=search&_coverDate=06%2F30%2F2010&_sk=9993
99993&view=c&wchp=dGLbVlb-zSkzk&md5=ad2b416d4eff4bf7e9a3a99eb6bbb9aa&ie=/sdarticle.pdf
Andersen, J.H., J. Dørge, H. Skov, A. Stock, J. Carstensen, K. Dahl, M. Hjorth, A.B. Josefsson, M.M. Larsen, J. Strand, P.
Andersson, P. Axe, J. Reker & S. Korpinen (2010): Delineation scenarios for the Kattegat, data availability and management support tools. DHI Technical Report to the Agency for Spatial and Environmental Planning, Denmark. 86 pp.
ETC/W (2010): Assessment of the European Marine Environment. Background Report for the SoER 2010, part B. By: J.-M.
Leppänen, A.-S. Heiskanen, M. Viitasalo, H. Rouse (Eds.), M. Peterlin, C. Emblow, N.W. Green, G. Coppini, J.
Dorandeu, G. Larnicol, S. Marullo, P. Lowe, N. Pinardi, J.N. Jensen, J.H. Andersen, H. Peltonen, M. Raateoja & A.
Räike. European Environment Agency. 111 pp.
Ferreira, J.G., J.H. Andersen, A. Borja, S.B. Bricker, J. Camp, M. Cardoso da Silva, E. Garcés, A.-S. Heiskanen, C. Humborg, L. Ignatiades, C. Lancelot, A. Menesguen, P. Tett, N. Hoepffner & U. Claussen (2010): Marine Strategy
Framework Directive. Task Group 5 Report. Eutrophication 49 pp.
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Andersen, J.H. & D.J. Conley (editors) (2009): Eutrophication in Coastal Ecosystems. Towards better understanding and
management strategies. Developments in Hydrobiology 207. 269 pp. Previously published in Hydrobiologia 629(1).
http://www.springer.com/environment/aquatic+sciences/book/978-90-481-3384-0.
Laamanen, M. & J.H. Andersen (2009): Eutrophication. In: HELCOM (2009): Biodiversity in the Baltic Sea. An integrated
thematic assessment of biodiversity and nature conservation in the Baltic Sea. Ed. by U.L. Zweifel. Baltic Sea
Environmental Proceedings No. 116B. Helsinki Commission. 188 pp.
Backer, H. & J.H. Andersen (2009): Towards an indicator-based assessment of the Baltic Sea biodiversity. In: HELCOM
(2009): Biodiversity in the Baltic Sea. An integrated thematic assessment of biodiversity and nature conservation in the
Baltic Sea. Ed. by U.L. Zweifel. Baltic Sea Environmental Proceedings No. 116b. Helsinki Commission. 188 pp.
Andersen, J.H. & D.J. Conley (guest editors) (2009): Eutrophication in coastal ecosystems. Hydrobiologia. 629(1), 269 pp.
http://www.springerlink.com/content/t8j727n2k266/?p=1b40a1097a1b47a9842b02ba104ea0b6&pi=5
Andersen, J.H. & D.J. Conley (2009): Eutrophication in coastal marine ecosystems: towards better understanding and
management strategies. Hydrobiologia 629(1):1-4. http://www.springerlink.com/content/w707717n84j65571/fulltext.pdf
Andersen, J.H., S. Korpinen & M. Laamanen (2009): Towards a holistic assessment of environmental status in the Baltic
Sea. HOLAS roadmap. DHI Technical Report to HELCOM. 43 pp.
Dahllöf, I. & J.H. Andersen (2009): Hazardous and Radioactive Substances in Danish Marine Waters. Status and
Temporal Trends. Danish Spatial and Environmental Planning Agency & National Environmental Research
Institute. 110 pp. http://www2.dmu.dk/pub/OSPAR_Hazardous_Substances_print.pdf
HELCOM (2009): Eutrophication in the Baltic Sea. An integrated thematic assessment of eutrophication in the Baltic Sea
region: Executive Summary. Edited by J.F. Pawlak, M. Laamanen & J.H. Andersen. Baltic Sea Environmental
Proceedings No. 115A. Helsinki Commission. 19 pp.
HELCOM (2009): Eutrophication in the Baltic Sea. An integrated thematic assessment of eutrophication in the
Baltic Sea region. Ed. by J.H. Andersen & M. Laamanen. Baltic Sea Environmental Proceedings No. 115B.
Helsinki Commission. 148 pp.
http://meeting.helcom.fi/c/document_library/get_file?p_l_id=79889&folderId=377779&name=DLFE-36818.pdf
Andersen, J.H. & H. Backer (2008): Development of an indicator-based tool for assessment of biodiversity in the Baltic
Sea. DHI Technical Report to HELCOM HABITAT and HELCOM BIO. 34 pp.
Andersen, J.H. & H. Kaas (2008): Danish assessment of eutrophication status in the North Sea, Skagerrak and
Kattegat: OSPAR Common Procedure 2001-2005. DHI Technical Report to the Danish Spatial and
Environmental Planning Agency. 86 pp.
ETC/Water (2008): Improving EEA marine indicators. A review of their performance and suggested ‘next steps’. Final Draft
Scoping Report edited by J.H. Andersen, P. Kuuppo, T. Christiansen & E.R. Gelabert. 77 pp.
Andersen, J.H., A. Erichsen, K. Garde, C. Murray & F. Møhlenberg (2007): Strengthening the Tools for Assessment of
Coastal Eutrophication in Russian Waters of the Baltic Sea. DHI Technical Report to the Danish Environmental
Protection Agency. 61 pp.
Hansen, I.S., N. Keul, J.T. Sørensen, A. Erichsen & J.H. Andersen (2007): Baltic Sea oxygen maps. BALANCE Interim
Report No. 17. 36 pp.
Andersen, J.H. & F. Møhlenberg (2007): Testing of the draft HELCOM Eutrophication Assessment Tool (HEAT) in 45
basins and coastal water bodies of the Baltic Sea. DHI Technical Report to HELCOM. 50 pp.
Andersen, J.H., & A. Erichsen (2007): Modelling of reference conditions in the Baltic Sea. DHI Technical Report to
HELCOM. 18 pp. + annexes.
Al-Hamdani, Z. & J. Reker (eds.), J.H. Andersen and 22 others (2007): Towards marine landscapes in the Baltic Sea
ecoregion. BALANCE Interim Report No. 10. 117 pp.
EEA & ETC/Water (2007): Towards a ‘converging’ framework for marine monitoring and assessment of European marine
waters. Synthesis of EEA-led workshops on Operational oceanography, Ecological processes and biological elements
and Chemical loads and burdens. Edited by E.R. Gelabert & J.H. Andersen. 43 pp. + annexes.
Andersen, J.H & J. Reker (2007): BALANCE Newsletter No. 3. 4 pp.
Andersen, J.H., C. Murray & H. Skov (2007): Interim overview of reporting obligations, monitoring activities and data
available for the HELCOM integrated thematic assessment of biodiversity and nature conservation in the Baltic Sea.
DHI Technical Report to HELCOM. 29 pp.
Andersen, J.H., H.B. Nielsen & H. Skov (2006): Getting the measure of biodiversity in the Baltic Sea. DHI Technical Report
to HELCOM. 21 pp.
HELCOM (2006): Development of tools for assessment of eutrophication in the Baltic Sea. Baltic Sea Environment Proceedings 104. 62 pp. Edited by J.H. Andersen.
Andersen, J.H. (2006): Project Proposal: Towards an integrated thematic assessment of eutrophication in the Baltic Sea.
DHI Technical Report to HELCOM. 46 pp.
DHI, Danish EPA, Fyn County & Swedish EPA (2006): Research and Management of Eutrophication in Coastal Ecosystems. An International Symposium, 20-23 June 2006, Nyborg, Denmark. Programme and Book-of-Abstracts. 94 pp.
Edited by J.H. Andersen.
Andersen, J.H & J. Reker (2006): BALANCE Newsletter No. 2. 4 pp.
Andersen, J.H & J.T. Pawlak (2006): Nutrients and Eutrophication in the Baltic Sea. Effects / Causes / Solutions. Booklet
produced for the Baltic Sea Parliamentary Conference. 32 pp.
Andersen, J.H., L. Schlüter & G. Ærtebjerg (2006): Coastal eutrophication: recent developments in definitions and implications for monitoring strategies. Journ. Plankt. Res. 28(7):621-628.
http://plankt.oxfordjournals.org/cgi/reprint/28/7/621
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Petersen, J.K., J.H. Andersen, K. Dahl, O.S. Hansen, A.B. Josefson, J. Karlsson, L.-O. Loo, J. Magnusson, F. Moy & P.
Nilsson (2006): Reference conditions and EQOs for aquatic vegetation and macrozoobenthos. TemaNord 2006:510.
138 pp.
Carstensen, J., D.J. Conley, J.H. Andersen & G. Ærtebjerg (2006): Coastal eutrophication and trend reversal: A Danish
case study. Limnology & Oceanology 51:398-408. http://aslo.org/lo/toc/vol_51/issue_1_part_2/0398.pdf
Andersen, J.H. & J. Reker (2005): BALANCE Newsletter No. 1. 4 pp.
Petersen, J.K., Hansen, O.S., Henriksen, P., Carstensen, J., Krause-Jensen, D., Dahl, K., Josefson, A.B., Hansen, J.L.S.,
Middelboe, A.L. & Andersen, J.H. (2005): Scientific and technical background for intercalibration of Danish coastal
waters. National Environmental Research Institute, Denmark. 72 pp. - NERI Technical Report No. 563.
Andersen, J.H, F. Møhlenberg, T. Uhrenholdt, M.H. Jensen, B. Sømod & P. Henriksen (2005): Testing of the HELCOM
Eutrophication Assessment Tool (HEAT) in Danish Marine Waters. DHI Technical Report. 32 pp.
Andersen, J. H. (Ed.) (2005) Marine waters. In: Svendsen, L. M., Bijl, L. van der, Boutrup, S. and Norup, B. (Eds.) 2005:
NOVANA: Nationwide Monitoring and Assessment Programme for the Aquatic and Terrestrial Environments. Programme Description - Part 2. National Environmental Research Institute, Denmark. 137 pp. - NERI Technical Report
No. 537.
Andersen, J.H., D.J. Conley & S. Hedal (2004): Palaeo-ecology, reference conditions and classification of ecological
status: The EU Water Framework Directive in practice. Mar. Poll. Bul. 49:282-290.
http://www.sciencedirect.com/science?_ob=MImg&_imagekey=B6V6N-4CPDF84-37&_cdi=5819&_user=684530&_orig=search&_coverDate=08%2F31%2F2004&_sk=999509995&view=c&wchp=dGLb
VlW-zSkWb&md5=a3554a13dd8d91c63139f6cc2593fc31&ie=/sdarticle.pdf
Christiansen, T., J. Andersen & J.B. Jensen (2004): Defining a typology for Danish coastal waters. Coastline Reports
4(2004):49-54.
Dahl, K., Larsen, M.M., Rasmussen, M.B., Andersen, J.H., Petersen, J.K., Josefson, A.B., Lundsteen, S., Dahllöf, I., Christiansen, T., Helmig, S.A. & Reker, J. (2004): Tools to assess the conservation status of marine habitats in special areas of conservation. Phase 1: Identification of potential indicators and available data. National Environmental Research
Institute. – Technical Report from NERI.
Andersen, J.H. & O.S. Hansen (2003): Background, definition, causes and effects. Pages 7-17 in: Ærtebjerg, G., J.H.
Andersen & O.S. Hansen (2003): Nutrients and Eutrophication in Danish Marine Waters. A Challenge to Science and Management. National Environmental Research Institute. 126 p.
Andersen, J.H. & O.S. Hansen (2003): Fish kills in coastal waters. Pages 80-83 in: Ærtebjerg, G., J.H. Andersen &
O.S. Hansen (2003): Nutrients and Eutrophication in Danish Marine Waters. A Challenge to Science and Management. National Environmental Research Institute. 126 p.
Andersen, J.H., J.B. Jensen & H. Karup (2003): Responses and adaptive management. Pages 85-99 in: Ærtebjerg,
G., J.H. Andersen & O.S. Hansen (2003): Nutrients and Eutrophication in Danish Marine Waters. A Challenge
to Science and Management. National Environmental Research Institute. 126 p.
Ærtebjerg, G., J.H. Andersen & O.S. Hansen (2003): Summary, conclusions and the future. Pages 101-109 in:
Ærtebjerg, G., J.H. Andersen & O.S. Hansen (2003): Nutrients and Eutrophication in Danish Marine Waters. A
Challenge to Science and Management. National Environmental Research Institute. 126 p.
Ærtebjerg, G., J.H. Andersen & O.S. Hansen (2003): Nutrients and Eutrophication in Danish Marine Waters. A Challenge to Science and Management. National Environmental Research Institute. 126 p.
http://www2.dmu.dk/1_Viden/2_Publikationer/3_ovrige/rapporter/Nedmw2003_alle.pdf
Conley, D., S. Markager, J. Andersen, T. Ellermann & L.M. Svendsen (2002): Coastal Eutrophication and the Danish
National Aquatic Monitoring and Assessment Program. Estuaries 25(4b): 848-861.
http://www.springerlink.com/content/95780457274837t6/fulltext.pdf
Conley, D., J. Andersen, J. Carstensen & P. Henriksen (2001): Long-term trends in nutrient loading, nutrient concentrations
and nutrient limitation in Danish estuaries. OSPAR MON 2001.
Aquatic Environment 1999. State of the Danish Aquatic Environment. Environmental Investigations, no. 3/2000.
Eds: J. Andersen & D. Barry. 138 p.
Christensen, P.B., F. Møhlenberg, L.C. Lund-Hansen, J. Borum, C. Christiansen, S.E. Larsen, M.E. Hansen, J. Andersen &
J. Kirkegaard (1998): The Danish Marine Environment: Has Action Improved its State? - Havforskning fra Miljøstyrelsen, nr. 62. 115 p.
Andersen, J. (1998): Draft Synopsis for the 4th Periodic Assessment of the State of the Marine Environment of the Baltic
Sea 1994-1998. Version 1.0. 28 p + annexes.
Aquatic Environment 1994. Overall Trend in Point-source Discharges and Status of the Danish Aquatic Environment. Environmental Investigations, no. 1/1996. Eds: J. Andersen & T. Christensen. 151 p.
Scientific Symposium on the North Sea Quality Status Report, 18-21 April 1994, Ebeltoft, Denmark. Proceedings. Eds.: J.
Andersen, H. Karup & U.B. Nielsen. Danish Environmental Protection Agency 1996. 346 p.
Progress Report. 4th International Conference on the Protection of the North Sea. Eds.: J. Andersen & T. Niilonen. Danish
Environmental Protection Agency, 1995. 247 p.
Andersen, J. & J. Kirkegaard (1992): The Marine Research Programme in Denmark. North Sea Task Force News 4:6-7. 2
p.
A Nordic Strategy for the Co-ordination and Enhancement of Marine Monitoring and Assessment. Ed.: J. Andersen. HELCOM EC2/INF.7, 1991. 8 p.
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Publications in Danish
= indikerer at publikationen er reviewet.
= indikerer at publikationen er en tilstandsrapport/assessment.
Fed
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2012
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Bach, H., I.K. Petersen, R.D. Nielsen, T. Fox, C. Topping, B. Nygaard, M. Elmeros, B. Søgaard, J. Kahlert, S. Sveegaard, R.
Dietz, J. Tougaard, J.N. Nielsen, J. Teilmann, A.B. Josefson, C. Mohn, J.L.S. Hansen, K. Timmermann, J.H. Andersen,
M. Maar, H.H. Jacobsen, E. Friis Møller, K. Dahl, L.C. Lund-Hansen, P. Grønkjær, M.B. Rasmussen, M. Winther, L. Martinsen, M. Zandersen, H.R. Olesen & B. Münier (accepted): Miljøfagligt review af VVM-redegørelsen for Femern forbindelsen. Miljø-beskrivelsen. Videnskabelig rapport fra DCE - Nationalt Center for Miljø og Energi. 86 pp.
Andersen, J.H. & C. Murray (accepted): Foreløbig integreret vurdering af miljøtilstanden i de danske farvande - en indikatorbaseret vurdering. Fagligt notat fra DCE - Nationalt Center for Miljø og Energi. 11 pp.
Andersen, J.H., M.M. Larsen, C. Murray & J. Strand (accepted): En integreret vurdering og klassifikation af den kemiske tilstand i de danske farvande - en indikatorbaseret vurdering. Fagligt notat fra DCE - Nationalt Center for Miljø og Energi. 11
pp.
Andersen, J.H., S. Korpinen & A. Stock (accepted): Foreløbig vurdering af kumulative påvirkninger og belastninger i de danske farvande. Fagligt notat fra DCE - Nationalt Center for Miljø og Energi. 14 pp.
Hansen, J.W., J.H. Andersen, J. Strand & T.K. Sørensen (accepted): Affald i havet. Fagligt notat fra DCE - Nationalt Center for
Miljø og Energi. 28 pp.
J.H. Andersen, J.W. Hansen & J. Carstensen (accepted): Væsentlige ændringer i temperatur- og salinitetsforholdene i de
danske farvande forårsaget af menneskelige aktiviteter. Fagligt notat fra DCE - Nationalt Center for Miljø og Energi. 13
pp.
J.H. Andersen, C. Göke & C. Murray (accepted): Klassifikation af af biodiversitetstilstanden i de danske farvande – en indikator-baseret statusvurdering. Fagligt notat fra DCE - Nationale Center for Miljø og Energi. 30 pp.
J.H. Andersen, C.D. Pommer, J.W. Hansen & P. Dolmer (accepted): Foreløbig karakterisering af fysisk skader forårsaget af
råstofindvinding og bundtrawling i de danske farvande. Fagligt notat fra DCE - Nationalt Center for Miljø og Energi. 27 pp.
J.H. Andersen, J.W. Hansen, C. Murray, C. Göke & D.LJ. Petersen (accepted): Klassifikation af eutrofieringstilstanden i de
danske farvande – en indikator-baseret statusvurdering. Fagligt notat fra DCE - Nationalt Center for Miljø og Energi. 42
pp.
Andersen, J.H. & J. Carstensen (2011): Gisp. Vandmiljøplanerne virker. Politiken, 15. oktober 2011. Debat-sektionen side 8.
Andersen, J.H. (2011): ’Havet omkring Danmark’ – et forslag til synopsis for havstrategidirektivets basisanalyser. Notat fra
DMU. 49 pp.
Josefson, A., D. Krause-Jensen, M.B. Rasmussen, J.H. Andersen & P. Henriksen (2009): Udvikling af indikatorer og tilstandsvurderingsværktøj for marine Natura 2000 områder. Faglig rapport fra DMU, nr. 701. 76 pp.
Møhlenberg, F., J.H. Andersen (Eds.), C. Murray, P.B. Christensen, T. Dalsgaard, H. Fossing & D. Krause-Jensen (2008):
Stenrev i Limfjorden: Fra naturgenopretning til supplerende virkemiddel. DHI Teknisk Rapport til By- og Landskabsstyrelsen. 41 p + bilag.
Hansen, J.W., M. Nedergaard & F. Skov (Eds.) (2008): IGLOO – Indikatorer for globale klimaændringer i overvågningen. DHI
rapport til Miljøcenter Ringkøbing, By- og Landskabsstyrelsen. 91 pp. Med bidrag fra J.H. Andersen.
Kaas, H. (Ed.) (2008): Ny teknologi i overvågningen. DHI Teknisk Rapport til By- og Landskabsstyrelsen. 96 pp. Med bidrag fra
J.H. Andersen.
Andersen, J.H. & H. Skov (2005): Synergi og overlap mellem Habitatdirektivet, Fuglebeskyttelsesdirektivet og Vandrammedirektivet – med fokus på kystvand. DHI Teknisk Rapport til Skov- og Naturstyrelsen og Miljøstyrelsen. 55 pp.
Andersen, J., S. Markager & G. Ærtebjerg (2005): Tekniske anvisninger for marin overvågning 2004 – 2009. Danmarks Miljøundersøgelser. Kan downloades via:
http://www.dmu.dk/Overvaagning/Fagdatacentre/Det+Marine+Fagdatacenter/Tekniske+anvisninger+NOVANA+20042009/
Andersen, J.H., Clarke, A., Conley, D.J., Dahllöf, I., Greve, T.M., Krause-Jensen, D., Larsen, M.M., Nielsen, K. & Reuss, N.
(2005): Eksempler på økologisk klassificering af kystvande. Vandrammedirektiv-projekt fase IIIa. – Faglig rapport fra DMU
nr. 530. 48 pp.
Svendsen, L.M., L. van der Bijl, S. Boutrup & B. Norup (2005): NOVANA. Det nationale program for overvågning af vandmiljøet
og naturen. Programbeskrivelse - del 2. Faglig rapport fra Danmarks Miljøundersøgelser. 128 pp. Med bidrag af J.H. Andersen.
Larsen, M.M., S. Foverskov & J.H. Andersen (2005): Havnesedimenter - Prøvetagning og analyser. Arbejdsrapport fra Miljøstyrelsen nr. 35, 2005. 77 pp.
Ærtebjerg, G. & Andersen, J.H. (red.) og mange flere (2004): Marine områder 2003 – Miljøtilstand og udvikling. – Faglig
rapport fra DMU nr. 513. 97 pp.
Andersen, J.H., J.B. Jensen, D. Krause-Jensen, H.B. Madsen & B. Riemann (2004): Fra vandmiljøplaner til vandplaner og
indsatsprogrammer – med kvælstof som eksempel. 9 pp. I: Ærtebjerg & Andersen (red.) og mange flere (2004): Marine
områder 2003 – Miljøtilstand og udvikling. - Faglig rapport fra DMU nr. 513. 97 pp.
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1994
Christensen, P.B, O.S. Hansen & G. Ærtebjerg (Eds.) (2004): Iltsvind. Med bidrag af J.H. Andersen, J. Carl, J. Carstensen, P.
Clausen, R. Dietz, J. Fenger, T.M. Greve, J.L.S. Hansen, O. Hertel, A.B. Josefson, D. Krause-Jensen, A. Branth Pedersen, I.K. Petersen, J.F. Steffensen & J. Teilmann. Forlaget Hovedland. 128 pp.
Ærtebjerg, G., Manscher O.H. & Andersen, J. (2004): Præliminær evaluering af marine data-værter i Sveriges marine overvågningsprogram. Evalueringsnotat til Naturvårdsverkets Miljöövervakningsenhet. 19 pp.
Riemann, B. J.K. Petersen, D. Conley, A.B. Josefson, D. Krause-Jensen, J.H. Andersen, P. Henriksen, M.B. Rasmussen & P.
Claussen (2004): Faglig udredning af problemer vedrørende tilstand og miljømål for Ringkøbing Fjord. Danmarks Miljøundersøgelser, 51 pp.
Fossing, H., Andersen, J.H. & Dalsgaard, T. (2003): Miljøtilpasset overvågning af saltvandsbaseret fiskeopdræt. Notat. 17 pp.
Rasmussen, M.B. & Andersen, J. (red.) og mange flere. (2003): Marine områder 2002 – Miljøtilstand og udvikling.
NOVA-2003. Danmarks Miljøundersøgelser. 98 pp. + bilag – Faglig rapport fra DMU nr. 467
Dahl, K., Larsen, M.M., Rasmussen, M.B., Andersen, J.H., Petersen, J.K., Josefsson, A.B., Lundsteen, S., Dahllöf, I. & Christiansen, T. (2003): Kvalitetsvurderingssystem for Habitatsdirektivets marine naturtyper. Fase I: Identifikation af potentielle indikatorer og tilgængelige data. Danmarks Miljøundersøgelser. 92 pp. – Faglig rapport fra DMU nr. 446.
Bendtsen, J., Andersen, J., Bendtsen, S.Å., Bruhn, B., Ellegaard, C., Rasmussen, J., & Vang, T. (2003): Kravspecifikation til
dele af det marine modelkompleks. Oktober 2003. 27 pp.
Conley, D.J., A. Clarke, S. Juggins, F. Adser, N. Reuss & J. Andersen (2003): Vandrammedirektivet, næringsstoffer i kystvande (3). Vand & Jord 2/2003: 52-56. 5 pp.
Ærtebjerg, G., Andersen, J., Carstensen, J., Christiansen, T., Dahl, K., Dahllöf, I., Fossing, H., Greve, T.M., Hansen,
J.L.S., Henriksen, P., Josefson, A., Krause-Jensen, D., Larsen, M.M., Markager, S., Nielsen, T.G., Pedersen, B., Petersen, J.K., Risgaard-Petersen, N., Rysgaard, S., Strand, J., Ovesen, N.B., Ellermann, T., Hertel, O., Skjøth, C.A.
2002: Marine områder 2001 - Miljøtilstand og udvikling. NOVA-2003. Danmarks Miljøundersøgelser. 94 pp. – Faglig rapport fra DMU nr. 419.
Pedersen, S., J. Andersen, J.G. Dannisøe, H. Kaas & F. Møhlenberg (2002): Vandrammedirektivet, konkretisering af miljømål
(2). Vand & Jord 1/2002: 25-29, 5 pp.
Andersen, J. (Red.) (2001): Fremtidens havovervågning. Revisionsscenarium med forslag til program for integreret overvågning af miljø- og naturforhold i de danske farvande 2004 – 2009. 54 pp.
Henriksen, P, J. Andersen og mange flere (2001): Marine områder 2000. Miljøtilstand og udvikling. 110 sider. Faglig
rapport fra Danmarks Miljøundersøgelser.
Dahl, K., J. Carstensen, C. Lundsgaard & J. Andersen (2001): Stenrev. 10 pp. In: Henriksen et al. (2001): Marine områder
2000. Miljøtilstand og udvikling. Faglig rapport fra Danmarks Miljøundersøgelser
Andersen, J., L.M. Munk & S. Pedersen (2001): Vandrammedirektivet, indhold og perspektiver (1). Vand & Jord 1/2001: 17-21.
5 pp.
Vandmiljø-2000. Status og perspektiver for indsatsen for et renere vandmiljø. Redegørelse fra Miljøstyrelsen, nr.
7/2000. Red.: J. Andersen, T. Christensen & S. Pedersen. 48 pp.
Andersen, J., J. Bielecki & M. Dam (2000): Bebyggelse i det åbne land. 23 pp inkl. bilag. I: Punktkilder 1999. Orientering fra
Miljøstyrelsen, nr. 16/2000.
Andersen, J. & J. Kirkegaard (2000): Overvågning af det danske vandmiljø, 1998-2003. Stads- & Havne-ingeniøren, august
8/2000:102-109. 5 pp.
NOVA-2003. Programbeskrivelse for det nationale program for overvågning af vandmiljøet 1998-2003. Redegørelse fra Miljøstyrelsen nr. 1. Red: J. Kirkegaard, T.M. Iversen & J. Andersen. 397 pp.
Vandmiljø-99. Status for vandmiljøets tilstand i Danmark. Redegørelse fra Miljøstyrelsen, nr. 1/1999. Red.: J. Andersen. 128 pp.
Evaluering af Vandmiljøplanens overvågningsprogram 1993-1997. Red: J. Andersen, T.M. Iversen & J. Kirkegaard. Notat fra
Miljøstyrelsen. 1998. 48 pp.
Andersen. J. (1998): Tilsyn med vandløb og søer. 5 pp. I: Miljøtilsyn 1997. Oversigt over kommunernes og amtskommunernes
miljøtilsyn. Orientering fra Miljøstyrelsen, nr. 4/1999.
Andersen, J. (1997): Fjorde, kyster og åbent hav. 6 pp. I: Vandmiljø-97. Miljøtilstanden i de ferske vande samt status for det
øvrige vandmiljøets tilstand i 1996. Redegørelse fra Miljøstyrelsen, nr. 4, 1997. 172 sider.
Andersen, J. (1997): Saltvandsbaseret fiskeopdræt. 4 pp. I: Punktkilder 1996. Vandmiljøplanens overvågningsprogram: Fagdatacenterrapport. Orientering fra Miljøstyrelsen nr. 16/1997.
Andersen, J. (1996): Saltvandsbaseret fiskeopdræt. 3 pp. I: Punktkilder 1995. Vandmiljøplanens overvågningsprogram: Fagdatacenterrapport. Orientering fra Miljøstyrelsen nr. 16/1996.
Christensen, P.B., F. Møhlenberg, L.C. Lund-Hansen, J. Borum, C. Christiansen, S.E. Larsen, M.E. Hansen, J. Andersen & J.
Kirkegaard (1996): Havmiljøet under forandring? Konklusioner og perspektiver fra Havforskningsprogram 90. - Havforskning fra Miljøstyrelsen, nr. 61. 120 pp.
Vandmiljø-95. Grundvandets miljøtilstand samt status for det øvrige vandmiljøs tilstand i 1994. Redegørelse fra Miljøstyrelsen, nr. 3, 1995. Red.: J. Andersen & J. Stockmarr. 156 pp.
Iltsvind i de danske farvande i oktober 1995. Red: J. Andersen & G. Ærtebjerg. 7 pp.
Iltsvind i de danske farvande i september 1995. Red: J. Andersen & G. Ærtebjerg. 7 pp.
Redegørelse om den miljømæssige og erhvervsmæssige betydning af saltvandsbaseret fiskeopdræt i Danmark. Redegørelse
fra Miljøstyrelsen 1995. Red.: J. Andersen & K. Hansen. 8 pp.
Andersen, J. (1995): Havbrug og saltvandsdambrug. 4 sider. I: Punktkilder 1994. Vandmiljøplanens overvågningsprogram:
Fagdatacenterrapport. Orientering fra Miljøstyrelsen nr. 10/1995.
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Vandmiljø-94. Udviklingen i belastningen fra punktkilder samt status for vandmiljøets tilstand. Redegørelse fra Miljøstyrelsen, nr. 2/1994. Red.: J. Andersen & T. Christensen. 160 pp.
Andersen, J. (1994): Havbrug. 5 pp. I: Punktkilder 1993. Vandmiljøplanens overvågningsprogram: Fagdatacenterrapport.
Orientering fra Miljøstyrelsen nr. 8/1994.
Nordisk havovervågningsprogram - forslag til koordinering af overvågningsaktiviteter. Ed.: J. Andersen. - Nord 1993:14. Nordisk Ministerråd. 153 pp.
Helmgaard, P. & J. Andersen (1991): Krebs og retningslinier for udsætning. - Vand & Miljø nr. 8/1991, side 415-417, 3 pp.
Andersen, J. & P. Helmgaard (1990): Populationsstruktur, vækstforhold og fødebiologi hos flodkrebs Astacus astacus L..
Specialerapport. Ferskvandsbiologisk Laboratorium, Københavns Universitet. 83 pp.
Publications in other languages
2
2009
1
2007
Andersen, H., J.H. Andersen, A. Erichsen, I.S. Hansen, F. Møhlenberg, & E.K. Rasmussen (2009): Danske erfarenhetar av
dynamiske modeller i den marina övervakningen. Organisation, model-lösningar og ”lessons learnt”, 1998-2008. DHI rapport til Naturvårdsverket. 75 pp. (In Swedish)
Андерсен, Э.Х., & Д. Паулак (2007): Биогенные вещества и эвтрофикация в Балтийском море: причины, последствия,
решения. Парламентская конференция Балтийского моря (ПКБМ). Russian translation of Andersen, J.H & J.T.
Pawlak (2006): Nutrients and Eutrophication in the Baltic Sea. Effects / Causes / Solutions. Booklet produced for the Baltic Sea Parliamentary Conference. 32 pp. (In Russian)
Oral presentations (since 1st of January 2011)
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2
1
“Økosystem-baseret forvaltning af de danske havområder: Hvordan går det? by J.H. Andersen – Danish Society for Marine
Biology, 26 October 2011. Invited speaker (In Danish)
“Towards eutrophication target setting in 5 simple steps. A status report from the HELCOM TARGREV project’ by J.H. Andersen & J. Carstensen – HELCOM MONAS Committee Meeting in Vilnius, Lithuania; 5 October 2011
“The HELCOM TARGREV: Status and perspectives” by J.H. Andersen & J. Carstensen – “HELCOM CORESET / TARGREV
Joint Advisory Board Meeting”, Warsaw (Poland); 28 June 2011.
“Whole system assessment – the fine art of converging indicators, quality elements and assessment principles” by J.H. Andersen & C. Murray – “WATERS Kick-Off Meeting”, Gothenburg (Sweden); 20 April 2011.
“The TARGREV project in support of the HELCOM BSAP implementation process” by J.H. Andersen, J. Carstensen & K.
Dormph – “HELCOM CORESET / TARGREV Joint Advisory Board Meeting”, Berlin (Germany); 20 March 2011.
“Towards modelling of connectivity and invasion scenarios” by J.H. Andersen, F.T. Hansen, T. Uhrenholdt & M. Potthoff –
“BWO Annual Meeting”, Newcastle (United Kingdom); 25 February 2011.
”Ålegræsværktøjets forudsætninger og usikkerheder” by J.H. Andersen & J. Carstensen – “Vandplaner - usikkerheder og
konsekvenser for dansk landbrug”, Nationalmuseet (Copenhagen); 8 February 2011. Invited speaker. (In Danish)
“Havstrategidirektivets basisanalyse – en faglig og administrativ udfordring” by J.H. Andersen – ”Havforskermøde 2011”,
Fuglsøcentret (Mols); 19 January 2011. (In Danish)
March 19, 2012
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Annex 4: Manuscripts
1. Conley, D.J., S. Markager, J.H. Andersen, T. Ellermann & L.M. Svendsen, 2002: Coastal
Eutrophication and the Danish National Aquatic Monitoring and Assessment Program. Estuaries 25:848-861.
2. Andersen, J.H., D.J. Conley & S. Hedal, 2004: Palaeo-ecology, reference conditions and
classification of ecological status: the EU Water Framework Directive in practice. Marine
Pollution Bulletin 49:283-290.
3. Andersen, J.H., L. Schlüter & G. Ærtebjerg, 2006: Coastal eutrophication: Recent developments in definitions and implication for monitoring strategies. Journal of Plankton Research
28(7):621-628.
4. Carstensen, J., D.J. Conley, J.H. Andersen & G. Ærtebjerg, 2006: Coastal eutrophication
and trend reversal: A Danish case study. Limnology & Oceanography 51(1-2):398-408.
5. Andersen, J.H., & D.J. Conley, 2009: Eutrophication in coastal marine ecosystems: towards
better understanding and management strategies. Hydrobiologia 621(1):1-4.
6. Andersen, J.H., C. Murray, H. Kaartokallio, P. Axe & J. Molvær, 2010: A simple method
for confidence rating of eutrophication status classifications. Marine Pollution Bulletin
60:919-924.
7. Andersen, J.H., 2010: Eutrophication. Baltic Sea Environmental Proceedings 122:16-17. In:
HELCOM, 2010: Ecosystem Health of the Baltic Sea. HELCOM Initial Holistic Assessment
2003-2007. Baltic Sea Environment Proceedings 122. 63 pp.
8. Andersen, J.H., P. Axe, H. Backer, J. Carstensen, U. Claussen, V. Fleming-Lehtinen, M.
Järvinen, H. Kaartokallio, S. Knuuttila, S. Korpinen, A. Kubiliute, M. Laamanen, E. LysiakPastuszak, G. Martin, F. Møhlenberg, C. Murray, G. Nausch, A. Norkko & A. Villnäs, 2011:
Getting the measure of eutrophication in the Baltic Sea: towards improved assessment principles and methods. Biogeochemistry 106:137-156.
9. Korpinen, S., L. Meski, J.H. Andersen & M. Laamanen, 2012: Human pressures and their
potential impact on the Baltic Sea ecosystem. Ecological Indicators 15:105-114.
10. Laamanen, M., S. Korpinen, U.-L. Zweifel & J.H. Andersen, submitted: Ecosystem health.
Textbook chapter in “Biological Oceanography of the Baltic Sea” (Eds: P. Snoeijs, H. Schubert & T. Radziejewska).
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Estuaries
Vol. 25, No. 4b, p. 848–861
August 2002
Coastal Eutrophication and the Danish National Aquatic Monitoring
and Assessment Program
DANIEL J. CONLEY1*, STIIG MARKAGER1, JESPER ANDERSEN2, THOMAS ELLERMANN3, and LARS
M. SVENDSEN4
1
Department of Marine Ecology, National Environmental Research Institute, P. O. Box 358,
DK-4000 Roskilde, Denmark
2 Danish Environmental Protection Agency, Strandgade 29, DK-1401 Copenhagen, Denmark
3 Department of Atmospheric Environment, National Environmental Research Institute, P. O. Box
358, DK-4000 Roskilde, Denmark
4 Environmental Monitoring Coordination Section, National Environmental Research Institute,
P. O. Box 358, DK-4000 Roskilde, Denmark
ABSTRACT: Nutrient over-enrichment and cultural eutrophication are significant problems in the Danish marine environment. Symptoms of eutrophication include periods of hypoxia and anoxia in bottom waters, death of benthicdwelling organisms during anoxia, long-term reductions in the depth distribution of macrophyte communities, changes
in the species composition of macrophyte communities, and increases in reports of harmful algal blooms. In 1987 the
Action Plan on the Aquatic Environment was adopted to combat nutrient pollution of the aquatic environment with the
overall goal of reducing nitrogen loads by 50% and point source phosphorus loads by 80%. The Danish Aquatic Nationwide Monitoring Program was begun in 1988 in order to describe the status of point sources (industry, sewage treatment
plants, stormwater outfalls, scattered dwellings, and fish farms), ground water, springs, agricultural watersheds, streams,
lakes, atmospheric deposition, and the marine environment. Another important aspect of the program was to document
the effects on the aquatic environment of the measures and investments taken for nutrient reduction as outlined in the
Action Plan. The monitoring program should determine if reductions in nutrients are achieved by the measures taken
and should help decision makers choose appropriate additional measures to fulfill the objectives. Coordination with
international programs and commissions is an important component of the monitoring program to meet internationally
agreed upon reductions in nutrient inputs. The future and direction of the Danish National Aquatic Monitoring and
Assessment Program will be to a large extent shaped by both the Water Framework Directive and Habitat Directive
adopted by the European Union.
and the death of benthic-dwelling organisms during anoxia (Fallesen et al. 2000).
During the 1980s a series of oxygen (O2) depletion events catalyzed the environmental movement
and led to the passage of the Action Plan on the
Aquatic Environment (Christensen et al. 1998). In
addition to creating a national agenda to reduce
the over-enrichment of natural waters with nutrients, the Action Plan established a coordinated
Aquatic Nationwide Monitoring Program (Kronvang et al. 1993) leading to the creation of the
National Environmental Research Institute
(NERI). The 1987 Action Plan on the Aquatic Environment I was only the first of a series of legislative actions in combination with international
agreements that have arisen to reduce nutrient
over-enrichment and the symptoms of eutrophication in Denmark with the overall objective of a
clean and healthy aquatic environment. Most of
the Action Plans are concerned with reduction of
discharges, loads, and emission of nutrients. The
current Danish National Aquatic Monitoring and
Introduction
Nutrient over-enrichment and cultural eutrophication of surface waters have been recognized as
being one of the most urgent problems to overcome in order to improve the environmental state
of lakes, streams, and marine waters in Denmark.
Nutrient loading from Denmark expressed on an
area basis ranks among the highest in Europe (Paaby and Møhlenberg 1996) and reflects the density
of the population and the intensity of agriculture.
Danish marine waters display all the classic symptoms associated with eutrophication including increased phytoplankton biomass and harmful algal
blooms (Kaas et al. 1999), reductions in the depth
distribution of macrophyte communities (SandJensen et al. 1994), changes in the species composition of macrophyte communities (Middleboe
and Sand-Jensen 2000), periods of hypoxia and anoxia in bottom waters (Conley and Josefson 2001),
* Corresponding author: tele: 145 4630 1200; fax: 145 4630
1114; e-mail: [email protected].
Q 2002 Estuarine Research Federation
848
Danish Monitoring and Assessment Program
Assessment Program which began in 1998 is designed to run through 2003 and is probably one
of the most comprehensive programs globally.
The purpose of this paper is to provide an overview of the current efforts in Denmark to reduce
nutrient over-enrichment of marine waters and to
monitor the aquatic environment. We present the
various National Action Plans and relevant parts of
derived legislation from the last 15 years that aim
to reduce nutrient inputs, we provide an overview
of the Danish National Aquatic Monitoring and Assessment Program focusing on the marine environment and the measurement of nutrient inputs, and
we briefly examine the documented changes in nutrient loading and the effects on the aquatic environment. We also address various aspects of the
monitoring program.
Strategies and Measures to Reduce Nutrient
Inputs to the Danish Aquatic Environment
POLITICAL OBJECTIVES
Since the mid-1980s a high priority has been given to the quality and protection of ground water
and surface water in Denmark with an overall goal
of ensuring that the waters are clean. The endeavors are described in the Environmental Policy
White Paper (Ministry of Environment and Energy
1999), which states that the Government will work
towards ensuring that streams, lakes, and marine
waters are clean and of a satisfactory quality with
regard to health and hygiene; that exploitation of
the water bodies and associated resources takes
place in a sustainable manner; and that the objectives of relevant international agreements will be
fulfilled. The central legal instrument to fulfill
these political objectives is the Consolidated Environmental Protection Act, which aims to safeguard
the environment, to support a sustainable social
development, and to conserve the flora and fauna
(Ministry of Environment and Energy 1998).
Overall objectives for streams state that water
flow must be adequate, that obstructions must not
hinder the dispersal of fish and macroinvertebrates, that there shall be 2-m wide borders free of
cultivation along natural streams, that streams have
good oxygen conditions, and that they contain a
varied and natural fauna and flora. Overall objectives for Danish lakes are that animal and plant
communities should be natural, that the water
should be clear and that submerged macrophytes
should be present in the shallow parts of the lakes
(Ministry of Environment and Energy 1999).
The overall objectives for Danish marine waters
are based on the 1992 Helsinki Convention on Protection of the Marine Environment in the Baltic
Sea Region (HELCOM 1992), the 1992 Conven-
849
tion for the Protection of the Marine Environment
of the Northeast Atlantic (OSPAR 1992), and the
1995 Declaration of the 4th International Conference on the Protection of the North Sea (Danish
Environmental Protection Agency 1995). The overall objectives are that the fauna and flora may only
be insignificantly or slightly affected by anthropogenic pollution and human activities, that nutrient
levels have to be at a natural level, the clarity of
the water has to be normal, unnatural blooms of
toxic phytoplankton or pollution-dependent macroalgae must not occur, and oxygen deficiency
may only occur in areas where it is natural, and
that the levels of hazardous substances have to be
at background levels in the case of naturally occurring substances and close to zero in the case of
hazardous substances. Commercial exploitation
(fisheries, navigation, drainage, offshore industry,
minerals extraction, marine dumping of seabed
material, recreational activities, and other uses of
surface water) has to be conducted in a manner
that respects environmental and natural wealth
and is sustainable (Ministry of Environment and
Energy 1999).
STRATEGIES AND MEASURES
The primary means of achieving the quality objectives for both ground water and surface waters
are reductions in nutrient loads. On November 28,
1986 the Danish Parliament adopted an agenda
that instructed the Government to reduce total
loads of nitrogen (N) and phosphorus (P) to the
aquatic environment by 50% and 80%, respectively. These reductions correspond to a change in annual loads from a level of 283,000 tonnes N and
9,120 tonnes P at the time the plan was adopted
to levels of about 141,600 tonnes N and 1,820
tonnes P (Folketinget 1987). During the political
process in 1987, these overall reduction targets
changed from being targets for loads to the aquatic
environment to reduction targets for the discharges and losses from three sectors: agriculture, municipal wastewater treatment plants (WWTP), and
point industrial discharges. It could be argued that
this change was rational since these three sectors
are the most relevant to reducing eutrophication.
As a result of the change, which was not in accordance with the November 28 Agenda, the adopted
Action Plan on the Aquatic Environment did not
include a reduction target for emissions of N to
the atmosphere, reduction targets for discharges
or losses from aquaculture, scattered settlements,
stormwater overflows, and offshore activities, and
a reduction target for P losses from agricultural
fields.
It is widely recognized that pollution of marine
waters crosses political boundaries. The countries
850
D. Conley et al.
around Denmark in the North Sea and Baltic Sea
adopted similar reduction targets through three
different conventions. The North Sea Conference
in London in 1987 (the countries of the North Sea
region excluding the United Kingdom) adopted
the goal of reducing N and P inputs to the sea by
50% over the period 1985–1995 in areas where
these could cause pollution. These reduction targets were reiterated at the conferences in The
Hague in 1990 and Esbjerg in 1995. The Paris
Commission in June 1988 adopted a 50% reduction target for nutrient inputs to marine waters susceptible to eutrophication and also adopted a program to achieve the reductions. In 1989, the reduction targets were specified in relation to specific sectors. In 1992, it was decided to integrate the
Oslo and Paris Conventions (the OSPAR Convention), both of which aimed to prevent marine pollution of the Northeast Atlantic region from dumping and land-based sources of pollution. In February 1988 the Helsinki Commission (HELCOM)
adopted a declaration specifying a 50% reduction
target for discharges of polluting substances, including nutrients, over a 10-yr period. In 1998, the
ministers confirmed their commitment to attaining the strategic goal from 1988 and defined specific objectives that must be achieved before the
year 2005.
Enacted Measures Taken to Achieve Nutrient
Reductions
SPECIFIC REDUCTION TARGETS FOR
AGRICULTURAL SECTOR
THE
Since the mid-1980s, a number of action plans
and strategies have been adopted by the Danish
Parliament to regulate development of the agricultural sector, one of the main sources of nutrients to the aquatic environment. The action plans
include the NPo (nitrogen, phosphorus, and organic matter) Action Plan in 1985, the Action Plan
on the Aquatic Environment I in 1987, the Action
Plan for Sustainable Agriculture in 1991, parts of
the Government’s 10-Point Program for Protection
of the Ground Water and Drinking Water in 1994,
follow-up on the Action Plan for Sustainable Agriculture in 1996, the Action Plan on the Aquatic
Environment II in 1998, and the Agreement on
May 2, 2001 on Supplementing Initiatives and
Preparations for the Action Plan III.
The reduction targets for N and P stipulated in
the Action Plan on the Aquatic Environment I are
an approximate halving of point source N loads
and a 80% reduction of point source P loads, including the elimination of the P farmyard load.
The reduction targets were to be attained by 1993
through the following measures carried out by the
agricultural sector: establishment of sufficient capacity to store 9 mo of manure production so that
manure can be stored until the crop growth season
begins, establish crop rotation and fertilization
plans to ensure that the N content of fertilizer is
optimally exploited, agricultural fields must have
green cover during the winter period, manure has
to be plowed in or in some other way deployed
into the soil within 12 h of application, and limits
on the amount of livestock manure applied to agricultural fields (Table 1).
It soon became clear that it would not be possible to attain the reduction targets by 1993 (Ministry of Agriculture 1991). The measures stipulated
in the Action Plan on the Aquatic Environment I
were tightened in 1991 in the Action Plan for Sustainable Agriculture. The reduction target was
maintained, but the time frame was extended to
the year 2000. The measures were fertilization accounts so that fertilizer application could be documented; more stringent and fixed requirements
on utilization of the N content of livestock manure;
all farms must establish sufficient capacity to store
9 mo of manure production; and a ban on the
application of liquid manure between harvest time
and February except on agricultural fields cultivated with winter rape or grass. Since the Action Plan
for Sustainable Agriculture there have been a number of follow-up plans for reducing the impact of
the agricultural sector, including the Government’s 1994 10-Point Program for Protection of
the Ground Water and Drinking Water in Denmark.
The need to further tighten the regulation of
agricultural loads of N has become even more necessary because Denmark must comply with the European Union (EU) Nitrates Directive by the year
2003. The directive restricts the application of livestock manure to 170 kg N ha21 yr21. In the case of
some farms this is less than the levels currently permitted. Denmark has sought permission to deviate
from the 170 kg N ha21 yr21 rule on cattle holdings
to enable the application of up to 230 kg N ha21
yr21 on a small number of these holdings.
In February 1998, the Danish Parliament adopted several new instruments aimed at achieving the
reduction targets. The Action Plan on the Aquatic
Environment II will reduce N leaching by a further
37,000 tonnes N yr21 to enable the reduction target
of 100,000 tonnes N yr21 to be achieved no later
than the end of the year 2003 (Table 1; Danish
Environmental Protection Agency 2000). Under
the Action Plan on the Aquatic Environment II,
16,000 ha of wet meadow will be re-established to
help reduce N leaching through denitrification,
20,000 ha forest will be planted before 2002, and
agri-environmental measures that include financial
851
Danish Monitoring and Assessment Program
TABLE 1. Summary of measures and estimated reductions (tonnes N yr21) in nitrogen loading from agriculture. The years 1993,
2000, and 2003 are the expected year of implementation of Action Plan on the Aquatic Environment I; Action Plan for Sustainable
Agricultural Development; and Action on the Aquatic Environment II, respectively. The reduction figures are not legally binding,
they should for that reason be considered merely as politically agreed upon goals. The goal of Action Plan II is 100,000 tonnes N
yr21; therefore the farmyard load (27,000 tonnes N yr21) is added to the estimated effect of Action Plan I and Action Plan for
Sustainable Agricultural Development (89,900 2 27,000 5 62,900 tonnes N yr21).
Action Plans and Measures:
Action Plan on the Aquatic Environment I (1987):
Optimal utilization of livestock manure
NPo Action Plan
NPo Subsidy Act
Further initiatives
Program for improved utilization of fertilizers
Systematic fertilization plans
Improved application methods
Winter green fields—catch crops and plowing down of straw
Winter green fields—further initiatives
Structural measures
Total
Action Plan for Sustainable Agricultural Development (1991):
Improved utilization of livestock manure
Reduction in commercial fertilizer consumption
Protection of groundwater in particularly vulnerable areas
Reduction in agricultural acreage
Structural development, other measures
Total
Action Plan on the Aquatic Environment II (1998):
Wetlands
Sensitive agricultural areas
Afforestation
Improved fodder utilization
Stricter harmonization criteria
Stricter requirements on utilization of N content of manure
Organic farming
Catch crops on a further 6% of the land
10% reduction in N norm
Total
Total:
support to farmers willing to cultivate sensitive agricultural areas in a more environmentally sound
manner will be implemented. Agricultural measures include using less fertilizer or completely refraining from cultivating the land (there has hitherto been very little interest in this scheme), improved fodder utilization and changes in feeding
practice, implementation of stricter harmony criteria governing livestock density, stricter requirements on utilization of the N content of livestock
manure, converting 170,000 ha to organic farming,
catch crops on a further 6% of a farmer’s land,
and reducing the N norm by 10%, e.g., farmers
may now only apply N in amounts corresponding
to 90% of the economically optimal level.
If the measures in the Action Plan on the Aquatic Environment II are fully implemented, it is expected that within several years it will result in a
100,000 tonnes N yr21 reduction in leaching from
agricultural land. N consumption in the form of
commercial fertilizer will decrease from 400,000
tonnes N yr21 in 1985 to 200,000 tonnes N yr21 in
2003 (Iversen et al. 1998).
1993
2000
2003
55,000
5,000
10,000
15,000
5,000
20,000
8,000
9,000
127,000
127,000
50,000
20,000–40,000
8,000–15,000
1,000–2,000
17,000–20,000
15,000
77,000
89,900
127,000
5,600
1,900
1,100
2,400
300
10,600
1,700
3,000
10,500
37,100
127,000
In connection with the Action Plan on the
Aquatic Environment I, it was estimated that N
loads could be reduced by a total of 127,000 tonnes
N yr21 by 1993. The reduction targets were 100,000
tonnes N yr21 for the N load from agricultural
fields and 27,000 tonnes N yr21 for the farmyard
load. In the Action Plan for Sustainable Agriculture it was estimated that by the year 2000 the measures stipulated in the Action Plan on the Aquatic
Environment I would only have reduced N loads
by 50,000 tonnes N yr21 and that further measures
were needed. The existing measures and targets
were re-evaluated in 1998 in connection with the
preparation of the Action Plan on the Aquatic Environment II, and it was concluded that by the year
2003 the existing measures would reduce N loads
by 89,900 tonnes N yr21. Together with the expected reduction under the Action Plan on the
Aquatic Environment II, it was concluded that N
loads would be reduced by 127,000 tonnes N yr21
by 2003.
Not all the measures in the Action Plan on the
Aquatic Environment II will have taken full effect
852
D. Conley et al.
by 2003. A mid-term evaluation in late 2000 indicated that further measures were needed to fulfill
the 100,000 tonnes N yr21 reduction of root zone
losses. It is assumed that an Action Plan III will be
passed in the Danish Parliament in 2003 or 2004
with regional reduction targets including measures
against diffuse P losses.
SPECIFIC REDUCTION TARGETS FOR MUNICIPAL
WASTEWATER TREATMENT PLANTS
Discharges from municipal WWTP are regulated
by the Environmental Protection Act and derivative statutory orders. The EU Council Directive
91/271/EEC concerning Urban Wastewater Treatment as amended by Commission Directive 98/
15/EU, commonly referred to as the Urban Wastewater Directive, is one of the most important legal
documents in the EU concerning water quality.
The purpose of the directive is to protect the environment against the negative effects associated
with the discharge of inadequately treated urban
wastewater and discharges of biologically degradable industrial wastewater from the food processing industry. According to the directive, wastewater
discharges must be subjected to a level of treatment appropriate to the environment in question
and to the designated use of the recipient water
body. Denmark implemented the provisions of the
directive in 1994 legislation.
The Action Plan on the Aquatic Environment’s
reduction targets for municipal WWTP were adjusted in 1990 on the basis of the results of the
Nationwide Aquatic Monitoring Program (Danish
Environmental Protection Agency 1991). In the
case of N, annual discharges in treated wastewater
were to be reduced from 18,000 tonnes N yr21 to
6,600 tonnes N yr21. P discharges were reduced
from 4,470 tonnes P yr21 to 1,220 tonnes P yr21.
The reduction in N corresponds to all new or upgraded plants exceeding 5,000 personal equivalents (PE) and all existing plants exceeding 1,000
PE having to implement biological treatment with
N removal down to an annual average of 8 mg N
l21. Municipal WWTP exceeding 5,000 PE have to
remove P down to an annual average of 1.5 mg P
l21.
SPECIFIC REDUCTION TARGETS FOR POINT
INDUSTRIAL DISCHARGES
Point discharges from industry are regulated by
the Consolidated Environmental Protection Act
and the EU Directive on Pollution Prevention and
Control (IPPC Directive) and derivative statutory
orders. The IPPC Directive aims at integrated prevention and control of pollution by major industries. The directive specifically regulates the energy
industry (e.g., power stations and refineries), pro-
duction and processing of metals, the mineral industry, the chemical industry, waste management
plus a number of other activities such as paper
manufacturers, textiles pre-treatment and dyeing,
slaughterhouses and dairies, and installations for
intensive rearing of poultry and pigs exceeding a
certain capacity. The IPPC Directive contains measures designed to prevent or, where that is not
practicable, to reduce emissions to the air, water,
and land. Because of the large differences between
individual enterprises and their discharges of
wastewater, the Action Plan on the Aquatic Environment I did not stipulate general discharge requirements for industry as it did for WWTP. Industry was to reduce its discharges through the application of Best Available Technology (BAT) at
the level of treatment that is technically attainable
and economically viable.
Costs of Action Plans
The costs of Action Plan on the Aquatic Environment I for the period 1985–1989 have been estimated in 1990 to be 1.2 billion 5
C (Danish Environmental Protection Agency 1991). This figure includes some investments agreed upon already in
the 1985 NPo Action Plan and reaffirmed in the
1987 Action Plan. The investments in agriculture
during the period 1985–1992 have been 400 milC with total investments in municipal WWTP
lion 5
C including both construction of new
of 1.1 billion 5
plants and enlargement and improvement of existing plants. In addition, industries with separate
treatment and discharge of wastewater invested
C in improved wastewater treatment.
135 million 5
The cost of Action Plan II is expected to be 135
million 5
C , half of which will be financed by the
agricultural sector (Iversen et al. 1998). The expenses of wetland restoration, groundwater protection areas, afforestation, and development of organic farming will be financed by different government agencies.
The total annual costs are 1.28 billion 5
C total for
planning and management of the aquatic environment by state, regional, and local authorities: 68
billion 5
C , maintenance and restoration of rivers,
streams and lakes: 81 million 5
C , national monitoring of the aquatic environment: 40 million 5
C , supervision by the counties and local communities:
11 million 5
C , supply of clean and healthy drinking
water: 405 million 5
C , and discharge and cleaning
of wastewater from households and industries: 676
C (Christensen personal communication).
million 5
The Danish National Aquatic Monitoring and
Assessment Program
MONITORING OF ATMOSPHERIC DEPOSITION OF
NUTRIENTS TO DANISH MARINE WATERS
The atmospheric component of the Action Plan
on the Aquatic Environment was initiated in 1989
Danish Monitoring and Assessment Program
with the focus of determining atmospheric N deposition to Danish waters. The first monitoring
program (1989–1994) expanded upon existing
monitoring stations for better geographical coverage. The program consisted of 6 stations for wet
and dry deposition of different N species and 12
stations where only wet deposition was measured.
About half of the stations were placed close to the
coast. Interpolations of measurements between stations were used to obtain the overall deposition to
Danish waters.
The program consists of sampling at the main
stations for wet deposition with bulk samplers on
a half-month basis. The precipitation samples are
analyzed for their content of nutrients (mainly ammonium and nitrate) and a number of other air
pollutants (including 9 heavy metals). The gas
phase and particulate phase air pollutants are collected on a daily basis on various types of filters,
which after extraction and analysis are used to determine the atmospheric content of N (ammonia,
particulate ammonium, and the sum of nitric acid
and particulate nitrate) and other important air
pollutants. Dry deposition at the monitoring stations is subsequently calculated by use of literature
data for the deposition velocities and measured
concentrations and meteorology.
During the early 1990s a comprehensive model
for calculation of deposition to Danish marine waters was developed, the ACDEP-model (Atmospheric Chemistry and Deposition; Hertel et al.
1995). The ACDEP model is a trajectory model
that calculates the atmospheric concentrations and
the wet and dry deposition of nutrients and other
important air pollutants to 30 3 30 km grids covering Danish marine waters and land. The transport of air pollutants is determined by using 96-h
back trajectories calculated by use of meteorological data from NERI (Brandt et al. 2000) and Cooperative Programme for Monitoring and Evaluation of the Long Range Transmission of Air Pollutants in Europe (EMEP). The model is supplied
with information of initial concentrations based on
a coarser long-range transport model (Brandt et
al. 2000) and emissions (from NERI and EMEP)
of the air pollutants, and simulates the vertical and
horizontal transport, chemical transformations (80
reactions), and the wet and dry deposition of 37
air pollutants.
In 1995 model calculations were implemented as
an integral part of the monitoring program and at
the same time the number of precipitation stations
were reduced to only two stations (Fig. 1a). The
aim was to improve the results through a combination of both measurements and model calculations. The measurements at the monitoring stations are used to determine the concentrations
853
Fig. 1. The geographical distribution of sampling stations in
various parts of the Danish Aquatic Monitoring and Assessment
Program. A) Atmospheric measurement stations. B) Catchment
distribution used to measure watershed input of nutrients. C)
Marine monitoring stations.
854
D. Conley et al.
Fig. 2. Long-term annual means for the concentration of
ammonia (closed squares), particulate ammonium (circles),
and sum of nitric acid and particulate nitrate (sum nitrate; open
squares) in the atmosphere at the monitoring station at Anholt.
and deposition at monitoring stations, while the
model calculations are used to calculate the deposition to both land surfaces and Danish marine
waters. Subsequently, measurements are used to
validate the results from model calculations. Seasonal and long-term trends are determined solely
on the basis of measurements due to lack of sufficiently updated emission data, while studies of
the origin of air pollution are based on model calculations. The influence of Danish emissions can
be estimated by model calculations with and without Danish emissions.
Long-term trends in concentrations at Anholt,
situated in the middle of Kattegat (Fig. 1a), show
a significant decrease of particulate ammonium
and sum nitrate, whereas long-term trends in ammonia concentration are insignificant (Fig. 2).
Similar patterns are observed at other Danish
monitoring stations (Ellermann et al. 2000). Reductions in emissions of N compounds have occurred in central Europe and Denmark from reductions in both combustion processes and farming. Calculation of the annual total deposition of
N to Danish marine waters in 1999 from the ACDEP model shows that N deposition varies between
1.0 and 1.8 tonnes N km22 with an average of 1.1
tonnes N km22 for all the Danish waters (Fig. 3).
Moreover, model calculations show that the high
N deposition rates to estuaries and coastal waters
are mainly due to the high dry deposition of ammonia from local animal husbandry.
MONITORING OF LAND-BASED NUTRIENT LOADS TO
DANISH MARINE WATERS
Denmark uses a load-oriented approach to quantify the load of nutrients from land to coastal areas.
The method is based on an extensive monitoring
program in rivers and streams and on all point
sources larger than 30 PE. WWTP, single industries
not connected to WWTP, aquaculture, and some
stormwater outfalls are also monitored. Nutrient
loads from scattered dwellings in rural areas, the
Fig. 3. ACDEP Model calculation of the annual total deposition of nitrogen (tonnes N km22) to Danish marine waters in 1999.
Note that the numbers are only valid for the part of the grid with water surface (Ellermann et al. 2000).
Danish Monitoring and Assessment Program
remaining stormwater outfalls, and other point
sources less than 30 PE are estimated by standard
values or models. A characteristic feature of the
Danish landscape is the lack of large rivers. Because it is economically and practically impossible
to monitor the several thousand small streams that
cover the landscape, the country has been divided
into monitored and unmonitored regions in relation to deriving riverine loads.
Nutrient loads are determined using approximately 130 river monitoring stations situated downstream in rivers. About 55% of the land surface
area is covered by riverine measurements. Water
samples are taken generally between 12 and 26 (average 19) times per year, the stage is continuously
recorded, and stage-discharge relationships are
used to calculate load. In addition to monitoring
stations in rivers located near the coast, 150 sampling stations are situated in small agricultural
catchments (5 to 60 km2) with only minor inputs
from point sources. The loads in the unmonitored
part of the country are calculated using flowweighted concentrations and discharge from agricultural catchments with corresponding climate,
soil type, geology, and agricultural practices as
found in the monitored catchments, and then adding monitored point sources.
The total loads to coastal areas are determined
by summing the total monitored load, the total unmonitored load, and nutrient loads from point
sources discharging directly to coastal areas (Fig.
1b). Source apportionment is performed on the
total load to evaluate the importance of different
nutrient sources. Nutrient loads from diffuse
sources, such as agricultural land and forested areas, are estimated as the difference between the
gross transport, including retention in rivers and
lakes, and the total load from point sources. Nutrient loads from cultivated areas include the estimated potential load from scattered dwellings entering the surface freshwater system.
Natural background loads of N and P constitute
a part of the total estimated nutrient inputs to surface water and include loads from unmanaged
land and that part of the loads of N and P from
managed land that would occur irrespective of anthropogenic activities. The natural background
loads are determined from measurements of nutrient loads in 9 small non-agricultural catchments
and subdivided into sandy and loamy catchments.
The means for these 9 natural watersheds from
1989 to 1999 were 1.50 6 0.15 mg N l21 (2.20 6
0.76 kg N ha21) and 50 6 6 Tg P l21 (0.078 6 0.04
kg P ha21).
Nutrient loads to freshwater are often greater
than the measured nutrient transport to coastal waters due to the retention of nutrients in lakes and
855
rivers. Retention plays a key role for the amount
and the composition of nutrient fluxes, especially
P, through river systems (Billen et al. 1991). Retention in the catchment must be added to the measured load to estimate the amount of diffuse
source (natural background 1 agriculture 1 scattered dwellings) loads to freshwater. In larger rivers, N loss through denitrification can have a significant influence on the total load from the river
system. In-stream river retention of N in Danish
streams can be high, primarily during the summer,
but over an annual cycle is negligible and therefore not included in the calculations. The retention of P is especially important in streams where
over bank flooding occurs. For the total load from
Denmark to coastal areas the net retention of N
and P in rivers and streams is less than 1% to 2%
(Svendsen et al. 1995). Retention in lakes constitutes 8% to 12% of the gross riverine N load and
0% to 2% of the gross riverine P load from 1989–
1999 (Svendsen et al. 1998). The retention in lakes
is calculated from mass balance calculations in
about 30 intensively monitored lakes.
The measures taken to reduce the load from
point sources in Denmark have been successful.
Since 1989 the N load from point sources has been
reduced by 66% and for P by 81% (Fig. 4). The
main part of the reduction has taken place by improved and extended purification of wastewater
from WWTP and from single industrial plants not
connected to treatment plants (SID). The N load
from WWTP has been reduced by 74% and for P
by 90% and the corresponding figures for SID are
85% and 95%, respectively. Further, the use of detergents without P reduced the P load from scattered dwellings. Thus, the original reduction targets for P have been fulfilled.
The total load of N to coastal waters is closely
related to runoff (Fig. 5). P loads have been dominated by inputs via wastewater load, but after 1994
the diffuse loading of P has been a significant portion of the load to coastal areas. The reduction in
the contribution of point sources means that P
loading is now correlated to runoff. Diffuse sources are the biggest N source. In 1999 81% of the
total N load originated from agricultural sources,
11% from natural background loads, and 8% from
point source loads. There has been a significant
reduction (Kendall trend test, p , 0.01) in the total P load to coastal areas. There is a tendency of
a minor reduction in diffuse N loads, but it is not
significant (p . 0.05).
Marine Monitoring
A marine component of the Aquatic Nationwide
Monitoring Program was initiated in October 1988
with implementation of the Action Plan for the
856
D. Conley et al.
Fig. 4. Freshwater discharge, total nitrogen load, and total
phosphorus load to coastal waters in Denmark divided up by
source. WWTP are wastewater treatment plants, SID are single
industries not connected to treatment plants, SWO are stormwater overflows, SD are scattered dwellings, and AQ are freshwater and coastal fish farms.
Fig. 5. Freshwater discharge, total nitrogen load, and total
phosphorus load to coastal waters in Denmark from different
sources. Diffuse loads also include natural background loads
and loads from scattered dwellings, freshwater point sources are
those discharging directly to freshwaters and direct point sources are those discharging directly into estuaries or coastal waters.
Aquatic Environment I. The program was based on
previous national and regional monitoring activities. Involvement of regional authorities in the
monitoring activities broadened the expertise
available to local decision makers and reduced
ship time for high frequency sampling. All data are
stored in a national database at the Marine Topic
Centre, National Environmental Research Institute, Roskilde, Denmark, whose staff is responsible
for writing technical guidelines, coordinating the
marine monitoring program, and producing an
annual national report about the State of the Marine Environment.
Marine monitoring consists of several components. Water column physical, chemical, and biological parameters are measured 26 to 46 times per
year at 16 open water and 38 estuarine stations.
Another 26 open water stations are monitored for
dissolved oxygen (DO) four times in the fall, and
twice a year at all 92 stations including the Danish
portion of the North Sea and Skagerrak (Fig. 1c).
Physical and chemical parameters include conductivity, temperature, depth, light, nutrients, and oxygen. Oxygen concentration data provide information about the distribution of oxygen depletion
as a consequence of eutrophication (Conley and
Josefson 2001). Species composition, coverage and
depth distribution of angiosperms and macroalgae
are assessed by scuba diving in 34 estuaries and 9
stone reefs in open waters (Middleboe et al. 1998).
Every third year the distribution in shallow waters
is assessed with the use of aerial photography. The
biomass and species composition of benthic fauna
is monitored at 24 stations in estuaries and 24 stations in open waters once a year ( Josefson and
Rassmussen 2000). Biomass, distribution, and condition of benthic filter feeders are monitored in
five estuaries. Measurements of sediment chemistry include pools of nutrients, H2S concentrations
at different depths, fluxes of nutrients between the
Danish Monitoring and Assessment Program
857
TABLE 2. Cost of the Danish Aquatic Monitoring and Assessment Program 1998–2003. Prices in million € per year (1996 prices).
Counties
Agricultural catchments
Ground water
Springs, watercourses, lakes
Point sources
Atmospheric deposition
Marine waters
Crosscutting research and development
Coordination and administration
Total
1.95
3.62
5.22
4.27
0
5.92
—
—
21.0
sediment and the overlying water, and measurements of denitrification. Hydrographic modeling is
conducted for the entire area with a three-dimensional model with the aim of calculating nutrient
fluxes between the Baltic Sea, the North Sea, and
the Kattegat. Separate models for water exchange
are run on six type area estuaries.
The marine program is divided into two parts:
estuaries and open marine areas. Six estuaries
serve as type areas for the categories of Danish estuaries (Horsens Fjord, Limfjorden, Odense Fjord,
Ringkøbing Fjord, Roskilde Fjord, Skive Fjord).
They are monitored intensively, up to 46 times per
year, for most water column parameters mentioned
above. A broader geographic coverage includes 34
additional estuaries but at a lower frequency of
sampling (Fig. 1c). The open marine area consists
of 16 intensively sampled water column stations, 76
additional stations at a lower frequency, a threedimensional hydrographic model, and monitoring
of macroalgae on 9 stone reefs. Data from the
more frequently sampled stations support a detailed understanding of the mechanisms relevant
to eutrophication in response to nutrient load
changes within a context of climatic variation. Data
from the broader geographic coverage are used to
document long-term changes in winter concentrations of nutrients and the fall distribution of oxygen depletion.
Costs of Monitoring Program
The total annual costs of the Danish National
Aquatic Monitoring and Assessment Program are
C (see Table 2), apportioned beabout 26 million 5
tween national agencies (5 million 5
C yr21) and regional authorities (21 million 5
C yr21). In parallel
to the activities of the Danish National Aquatic
Monitoring and Assessment Program, the Danish
counties have a number of compliance monitoring
activities in order to document that legally binding
discharge permits are kept and assess the development and fulfillment of normative ecological
quality objectives. The annual costs of compliance
monitoring are about 11 million 5
C yr21 (Andersen
et al. 2001).
State Institutions
0.20
0.23
0.60
0.30
1.28
1.30
0.70
0.31
4.92
Total
2.15
3.85
5.82
4.57
1.28
7.22
0.70
0.31
25.9
Effects of the Action Plan on Marine Waters
The activities of the Aquatic Nationwide Monitoring Program (1988–1997) and the follow-up
program, the Danish National Aquatic Monitoring
and Assessment Program (1998–2003) have been
instrumental in documenting eutrophication effects in Danish waters (e.g., Ophelia 1995a,b;
Jørgensen and Richardson 1996; Christensen et al.
1998; Conley et al. 2000). New results regarding
long-term trends in the marine environment from
the Danish National Aquatic Monitoring and Assessment Program are presented here.
Nutrient management actions have resulted in a
significant decrease in the P load from point sources and a decrease (not significant) in the N load
(Fig. 5). The reductions come as a result of gradual
increases in the tertiary treatment of sewage first
with implementation of local plans in the 1980s to
reduce the effects of eutrophication in inland waters, and later with implementation of the 1987 Action Plan on the Aquatic Environment. Consequently, the potential for nutrient limitation has
increased, especially in the spring for P and significant reductions have occurred in nuisance growth
of macrophytes, e.g., Ulva lactua (Conley et al.
2000).
Significant relationships were found between
loads of N and P and the mean Secchi disk depth
from 1989 to 1999 (Fig. 6). There is a significant
linear correlation between both N loading and P
loading in estuaries with Secchi disk depth (p 5
0.006 for P and p 5 0.026 for N). In the open areas
only N is significantly correlated to Secchi disk
depth (p 5 0.14 for P and p 5 0.001 for N). The
relationship for P appears somewhat curvilinear
perhaps indicating that the Danish estuaries were
saturated with P at loadings above 4,000 tonnes
yr21. The data suggest that P is limiting for primary
production in estuaries, at least during part of the
season, and extensive P limitation is supported by
experimental data from Danish estuaries (Holmboe et al. 1999). Secchi depth increased 30% from
a level of 3 m between 1984 and 1988 to 3.9 m
between 1996 to 1999, while over the same time
858
D. Conley et al.
Fig. 6. The relationship between Secchi disk depth and the total land-based loads of phosphorus and nitrogen from Denmark
between 1989 and 1999 in estuaries and open sea areas. Estuaries also include near coastal areas with a water depth less than 10 m.
Open waters are all areas outside the estuaries with a water depth greater than 10 m.
period areal primary production decreased by 28%
(Fig. 7).
While implementation of nutrient reductions
mandated in the Action Plans are legally enforceable, the effects of the Action Plans on nutrient
loading are not expected to occur within the time
frame of the implementation of the measures taken. This is due to the fact that there are N and P
stores in soils, ground waters, and marine sediments that must be depleted before actual nutrient
reductions may be observed in the environment.
Reductions in nutrient loads are expected to eventually occur when all measures are implemented,
but it is doubtful whether the target of a 50% reduction in N load will be achieved. Positive effects
in the marine environment are currently documented, and further improvements are expected.
A practical example of the effects of nutrient reductions occurred during the two dry years of 1996
and 1997 with a 50% reduction in N loading to the
marine environment. Effects on the marine environment included lower than average water column chlorophyll concentrations, increased depth
penetration of aquatic macrophytes, and reduced
extent of oxygen depletion (Markager et al. 1999;
Rask et al. 1999).
General Aspects of the Monitoring Program
The Danish National Aquatic Monitoring and
Assessment Program illustrates the importance of
monitoring data feedback upon regulatory actions
for the successful management of the aquatic environment. As shown earlier, evaluation of the Action Plan on the Aquatic Environment I demonstrated that the measures taken were not sufficient
to reduce nutrient levels to the targeted amounts.
Additional measures to reduce nutrient loads were
taken with the Action Plan on the Aquatic Environment II. The process is also an example of strategic environmental planning (Table 1), whereby
a target is set, measures are implemented, the effects are monitored and evaluated, and supplementary measures are implemented if, as in the
present case, the original targets are not attained
as expected.
Future revisions of the Danish National Aquatic
Monitoring and Assessment Program will be greatly influenced and guided by adoption of the European Union’s Water Framework Directive (WFD;
European Union 2001). The legislation must be
implemented in national law before December 22,
2003 by the member countries. The monitoring
obligations according to Annex V of the Directive
should be fully implemented and fully operational
Danish Monitoring and Assessment Program
Fig. 7. Long-term data for Secchi disk depth and areal primary production in Danish estuaries between 1977 and 1999.
The line indicates the long-term mean.
by December 22, 2006. One of the many important
goals of the WFD is to classify the state of a particular environment and identify measures needed to
reach stated environmental goals. Goal setting
should be an integral part of a successful monitoring program as exemplified by the Chesapeake Bay
Program, where goals have been defined for a long
list of parameters. The goals set for the Danish Action Plan were originally defined only in terms of
reduction in loads relative to the situation in 1987
without regard to the state of the environment.
Those numbers were set somewhat arbitrarily in a
situation where there was an immediate need for
political action due to severe hypoxia and anoxia
in many areas. The lack of quantitative goals and
operational goals for ecological quality is probably
one of the major weaknesses of the Danish National Aquatic Monitoring and Assessment Program
(Markager 2001). In the coming years with imple-
859
mentation of the WFD, operational goals have to
be formulated that relate to the state of the environment. Monitoring data will be essential in defining quality classes and quality criteria in the
WFD. The EU Habitat Directive will also set conservation goals that are relevant to marine areas.
The scientific challenge for the monitoring program will be to define operational definitions for
the state of the marine environment and to establish quantitative relationships between load and
the state of the environment (Markager 2001).
It is our belief that more time and money needs
to be allocated for assessment activities in the current Danish National Aquatic Monitoring and Assessment Program. The goal of the monitoring
programs are to determine the state of the marine
environment, determine temporal trends in the
marine environment, and document the effects of
the Action Plans and other relevant measures on
the marine environment. It is implicit that within
this framework lies knowledge about the marine
environment (Markager 2001). Data must be processed to provide information and then further analyzed to gain knowledge about how systems operate. This level of data analysis of the Danish marine monitoring data has not been conducted with
sufficient strength, with the consequence that our
understanding of the interactions between loads of
nutrients and the state of environment is less developed than possible from the existing data.
In connection with the Action Plan on the
Aquatic Environment, a number of concurrent research programs were established. The first program, the NPo Program ran from 1985 to 1990
C.
with the marine part amounting to 1.2 million 5
The follow-up program, the Marine Research Program (HAV90) in Denmark was a significant research program targeted for the marine environment in 1990–1995 for 10.8 million 5
C . The products from the program included a series of scientific reports and 300 peer reviewed publications
including those listed earlier. The Strategic Environmental Program funded an aquatic section
from 1995–1998 and, in addition to publishing
their findings in the literature, the project finished
with an undergraduate textbook about the marine
environment (Lomstein 1999). These research
programs have helped propel Denmark into a
leading role in research in marine ecology, contributed to an understanding of the Danish marine
environment, and provided critical support for the
improved management of marine areas. Priorities
for funding scientific research in Denmark have
since changed. These programs have been completed with no new directed research money forthcoming regarding research in the marine environment. The European Union has recognized the
860
D. Conley et al.
importance of using the wealth of data collected
within national monitoring programs, and research funding within the European Union (http:
//www.cordis.lu) has encouraged the utilization of
monitoring data within research programs.
The Danish National Aquatic Monitoring and
Assessment Program probably holds one of the
best databases in the world for evaluation of the
effects of nutrient loading on the marine environment. The program has been criticized within Denmark for not being cost-effective, not providing
sufficient data for calibration of modeling efforts,
and not including enough modeling efforts into
the data analysis (Dahl-Madsen 2000). Some of the
criticisms were valid (Riemann et al. 2001), and
modeling efforts are currently being directed at
several levels (empirical models, simple budgets,
and more advanced models). The recent experience from the Chesapeake Bay Program WaterQuality Model review of the three-dimensional
coupled hydrodynamic-water quality model (Scientific and Technical Advisory Committee 2000)
highlights the danger of relying on only one model
in order to develop management alternatives.
The Danish National Aquatic Monitoring and
Assessment Program has been modified since its
inception based on internal and external evaluations and reflects compromises among political
goals, government institutions, local authorities,
and available funding. Modifications will occur
with the next phase planned to run from 2004–
2009. While monitoring programs by their nature
are essentially conservative, in order to maintain
long time series essential to document long-term
changes, the program must also have the flexibility
to adapt to changing needs. Future revisions of the
Danish National Aquatic Monitoring and Assessment Program will encounter increased responsibilities especially regarding European Union legislation and the Water Framework Directive and
the Habitat Directive, and will need to adapt to
proposed decreases in funding levels and the addition of emerging technologies, such as ships-ofopportunity and instrument buoys.
ACKNOWLEDGMENTS
We express our thanks to Jørn Kirkegaard and Tony Christensen, Danish Environmental Protection Agency and Kitt Bell
Andersen, Department of Agriculture and Biotechnology, Danish Forest and Nature Agency.
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Received for consideration, July 6, 2001
Accepted for publication, February 19, 2002
Marine Pollution Bulletin 49 (2004) 283–290
www.elsevier.com/locate/marpolbul
Viewpoint
Palaeoecology, reference conditions and classification of
ecological status: the EU Water Framework Directive in practice
Jesper H. Andersen
a
a,*
, Daniel J. Conley
a,b
, Søren Hedal
c
Department of Marine Ecology, National Environmental Research Institute, Frederiksborgvej 399, P.O. Box 358, DK-4000 Roskilde, Denmark
b
Department of Marine Ecology, University of Aarhus, Finlandsgade 14, DK-8200 Aarhus, Denmark
c
Roskilde County, Aquatic Environment Division, Køgevej 80, DK-4000 Roskilde, Denmark
Abstract
The European Union’s Water Framework Directive (WFD) requires that all Member States within the European Union
determine reference conditions for aquatic ecosystems to provide a baseline against which to measure the effects of past and present
activities. Reference conditions are subsequently used to classify the ecological status of European waters. The decisions regarding
environmental status will be important future elements in the management of European coastal waters. We have developed a
number of classification scenarios for total nitrogen (TN) in the overlying waters of the southern part of Roskilde Fjord, Denmark,
taking as our basis a palaeoecological reconstruction of fluctuations in TN between 1850 and 1995. We present a provisional
classification scheme for the ecological status of Roskilde Fjord, sensu the WFD. Decision(s) regarding the deviation from reference
conditions will give a wide range of apparent ecological status from good, through moderate and poor, to bad depending upon the
definition of an acceptable deviation from reference conditions. The determination of an acceptable deviation will ultimately be a
political decision, and will result in a wide range in the protection of coastal waters in Europe. There is still, however, an urgent need
for a sound scientific documentation of the various scenarios for the implementation of the WFD.
Ó 2004 Elsevier Ltd. All rights reserved.
Keywords: Eutrophication; Nitrogen; Palaeoecology; Reference conditions; Classification; The EU Water Framework Directive
1. Introduction
For more than 30 years nutrient enrichment has been
considered a major threat to the health of coastal marine
waters (Ryther and Dunstan, 1971; Nixon, 1995; Elmgren, 2001, de Jonge et al., 2002). Many national and
international initiatives have been implemented in order
to reduce the inputs and effects of nutrients in coastal
waters, e.g. Conley et al. (2002). Within Europe, the
most recent legislation is the European Union’s Water
Framework Directive (Anon., 2000; Elliot et al., 1999).
The Water Framework Directive (WFD) provides
a framework for the protection of ground water, inland surface waters, transitional waters (estuaries) and
coastal waters (Anon., 2000). The overall aim of the
WFD is (1) to prevent further deterioration, protect and
enhance the environmental status of aquatic systems,
*
Corresponding author. Tel.: +45-4630-1280; fax: +45-4630-1211.
E-mail address: [email protected] (J.H. Andersen).
0025-326X/$ - see front matter Ó 2004 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2004.04.014
and (2) to promote the sustainable use of water, while
progressively reducing or eliminating discharges, losses
and emissions of pollutants and other pressures for the
long-term protection and enhancement of the aquatic
environment. The WFD provides national and local
authorities with a legislative basis for the maintenance
and recovery of water quality to achieve good ecological
and chemical status for all surface waters and good
chemical status for groundwater. Accordingly, the WFD
is considered to be the most significant piece of legislation regarding water policy produced in the last 20
years.
The coastal waters covered by the WFD with respect
to biological features are generally limited to surface
waters one nautical mile from the coast. With respect to
chemical features the areas in question are limited to the
surface waters 12 nautical miles from the coast. Open
marine waters are not covered, but the WFD is likely to
influence management of all marine ecosystems because
all land-based inputs of pollutants pass through the
coastal zone to the open waters.
284
J.H. Andersen et al. / Marine Pollution Bulletin 49 (2004) 283–290
The WFD requires EU Member States to develop
classification systems to describe the ecological status of
a given water body at a given time. We use here a palaeoecological method to establish reference conditions
for TN concentration in a given coastal water (Clarke
et al., 2003). Using these reference conditions we present
examples of how the ecological status of the southern
part of Roskilde Fjord, Denmark, could be classified
according to the principles of the WFD. In this viewpoint, we show that a wide range of environmental
protection will be achieved based upon decision(s)
regarding the deviation from reference conditions and
our objective is to promote a general discussion of
acceptable deviations from reference conditions.
The WFD is not the only directive seeking to improve
the eutrophication status of European coastal waters
and there are other directives seeking to manage both
inputs and the ecological structure and functioning of
coastal ecosystems (Elliot et al., 1999; Elliot and de
Jonge, 2002). Inputs of nutrients from point sources and
losses of nitrogen from agricultural practices are managed via the Urban Waste Water Directive and the
Nitrates Directive, respectively (Anon., 1991a,b). Other
directives, e.g. the Habitats Directive (Anon., 1992) and
the Birds Protection Directive (Anon., 1979), indirectly
influence management practices via conservation
objectives (i.e. abundance of species or food availability), which are influenced by nutrient enrichment and
eutrophication. The objectives of these directives are
summarised in Table 1.
2. Determination of reference conditions
Establishment of reference conditions in aquatic
systems can be made in a number of different ways for
the WFD. The methods described in the legislation for
the WFD are (1) spatially based reference conditions
(including historical data), (2) modelling (empirical or
dynamic), (3) combinations of (1) and (2), and (4) expert
judgement.
If undisturbed or minimally disturbed sites are
available, reference conditions can readily be defined by
comparing other sites to these undisturbed sites. However, in marine environments, where waters are easily
transported from one area to another through currents
and mixing, isolated or undisturbed sites are generally
not available, especially in the nutrient impacted enclosed seas and coastal areas of Europe. Therefore,
alternative methods must be used to define reference
conditions in coastal waters.
The first and most widely used method to determine
reference conditions is the use of monitoring data
combined with historical data on the system to estimate
conditions prior to large-scale disturbance. While historical data are not abundant, there is a surprising
amount of data collected from earlier surveys that can
be mined and used to help estimate reference conditions.
Another important methodology that is widely used
to estimate reference conditions is to make predictive
watershed models to estimate long-term trends in
nutrient loading with changes in land-use and agricul-
Table 1
Summary of EU Directives other that the WFD of relevance to management of coastal eutrophication
Directive 91/676/EEC of 12 December 1991 concerning the protection of waters against pollution caused by nitrates from agriculture
The objective of the Nitrates Directive is to reduce water pollution caused or induced by nitrates from agricultural sources and to prevent further
such pollution. EU Member States shall designate vulnerable zones, which are areas of land draining into waters affected by pollution and which
contribute to pollution. Member States shall set up where necessary action programmes promoting the application of the codes of good
agricultural practices. Member States shall also monitor and assess the eutrophication status of freshwater, estuaries and coastal waters every four
years.
Directive 91/271/EEC of 21 May 1991 concerning urban waste water treatment
The objective of the Urban Waster Water Directive is to protect the environment from the adverse effects of discharges of waster-water. The
directive concerns the collection, treatment and discharge of urban waste water and the treatment of discharges of waste water from certain
industrial sectors. The degree of treatment (i.e. emission standards) of discharges is based on assessment of the sensitivity of the receiving waters.
Member States shall identify areas, which are sensitive in terms of eutrophication. Competent authorities shall monitor discharges and waters
subject to discharges.
Directive 92/43/EEC of 21 May 1992 on the conservation of natural habitats and of wild fauna and flora
The objective of the Habitats Directive is to contribute towards ensuring biodiversity through the conservation of natural habitats and of wild flora
and fauna in European territory of the Member States. Measures shall be designed to maintain or restore, at favourable conservation status,
natural habitats and species of wild flora and fauna of interest. The habitats and species protected are identified and defined in Annex I and II,
respectively. Many coastal waters in the Northern Europe are identified as eutrophic due to anthropogenic inputs. Member States are required to
implement management plans in order to restore these coastal waters and to achieve a favourable conservation status. Member States shall
monitor habitats and species with particular regard to priority habitat types and priority species.
Directive 79/409/EEC of 2 April 1979 on the conservation of wild birds
The objective of the Birds Protection Directive is the long-term protection and conservation of all birds naturally living in the wild within the
European territory of the Member States. The directive seeks to protect, manage and regulated all bird species, including eggs, nests and habitats.
Member States must conserve, maintain or restore the biotopes and habitats of these birds. As many birds lives in coastal waters subject to
eutrophication Member States are required to restore destroyed or impaired biotopes and to maintain undisturbed habitats.
J.H. Andersen et al. / Marine Pollution Bulletin 49 (2004) 283–290
tural practices (Billen and Garnier, 1997). These models
can also be used to estimate the effects of changes in
management practice on improvements in water quality.
An alternative methodology is the use of palaeoreconstruction using relationships between fossil remains and environment to infer past conditions (ter
Braak and Juggins, 1993). Palaeoecological methods
based on quantitative transfer functions have received
widespread use in freshwaters (e.g. Bennion et al., 1996),
but have only recently been applied to coastal waters
(Clarke et al., 2003; Weckstr€
om et al., in press). All
methodologies require the use of expert judgement and
the final determination of reference conditions benefits
from a combination of methodologies.
A quantitative palaeoecological inference-model (a
weighted averaging-partial least squares diatom-based
transfer function) has recently been developed for
Danish coastal waters (Clarke et al., 2003). The transfer
function can be used to estimate historical changes in
TN concentrations in overlying waters from the biological record of environmental change, preserved in
fine-grained coastal sediments as fossil diatom assemblages. Clarke et al. (2003) describe the methodology
and present diatom-inferred TN estimates from a 1 m
sediment core from the southern part of Roskilde Fjord
(location: 55°400 92N and 11°580 09E) that represents the
time period from 1850. The reconstruction by Clarke
et al. (2003) indicated that historical annual mean TN
concentrations fluctuated between 58 lmol l1 (1850)
and 50 lmol l1 (1950). A rapid increase in annual
mean TN concentration is seen after the mid-1950s
with the highest diatom-inferred TN concentration (91
lmol l1 ) corresponding to the surface sample collected
in 1995.
The palaeo-reconstruction by Clarke et al. (2003) can
be compared to other methodologies used to establish
reference conditions in Denmark. The 1983 Management Plan for Roskilde Fjord and its catchment estimated annual mean TN concentrations in reference
conditions to be on the order of 60–65 lmol l1 using
modelling (Hovedstadsr
adet, 1983). By comparison, in
Odense Fjord, a relatively small (65 km2 ) and nutrientimpacted Danish estuary (Conley et al., 2000), reference
conditions in terms of annual TN means have been
estimated to 48–51 and 19–23 lmol l1 for a station in
the inner estuary (55°270 11N, 10°280 69E) and one in the
middle of the estuary (55°280 75N, 10°310 15E), respectively, also using a modelling approach (DHI, 2002,
2003) that combines hydrodynamic transport and biological features in a 3-D fashion (MIKE 3 model). The
actual numbers have been estimated by using the climatic and meteorological conditions in 1998 and 2002.
However the scenarios radically altered various input
data (e.g. nutrient loads and concentrations in rivers,
atmosphere and the connecting coastal sea), and process
specifications (e.g. growth characteristics of phyto-
285
plankton and macroalgae) to fit into a so-called ‘‘natural
state’’ concept from which the reference conditions have
been drawn. In a third example from Denmark, Randers
Fjord, the reference concentrations of TN have been
estimated to be 57 lmol l1 , using dynamic, deterministic modelling (Nielsen et al., 2003).
On a cautionary note, we know that European estuarine and coastal systems have experienced massive
changes in nutrient loading from anthropogenic activities during the history of human occupation (Billen and
Garnier, 1997). Increases in nutrient loading between
pristine conditions to present are certainly much larger
than those that have occurred during the last 100–150
years (Conley, 1999). For example, a historical reconstruction of P concentrations in Dallund Lake on the
island of Fyn, Denmark has shown that the largest
changes in P concentrations occurred during the Middle
Ages associated with changes in land-use, e.g., forest
clearance, increasing in concentration 6–8 times from
the year 1000 to 1200 AD (Bradshaw, 2001). Therefore,
reference conditions determined for the time prior to the
intensification of agriculture will not necessarily reflect
an unimpacted state.
Variations in TN concentration through time are
observed in the palaeoecological reconstruction in
Roskilde Fjord, Denmark, by Clarke et al. (2003), with
the lowest nutrient concentrations observed in the 1950s.
Changes in nutrient loading with changing land use
practices (Billen and Garnier, 1997), industrialisation
(Billen et al., 1999), conversion from animal derived
locomotion to the combustion engine (Nixon, 1997),
development of sewage treatment practices, and climate,
combine to influence nutrient concentrations. In addition, estuarine systems are dynamic entities where
freshwater and associated nutrients from the land are
mixed with water from coastal waters, such that the
average nutrient concentration will be to a great extent
dependent upon location in the estuary.
3. Classification of ecological status
The WFD requires classification, in terms of ecological status, for all European surface waters. The classification should be based on reference conditions, which
are intended to represent minimal anthropogenic impact, and observed deviation from these conditions.
Ecological status is to be expressed as a numerical value
(the ecological quality ratio) between 1 (high ecological
status) and 0 (bad ecological status) with intervals
equating to: high, good, moderate, poor and bad ecological status (Fig. 1). The lower value in the good
ecological status bracket is equivalent to the higher
value in the moderate ecological status bracket and has
been set as the management quality standard. The WFD
requires that all waters must be restored to above this
286
J.H. Andersen et al. / Marine Pollution Bulletin 49 (2004) 283–290
Fig. 1. Basic principles for the classification of ecological status based on ecological quality ratios (adapted from Anon., 2003). In this classification,
equal intervals between the different classes of ecological status are assumed.
value by 2017 (Anon., 2000). The ecological quality
ratio is defined as the relation between the observed
condition and the reference condition. The management
quality standard is described as an ‘‘ecological status
that shows a low level of distortion resulting from human activity, but deviates only slightly from those normally associated with the water body type under
undisturbed conditions’’ (Anon., 2000). The legislation
defines these conditions to be pristine with no or very
minor deviations from undisturbed conditions. Practically they are being defined as conditions prior to the
intensification of agriculture 100–150 years ago as the
post-war intensification of agriculture and urban pollution are believed to have had the largest impacts on
coastal waters. Regarding nutrients it means ‘‘concentrations do not exceed the levels established so as to
ensure the functioning of the ecosystem and the
achievement of the values specified for biological quality
elements’’ (Anon., 2000).
The classification of ecological status sensu the WFD
and the management quality standard of ‘‘good ecological status’’ is likely to match the sensitive (and less
sensitive) areas sensu the Urban Waste Water Treatment
Directive and the vulnerable (and less vulnerable) zones
sensu the Nitrates Directive (Fig. 2). Correspondence of
management practices and standards between directives
is important and a situation with different objectives and
standards is avoided. If the correspondence is not
identical, management of coastal eutrophication could
become complicated beyond reason.
We consider the estimated annual mean TN concentrations from Clarke et al. (2003) using a palaeoecological
methodology to provide an independent, robust estimate
of past nutrient concentrations consistent with other
independent methods of estimation. The method could be
used to determine reference condition(s), these being the
best condition(s) within ‘‘high ecological status’’.
The selection of class boundaries will be one of the
most politically sensitive issues in setting ecological
quality standards. Once a reference value is decided, a
deviation from this value must be set in order to determine the border between a high and a good ecological
status. It is expected that a suitable boundary will be
assigned from evaluation of data or historical records
(Anon., 2003). The critical value will be the border between good and moderate ecological status. Since the
WFD requires that all waters be restored to be above
this value, the threshold value between good and moderate status will have significant economic consequences.
The remaining class boundaries below this value will be
set to account for the remainder of the scale. Initially
member states may set their own class boundaries,
however harmonisation into a common scale will be
made within Europe through intercalibration exercises.
At present there is no commonly accepted panEuropean guidance on what an acceptable deviation
Fig. 2. Proposed correspondence between eutrophication related ecological quality objectives (EQOs) according to the EU Water Framework
Directive (WFD) and other direct or indirect related EC Directives (Urban Waste Water Directive, Nitrates Directive, Habitats Directive and Birds
Protection Directive), the OSPAR Common Procedure and the present Danish quality objectives. Modified from Ærtebjerg et al. (2003).
J.H. Andersen et al. / Marine Pollution Bulletin 49 (2004) 283–290
from reference conditions is and it is unclear what level
of deviation is to be deemed slight or moderate. To our
knowledge no country in Europe has at this stage tried
to ‘‘translate’’ the WFD lingo of what ‘‘no’’, ‘‘slight’’ or
‘‘moderate’’ deviation from reference conditions are into
a meaningful classification scheme based on reference
conditions. We believe that this paper is the first attempt
to do so and we emphasise that this Viewpoint aims to
open or facilitate the discussion, not to prejudge the
outcome. Classification of ecological status for the
WFD and the development of classification schemes for
eutrophic European coastal waters appear to be in an
early stage (SFT, 1997; Swedish EPA, 2000; OSPAR,
2001). Further development is needed to meet the
requirements of the WFD.
To our knowledge, only a few classifications of
European coastal waters based on reference conditions
have been published (Krause-Jensen et al., in press).
Most existing Danish eutrophication assessment criteria
have defined a 25% deviation from reference conditions
as being acceptable (Ærtebjerg et al., 2003). KrauseJensen et al. (in press) discuss acceptable deviations
from reference conditions (both 25% and 15%). These
authors demonstrate that type specific classification
sensu the WFD could result in both false-positive and
false-negative results. They suggest site-specific classification might be an alternative.
4. Application to a Danish estuary
We have combined these reference values for TN in
the south-eastern part of Roskilde Fjord (from Clarke
et al., 2003) with the Danish management suggestions
287
from Ærtebjerg et al. (2003) and Krause-Jensen et al. (in
press). In practice we have selected three reference
conditions representing a range in nutrient concentrations (50, 54 and 58 lmol l1 ) and assumed that 25% and
15% (from Ærtebjerg et al., 2003; Krause-Jensen et al.,
in press) may be considered a first and provisional definition of what an acceptable deviation from the reference conditions might be. This results in six
classification scenarios (Tables 2 and 3).
According to the six scenarios (A–F), the current
ecological state of the southern parts of Roskilde Fjord,
when using a 5-year weighted mean (78 lmol l1 ) can be
characterised as being at best ‘‘moderate’’ or at worst as
‘‘poor’’. Focusing on a single year will change the outcome of the assessment as shown in Fig. 3.
Which of the six scenarios (A–F) should be used for
the future management of Roskilde Fjord? The decision
will be political and will ultimately have a large impact
upon the level of protection of the marine environment.
Recently, Roskilde County Council has agreed on ecological objectives for submerged aquatic vegetation
(Zostera marina) (HUR, 2001). This objective is likely
to be fulfilled when the annual mean TN concentrations are reduced below 57 lmol l1 (Roskilde County,
unpublished data). This is more or less equal to the most
restrictive scenario (scenario F) with reference conditions set at 50 lmol l1 and an acceptable deviation
assumed to be 15%.
The Swedish EPA has developed a comprehensive
classification system for coastal and open marine waters
(Swedish EPA, 2000). This system, which include five
status classes and uses TN concentrations in surface
water during winter and summer as indices, is not based
on reference conditions as required by the WFD, but on
Table 2
Draft classification scenarios (A–C) of ecological status of TN in seawater in the southern Roskilde Fjord
Scenario
Acceptable
deviation (%)
Reference condition
(lmol l1 )
High
(100–90%)
Good
(90–75%)
Moderate
(75–55%)
Poor
(55–30%)
Bad
(30–0%)
A
B
C
25
25
25
58
54
50
58–62
54–59
50–55
63–72
60–67
56–62
73–84
68–78
63–72
85–99
79–92
73–85
>99
>92
>85
The acceptable deviation from the reference condition (being the value defining the border between good and moderate ecological status) is assumed
to be 25% of the reference value. The vertical line indicates this management standard. The acceptable deviation is expressed in %, reference
conditions in lmol l1 .
Table 3
Draft classification scenarios (D–F) of ecological status of TN in seawater in the southern Roskilde Fjord
Scenario
Acceptable
deviation (%)
Reference conditions
(lmol l1 )
High
(100–95%)
Good
(95–85%)
Moderate
(85–65%)
Poor
(65–35%)
Bad
(35–0%)
D
E
F
15
15
15
58
54
50
58–61
54–57
50–52
62–68
58–62
53–57
69–78
63–73
58–67
79–96
74–89
68–83
>96
>89
>83
The acceptable deviation from the reference condition (being the value defining the border between good and moderate ecological status) is assumed
to be 15% of the reference value.
288
J.H. Andersen et al. / Marine Pollution Bulletin 49 (2004) 283–290
Fig. 3. Ecological status of the southern part of Roskilde Fjord using seasonally weighted mean TN concentrations for the period 1988–2002
according to scenarios A and F. Scenario A represents an acceptable deviation of 25% and a reference condition of 58 lmol l1 . Scenario F represents
an acceptable deviation of 15% and a reference condition of 50 lmol l1 . The size of the quality classes varies and is determined by the percent
deviation from reference conditions. The ecological quality ratio (EQR) is shown to illustrate the relation between TN and EQR.
available monitoring results for both open waters
and coastal waters. A direct comparison should therefore be done with caution, however, if the Swedish
assessment criteria are applied to Roskilde Fjord, the
TN concentrations are to be characterised as ‘‘very
high’’ equivalent to ‘‘bad ecological status’’. The same
results would occur using nutrient concentrations in the
Mediterranean Sea as the standard for Danish coastal
waters. A concentration determined for one location or
Member State cannot be applied on a pan European
basis.
Roskilde Fjord is located within the area of the
OSPAR Convention (see www.ospar.org) and is therefore also subject to the OSPAR Common Procedure for
the Identification of the Eutrophication Status of the
Maritime Area (OSPAR, 1998, 2001). In this context,
the generally acceptable deviation from background
concentrations or reference conditions is 50%. If this
general principle is applied to Roskilde Fjord, the
acceptable concentrations will be in the range of 75–87
lmol l1 , indicating that the area in question is likely to
be a ‘‘non-problem area’’ with no eutrophication effects.
This conclusion disagrees with our assessment presented
above and with the annual assessment by Roskilde
County. Parts of the OSPAR principles need a scientific
review in order to test the principles and to ensure that
the strategy to combat eutrophication is based on verified information.
J.H. Andersen et al. / Marine Pollution Bulletin 49 (2004) 283–290
5. Conclusions and perspectives
The definition of reference conditions and classification of the ecological status represents an important
aspect in the implementation of the WFD. The determination by Clarke et al. (2003) of reference conditions
in a Danish estuary demonstrate that the use of palaeoecological methods produces results with a reasonable
degree of accuracy. However the subsequent decision on
the deviation from the reference conditions actually
controls the classification of the ecological status, and
complicates an objective classification. The palaeoecological approach applied here is a useful tool to reconstruct reference conditions in Roskilde Fjord and
possibly also in other estuaries and coastal marine areas.
There is an urgent need to discuss and define procedures to establish reference conditions and quality classes in a variety of environmental conditions. It is
important to translate ‘‘slight’’ and ‘‘moderate’’ deviations into acceptable deviations from reference conditions. However, the implementation of the WFD is still
only in its beginning phase. The coming years will,
therefore, be a learning process during which our proposed classification will be challenged and tested. In this
process it is important to integrate science and management (Gray, 1999; Elliot et al., 1999) in order to ensure
that the WFD will be a strong instrument to manage the
ecological quality of European coastal waters.
Acknowledgements
Thanks to our colleagues Annemarie Clarke,
Bo Riemann, Jens Brøgger Jensen (Danish EPA) and
Michael Hjort Jensen (Fyn County) for valuable discussion and comments on earlier versions of the manuscript. This research was partially supported by
European Union funding of the MOLTEN (contract
EVK3-CT-2000-00031) and CHARM (contract EKV3CT-2001-0065) projects.
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HORIZONS
Coastal eutrophication: recent
developments in definitions and
implications for monitoring strategies
JESPER H. ANDERSEN1*, LOUISE SCHLÜTER1 AND GUNNI ÆRTEBJERG2
1
DEPARTMENT OF ECOLOGY AND ENVIRONMENT, DHI WATER & ENVIRONMENT, AGERN ALLÉ 5,
2790 HØRSHOLM, DENMARK AND 2DEPARTMENT OF
4000 ROSKILDE, DENMARK
MARINE ECOLOGY, NATIONAL ENVIRONMENTAL RESEARCH INSTITUTE, FREDERIKSBORGVEJ 399,
*CORRESPONDING AUTHOR: [email protected]
Received September 15, 2005; accepted in principle January 12, 2006; accepted for publication March 22, 2006; published online March 29, 2006
Communicating editor: K.J. Flynn
The word ‘eutrophication’ has its root in two Greek words: ‘eu’ which means ‘well’ and ‘trope’ which
means ‘nourishment’. The modern use of the word eutrophication is related to inputs and effects of
nutrients in aquatic systems. Despite a common understanding of its causes and effects, there is no
agreed definition of coastal eutrophication. This communication aims to review recent developments in
the definitions of coastal eutrophication, all of which focus on ‘accelerated growth’, and to discuss the
implications in relation to monitoring and assessment of ecological status. It is recommended that
measurements of primary production, being a sensitive and accurate indicator of eutrophication, should
be mandatory when monitoring and assessing the ecological status of coastal waters.
INTRODUCTION
Eutrophication of coastal waters has been considered one
of the major threats to the health of marine ecosystems
for more than 30 years (Ryther and Dunstan, 1971;
Nixon, 1995; Elmgren, 2001; Bachmann et al., 2006).
The different processes and effects of coastal eutrophication are well known and documented (Cloern, 2001;
Conley et al., 2002; Rönnberg and Bonsdorff, 2004).
In 2000, the European Parliament and the Council
adopted the European Union (EU) Water Framework
Directive (WFD), which provides a framework for the
protection of groundwater, inland surface waters, transitional waters (estuaries) and coastal waters (Anonymous,
2000). The overall aim of the WFD was: (i) to prevent
further deterioration, protect and enhance the environmental status of aquatic systems and (ii) to promote the
sustainable use of water while progressively decreasing or
eliminating discharges, losses and emissions of pollutants
and other pressures for the long-term protection and
enhancement of the aquatic environment. The WFD
is intended to improve the ecological status, including
eutrophication status, of all European surface waters of
which many are considered to be eutrophic (European
Environment Agency, 2001, 2003). The directive provides
national and local authorities with a legislative basis for the
maintenance and recovery of water quality to achieve good
ecological and chemical status for all surface waters and
good chemical status for groundwater. Accordingly, the
directive can be considered the most significant piece of
legislation of the last 20 years, in regard to water policy not
only in Europe but also in non-European countries seeing
EU legislation as a benchmark for their own legislation.
Written responses to this article should be submitted to Kevin Flynn at [email protected] within two months of publication. For
further information, please see the Editorial ‘Horizons’ in Journal of Plankton Research, Volume 26, Number 3, Page 257.
doi:10.1093/plankt/fbl001, available online at www.plankt.oxfordjournals.org
Ó The Author 2006. Published by Oxford University Press. All rights reserved. For Permissions, please email: [email protected]
JOURNAL OF PLANKTON RESEARCH
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However, the WFD lacks a definition of eutrophication. The directive’s treatment of eutrophication is indirect, with the boundary between good and moderate
ecological status being defined as an environmental management objective. For waters failing to meet the objective of at least good ecological status, the directive
requires that competent authorities establish programmes of measures and river basin management
plans to secure this status. The measures to be implemented in the context of eutrophication are already
required under other existing directives, for example,
the Urban Waste Water Treatment (UWWT) Directive
(Anonymous, 1991a) and the Nitrates Directive
(Anonymous, 1991b). If these are insufficient, then the
implementation of supplementary measures is required.
The WFD thus acts as an umbrella for the UWWT
Directive and the Nitrates Directive, and as such it has
to respect the definitions of eutrophication in these
directives.
HOW IS EUTROPHICATION
DEFINED?
Within the EU, there has been a sound tradition of
focusing measures on the sources causing eutrophication
(Elliot et al., 1999; Elliot and de Jonge, 2002).
Consequently, eutrophication has been defined in relation to sources and/or sectors. For example, the
European Commission (EC) UWWT Directive defines
eutrophication as ‘the enrichment of water by nutrients,
especially nitrogen and/or phosphorus, causing an accelerated growth of algae and higher forms of plant life to
produce an undesirable disturbance to the balance of
organisms present in the water and to the quality of
water concerned’ (Anonymous, 1991a).
According to the EC Nitrates Directive, eutrophication is defined as ‘the enrichment of water by nitrogen
compounds causing an accelerated growth of algae and
higher forms of plant life to produce an undesirable
disturbance to the balance of organisms present in the
water and to the quality of water concerned’
(Anonymous, 1991b). The difference between the two
definitions can be explained by the focus of the Nitrates
Directive which, perhaps unsurprisingly, rests on losses of
nitrogen from agriculture.
There has been some justifiable discussion of these
definitions, in particular their focus on nutrients, and
also the need to clarify what constitutes an ‘undesirable
disturbance’ and an ‘accelerated growth’. Is ‘accelerated’
the right word to use in this context? No, accelerated,
meaning speed up, is in our opinion the wrong word and
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should be replaced by ‘increased’. Nixon (Nixon, 1995)
defines eutrophication as ‘an increase in the rate of
supply of organic matter to an ecosystem’. This definition
is short and consistent with historical usage and emphasizes eutrophication as a process rather than a trophic
state. Nixon also notes that the increase of the supply of
organic matter to coastal systems may have various
causes, but the common factor is clearly nutrient enrichment. The supply of organic matter to an ecosystem is
not restricted to pelagic primary production, even
though such an interpretation leads to a convenient
operational definition. It also includes primary production of higher plants and benthic microalgae as well as
inputs of organic matter from adjacent waters or from
land, via rivers or point sources. Having such a broad
interpretation of the term ‘supply’ makes the definition,
despite its obvious strengths, difficult to apply in a monitoring and management context.
Eutrophication and definition(s) of eutrophication are
much discussed topics as indicated above and also
pointed out by Jørgensen and Richardson (Jørgensen
and Richardson, 1996). The most common use of the
term is related to inputs of mineral nutrients, primarily
nitrogen and phosphorus, to specific waters.
Consequently, eutrophication deals with both the process
as such, the associated effects of nutrient enrichment and
natural versus cultural caused eutrophication. And as
prudently pointed out by Jørgensen and Richardson,
when we speak of eutrophication, it is anthropogenic
eutrophication that is of interest.
Within the OSPAR Convention for the Protection of
the Marine Environment of the North-East Atlantic, the
definition of eutrophication follows the above definitions
and thoughts and defines eutrophication similar to the
UWWT Directive and continues ‘and therefore refers to
the undesirable effects resulting from anthropogenic
enrichment by nutrients described in the Common
Procedure’ (OSPAR, 2003).
The implementation of the WFD has revealed the
need for a common understanding and definition of
eutrophication as well as a need for stronger coordination between directives dealing directly or indirectly with
eutrophication. The EC has initiated a process with the
aim of developing a pan-European conceptual framework for eutrophication assessment in the context of all
European waters and policies. At a workshop in
September 2004, hosted by the EC and Joint Research
Centre in Ispra, Italy, draft guidance on a pan-European
framework for assessment of eutrophication was presented and discussed. The objective of the workshop
was to coordinate different activities under the EU
WFD and other eutrophication-related directives (e.g.
UWWT Directive and Nitrates Directive). The
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workshop concluded that a draft pan-European definition of eutrophication could use the UWWT Directive as
a starting point for further developments on the issue of
eutrophication. Taking the comments put forward at the
workshop into consideration, eutrophication can be
defined as ‘the enrichment of water by nutrients, especially nitrogen and/or phosphorus and organic matter’
(Anonymous, 2004). Work is ongoing and expected to be
reported in the spring 2006 in the form of an interim
guidance document. Revision of the guidance is planned
in 2007, following the WFD inter-calibration exercise
and some on-going activities by the conventions for the
protection of the marine environment of Baltic Sea and
the North-East Atlantic.
TOWARDS A PROCESS-ORIENTED
MONITORING AND ASSESSMENT
STRATEGY
How are member states of the EU obliged to monitor
and assess the ecological status of coastal waters?
Monitoring networks should be established to create a
coherent and comprehensive overview of ecological and
chemical status and ecological potential. The networks
should be operational by 20 December 2006 or by 1
January 2007 at the latest. Monitoring networks should
in principle be based on variables/indicators that are
indicative of the status of each relevant quality element
[biological (e.g. phytoplankton, submerged aquatic vegetation and invertebrate benthic fauna), hydromorphological or physiochemical]. In addition, the networks
should permit classification of water bodies in five classes
consistent with the normative definitions of ecological
status.
In a North European perspective, there are at least
two or three important drivers for the design, execution
and reporting of monitoring activities. These are the
WFD including the WFD Common Implementation
Strategy guidance on monitoring (Anonymous, 2000,
2003a), the HELCOM COMBINE Programme (Cooperative Monitoring in the Baltic Sea Environment)
(HELCOM, 2003) and the OSPAR Joint Assessment
and Monitoring Programme (JAMP), including the
Eutrophication Monitoring Programme, which describes
the indicators and sampling methods (OSPAR, 2004,
2005). So far, the pan-European process for development
of a conceptual framework for eutrophication assessment
has not included discussion of specific monitoring guidance. This will take place at a later stage. The only
available guideline for selection of indicators is a draft
holistic checklist (Anonymous, 2004).
The requirement relating to the monitoring of pelagic
biological and chemical indicators in EU WFD,
HELCOM COMBINE, OSPAR JAMP/Coordinated
Environmental Monitoring Programme (CEMP) and
the ongoing pan-European process is summarized in
Table I. Measurements of phytoplankton species abundance, composition and biomass are mandatory in most
monitoring networks. Measurements of chlorophyll a
(Chl a) and nutrients are mandatory within HELCOM
and OSPAR but considered a recommended supporting
indicator by European drivers. Measurements of primary
production are not mandatory at present.
How to assess ecological status?
The WFD requires EU member states to develop classification systems to describe the ecological status of a given
water body at a given time. The results of the monitoring
programmes are the basis for an assessment of ecological
status of a given water body that according to the directive
will fall into one of five classes (categories): high, good,
moderate, poor or bad. The status classes high and good
are in general considered to be acceptable.
An important step in assessing ecological status is the
setting of reference condition standards with the objective of enabling the assessment of ecological quality
against these standards. Reference condition is in this
context defined as a description of the biological quality
elements that exist, or would exist, at high status, that is,
with no, or very minor, disturbance from human activities (Anonymous, 2003b).
Another important step is to define what constitutes an
acceptable deviation. An acceptable deviation sensu the
WFD is to us equivalent to high and good ecological
status, the latter defined as a status where the values of
the biological quality elements show low levels of distortion resulting from human activity. An unacceptable
deviation is in our understanding equivalent to bad,
poor or moderate ecological status, where values of the
biological quality elements deviate moderately or more
from those normally associated with the coastal water
body type under undisturbed conditions sensu the WFD
definition of reference conditions.
The approach employed in the so-called OSPAR
Comprehensive Procedure (COMPP) is very pragmatic
and straightforward. On the basis of background values,
in practice identical to reference conditions, a water body
is considered an ‘Eutrophication Problem Area’ if actual
status deviates 50% or more from reference conditions
(OSPAR, 2003). It should be noted that the choice of
50% is arbitrary, not based on any scientific considerations about ecological changes caused by nutrient enrichment. The application of percentages lower than
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Table I: Selection of relevant quality elements and indicators by WFD, HELCOM COMBINE,
OSPAR COMPP and the draft holistic checklist of the pan-European conceptual framework for
eutrophication assessment
Quality elements and indicators
EU WFD
HELCOM
OSPAR
pan-European
Abundance
M
M
(R)
(R)
Composition
M
M
M
(R)
Phytoplankton
Diversity
M
(R)
(R)
(R)
Biomass
M
M
(R)
R
Primary production
n.i.
R
n.i.
R
Chlorophyll a
R
M
M
R
Fluorescence
n.i.
R
n.i.
n.i
Secchi depth
R
M
R
n.i.
Light attenuation
n.i.
Ma
n.i.
R
Turbidity
R
n.i.
n.i.
R
Color
R
R
n.i.
n.i.
Transparency
Nutrients
Total P
R
M
M
R
Soluble reactive P
R
M
M
R
Total N
R
M
M
R
Nitrate + nitrite
R
M
M
R
Ammonium
R
M
M
R
Silicate
n.i.
M
n.i.
R
COMPP, Comprehensive Procedure; EU WFD, European Union Water Framework Directive; M, mandatory; R, recommended; (R), recommended
indirectly; n.i., no information. Compiled from Anonymous (Anonymous, 2003a; Anonymous, 2003b; Anonymous, 2004), HELCOM (HELCOM, 2003) and
OSPAR (OSPAR, 2005).
a
Mandatory when primary production is measured.
50% has been discussed, for example, by Ærtebjerg
et al. (Ærtebjerg et al., 2003), Andersen et al. (Andersen
et al., 2004) and Krause-Jensen et al. (Krause-Jensen et al.,
2005). Recently, the OSPAR Eutrophication Committee
amended the procedures of the next application of
the Comprehensive Procedure, so that the acceptable
deviation should be justified but not exceed 50%
(OSPAR, 2005).
How can primary production be estimated?
With the development in relation to a pan-European
definition of eutrophication, it would be logical to focus
monitoring on relevant biological indicators including
measurement of ‘increased growth’. In our understanding, measurement of primary production is a relevant
indicator that can indicate if algal growth is increased.
Primary production is a fundamental ecological indicator (variable), because it is a measure of the extent to
which primary energy input (solar energy) to the aquatic environment is transformed into the biological/
ecological sphere. It is defined as the flux of inorganic
carbon into planktonic algae per unit time. It has significant capability to indicate and characterize the status of a
particular water body. Primary production can conveniently be measured using the so-called 14C method
(Steemann Nielsen, 1952). When adding a known quantity of the radioactive isotope 14C to a water sample, the
planktonic algae will take up 14C along with ‘native’ 12C
present in water. After a short incubation period (2 h), the
14
C incorporated into the algal cells can be measured by
liquid scintillation counting. The total carbon uptake,
which is a good approximation of net production
(Jespersen et al., 1995), can then be calculated by:
12
CO2 uptake ¼ ð14 CO2 uptake=14 CO2 addedÞ
12 CO2 concentration
Primary production can either be determined as
particulate production or total production. For
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particulate production, only the 14C uptake in the
algae cells is determined, whereas total production
also includes the 14C incorporated into the organic
matter, which can be lost to the environment outside
the cell during incubation. The method is very sensitive, and primary production is a widely used
method when assessing eutrophication effects in
coastal waters (e.g. Pinckney et al., 1999; Prins
et al., 1999; Bonsdorff et al., 2002). Primary production is also used as an important indicator when
modelling how changes in loads impact upon the
environment.
Various research activities and monitoring networks
have made use of the 14C method and have documented considerable changes in the levels of the primary production since the 1950s (e.g. Richardson and
Heilmann, 1995; Bonsdorff et al., 2002). In the central
Great Belt, Denmark (558 220 3600 N, 118000 E), the
annual primary production, averaged over each decade, has roughly doubled from the 1950s to the 1980s
and 1990s (Fig. 1). In the central Kattegat, the average monthly primary production at four different
depths in the water column through the year is compared for the two periods 1954–60 and 1984–93 (Fig.
2). It can be seen that both the spring bloom and the
algal production during the summer months increased
significantly from the 1950s to 1984–93, as a consequence of eutrophication (Jørgensen and Richardson,
1996).
How to link the definition with monitoring
and assessment activities?
Despite positive pan-European developments in defining
eutrophication, it is still unclear what an ‘undesirable
disturbance’ is. The phrase is open to interpretation
Primary production, g C m–2 month–1)
200
150
100
50
0
1950s
1960s
1970s
1980s
1990s
Decades
Fig. 1. Examples of observed changes in the primary production in the
central Great Belt, Denmark, depicted as averages of the annual primary production of the decades [unpublished data from G. Ærtebjerg,
NERI, Denmark].
Primary production (g C m–2 month–1)
50
1954–1960
1984–1993
40
30
20
10
0
J
F
M
A
M
J
J
A
S
O
N
D
Month
Fig. 2. Primary production in the Kattegat, Denmark, through the
year as estimated by Steemann Nielsen (Steemann Nielsen, 1964) and
Richardson and Heilmann (Richardson and Heilmann, 1995) [From
Jørgensen and Richardson (1996). Copyright 1996 American
Geophysical Union. Modified by kind permission of American
Geophysical Union].
and should be reconsidered. We suggest that an ‘undesirable disturbance’ in ecological terms is understood as an
‘unacceptable deviation from reference conditions’. We
realize that an ‘unacceptable deviation’ is also open to
interpretation, but the advantage is 2-fold. First, the
definition will be linked to the WFD implementation
process, and second, reference conditions sensu the
WFD will be the starting point.
We also suggest inclusion of primary production measurements in monitoring systems. These should be based
on a reasonable and cost-effective approach, that is,
monitoring networks should be stratified and based on
two types of stations: (i) intensive stations/areas where
many indicators are monitored with high frequency and
(ii) mapping stations where a few indicators are monitored with lower frequency. This kind of stratification has
been used in the HELCOM COMBINE Programme
(HELCOM, 2003) and in Danish National Marine
Monitoring and Assessment Programme 2003–09
(DNAMAP) (Andersen, 2005).
In our opinion, measurements of primary production
should be carried out at all intensive stations or at least
one coastal station per type of coastal water or river basin
district. Sampling frequency should be based on information on the ecological status and take seasonal variations
at the station into account.
We also recommend that primary production measurements should follow the methodology developed
within International Council for the Exploration of the
Sea (ICES) and currently described in the HELCOM
COMBINE Manual (HELCOM, 2003). However, existing time series on primary production should be continued using the original measurement method.
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We are of the opinion that the 14C method allows
precise determination of phytoplankton production.
However, these measurements are not mandatory in monitoring programmes coordinated on an international level
(e.g. HELCOM COMBINE, OSPAR JAMP and WFD
related monitoring activities). If our suggestion of including estimates of primary production in the monitoring
programmes is followed, then these programmes will be
linked directly to both the definition and process of eutrophication. Other methods for determining primary production could be employed, for example, non-isotope
method, that is, the oxygen method (Hall and Moll,
1975; Reid and Shulenberger, 1986; Olesen et al., 1999).
An indicator often used for assessment of eutrophication and as a proxy for primary productivity, nutrient
status or phytoplankton biomass is Chl a. Some caution
is recommended when using this indicator, and the information inherent in Chl a measurements should be interpreted as what it is: a Chl a concentration and nothing
more, cf. Kruskopf and Flynn (Kruskopf and Flynn, 2006).
DNAMAP 2003–09, which implements the monitoring
requirements of the WFD, was designed according to a
principle stating: ‘No monitoring without Ecological
Quality Objectives, no Ecological Quality Objectives without monitoring’ (Svendsen and Norup, 2005). We completely agree with this principle and present a total of nine
draft classification scenarios on the basis of percentage
deviations for the various boundaries between the classes
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high, good, moderate, poor and bad (Table II). The scenarios are site specific (The Great Belt, Denmark) and not
directly applicable to other coastal waters. They are also
specific for the results of primary production measurements and may not be applicable for other indicators. As
a cautionary note, we acknowledge that the decision on
which of the presented scenarios to implement as an
environmental management standard will be political.
CONCLUSIONS
Our mission is to propose a better definition of eutrophication and to link the definition with monitoring and
assessment systems. By understanding in ecological
terms an ‘undesirable disturbance’ as an ‘unacceptable
deviation from reference conditions’, we arrive at a definition that is consistent with the normative definitions of
moderate (and poor/bad) ecological status sensu the
WFD. Consequently, an acceptable deviation will correspond to the normative definition of high and good
ecological status.
Accepting the above suggestions allows a definition of
eutrophication as ‘the enrichment of water by nutrients,
especially nitrogen and/or phosphorus and organic matter, causing an increased growth of algae and higher
forms of plant life to produce an unacceptable deviation
in structure, function and stability of organisms present
Table II: Scenarios for ecological classification in the Great Belt, Denmark using primary production
as an indicator and assuming that deviations of 15% (restrictive), 25% (intermediate) and 50%
(non-restrictive) from reference conditions are acceptable deviations
Scenarios
Reference conditions
Restrictive
Primary production
High (%)
Good (%)
Moderate (%)
Poor (%)
Bad (%)
<5
5–15
15–35
35–65
A1
48
<50
50–55
55–65
65–79
>65
>79
A2
67
<70
70–77
77–90
90–111
>111
A3
86
<90
90–99
99–116
116–142
>142
>70
<10
10–25
25–45
45–70
B1
48
<53
53–60
60–70
70–82
>82
B2
67
<74
74–92
92–97
97–114
>114
B3
86
<95
95–108
95–125
125–146
>146
>90
Intermediate
<20
20–50
50–70
70–90
C1
48
<58
58–72
72–82
82–91
>91
C2
67
<80
80–100
80–114
114–127
>127
C3
86
<103
103–129
103–146
146–163
>163
Non-restrictive
The primary production is expressed as g C m–2 year–1. Reference conditions in scenarios A1, B1 and C1 are defined by Hansen et al. (Hansen et al.,
2003). Reference conditions in scenarios A3, B3 and C3 are defined by Ærtebjerg (unpublished data). Scenarios A2, B2 and C2, where the reference is
67 g C m–2 year–1, are an average of 48 and 86 g C m–2 year–1. The approach used for division in five quality classes is based on Andersen et al. (Andersen
et al., 2004) and Krause-Jensen et al. (Krause-Jensen et al., 2005).
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in the water and to the quality of water concerned,
compared to reference conditions’.
In our opinion, the proposed definition of eutrophication will lead to revision of existing monitoring strategies.
Measurement of primary production, being an indicator
of ‘increased growth’, should be mandatory in monitoring networks and should consequently be included as a
monitoring or an assessment indicator in the panEuropean guidance on a conceptual framework for
eutrophication assessment.
We have raised many rhetorical questions and believe
we have answered most of the questions and by doing so
promoted the idea of having a process-oriented approach
to monitoring and assessment of coastal eutrophication.
However, one important question is still to be answered:
‘How should primary production be measure or estimated?’ Such question requires thorough scientific analyses as well as coordination, otherwise the answer would
be up to individual member states meaning that there will
be only limited coordination.
The approach to be employed in setting up classifications scenarios is a topic for discussion. Our intention is
simply to present some examples of how ecological
classification scenarios could be constructed on the
basis of measurements of primary production. Further
work is needed to verify both the approach and the
scenarios. However, we consider it vital that science
and management are integrated to ensure that the
WFD will be a strong legal instrument for the protection
and, where needed, restoration of the ecological status of
European waters. Implementation of the WFD is still in
its initial phases. The coming years will, therefore, be a
learning process. Agreement on a pan-European definition of eutrophication and putting emphasis on primary
production will be a good start to this process.
Andersen, J. H. (ed.) (2005) Marine waters. In Svendsen, L. M., Bijl, L. van
der, Boutrup, S. and Norup, B. (eds), NOVANA: Nationwide Monitoring and
Assessment Programme for the Aquatic and Terrestrial Environments. Programme
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Denmark, 137 pp. NERI Technical Report No. 537. http://
www2.dmu.dk/1_Viden/2_Publikationer/3_Fagrapporter/rapporter/
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ACKNOWLEDGEMENTS
The authors acknowledge discussions with and inputs
from Jens Brøgger Jensen, Bo Guttmann, Juha-Markku
Leppänen and Ciarán Murray. The manuscript was
improved from comments by the referees and the communicating editor. J.H.A. is partly funded by the
Helsinki Commission (HELCOM EUTRO, Project
No. 11.24/05–06).
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628
Limnol. Oceanogr., 51(1, part 2), 2006, 398–408
q 2006, by the American Society of Limnology and Oceanography, Inc.
Coastal eutrophication and trend reversal: A Danish case study
Jacob Carstensen1
Department of Marine Ecology, National Environmental Research Institute, P.O. Box 358, DK-4000 Roskilde, Denmark;
European Commission, Joint Research Centre, Institute for Environment and Sustainability, TP 280, I-21020
Ispra (VA), Italy
Daniel J. Conley
Department of Marine Ecology, National Environmental Research Institute, P.O. Box 358, DK-4000 Roskilde, Denmark;
Department of Marine Ecology, Aarhus University, Finlandsgade 14, DK-8200 Aarhus, Denmark
Jesper H. Andersen 2 and Gunni Ærtebjerg
Department of Marine Ecology, National Environmental Research Institute, P.O. Box 358, DK-4000 Roskilde, Denmark
Abstract
In the past 2 decades significant measures have been taken to reduce nitrogen and phosphorus discharges from
Denmark by 50% and 80%, respectively. These nutrient reduction targets now appear within reach after several
consecutive reduction measures are fully implemented, particularly toward diffuse discharges, and reduced nutrient
concentrations are beginning to be observed in estuaries and the Danish straits. Phosphorus concentrations have
declined by 22% to 57% from the early 1990s, mainly owing to improved treatment of urban and industrial
wastewater. Changes in nitrogen concentrations, following reduction measures toward diffuse sources, were more
recent and partly masked by large interannual variations in freshwater discharge. The response in marine nitrogen
concentrations was delayed relative to the decline in riverine concentrations, most likely owing to large internal
loading from the sediments. Two consecutive dry years appeared to be the triggering mechanism for nitrogen
concentrations to decline. In the last 5 yr, nitrogen levels in estuaries and coastal waters have decreased up to 44%
when interannual variations in freshwater discharge were accounted for. These first signs of environmental recovery
were most pronounced in estuaries and coastal waters but also were apparent in open waters of the Kattegat, the
Sound, and the Belt Sea. This case study is the first to document significant decreases in nutrient concentrations on
a large regional scale resulting from an active management strategy to reduce nutrients from both diffuse and point
sources.
Eutrophication of coastal ecosystems from nutrient overenrichment is widespread (Nixon 1995), with the effects
manifested in a myriad of direct and indirect responses
(Cloern 2001). Although the sources and pathways of nutrient inputs to aquatic ecosystems can be estimated with reasonable certainty, it has been difficult to achieve reductions
in the different sources (Boesch 2002). However, some
coastal ecosystems have experienced reductions in inputs of
phosphorus and nitrogen primarily through improvement in
treatment of wastewater and reductions in point sources
(Butt and Brown 2000; Conley et al. 2000), although relatively little progress has been made in reducing diffuse
sources of nutrients (Butt and Brown 2000; Boesch 2002).
In eastern Europe a number of studies have shown decreasing nutrient concentrations in rivers and streams from reduced application of fertilizers after the economic collapse
in eastern Europe and the Soviet Union in the early 1990s
(for overview, see Stålnacke et al. 2003).
In 1987 a National Action Plan on the Aquatic Environment was enacted in Denmark to reduce nutrient inputs to
the aquatic environment. This action plan was based on an
agenda adopted in 1986 that aimed to reduce nitrogen and
phosphorus discharges by 50% and 80%, respectively (Kronvang et al. 1993). The commitment to nutrient reductions
was also made in multijurisdictional agreements with both
Helsinki Commission for the Protection of the Marine Environment of the Baltic Sea Area (HELCOM) and Oslo-Paris
Commission for the Protection of the Marine Environment
of the North-East Atlantic (OSPAR) (Conley et al. 2002b).
This first action plan was most effective toward nutrient reductions from municipal wastewater, and it was soon recognized that new actions had to be taken toward the diffuse
loading, in particular nitrogen. Another action plan for sustainable agricultural production followed in 1991 and a second National Action Plan on the Aquatic Environment in
Corresponding author ([email protected]).
Present address: DHI Water & Environment, Agern Allé 5, DK2970 Hørsholm, Denmark.
1
2
Acknowledgments
The present work is a contribution of the CHARM (EVK3-CT2001-00065) and REBECCA (SSPI-CT-2003-502158) projects
funded by the European Commission. We gratefully acknowledge
the Danish counties responsible for data collection under the Danish
Nationwide Aquatic Monitoring and Assessment Program and the
Swedish Hydrological and Meteorological Institute for providing
data from the Swedish monitoring programs. German loading data
were provided by Heike Herata, Federal Environmental Agency in
Berlin, and Thorkild Petenati from the federal state of SchleswigHolstein. We thank Ole Hertel for providing atmospheric deposition
data and Bo Riemann for comments on the manuscript. The manuscript was improved by valuable comments made by three anonymous reviewers.
398
Eutrophication trend reversal
399
measures that have been taken in Denmark to reduce the
load of nutrients from both point and nonpoint sources are
successful. We have analyzed the trends in nutrient loading
and concentrations after the first Action Plan on the Aquatic
Environment and the establishment of DNAMAP in 1989 up
to the most recent data from 2002, a period of 14 yr with
large changes in nutrient loading that has allowed us to identify responses in the ecosystem to nutrient management. This
analysis was possible owing to the extensive data set collected under DNAMAP, and our report is the first documentation of significant effects in the marine ecosystem that can
be traced to nutrient reductions resulting from an active management strategy.
Fig. 1. Map of the Kattegat, the Sound, and the Belt Sea showing location of monitoring stations used in the study, partitioned
into estuarine and coastal stations (squares) and open-water stations
(circles). Boundaries of the study area are marked by dashed lines.
1998. The two latter action plans included a variety of strategies and measures to reduce diffuse nitrogen inputs, including fertilizer reductions, buffer strips, and restoration of
wetlands (for further details, see Conley et al. 2002b).
The Danish National Aquatic Monitoring and Assessment
Program (DNAMAP) was established in 1988 to monitor
nutrient loading and ecological responses to the nutrient reduction targets. DNAMAP was organized with the aim of
obtaining information on a wide range of eutrophicationrelated variables (e.g., nutrients, chlorophyll a, macrophytes,
benthic macrofauna) covering many estuaries and coastal
zones in Denmark. Monitoring in the open waters and selected coastal waters is a requirement of HELCOM and OSPAR, with DNAMAP and the Swedish national monitoring
program operating in a coordinated joint effort.
The objective of this article is to demonstrate that the
Study area—The Kattegat, the Sound, and the Belt Sea
(the Danish straits) comprise a shallow transition area between the brackish Baltic Sea and the more saline Skagerrak/
North Sea (Fig. 1). It is a coastal ecosystem with estuarine
character dominated by advective transports and an almost
permanent halocline located at ;15-m depth (Andersson and
Rydberg 1993; Jakobsen and Trébuchet 2000). Transport in
the surface layer is generally northward, whereas Skagerrak
water penetrates southward into the Danish straits as a bottom current that gradually mixes with the surface layer. The
residual flow from the Baltic Sea is ;500 km3 yr21 (Stigebrandt and Gustafsson 2003), and bottom water exchanges
with the Baltic Sea occur over the sills at Drogden (8-m
depth) and Darss (15-m depth). The area and volume of the
Danish straits are 41,000 km 2 and 810 km3, respectively
(Gustafsson 2000).
A total land area of 64,135 km 2 discharges directly into
the Danish straits with 47%, 37%, and 16% belonging to
Denmark, Sweden, and Germany, respectively (Table 1). The
Göta River (Fig. 1), which is the sixth largest river in the
entire Baltic Sea area, is not included in the figures for the
Kattegat, because discharges from this river mainly are carried northward out of the Kattegat into the Skagerrak. The
Danish straits receive discharges from 70% of Denmark. Approximately 9 million people inhabit the catchment, the majority of these living in urban settlements. Land use is generally dominated by agriculture (Table 1), except for the
Swedish catchment discharging into the Kattegat with 72%
forest.
Local inputs of freshwater and nutrients are primarily discharged through productive estuaries and coastal regions
(Carstensen et al. 2003). The Danish estuaries are for the
Table 1. Catchment area and land use for the Kattegat, the Sound, and the Belt Sea. Areas are from HELCOM (2002), and land uses
were compiled from the GRID-Arendal database (Sweitzer et al. 1996).
Sweden
Denmark
Kattegat
The Sound
Belt Sea
Total
Catchment
area (km 2)
Arable and
pasture (%)
15,850
1,740
12,340
29,930
63
47
68
64
Catchment
area (km 2)
20,920*
2,885
—
23,805
* Excluding the catchment of the Göta River (50,233 km 2).
† The catchment area for the German federal state Schleswig-Holstein is 5,450 km 2.
Germany
Arable and
pasture (%)
18
64
—
23
Catchment
area (km 2)
—
—
10,400†
10,400
Arable and
pasture (%)
—
—
65
65
400
Carstensen et al.
most part shallow (,3 m deep) with relatively short residence times (Conley et al. 2000). The majority of estuaries
have a well-mixed water column with intermittent periods
of stratification during periods of calm winds or inflow of
saline bottom water. Agricultural production in Denmark is
very specialized and highly and consistently productive both
per unit land and per unit resource (Porter and Petersen
1997). Over the past decade, there has been an increase in
animal husbandry, which together with the measures in the
action plans has precipitated a shift from chemical fertilizers
to manure for crop production. Denmark is now the world’s
largest exporter of pork meat, with a standing stock of 13
million pigs in addition to 1.7 million cattle (2003 data from
www.ddl.dk). For comparison, the human population of
Denmark is 5.4 million.
Materials and methods
Detailed load compilations have been carried out in Denmark since 1989 as part of the first Action Plan for the
Aquatic Environment. Data on the freshwater discharges
from the Swedish catchment were obtained from the Swedish Meteorological and Hydrological Institute (SMHI), and
nutrient loading figures were compiled from Stålnacke et al.
(1999) and data from the Swedish University of Agricultural
Sciences (www.slu.se). A long time series was available for
nutrient and freshwater inputs from the German federal state
of Schleswig-Holstein (1977–2002), whereas total inputs
from Germany to the Belt Sea were available from 1994
onward (data source, Federal Environmental Agency, Berlin,
Germany). We calculated the average ratio between total
German input and that from Schleswig-Holstein for 1994–
2002 and used this value for scaling up the inputs from
Schleswig-Holstein during 1989–1993. Danish nutrient loading was partitioned into riverine and point source contributions, the latter combined of discharges directly to marine
waters and discharges to freshwater carried with the riverine
input. The diffuse nutrient loading was calculated as riverine
input minus the point source input to freshwater. In this calculation we assumed that the retention of nutrients from
point sources to freshwater was negligible, because freshwater point sources generally discharge in the downstream
area. It should be recognized that certainty in the loading
compilations is likely to have increased with time. The ratio
between the diffuse nutrient loading and the freshwater discharge will hereafter be referred to as flow-weighted concentrations of TN and TP.
Atmospheric nitrogen deposition was calculated by means
of a Lagrangian model with 96-h trajectories of air parcels
to a net of receptor points having a resolution of 25 3 25
km (Hertel et al. 1995). The model was calibrated to deposition rates measured at two coastal gauges located in the
northern and southern part of the study area. Atmospheric
deposition of phosphorus has not been calculated on an annual basis, but Andersen et al. (2004) estimated it to be ;8
kg P km22 for the study area and contended that temporal
trends were unlikely. Based on these results a constant atmospheric phosphorus input of 328 3 103 kg yr21 was assumed in this study.
Measurements of nutrient concentrations (NH14 , NO22 ,
NO23 , PO32
4 , TN, and TP), collected within the framework
of DNAMAP and the Swedish national monitoring program,
were investigated in the present study. A total of 46 stations
located in estuaries and the coastal region and 27 open-water
stations (Fig. 1) were sampled with varying frequencies from
one up to 103 times per year, however unevenly distributed
both within and between years. Dissolved inorganic nitrogen
(DIN) was calculated as the sum of ammonia, nitrite, and
nitrate, whereas dissolved inorganic phosphorus (DIP) comprised phosphate only. For estuarine and coastal stations, we
calculated the average concentration of all nutrient constituents over the entire water column, whereas average concentrations of samples #10 m and samples $20 m were
used to characterize the surface and the bottom layer at openwater stations.
Yearly means of DIN, DIP, TN, and TP were computed
through use of a general linear model that described variations between stations, years, and months after log-transformation of the variables. Thus, if Yijkl described the observations of any of the four considered variables, then
Y ijkl 5 station i 3 year j 3 month k 3 « ijkl
⇑
⇓
log(Y ijkl ) 5 log(station i ) 1 log(year j ) 1 log(month k )
1 e ijkl
(1)
where eijkl was a normal distributed random error with zero
mean and variance s 2, station i described the station-specific
mean levels, yearj described the year-specific mean level,
and month k described the seasonal variation by monthly
means. It was assumed that interannual and seasonal variations were multiplicative factors to each other and to the
station-specific mean level, and that the error term on the
original scale was lognormaly distributed.
The monitoring data were not balanced, i.e., uneven number of observations for different combinations of station,
year, and month, and averaging all observations for a given
year would result in values that depended on the differences
in sampling frequencies. Comparable yearly means were calculated by computing the marginal distributions of yearj as
linear combinations of the parameter estimates in Eq. 1
(Searle et al. 1980) to account for differences in sampling
frequencies. Yearly means obtained from Eq. 1 were backtransformed to the original scale by
5
E(year j ) 5 exp E[log(year j )] 1
s2
2
6
In the following we shall refer to the back-transform of the
marginal means computed from the model in Eq. 1 as the
yearly nutrient means or levels.
Nitrogen and phosphorus concentrations were related to
freshwater discharge, and point source nitrogen and phosphorus loading, respectively, by means of multiple linear regression models using yearly means from 1989–1997. The
last 5 yr (1998–2002) were used to investigate deviations
from those established relationships that could potentially
accrue from changes in agricultural practices. Mean nutrient
concentrations based on estuarine and coastal stations were
Eutrophication trend reversal
401
Finally, the mean nutrient levels were adjusted to variations in freshwater discharge and in point source loading
(denoted DINADJ, TNADJ, DIPADJ, and TPADJ ) by means of the
multiple regressions models described above. Differences
between the yearly means and predicted values from the
multiple regressions were added to the predicted mean levels
for DIN, TN, DIP, and TP corresponding to the average
freshwater discharge over the entire period (1989–2002) and
the average point source loading for the five most recent
years (1998–2002). The adjusted nutrient levels would indicate changes in the diffuse inputs when random variations
in freshwater discharge were taken into account. These adjusted means described the nutrient level in a given year if
the loading from point sources had been low and if the freshwater discharge had an average level. Nutrient inputs, nutrient levels, and adjusted nutrient levels were analyzed for
trends by means of linear regression.
Results
Fig. 2. Annual discharges of (A) freshwater, (B) total nitrogen,
and (C) total phosphorus to the Danish straits (1989–2002). Atmospheric nitrogen and phosphorus depositions to the study area
are shown by diamonds connected with a thin line. Danish nitrogen
and phosphorus input has been partitioned into point sources (below
dashed line) and diffuse sources (above dashed line). Inputs from
the Göta River were not included in the Swedish figures.
related to input from Denmark only. In contrast, a combined
freshwater discharge from Denmark, Sweden, and Germany
was used to relate to concentrations found at open-water
stations. We used point source loading from Denmark alone
for the open-water stations, because annual figures of nutrient loading partitioned into point and diffuse sources were
not available from Sweden and Germany.
Nutrient loading—There were strong interannual variations in freshwater discharge as well as nutrient loading (Fig.
2). The total freshwater discharge varied from 12–29 km3
yr21, and interannual variations in the Danish, Swedish, and
German values were highly correlated (rDK,SE 5 0.82; rDK,GE
5 0.86; rSE,GE 5 0.64). Particularly, 1989, 1996, and 1997
were ‘‘dry’’ years, whereas the other years had freshwater
discharges .20 km3 yr21. During the entire period the freshwater discharge from Sweden was the largest (49%), followed by discharge from Denmark (40%) and Germany
(11%). However, the loading from Denmark was clearly the
largest for both total nitrogen (62%) and total phosphorus
(74%). Interannual variation in the Danish, Swedish, and
German contributions were highly correlated for nitrogen
loading (rDK,SE 5 0.87; rDK,GE 5 0.79; rSE,GE 5 0.86) and less
correlated for phosphorus loading (rDK,SE 5 0.77; rDK,GE 5
0.22; rSE,GE 5 0.68). All freshwater and nutrient discharges
from the three countries were significantly correlated (p ,
0.05), except rDK,GE for phosphorus loading (p 5 0.4552).
Interannual variations in total nitrogen loading from land
did not reflect any trends (Table 2) and were clearly linked
to freshwater discharge (Fig. 2). Over the entire study period,
the input from Danish point sources declined significantly,
comprising ;50% in the dry year of 1989 to ,10% of the
total Danish contribution in the most recent years. This corresponded to a reduction of ;20,000 3 103 kg of nitrogen.
The total phosphorus loading decreased significantly from
Table 2. Trend analysis of nutrient inputs to the Danish straits (1989–2002, n 5 14) by linear regression (F1,12). Significant trends
(103 kg yr21) at the 5% significance level are highlighted by boldface type.
Denmark
Sweden
Germany
Source
Trend
p
Trend
p
Trend
p
TN diffuse sources
TN point sources
TN total input
TP diffuse sources
TP point sources
TP total input
354
21537
21184
27
2281
2253
0.7558
,0.0001
0.3088
0.3933
,0.0001
0.0001
—
—
212
—
—
216
—
—
0.9718
—
—
0.0761
—
—
193
—
—
2
—
—
0.5919
—
—
0.7048
402
Carstensen et al.
Fig. 3. The ratio between the diffuse nutrient loading and the
freshwater discharge from Denmark, referred to as flow-weighted
TN and TP concentrations in the text.
;6000 3 103 kg yr21 in the beginning of the period to
;3000 3 103 kg yr21 after 1994. This trend was attributed
to reductions in Danish point sources from ;4500 3 103 kg
yr21 in 1989 to 600–800 3 103 kg yr21 in recent years. Point
source reductions have made the diffuse input of phosphorus
the dominating source in Denmark from 41% in 1989 to
;80% in recent years. This change in dominating inputs
from point to diffuse sources has also resulted in a gradual
covariation of phosphorus loading with the freshwater discharge. There were no significant correlations between the
freshwater discharges and point source loading data used in
the multiple regression analysis below (t-test, all p . 0.05).
Atmospheric deposition of nitrogen was relatively stable
(average of 55,000 3 103 kg yr21) over the entire study period, ranging from 47,000 3 103 kg yr21 to 65,000 3 103
kg yr21 with actually only 1 yr (1990) exceeding 60,000 3
103 kg yr21 (Fig. 2). There was no trend in atmospheric deposition (F1,12 5 3.66, p 5 0.0799), particularly if 1990 was
excluded in the linear regression (F1,11 5 1.55, p 5 0.2394).
Consequently, atmospheric nitrogen deposition was considered constant over the study period and not included in the
multiple regression. Atmospheric deposition of phosphorus
was on average ,10% of the land-based inputs and was not
included in the multiple regression for the same reason.
The ratio between nitrogen and phosphorus diffuse input
and freshwater discharge had the largest variations in the
beginning of the period when loading compilations were
considered more uncertain (Fig. 3). There were no significant
trends in the flow-weighted concentrations for either TN
(F1,12 5 2.28, p 5 0.1573) or TP (F1,12 5 0.06, p 5 0.8136)
over the entire period, but TN decreased significantly (F1,11
5 10.90, p 5 0.0071) if the first year with more uncertain
loading figures was excluded. Flow-weighted TN concentrations from diffuse sources was about 7 mg L21 in the beginning of the 1990s, decreasing to ;5.5 mg L21 in recent
years. Low levels were observed in the three dry years of
1989, 1996, and 1997. The flow-weighted TP concentration
from diffuse sources was relatively constant, ;0.11 mg L21
from 1993 and onward. There was no trend in TP levels
during this period (F1,8 5 0.95, p 5 0.3594).
Nutrient concentrations—The two wet years in 1994 and
1995 and the two dry years in 1996 and 1997 were clearly
visible in the mean nitrogen levels in estuaries and coastal
areas as well as for the open-water stations (Fig. 4A,C). DIN
levels in surface waters decreased by ;30% from the two
first years to the two last years, whereas there was no change
in bottom water concentrations. TN levels decreased by 12–
18% during the study period. However, nitrogen means during the last couple of years were almost at the same level
as in 1996 and 1997 although the freshwater discharge was
considerably higher. Trends were not significant for DIN,
whereas TN levels decreased significantly by 8 mg L21 yr21
in estuaries and coastal waters and approximately at half the
rate in open waters (Table 3).
The effect of dry and wet years on phosphorus levels was
less pronounced; the phosphorus means decreased in the beginning of the study period and were more or less stationary
in the last part of the investigated period (Fig. 4B,D). DIP
levels decreased with significant changes in surface waters
of 48–57% observed between 1989–1990 to 2001–2002 (Table 3), whereas the decline in the open-water bottom layer
was more moderate (22%). Trends in TP levels were also
significant with more similar changes using the same periods
(30–39%) for the three considered water types.
The yearly means of DIN and TN were significantly related to freshwater discharge for estuaries and coastal stations as well as for the surface and bottom layer at openwater stations (Table 4). Nitrogen loading from point sources
did not show a consistent pattern for explaining interannual
variations in nitrogen concentrations, with significant relationships observed only for DIN levels in the surface layer
of open-water stations and TN levels in the bottom layer of
open-water stations. The yearly means of DIP and TP were
clearly linked to point source loadings of phosphorus (Table
4), whereas interannual variations in freshwater discharges
did not correlate significantly. The most significant relationships were obtained for estuarine and coastal stations, all
having R 2-values .0.85.
In the last 5 yr of the study period, nitrogen did not exhibit
the same behavior with respect to freshwater discharge as
for the earlier period of 1989–1997 (Fig. 5). All the yearly
means for both DIN and TN were below the regression lines
except for the DIN level in the bottom layer at open-water
stations that had 2 yr above the regression line (1999 and
2002). Although the phosphorus point source loading from
Denmark was lower in 1998–2002 than in all the previous
years, the phosphorus levels in these years were mostly
above the extrapolation of the regression lines, particularly
for DIP in estuaries and the coastal area as well as the openwater surface layer (Fig. 6).
Nitrogen levels adjusted for variations in freshwater discharge and nitrogen point source loading showed consistent
decreasing trends (Fig. 7A,C), all significant but the adjusted
DIN means at open-water stations (Table 3). Removing the
interannual variation related to freshwater discharge improved the significance of the trends. The relative change in
adjusted DIN levels from 1989–1990 to 2001–2002 varied
from 23% in the open-water bottom layer to 14% in the
open-water surface layer and 44% in estuaries and coastal
waters. These trends were more similar for adjusted TN levels (15–18%). Adjusted phosphorus levels were generally
more variable than were adjusted nitrogen levels (Fig.
7B,D), however, without any significant trend (Table 3). For
both nitrogen and phosphorus, the highest rate of change and
Eutrophication trend reversal
403
Fig. 4. Yearly means for estuarine and coastal stations, surface layer of open-water stations, and bottom layer of open-water stations.
Error bars show the 95% confidence limits of the mean values.
the most significant trends were observed in estuaries and
coastal regions.
Discussion
In this study we have identified strong relationships between land-based discharges and nutrient concentrations in
the marine environment. This was possible owing to several
reasons. First, yearly means of nutrient concentrations with
a high precision were obtained by pooling data from a large
number of stations, assuming that all stations had the estimated interannual variation in common. Second, the investigated period had large variations in both freshwater discharges and point source phosphorus loadings, yielding a
high power for the multiple regression analysis. Third, interannual variations in freshwater discharge and point source
loading were not correlated, and the estimates resulting from
the multiple regression were consequently not biased. Finally, changes in nutrient concentrations attributed to management actions occurred at different times for nitrogen (end of
Table 3. Trend analysis of the nutrient means (1989–2002, n 5 14) and means adjusted for variations in freshwater discharge and point
source loading. Significant trends (mg L21 yr21) at the 5% significance level are highlighted by boldface type.
Estuarine and coastal
Open-water surface
Variable
Trend
p
Trend
DIN
TN
DIP
TP
DINADJ
TNADJ
DIPADJ
TPADJ
23.93
28.17
21.78
22.05
26.36
211.02
0.16
0.22
0.1084
0.0321
,0.0001
,0.0001
0.0006
0.0003
0.4023
0.2783
20.60
22.94
20.42
20.87
20.13
23.70
0.11
20.17
p
0.0723
0.0143
0.0040
,0.0001
0.4899
0.0005
0.2753
0.1970
Open-water bottom
Trend
20.22
24.55
20.70
21.15
20.34
24.04
0.23
20.08
p
0.8651
0.0134
0.0032
,0.0001
0.6851
0.0005
0.1435
0.5514
404
Carstensen et al.
Table 4. Nutrient means (1989–1997, n 5 9) related to freshwater discharge and nutrient loading from point sources by multiple
regression. Freshwater discharge included only Danish data for estuarine and coastal stations, and the contribution from Denmark, Sweden,
and Germany for open-water stations. Total nitrogen loading from Denmark was used for DIN and TN levels, and total phosphorus loading
from Denmark was used for DIP and TP levels. Significant relations (F1,6 at 5% significance level) are highlighted by boldface type.
Freshwater discharge
Point source loading
Variable
R2
Intercept
Estimate
(mg L21 km23)
p
Estimate
(mg L21 1026 kg21)
p
DIN
Estuarine and coastal stations
Open-water stations (surface)
Open-water stations (bottom)
0.9535
0.7567
0.6610
33.86
477
42.74
16.83
0.604
2.615
,0.0001
0.0320
0.0211
0.805
0.439
0.510
0.1707
0.0355
0.4491
TN
Estuarine and coastal stations
Open-water stations (surface)
Open-water stations (bottom)
0.9509
0.8784
0.9243
416.9
237.5
190.8
25.556
3.173
5.039
,0.0001
0.0007
0.0003
1.763
0.214
1.463
0.0750
0.5900
0.0289
DIP
Estuarine and coastal stations
Open-water stations (surface)
Open-water stations (bottom)
0.8701
0.6977
0.7676
12.88
4.59
18.13
20.2047
20.0350
20.0735
0.7297
0.5795
0.5294
6.77
1.84
3.25
0.0007
0.0100
0.0044
TP
Estuarine and coastal stations
Open-water stations (surface)
Open-water stations (bottom)
0.9228
0.7637
0.8650
34.34
16.40
28.62
0.3019
0.1222
0.0313
0.5844
0.4497
0.8424
8.32
2.63
3.86
0.0002
0.0061
0.0009
Fig. 5. Yearly (A) DIN and (B) TN levels for estuarine and
coastal stations, surface layer of open-water stations, and bottom
layer of open-water stations versus freshwater discharge. Freshwater
discharges related to nitrogen levels for open-water stations included contributions from Denmark, Sweden, and Germany; coastal nitrogen levels were related to Danish discharges only. Open symbols
show the recent levels (1998–2002) not included in the multiple
regression.
Fig. 6. Yearly (A) DIP and (B) TP levels for estuarine and
coastal stations, surface layer of open-water stations, and bottom
layer of open-water stations versus phosphorus point source loading
from Denmark. Open symbols show the recent levels (1998–2002)
not included in the multiple regression.
Eutrophication trend reversal
Fig. 7.
405
Nutrient means adjusted for variations in both freshwater discharge and point source loading from the regression analysis.
the 1990s) and phosphorus (beginning of the 1990s), allowing us to separate out these different sources of variation
(Conley et al. 2000)
Trends in nutrient concentrations—Nitrogen concentrations showed decreasing trends (Table 3) when variations in
freshwater discharge were accounted for, and this decline can
primarily be attributed to changes in diffuse loading. This is
also shown by decreasing flow-weighted TN concentrations
(Fig. 3), taking into account that nitrogen loading in the beginning of the period was more uncertain and that the high
value in 1998 may be a consequence of the two dry years
of 1996 and 1997 with nitrogen accumulating in the catchment.
Although flow-weighted concentrations of TN began to
decline already in 1994 (Fig. 3), the response in nitrogen
levels was not clearly identifiable until 1998 in the mean
levels adjusted for variations in freshwater discharge (Fig.
7A,C). This delayed response may be partly owing to a substantial reduction in the internal recycling of nitrogen following the two dry years. In 1994 and 1995, nitrogen levels
in the water column were, most likely, kept high by a large
internal nitrogen release from the sediments. This may potentially also have been the case in 1996 and 1997, although
these two dry years constitute one end of the scale in the
relationship between nitrogen concentrations and freshwater
discharge (Fig. 5). Coupling between nutrient loading, watercolumn production of organic matter, and recycling of nutrients from sediments occurs over time scales of several
years or less (Boynton et al. 1995). Attempts to find significant correlations between nutrient load and system level responses in estuaries often succeeds only when annual nutrient loads are combined with some fraction of the previous
year’s nutrient load (Boynton et al. 1995; Conley et al.
2000), suggesting that an internal load is important.
Phosphorus concentrations declined substantially in estuaries and coastal areas as well as in the open-waters of the
Danish straits during the beginning of the investigated period
(Fig. 4B,D). This trend was clearly linked to reductions in
point source loading, mainly through improved wastewater
treatment. Changes in diffuse phosphorus loading should in
principle, as shown for point source loading, show a similar
variation in phosphorus concentrations in the water column,
but these small-scale variations related to freshwater discharge are masked by larger fluxes such as phosphorus release during anoxic conditions from sediments. In the Baltic
Sea, for example, annual variations in phosphorus release
from sediments with variations in anoxia are over an order
of magnitude larger than annual phosphorus loading (Conley
et al. 2002a).
Flow-weighted TP concentrations in freshwater discharge
corrected for point sources have remained almost constant,
406
Fig. 8.
(B).
Carstensen et al.
The nitrogen/phosphorus molar ratio for the nutrient means (A) and nutrient means adjusted to variations in freshwater discharge
whereas flow-weighted TN concentrations have declined.
This has naturally altered the nitrogen/phosphorus ratio of
diffuse loading and consequently that of the total loading.
The increase in the DIN/DIP ratio in the beginning of the
1990s (Fig. 8) signals a change from potential nitrogen limitation in favor of phosphorus limitation. In the last 5 yr, the
DIN/DIP ratio has declined in estuaries and coastal waters,
which may potentially have led to more nitrogen limitation
in favor of phosphorus limitation. Several Danish estuaries
have shown spring phosphorus limitation switching to nitrogen limitation in early summer (Holmboe et al. 1999). There
can still be large interannual variations in nitrogen and phosphorus limitation owing to changes in freshwater discharge,
but nitrogen has become more important as the limiting nutrient over the last 5 yr (Fig. 8B). Although the decreasing
trends in DIN and DIP may have resulted in changing patterns of nutrient limitation, it should be acknowledged that
the combined effect has increased overall nutrient limitation
(Ærtebjerg et al. 2003).
Bottom water concentrations—Advective transport dominates the open-water of the Danish straits, and the origin of
the inflowing water from bottom waters in the Skagerrak
may originate from different regions of the North Sea, with
large variations in nutrient levels (Rydberg et al. 1996). The
main source of inflowing bottom water is from the central
North Sea having moderate nutrient levels and salinity of
;34, but occasionally nutrient-rich water with salinity of
;32 originating from the German Bight and carried with the
Jutland Coastal Current spills into the Kattegat (Rydberg et
al. 1996). This has happened to varying degrees in 1989,
1995, 1999, and 2002 (observed in winter–spring monitoring
data of DIN vs. salinity) (data not shown), which could explain the relatively high values of adjusted DIN levels in
these years (Fig. 7A). Another phenomenon that may influence bottom water concentrations in the Danish straits, particularly in the Kattegat, is the inflow of DIN-depleted surface water from the Skagerrak into the bottom layer
(observed in 1990 and 1997) (data not shown).
The phosphorus pool in the bottom layer is considered to
depend on local loading, water exchanges, and oxygen con-
ditions (Rasmussen et al. 2003b). Oxygen depletion is a reoccurring phenomenon in the Danish straits (Andersson and
Rydberg 1988, 1993; Babenerd 1991), and 2002 was the
worst year ever recorded, with 21% of the area having oxygen concentrations ,2 mg L21 for extensive periods (HELCOM 2003). On the other hand, oxygen conditions were
generally good in 1997 (Rasmussen et al. 2003a). These 2
yr, corresponding to the highest and lowest values for the
adjusted DIP means, demonstrate that sediment phosphorus
release during anoxic conditions increases the DIP levels in
the bottom layer in open waters (Mortimer 1941; Conley et
al. 2002a).
Nutrient levels in the open-water bottom layer were mainly determined by local inputs of both nitrogen and phosphorus, with relationships between concentrations and loads as
strong as those in the open-water surface layer (Table 3).
These interannual variations in nutrient levels cannot be explained by inflow from the central part of the North Sea,
where nutrient levels are low (OSPAR Commission 2000)
and presumably not directly influenced by land-based loading. This suggests a substantial vertical exchange of nutrients
over the pycnocline through upwelling and entrainment
(Gustafsson 2000) and remineralization of sedimenting particulate matter, particularly after the diatom spring bloom
(Josefson and Hansen 2003). This supports the idea of the
Danish straits being a marginal sea with estuarine character.
Thus, interannual variations in bottom water nutrient levels
were mainly determined by discharges from local sources,
whereas changes in Skagerrak inflow and episodes of oxygen depletion only caused minor deviations from this pattern.
Nutrient management—Over the past 2 decades, coastal
eutrophication of Danish marine waters has been a major
concern, and substantial nutrient reductions have been
achieved through national action plans, international marine
conventions, and European Union legislation (Iversen et al.
1998; Conley et al. 2000; Ærtebjerg et al. 2003). The declining trends in nutrient concentrations documented here are
to our knowledge the first successful large-scale effort to
reduce inputs from both point and diffuse sources.
Eutrophication trend reversal
407
Table 5. Nitrogen and phosphorus inputs to the aquatic environment for the baseline (mid-1980s) with reduction targets partitioned into
sectors. UWTPs, urban wastewater treatment plants; IDs, industrial discharges. Phosphorus discharges from the agricultural sector include
losses from farmyards only. For details, please see Ærtebjerg et al. (2003) and Grant and Waagepetersen (2003).
Sector
Agriculture
UWTPs
IDs
Total
Nitrogen
Phosphorus
Baseline 4 Reduction (%) 5 Target
Baseline 4 Reduction (%) 5 Target
311,000 4 152,400
18,000 4 11,400
5,000 4 3,000
334,000 4 166,800
(49)
(60)
(60)
(50)
5
5
5
5
158,600
6,600
2,000
167,200
The first national initiative was the 1985 NPo Action Plan
with a suite of measures implemented in relation to the discharge of nitrogen (N), phosphorus (P), and organic matter
(o) from agriculture and wastewater; however, specific reduction targets were not set. An event of widespread hypoxia
in the Danish straits in 1986 led to the adoption of an agenda
urging the government to reduce discharges and losses of
nitrogen (by 50%) and phosphorus (by 80%) from agriculture, municipal wastewater treatment plants, and individual
industrial discharges (Kronvang et al. 1993; Conley et al.
2002b). This strategic aim was formulated into sector-specific reduction objectives and targets for (1) discharges and
losses from agriculture, (2) discharges from municipal
wastewater treatment plants, and (3) direct discharges from
industries (Table 5).
The reduction targets for both municipal wastewater treatment plants and industries were met in 1995, whereas the
specific objectives and targets for the agricultural sector were
difficult to meet within the original time frame. The Action
Plan on Sustainable Agriculture adopted in 1991, focusing
on reduction of losses from cultivated fields, was followed
by the second Action Plan on the Aquatic Environment with
additional measures in 1998 to fulfill the requirements of the
European Union Nitrates Directive (Anonymous 1991). The
decision for the second action plan was influenced by collapse of the Mariager Fjord estuary, which went completely
anoxic in 1997 (Fallesen et al. 2000).
Total expected reductions in nitrogen root zone losses
from agriculture in 2002 were estimated at 149,000 3 103
kg, corresponding to a reduction of 48% in the most recent
assessment on the effectiveness of the measures (Grant and
Waagepetersen 2003). However, the reduction estimates are
associated with substantial uncertainty, and there are considerable time lags between changes in agriculture practice and
water quality responses (Stålnacke et al. 2003; Tomer and
Burkart 2003). These reductions would not have been
achieved if the periodic assessment for reduction compliance
had not been carried out and if the national monitoring program had not maintained focus on eutrophication. Reductions in nutrient overenrichment of coastal ecosystems will
rely on implementation of an adaptive management framework (Boesch 2002). Further reductions in diffuse nitrogen
loading may be needed, considering that Denmark has had
one of the highest area-specific nitrogen loss rates (Conley
et al. 2000) and that the Danish straits with an almost permanent stratification are prone to hypoxia.
4,400
4,470
1,250
10,120
4
4
4
4
4,000
3,250
1,050
8,300
(91)
(72)
(82)
(80)
5
5
5
5
400
1,220
1,820
3,440
Ecosystem perspectives—Decreasing nutrient concentrations is the first step toward reducing the adverse effects of
eutrophication. Danish estuaries and coastal areas have had
a long record of eutrophication symptoms from nutrient enrichment, including increased primary production (Richardson and Heilmann 1995), decreasing bottom oxygen concentrations (Andersson 1996), and loss of macrophytes
(Borum and Sand-Jensen 1991). Interannual variations in nitrogen loading are reflected in summer chlorophyll a concentration and bloom frequency in the Kattegat (Carstensen
et al. 2004) as well as primary production (Carstensen et al.
2003). Nielsen et al. (2002b) found that chlorophyll a and
water transparency were significantly related to total nitrogen concentrations in Danish estuaries and coastal waters
and, further, that the depth colonization of eelgrass and macroalgae were significantly related to water transparency
(Nielsen et al. 2002a). These small components of the large
complex ecosystem show that reduced nutrient concentrations are likely to improve the ecological status of estuaries
and coastal and open waters in Denmark through direct or
delayed responses or alternatively through threshold mechanisms (Scheffer et al. 2001). Some signs of ecosystem recovery have already been observed (Ærtebjerg et al. 2003).
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Received: 19 March 2004
Accepted: 29 April 2005
Amended: 1 May 2005
Hydrobiologia (2009) 629:1–4
DOI 10.1007/s10750-009-9758-0
EUTROPHICATION IN COASTAL ECOSYSTEMS
Eutrophication in coastal marine ecosystems: towards better
understanding and management strategies
J. H. Andersen Æ D. J. Conley
Published online: 27 April 2009
Ó The Author(s) 2009. This article is published with open access at Springerlink.com
The Second International Symposium on Research
and Management of Eutrophication in Coastal Ecosystems took place 20–23 June 2006 in Nyborg,
Denmark. The Symposium was attended by more
than 200 persons with a specific interest in eutrophication processes as well as a common interest in
science-based management and implementation of
nutrient reduction strategies. More than 120 oral
presentations were made, mostly focussing on both
science and management of nutrient enrichment and
eutrophication. The papers in this Special Issue of
Hydrobiologia are all based on presentations made at
the Symposium.
Electronic supplementary material The online version of
this article (doi:10.1007/s10750-009-9758-0) contains
supplementary material, which is available to authorized users.
Guest editors: J. H. Andersen & D. J. Conley
Eutrophication in Coastal Ecosystems: Selected papers from
the Second International Symposium on Research and
Management of Eutrophication in Coastal Ecosystems, 20–23
June 2006, Nyborg, Denmark
J. H. Andersen (&)
DHI, Agern Allé 5, 2970 Hørsholm, Denmark
e-mail: [email protected]
D. J. Conley
Department of Geology, GeoBiosphere Science Centre,
Lund University, Sölvegatan 12, 223 62 Lund, Sweden
About the symposium
The symposium focused on the following four topics:
(1) new and existing knowledge regarding coastal
eutrophication, (2) specific eutrophication issues such
as: (a) definition(s) and causes, (b) nutrient cycling
and nutrient limitation, (c) reference conditions and
(d) linkages to other pressures (climate change and
top/down control), (3) summaries of existing knowledge in relation to monitoring and modelling coastal
eutrophication and (4) adaptive environmental management strategies in relation to coastal
eutrophication.
The symposium was jointly organised by the
Danish Environmental Protection Agency (EPA), the
Swedish EPA, Fyn County and DHI Water &
Environment and received financial support from
the organising institutions. In addition, the Symposium has been kindly sponsored by: (1) Baltic Sea
2020, (2) Danish Agriculture, (3) the International
Agency for 14C Determination, (4) MARE—the
Swedish Marine Eutrophication Research Programme
and (5) the University of Southern Denmark. Further,
the symposium received support from the European
Commission’s Joint Research Centre, Hotel Nyborg
Strand and Scandinavian Airlines Systems (SAS).
The planning of the symposium was coordinated
by an Organising Committee with the overall
responsibility and a Scientific Committee which
compiled a broad programme focussing on both
science and management. A list of members of the
123
2
committees is available as supplementary online
material.
Eutrophication research and management—the
Danish connection
The symposium was a follow up to the highly
successful 1993 Symposium Nutrient Dynamics in
Coastal and Estuarine Environments, organised by
the Danish EPA in collaboration with the European
Commission, Directorate-General for Science,
Research and Development. The Symposium Proceedings were published in the journal Ophelia with
several seminal papers, for example, Duarte (1995),
Nixon (1995) and Richardson & Heilmann (1995).
There was great regional and international interest
for a follow-up symposium with a focus on both
science and management. This interest in science and
management has been stimulated by legislative
settings, particularly the EU Water Framework
Directive, in which coastal eutrophication problems
are important issues in adaptive management plans
(Anon., 2000).
During recent decades, Denmark and Sweden have
been at the forefront of research on and management
of eutrophication in coastal marine ecosystems
(Jørgensen & Richardson, 1996; Christensen et al.,
1998; Carstensen et al., 2006; Table 1), partly
because the straits between Denmark and Sweden
connecting the Baltic Sea to the North Sea are
vulnerable to nutrient enrichment. Denmark and
neighbouring countries have made substantial efforts
to improve the marine environment through nutrient
reductions both at the national level and through
decades of regional cooperation regarding the Baltic
Sea under the Helsinki Convention (www.helcom.fi)
and the North Sea through the OSPAR Convention
(www.ospar.org).
Both Denmark (Fig. 1) and Sweden have made
large reductions in the discharge of nutrients. Billions
of Euros have been spent, and they have not been
spent in vain. The point source inputs of nutrients to
the marine environment are significantly lower than
they were 20 years ago. However, these reductions
have not been sufficient to reduce the harmful effects
of eutrophication and the targets for improved
ecological status have not been reached.
123
Hydrobiologia (2009) 629:1–4
Three Danish Action Plans for the Aquatic Environment over the past two decades (Conley et al.,
2002) have resulted in significant reductions in the
loss of nutrients to the environment (Conley et al.,
2002; Carstensen et al., 2006). Point source inputs of
phosphorus have decreased by more than 80%.
Losses of nitrogen are expected to be reduced by
*50% when changes in agricultural practises that
have already been implemented result in reduced
loads to the marine environment. Figure 1 shows the
temporal trends in total nitrogen loading to the
Kattegat and Danish Straits over a 100-year period,
with a peak in total nitrogen loading in the 1980s.
Since the late 1970s, loads originating from both
point sources and diffuse sources have been declining. However, more than three decades since the first
measures were implemented and more than a decade
after the First International Danish Symposium on
eutrophication, the problems associated with eutrophication are still far from being resolved. There has
been a major development in scientific knowledge
and in the conceptual understanding of nutrient
enrichment and eutrophication in coastal waters.
New questions and challenges have emerged, especially in relation to modelling and management of
coastal eutrophication. In parallel, new legal and
management settings have emerged or will emerge in
the near future, for example, the EU Water Framework Directive and the process in relation to the
implementation of the European Marine Strategy.
Therefore, it was proposed and agreed in 2004 that a
follow-up symposium focussing on both science and
management of coastal eutrophication should be
organised for June 2006.
About this Special Issue
The 21 papers in this Special Issue are a mixture of
Research Papers, Opinion Papers and Short Notes,
which reflect the broad range of presentations at the
June 2006 symposium. Each manuscript was
reviewed by at least two independent reviewers and
by one of the guest editors. Copy editing was
conducted by Janet F. Pawlak and Carolyn Symon.
The Special Issue has received direct financial
support from the Nordic Council of Ministers via
the working group on the Sea and Atmosphere
Hydrobiologia (2009) 629:1–4
3
Table 1 Danish nutrient reduction targets sensu the Action Plans for the Aquatic Environment I, II and III. Baseline is 1987;
reductions and targets were agreed by the Danish Parliament in 1987 and subsequently adjusted in 1990, 1999 and 2004
Sector
Total nitrogen loads (tonnes)
1987
7
Reduction
Total phosphorus loads (tonnes)
%
=
Target
7
1987
Reduction
%
=
Target
Agriculture
311,000
7
152,400
49
=
158,600
4.400
7
4,000
91
=
400
UWWTPs
18,000
7
11,400
63
=
6,600
4.470
7
3,250
73
=
1,220
Industries
Total
5,000
7
3
60
=
2,000
1.250
7
1,050
84
=
200
334,000
7
166,800
50
=
167,200
10.120
7
8,300
82
=
1,820
See Carstensen et al. (2006) for details
UWWTPs: urban wastewater treatment plant effluents
140
120
Diffuse sources
Point sources
6
Total nitrogen input (10 kg)
Fig. 1 Trends in estimated
total nitrogen inputs (solid
line) from Denmark to the
Danish Straits including the
Kattegat since 1900, with 5year averages of point and
diffuse sources. Used with
the kind permission of
Jacob Carstensen, NERI;
based on Conley et al.
(2007)
100
80
60
40
20
0
1900 1910
1920
(Hav- og Luftgruppen) as well as indirect financial
support from the Danish EPA, Swedish EPA and DHI.
This Special Issue as well as others (Kononen &
Bonsdorf, 2001; Rabalais & Nixon, 2002; Bachmann
et al., 2006) demonstrate that considerable knowledge
has been generated since the First Danish Symposium
in 1993. We, as guest editors, are pleased with the
Special Issue as compiled and hope that the readers
will share this opinion.
Despite the vast knowledge and common understanding of eutrophication, some important gaps still
remain, especially with regard to regime shifts,
thresholds and multiple stressors. In addition, climate
change needs to be taken into account. A fundamental
problem that needs to be addressed is the lack of
political will to implement adequate nutrient management strategies. A broader acceptance of the need
1930
1940
1950
1960
1970
1980
1990 2000
to use the best scientific information we have (whilst
still seeking to improve knowledge = ‘‘moving
whilst improving’’) rather than wait for ‘perfection’
is recommended. Finally, it should be kept in mind
that we do not manage eutrophication as such, we
manage humans with the aim of reducing the effects
of eutrophication.
Acknowledgements Thanks are expressed to the Nordic
Council of Ministers and the organisers and sponsors of the
Second International Symposium on Research and
Management of Eutrophication in Coastal Ecosystems. The
Preface improved as a result of comments from Jacob
Carstensen and Scott W. Nixon. Special thanks go to Sif
Johansson, Jørgen Dan Petersen, Jens Brøgger Jensen, Torkil
Jønch Clausen, Henning Karup, Jørgen Magner and Morten
Søndergaard. We are indebted to the reviewers and to Janet F.
Pawlak and Carolyn Symon; this Special Issue would not have
been possible but for their helping hands.
123
4
Open Access This article is distributed under the terms of the
Creative Commons Attribution Noncommercial License which
permits any noncommercial use, distribution, and reproduction
in any medium, provided the original author(s) and source are
credited.
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to different nutrient regimes. Ophelia 41: 87–112.
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Studies, 52. American Geophysical Union, Washington,
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Kononen, K. & E. Bonsdorf (eds), 2001. Man and the Baltic
Sea. Ambio 30: 171–326 (Special Issue).
Nixon, S. W., 1995. Coastal marine eutrophication: a definition, social causes, and future concerns. Ophelia 41: 199–
219.
Rabalais, N. N. & S. W. Nixon (eds), 2002. Nutrient overenrichment in coastal waters: global patterns of cause and
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the Kattegat: past and present. Ophelia 41: 317–328.
Marine Pollution Bulletin 60 (2010) 919–924
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
Note
A simple method for confidence rating of eutrophication status classifications
Jesper H. Andersen a,*, Ciarán Murray a, Hermanni Kaartokallio b, Philip Axe c, Jarle Molvær d
a
DHI Water Environment Health, Department of Ecology and Environment, Agern Allé 5, DK-2970 Hørsholm, Denmark
Finnish Environment Institute, Marine Research Centre, Helsinki, Finland
c
Swedish Meteorological and Hydrological Institute (SMHI), Västra Frölunda, Sweden
d
Norwegian Institute for Water Research (NIVA), Oslo, Norway
b
a r t i c l e
i n f o
Keywords:
Eutrophication
Classification
Indicators
HEAT
Confidence rating
Water Framework Directive
a b s t r a c t
We report the development of a methodology for assessing confidence in ecological status classifications.
The method presented here can be considered as a secondary assessment, supporting the primary assessment of eutrophication or ecological status. The confidence assessment is based on scoring the quality of
the indicators on which the primary assessment is made. This represents a first step towards linking status classification with information regarding their accuracy and precision. Applied to an existing data set
used for assessment of eutrophication status of the Baltic Sea (including the Kattegat and Danish Straits)
we demonstrate that confidence in the assessment is Good or High in 149 out of 189 areas assessed (79%).
Contrary to our expectations, assessments of the open parts of the Baltic Sea have a higher confidence
than assessments of coastal waters. We also find that in open waters of the Baltic Sea, some biological
indicators have a higher confidence than indicators representing physical–chemical conditions. In coastal
waters, phytoplankton, submerged aquatic vegetation and indicators of physical–chemical conditions
have a higher confidence than indicators of the quality of benthic invertebrate communities. Our analyses
also show that the perceived weaknesses of eutrophication assessments are due more to Low confidence
in reference conditions and acceptable deviations, rather than in the monitoring data.
Ó 2010 Elsevier Ltd. All rights reserved.
1. Introduction
The purpose of this article is to present for the first time a methodology for rating the confidence of classifications of marine eutrophication status. The evaluation and testing of the presented
methodology for the confidence assessment has been conducted
with the dataset used in the recent HELCOM integrated thematic
assessment of eutrophication in the Baltic Sea (HELCOM, 2009).
The results provide valuable information and show a possible
way forward for the assessment of confidence in future environmental status assessments.
The three key terms discussed in the present paper are: accuracy,
precision, and confidence. If accuracy is the degree of closeness of a
measured or calculated quantity to its actual (true) value, precision
can be considered to be the degree of reproducibility of that measurement. Calculations or measurements can be accurate though
not precise, precise but not accurate, neither, or both. A measurement system or computational method is considered valid if it is
both accurate and precise. Valid methods inspire confidence. Accuracy and precision are statistical terms, when we use the word confidence on its own, it has a non-statistical meaning.
* Corresponding author. Tel.: +45 4516 9235.
E-mail address: [email protected] (J.H. Andersen).
0025-326X/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2010.03.020
Numerous assessments of eutrophication status are being produced in Europe, ranging from local assessments to national and
regional ones. Nowadays local assessments are tightly linked to
the European Water Framework Directive (WFD). Other types of
assessments can be found in EEA (2001), Ærtebjerg et al. (2003),
OSPAR (2008) and HELCOM (2009). The majority of the assessments are perceived to be based on sound science and the use of
indicators, but relatively few are based on the use of multi-metric
assessment tools. Only few classification tools in accordance with
the WFD exist (e.g. Borja et al., 2009), but most European countries
are developing or testing indicators and/or tools that can be used in
the future.
Currently, the most widely used multi-metric assessment tools
for assessing marine eutrophication in Europe are those of the OSPAR Comprehensive Procedure (see OSPAR, 2008; Claussen et al.,
2009) and the HELCOM Eutrophication Assessment Tool (HEAT)
(see HELCOM, 2009; Andersen et al., submitted for publication).
To our knowledge, only one assessment of marine eutrophication
status has included an attempt at confidence assessments: the recently published HELCOM integrated thematic assessment of
eutrophication in the Baltic Sea region (HELCOM, 2009). The main
cause of the apparent lack of this form of secondary assessment is
the limited used of multi-metric indicator-based assessment
tools.
920
J.H. Andersen et al. / Marine Pollution Bulletin 60 (2010) 919–924
2. Materials and methods
2.1. Data and information sources
This work is based on the existing data used in the HELCOM
assessment of Baltic Sea eutrophication 2001–2006 (HELCOM,
2009). Most of the monitoring data representing actual status originate from the HELCOM Cooperative Monitoring in the Baltic Marine Environment (HELCOM COMBINE) Programme although some
come from national monitoring and assessment activities. HELCOM COMBINE is a cooperative monitoring programme shared
by the Baltic countries. Andersen et al. (submitted for publication)
describes the data origin in more detail.
2.2. Primary assessment: Eutrophication status
The assessment tool used to illustrate how confidence rating
can be done is the HELCOM Eutrophication Assessment Tool
(HEAT). This is a multi-metric indicator-based tool developed for
the assessment and classification of the eutrophication status of
the entire Baltic Sea. A total of 189 areas were assessed using indicators where information on reference conditions (RefCon) and
acceptable deviation from reference conditions (AcDev) could be
combined with national monitoring data describing the actual status (AcStat) for the period 2001–2006 (Andersen et al., submitted
for publication). Using the described RefCon, AcDev, and AcStat
concepts, the basic assessment principle is:
EutroQO ðindicatorÞ ¼ RefCon AcDev
ð1Þ
where EutroQO is an ‘‘eutrophication quality objective” (or target),
RefCon is an ‘anchor’ for the assessment while AcDev is the
‘yardstick’.
For indicators which have a positive response to nutrient inputs,
the classification is determined by the following:
If AcStat < RefCon þ AcDev; then the EutroQO ðor targetÞ is met
ð2Þ
Similarly for indicators having a negative response to nutrient
inputs:
Good are considered acceptable, while Low indicates a problem related to the quality of the input parameters. We acknowledge that
the system has a degree of subjectivity since it relies on expert
judgment. Details of the scoring principles are described in Section
2.4.
A so-called Final Confidence Rating (FCR) is then calculated in
three steps: after combining the RefCon, AcDev, and AcStat scores
into an Interim Indicator confidence (see Table 1), then the quality
element interim confidence (QE-IC) is calculated by taking the
weighted arithmetic mean of the confidences of the indicators
within the quality element (QE). The Final Confidence Rating
(FCR) for a station/water body is then obtained from the arithmetic
mean of quality element interim confidences (QE-IC). In calculating
the FCR, the quality elements are weighted equally, though quality
elements not having any indicators are ignored. For example, at a
station where only two quality elements have indicators, the final
confidence is arrived at by giving a weighting of 50% to each of
these two QE’s. For each station, separate Interim Confidence ratings are also calculated for RefCon, AcDev, and AcStat (respectively
RefCon-IC, AcDev-IC, AcStat-IC) by taking the arithmetic mean of
the values for all indicators from all quality elements, without
any weighting. All HEAT spread sheets including the secondary
assessment of confidence can be found in the Electronic Supplementary material.
Experiences from the HEAT classifications have lead to a principle where the final classification of eutrophication status has to be
based on at least two, but preferably at least three QÉs, with ideally
a minimum of two indicators per QE (data not shown). This has
been incorporated in two ways. Firstly, a QE with only one indicator has its QE-IC reduced by 25%. Secondly, if the assessment is
based on only a single QE, its FCR is reduced by 50%.
The FCR has three quality classes: High (100–75%), Good (75–
50%), and Low (50–0%). This is comparable to the method used
for analysis of Data Completeness and Reliability (DCR) in the ASSETS tool developed for assessment of eutrophication status of
estuaries in the United States (NOAA, 2007). In HEAT, High and
Good confidence ratings are considered acceptable, while Low
indicates a problem related to the quality of the eutrophication
classification.
If AcStat > RefCon AcDev; then the EutroQO ðor targetÞ is met
ð3Þ
The HEAT tool integrates the elements described above and is
based on: (1) indicators representing well documented eutrophication effects with synoptic information on reference conditions
(RefCon), acceptable deviations (AcDev) and actual status (AcStat),
(2) quality elements sensu the EU Water Framework Directive (see
Anon., 2000), (3) HELCOM Ecological Objectives (see HELCOM,
2009), (4) the relative weighting of indicators within quality elements, and (5) integration of the quality elements used into a final
assessment based on the ‘One out – all out’ principle sensu the
Water Framework Directive.
The primary assessment calculated by HEAT is a classification of
eutrophication status in five classes: High, Good, Moderate, Poor,
and Bad. The EutroQO (or target) corresponds to the boundary between Good and Moderate status. Assessment results are described
in HELCOM (2009) and Andersen et al. (submitted for publication).
2.3. Secondary assessment: Confidence rating
HEAT also produces an overall confidence rating for each indicator, where scores are assigned to each RefCon, AcDev, and AcStat
value. This scoring is based on expert judgment, where the quality
of RefCon, AcDev, and AcStat is assigned to one of three classes:
High (score = 1), Good (score = 2), and Low (score = 3). High and
2.4. Scoring principles
Initially, indicators used to assess open water eutrophication
status were provisionally scored by a group of national experts
(six persons) from Finland (lead country), Denmark, Germany, Poland and Sweden. For coastal waters, a provisional scoring was
made by 2–3 national experts from each of the Baltic Sea countries.
Table 1
Scoring matrix. A score of 3 indicates the best possible classification where a ‘‘High”
quality rating has been assigned to RefCon, AcDev and to AcStat. A score of 9 indicates
the worst possible classification, where all scores are rated ‘‘Low”. The indicator
scores from 3 to 9 are converted to an indicator confidence rating ranging from 0 to
100%. The worst possible result is assigned a confidence score of 0% and the best
possible score is equivalent to a confidence score of 100%. The confidence score for
other combinations is arrived at by interpolating linearly.
P
Scores
Ind_conf
RefCon
AcDev
AcStat
1
1
1
2
2
2
3
1
1
2
2
2
3
3
1
2
2
2
3
3
3
3
4
5
6
7
8
9
100%
83%
67%
50%
33%
17%
0%
921
J.H. Andersen et al. / Marine Pollution Bulletin 60 (2010) 919–924
The type of data used for setting RefCon values, whether it be
historical data, modelling or expert judgment (HELCOM, 2009),
has implications for the scoring. In many cases, historical or modelled data are used directly, especially if the data are already published or if the methods are in line with the monitoring methods
used for AcStat. Because RefCons are commonly based on a combination of methods, e.g. (1) historical data and statistical modelling,
(2) historical data and dynamical modelling, or (3) historical data
and expert judgment, the level of confidence in the RefCon has to
be taken into account.
The sources of information on AcDevs, and hence the scoring,
differ slightly for open and coastal waters. For open waters, we
generally use +50% and 25%, but other values are used if justified
(HELCOM, 2006; OSPAR, 2008; HELCOM, 2009; Andersen et al.,
submitted for publication). For coastal waters, we use AcDevs originating from the WFD implementation process as far as possible
(e.g. Anon., 2008a, and also Andersen et al., 2004; Krause-Jensen
et al., 2005; Henriksen, 2009; Lysiak-Pastuszak et al., 2009a,b,c).
If no coast-specific AcDevs are available, we use the AcDevs derived for open waters (see HELCOM, 2006; Andersen et al., submitted for publication). These different approaches are taken into
account where scoring the indicators, since coastal AcDevs in most
cases are better justified and documented compared to those for
open waters. In addition, AcDevs larger than +50% and 25% are
in general considered outside the range of minor or slight deviations from RefCons and are therefore given a lower confidence
score.
The scoring of AcStat, where the information sources are mostly
in situ monitoring (and in a few cases also modelled data) is based
on the reliability of the observations, their spatial coverage and frequency. The confidence assessment could also consider the number of years of adequate data, QA/QC procedures, and whether
data have been reviewed and published. A non-exhaustive list of
issues to consider when scoring the quality of an indicator is pre-
sented in Table 2. The issues of concern and the questions to be
considered are closely interconnected and they can sometimes
overlap.
In the second round of the confidence assessment, the scoring
for open and coastal waters was tentatively checked by the convener of the assessment work. The results were then, as a third
round, presented, argued and finally agreed collectively to ensure
that each expert’s assessment methodology was consistent, producing a harmonized, Baltic Sea-wide assessment (see HELCOM,
2008). The final and fourth round in the assessment process
included verification by the convener and final approval by
national contact points (who were not necessarily the experts
who carried out the assessment).
3. Results and discussion
Confidence ratings were made for 189 areas of the Baltic Sea
including the Kattegat and Danish Straits. These areas consisted
of 172 coastal areas and 17 open water bodies. A summary of all
confidence ratings is presented in Table 3.
3.1. RefCon, AcDev, and AcStat
For RefCons, the average confidence score was 0.75, indicating
that the experts had High confidence in them. Open waters had
an average RefCon confidence of 0.67, while coastal waters had
an average RefCon confidence of 0.76 (Fig. 1). For AcDev, the pattern was the same: The average confidence of AcDev was 0.46,
while open waters and coastal waters had average AcDev confidence of 0.53 and 0.46, respectively. The AcStat confidence was
somewhat higher with an average AcStat confidence of 0.86. Open
waters had an average AcStat confidence of 0.96, while the AcStat
confidence for coastal waters was only 0.85.
Table 2
Key issues of concern and questions to be addressed when scoring an indicator in HEAT calculations of eutrophication status and confidence. ‘""’ indicates a large positive
influence on the score; ‘"’ indicates a positive influence; ‘;’ indicate a negative influence; and ‘;;’ indicates large negative influence.
Indicator
part
Concerns and questions
Yes
/
No
RefCon
Are reliable historical data used?
Is the method used then comparable with today monitoring methods?
Is the spatial and temporal coverage for historical data representative?
Are RefCon data modelled and justified?
Are RefCon data from so-called reference/undisturbed sites being used?a
Is the influence of expert judgment low?
Are the RefCon values published in a peer reviewed paper or report series?
"
"
"
"
;;
""
""
or
or
or
or
or
or
or
;
;
;
;
"
;;
;
AcDev
Open waters:
Is the indicator a HELCOM target indicator?b
Is the AcDev value lower than +50% (for a positive indicator response to nutrient concentrations, such as chlorophyll
concentration)?
Is the AcDev value lower than 25% (for a negative indicator response, such as Secchi depth)?
Is a functional relation established between nutrient concentrations and indicator response?
Is the influence of expert judgement low?
Is the AcDev value published in a peer reviewed paper or report series?
Coastal waters:
Does the AvDev value originate from the WFD implementation?
Is the AcDev value lower than +50%?
Is the AcDev value lower than 25%?
Is a functional relation established?
Is the degree of expert judgment low?
Is the AcDev value published in a peer reviewed paper or report series?
"
"
or
or
;
;
"
"
"
""
or
or
or
or
;
;
;
;
"
"
"
"
"
""
or
or
or
or
or
or
;
;;
;;
;
;
;
Does the indicator represent eutrophication well?
Is the indicator reported in a HELCOM indicator fact sheet?b
Is the spatial coverage adequate?
Is the temporal coverage (frequency) adequate and does it match seasonality?
Is the uncertainty known and low?
Are AcStat data published in a peer reviewed paper or report series?
"
""
"
"
""
""
or
or
or
or
or
or
;;
;
;;
;;
;
;
AcStat
a
b
RefCon sites do no longer exist within the Baltic Sea.
More information is available via http://www.helcom.fi.
922
J.H. Andersen et al. / Marine Pollution Bulletin 60 (2010) 919–924
Table 3
Rating of confidence in 14 Baltic Sea basins including the Kattegat and the Danish Straits. High and Good are considered acceptable, while Low indicate a problem related to the
quality of the eutrophication classification. Information about the eutrophication status of these basins/areas can be found in HELCOM (2009) and Andersen et al. (submitted for
publication).
Basin
High
Good
Low
Total
Bothnian Bay and the Quark
Bothnian Sea
The Archipelago and Åland Seas
Baltic Proper, northern parts
Gulf of Finland
Baltic Proper, Eastern Gotland Basin
Gulf of Riga
Western Gotland Basin
Gulf of Gdansk
Bornholm Basin
Arkona Basin
Kiel Bight and Mecklenburg Bight
Danish Straits including the sound
Kattegat
Total
0 (0%)
0 (0%)
0 (0%)
0 (0%)
0 (0%)
1 (11.1%)
0 (0%)
0 (0%)
1 (20.0%)
2 (14.3%)
0 (0%)
0 (0%)
6 (60.0%)
5 (31.3%)
15 (7.9%)
6 (55.4%)
9 (40.9%)
5 (83.3%)
32 (78.0%)
18 (90%)
7 (77.8%)
5 (83.3%)
17 (85.0%)
3 (60.0%)
9 (64.3%)
4 (100%)
5 (100%)
4 (40.0%)
10 (62.5%)
134 (70.9%)
5 (45.5%)
13 (59.1%)
1 (16.7%)
9 (22.0%)
2 (10%)
1 (11.1%)
1 (16.7%)
3 (15.0%)
1 (20.0%)
3 (21.4%)
0 (0%)
0 (0%)
0 (0%)
1 (6.3%)
40 (21.2%)
11 (100%)
22 (100%)
6 (100%)
41 (100%)
20 (100%)
9 (100%)
6 (100%)
20 (100%)
5 (100%)
14 (100%)
4 (100%)
5 (100%)
10 (100%)
16 (100%)
189 (100%)
AcStat and RefCon generally had an average High confidence.
We believe there are two reasons for this. Firstly, the monitoring
activities carried out by the Baltic Sea countries through the HELCOM COMBINE Programme generally hold a high-quality in terms
of the methods used, temporal and spatial resolution as well as QA/
QC procedures. Secondly, much effort has been put into establishing RefCon values, both for open (HELCOM, 2006) and coastal
waters (Anon., 2008a).
The Low average confidence of AcDev, which are probably more
difficult to establish than RefCon, leave room for improvements,
especially for coastal waters. Despite Baltic EU Member States having spent considerable resources on implementation of the WFD,
there seems to be an urgent need for improving the boundary setting, since this sets the target for having an acceptable or unacceptable ecological status.
concentrations, phytoplankton and benthic communities should
always be used. The benthic communities to be included in assessments of open (deep) waters will in practice be benthic invertebrates. For coastal waters, both submerged aquatic vegetation
and benthic invertebrates should be included. However, based on
this study, it appears that vegetation (0.64) is considered a more
reliable quality element than invertebrate fauna (0.30) at least in
terms of QE-IC. We assume there are two reasons for this: firstly,
the monitoring of SAV is mostly straightforward and based on
the depth limits of the dominating SAV species. Secondly, indices
currently used to assess the status of benthic invertebrate communities in coastal waters are not completely developed and perhaps
not applicable in brackish or near-limnic coastal waters such as the
Baltic Sea.
3.3. Areas
3.2. Quality elements
The QE based on phytoplankton indicators had an average confidence of 0.71, indicating Good confidence (Table 1). The indicators used for the phytoplankton quality element were mostly,
though not exclusively, based on chlorophyll-a. Assessments of
open waters and coastal waters had an average confidence of
0.74 and 0.71, respectively. Submerged aquatic vegetation (SAV)
is used only for assessment of eutrophication status of coastal
waters because of depth limitation. The average confidence for
SAV was 0.64, indicating Good confidence. For the QE on benthic
invertebrate communities, the average confidence was 0.33. For
open waters it was 0.54, whilst it was only 0.30 for coastal waters.
For the QE including physical–chemical indicators, the average
confidence was 0.54. Open waters and coastal waters had an average confidence of 0.62 and 0.53, respectively.
For open waters, the ranking of quality elements in terms of
average confidence went from phytoplankton (0.74) > physical–
chemical conditions (0.62) > benthic invertebrate communities
(0.54), all holding a Good confidence. For coastal water the picture
was different, here the rank was phytoplankton (0.71) > submerged aquatic vegetation (0.64) > physical–chemical conditions
(0.57) > benthic invertebrates (0.30), the very last being a Low confidence QE.
To say that one or more quality element(s) is superior to the
others is tempting. However, it should be kept in mind that all
high-quality assessments of eutrophication status have to be based
on a range of QEs and indicators. Causative factors, as well as primary and secondary eutrophication effects should be included in
any assessment, meaning that indicators dealing with nutrient
A total of 149 out of the 189 areas assessed were rated as having
High or Good confidence. Forty areas were rated as having Low
confidence (Table 3). In the Arkona Basin, Kiel Bight and Mecklenburg Bight as well as the Danish Straits including The Sound, none
of the areas assessed hold Low confidence.
Discouragingly, some parts of the Baltic Sea had a high proportion of areas with Low confidence ratings (see Table 3), e.g. the
Bothnian Bay (5 out of 11 or 45.5%), and the Bothnian Sea (13
out of 22 or 59.1%). The underlying reasons for this are difficult
to deduce. Meaningful explanations could be that the Gulf of Bothnia, surrounded by areas of low population density, is normally regarded as being unaffected or only slightly affected by
eutrophication. Hence, despite about 40% of the runoff to the Baltic
draining into the gulf, monitoring activities are relatively modest.
This may result in a limited number of available indicators, low
sampling frequencies, indicators with inaccurate setting of RefCons, or AcDevs that are outside the range of what is normally
interpreted as minor or slight deviation from RefCons. Combinations of any or all of these are likely to result in a low FCR.
An important result is that none of Baltic Sea basins had monitoring and assessment programs which averaged High confidence
(see Table 3).
For all areas, the rank of the confidence was as follows: the Danish Straits including The Sound (0.75) > Kattegat (0.69) > Arkona
Basin (0.67) > Kiel Bight and Mecklenburg Bight (0.64) > Western
Gotland Basin (0.61) > Gulf of Finland (0.59) > Baltic Proper
(0.58) > Bornholm Basin and Gulf of Riga (both 0.57) > Gulf of
Gdansk (0.55) > Eastern Gotland Basin (0.54) > Archipelago and
Åland Seas (0.52) > Bothnian Bay and the Quark (0.50) > Bothnian
J.H. Andersen et al. / Marine Pollution Bulletin 60 (2010) 919–924
923
Fig. 1. Confidence ratings of open waters (n = 17) and coastal waters (n = 172) in the Baltic Sea in regard to indicators, quality elements and Final Confidence Ratings (FCR).
For the indicators, confidence ratings of reference conditions (RefCon-IC), acceptable deviation (AcDev-IC) and actual status (AcStat-IC) are compared. For the quality element,
we compare the confidence rating of phytoplankton (PHY-IC), submerged aquatic vegetation (SAV-IC), benthic invertebrate communities (BIC-IC) and physical–chemical
indicators (PC-IC). The horizontal lines in regard to indicators and quality elements represent the boundary between Good and Low confidence. Please note the y-axes differ in
the FCR figures.
Sea (0.46). The monitoring activities in the majority of areas can
hardly be reduced without compromising the quality of future
eutrophication assessments. The monitoring of the Bothnian Sea,
and perhaps also some of the other areas (e.g. coastal parts of
the northern parts of the Baltic Proper, Gulf of Gdansk and the
Bornholm Basin), ought to be improved in order to improve any
future assessment of eutrophication status in these areas.
Of the 189 areas assessed, 13 were considered to be unaffected
by eutrophication (HELCOM, 2009; Andersen et al., submitted for
publication). However, only four of these were rated as having
High or Good confidence. This does not imply that the remaining
nine areas are mis-classified, but it could indicate that the quality
of these status classifications can be questioned. False positive
classifications (areas mis-classified as being affected by eutrophication) are perhaps a more worrying scenario, at least in budgetary
terms, since investment in load reductions might not be scientifically based. Here, it would be prudent to state that the costs of acting on incomplete information or knowledge are generally
believed to be significantly greater than the costs of obtaining
the information.
3.4. Countries
When looking at individual countries (data not shown), it is
clear that some countries (Denmark, Estonia, Germany) had better
coastal monitoring programs (providing data on AcStat) and indicators (providing data on RefCon and AcDev values) than others,
leading to Good confidence in most places. Finland, Lithuania, Po-
land and Sweden had acceptable programs, but there might be
areas of concern such as in the western coastal waters of the Gulf
of Bothnia (see Table 3). Poland might also have reason for concern
in some lagoons. The FCR’s made for Latvian and Russian coastal
waters indicate monitoring activities below par, mostly caused
by insufficient monitoring activities, both in terms of spatial and
temporal coverage as well as numbers of indicators.
None of the Baltic Sea countries had monitoring programs
which averaged High confidence. For coastal waters, the rank of
the countries was Denmark (0.70) > Estonia (0.65) > Germany
(0.64) > Sweden (0.57) > Finland (0.55) > Lithuania (0.54) > Poland
(0.51) > Russia (0.47) > Latvia (0.42). The programs in Denmark,
Estonia, Finland, Lithuania, Sweden and Poland cannot be reduced
without compromising the requirements of the WFD.
An interesting result was that the open waters, often being
monitored collectively by 2–3 countries per station, had a better
average confidence (0.66) than coastal waters (0.57). There are
probably several explanations. Firstly, the monitoring of open
waters has been harmonized and coordinated by HELCOM for almost three decades, leading to a good understanding of the monitoring system. Secondly, HELCOM requirements have traditionally
put more focus on the shared open waters rather than the national
coastal waters. Thirdly, cruises, even in coastal waters, are expensive and while monitoring of offshore regions has benefitted from
the cost-sharing of the cooperative HELCOM monitoring, some Baltic Sea states have had difficulty monitoring their local archipelagos. Furthermore, at least for some parameters, variability
(particularly spatially) is higher in coastal than open waters.
924
J.H. Andersen et al. / Marine Pollution Bulletin 60 (2010) 919–924
3.5. Perspectives
The reported method has several uses, the following five being
the most obvious: – it can be used to implement or strengthen
nutrient management strategies in the upstream catchment where
the primary assessment indicates that the ‘downstream’ area in
question is ‘affected by eutrophication’ and where the FCR is High
or Good, (2) it can be used to improve the scientific knowledge of
the basis on which load reductions are being decided or it can prevent mistaken investments in load reductions where the primary
assessment indicates that the area in question is an ‘area affected
by eutrophication’ but with a low FCR, (3) it can be used to improve
setting of RefCons and AcDevs, e.g. by encouraging applied research in regard to boundary/target setting, (4) it can be used to
improve monitoring activities in areas where the current (2001–
2006) confidence is low, and (5) it can be used to warn against
unsupported reductions in monitoring activities, particularly
where a reduction would change the FCR from Good to Low.
The approach could also be useful in regard to assessments
according to the Water Framework Directive (Anon., 2000) and
the Marine Strategy Framework Directive (Anon., 2008b), since
EU Member States already are required not only to assess ecological/environmental status of transitional and coastal marine waters,
but also provide information about levels of confidence and precision in assessments (Irvine, 2004).
The guiding principle for the use of the scoring results and confidence estimates in managing and developing marine environmental assessments should be ‘‘To strengthen and/or maintain that
which already has an appropriate quality – Improve what does not
have a good quality”. By doing so, we are likely to, in the longer
term, end up with better monitoring programmes, better indicators, better environmental targets, better assessments, and ultimately better science-based nutrient management strategies.
4. Conclusions
We have shown that confidence rating of marine eutrophication
(and other status assessments) can be performed using a practical,
indirect and non-statistical approach. The approach should at this
stage be seen as a first step only, paving the way for more sophisticated and formal methodologies. We do not yet intend to promote the suggested methodology, but do stress the need for
confidence rating eutrophication status classifications. However,
we consider the methodology reported here to be a useful tool in
regard to: (1) indirect assessment of the accuracy and precision
of the primary assessment of eutrophication status, (2) future revision of marine monitoring networks as well as (3) implementation
of science-based nutrient management strategies, especially those
pursuant to the EU Water Framework Directive and the EU Marine
Strategy Framework Directive.
Acknowledgements
This study is funded by Nordic Council of Ministers (CONFIRM).
JHA has been partly funded by HELCOM (via the EUTRO-PRO project). We would like to thank the following colleagues in the HELCOM EUTRO-PRO project: Mats Blomkvist, Ulrich Claussen, Vivi
Fleming-Lehtinen, Pirkko Kauppila, Aiste Kubiliute, Elisabeth Lysiak-Pastuszak, Georg Martin, Günther Nausch, Alf Norkko and
Anna Villnäs for providing the data on which this study is based.
Special thanks to the DHI-NTU Water and Environment Research
Centre in Singapore and to Uwe Brockmann, David Connor, Maria
Laamanen and Flemming Møhlenberg for stimulating discussions.
Appendix A. Supplementary data
Supplementary data associated with this article can be found, in
the online version, at doi:10.1016/j.marpolbul.2010.03.020.
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Ærtebjerg, G., Andersen, J.H., Hansen, O.S., (Eds.), 2003. Nutrients and
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HELCOM, 2006. Development of tools for assessment of eutrophication in the Baltic
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HELCOM, 2008. Minutes of the eight meeting of HELCOM project to elaborate the
HELCOM Baltic Sea-wide integrated thematic assessment on eutrophication
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HELCOM, 2009. Eutrophication in the Baltic Sea – an integrated thematic
assessment of eutrophication in the Baltic Sea region. Baltic Sea
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Henriksen, P., 2009. Reference conditions for phytoplankton at Danish Water
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Irvine, K., 2004. Classifying ecological status under the European Water Framework
Directive: the need for monitoring to account for natural variability. Aquatic
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2.2 Eutrophication
Eutrophication has its roots in Greek: ‘eu’ meaning ‘well’
and ‘trope’ meaning ‘nourished’, but the translation
trivializes the impact of this very serious and expensive
ecological syndrome gripping the Baltic. Algal blooms,
turbid waters, loss of submerged aquatic vegetation,
and dead zones spreading on the sea floor – the consequences of nutrient inputs and nutrient enrichment in
the Baltic are manifold. They have changed the structure
and functioning of the marine ecosystem and continue
to impair our uses of the ecosystem services.
Eutrophication is triggered by excessive amounts of nutrients washed into the sea. Although nutrient chemicals
are themselves harmless, in large quantities they cause
eutrophication. The nutrients come from our farmlands,
homes and gardens, cars, cities and industries. In the
sea, the nutrients first foster the production of planktonic algae forming algal blooms, which in the worst
case are so large and dense that they are visible even to
astronauts in space.
This increased production of organic matter often has
secondary and drastic negative consequences: the water
becomes murkier and less transparent, the sedimentation of organic material to the sea floor increases,
decomposition of organic matter increases and oxygen is
consumed, thus depleting the bottom waters of oxygen.
Benthic communities such as meadows of submerged
aquatic vegetation are deprived of light, and benthic
invertebrate communities and fish are affected by
oxygen depletion, ultimately suffocating (Fig. 2.3).
Over the years, HELCOM has put considerable efforts
into monitoring and assessment of the eutrophication
status of the Baltic Sea. Special focus has been on indicators in the following groups: (1) phytoplankton, (2)
submerged aquatic vegetation, (3) benthic invertebrates,
and (4) supporting features, e.g., nutrient concentrations
and water transparency. HELCOM has also focused on the
development of tools for the assessment of eutrophication status (HELCOM 2006, 2009a). Combining indicators
into a final classification of ‘areas unaffected by eutrophication’ and ‘areas affected by eutrophication’ is carried
out using the HELCOM Eutrophication Assessment Tool
(HEAT, see Section 1.6, HELCOM (2009a) and Andersen
et al. (2010b) for details). HEAT calculates the integrated
classification of ‘eutrophication status’. HEAT also calculates a secondary assessment of the confidence in the
eutrophication assessment.
To determine the current status of eutrophication in the
Baltic marine ecosystem, the conditions at 17 open-water
areas and 172 coastal areas were assessed using data collected between 2001 and 2006.
All open waters in the basins of the Baltic Sea, including
the open parts of the Bothnian Sea, were found to be
‘affected by eutrophication’. The only open-water areas
‘not affected by eutrophication’ included the open waters
of the Bothnian Bay and the Swedish parts of the northeastern Kattegat, the latter being renewed by oxygen-rich
Atlantic waters (Fig. 2.4). The open parts of the Bothnian
Sea were labelled ‘affected’ due to increased chlorophylla concentrations (see HELCOM 2009a for details).
In most of the coastal waters, nutrient concentrations and
chlorophyll-a concentrations generally are elevated compared to both target values and reference conditions. In
most open basins, mussels, clams, crustaceans and other
invertebrates living at the sea floor are outside the range
of what is considered as being in a ‘good status’.
Only 11 out of 172 coastal areas were found to be ‘unaffected by eutrophication’; all of these were located in the
Gulf of Bothnia. Outside the Gulf of Bothnia, not a single
coastal area in the Baltic achieved this status. Thus, all
161 coastal areas assessed outside the Gulf of Bothnia
received the classification ‘affected by eutrophication’.
The impaired conditions included elevated levels of nutrients and chlorophyll-a, loss of submerged aquatic vegetation, as well as periods of oxygen depletion particularly
affecting benthic invertebrates.
16
Figure 2.3 Conceptual model of eutrophication. The arrows indicate the interactions between different ecological
compartments. A balanced coastal ecosystem in the Baltic Sea is supposedly characterized by: (1) a short pelagic food
chain (phytoplankton > zooplankton > small fish > large fish), (2) natural species composition of plankton and benthic
organisms, and (3) a natural distribution of submerged aquatic vegetation. Nutrient enrichment results in changes
in the structure and function of marine ecosystems, as indicated with bold lines. Dashed lines indicate the release of
hydrogen sulfide (H2S) and phosphorus, which both occur under conditions of oxygen depletion. Abbreviations: N =
nitrogen; P = phosphorus; Si = silicon; DIN = dissolved inorganic nitrogen; DIP = dissolved inorganic phosphorus.
The accuracy of the classification results was generally
good, although there is some room for improvement. This
has been documented indirectly by the rating of confidence, in which the data on which the classification was
based was scored in terms of accuracy. 145 of 189 areas
had an acceptable confidence level, while the remaining
44 areas had low confidence. Low confidence is generally
a consequence of mediocre monitoring activities or the use
of too few or low quality indicators or targets. The interim
assessment of confidence is summarized in Figure 2.4,
Panel C. The areas with low confidence are generally found
in the southeastern or northern parts of the Baltic Sea.
Assessing the eutrophication status in an integrated
manner for the whole Baltic Sea provides a good basis
for evaluating the effectiveness of the implementation
of the eutrophication segment of the HELCOM Baltic Sea
Action Plan. The assessment clearly documents that nutrient inputs need to be further reduced, even though the
Baltic Sea countries have successfully reduced nutrient
inputs to a certain degree (see Section 3.1.7 and HELCOM
2009a). The eutrophication status of the Baltic Sea will
only improve if inputs of both nitrogen and phosphorus are
significantly further reduced (Conley et al. 2009b, HELCOM
2009a).
The limited water exchange with the North Sea and the
long residence time of water are the main reasons for
the sensitivity of the Baltic Sea to eutrophication. High
nutrient loads in combination with a long residence time
means that nutrients discharged to the sea will remain in
the basin for a long time. In addition, the vertical stratification of the water masses increases the vulnerability
of the Baltic Sea to eutrophication. The most important
effect of stratification in terms of eutrophication is that
it hinders or prevents ventilation and oxygenation of the
bottom waters and sediments by vertical mixing of the
water, a situation that often leads to oxygen depletion.
Furthermore, hypoxia and anoxia worsen the situation
by affecting nutrient transformation processes, such as
nitrification and denitrification, as well as the capacity
of the sediments to bind phosphorus. In the absence of
oxygen, reduced sediments release significant quantities
of phosphorus to the overlying water.
Large parts of the Baltic marine ecosystem are trapped
in a vicious circle that encourages algal blooms,
although the inputs of nitrogen and phosphorus to the
sea have been reduced in significant amounts since
the late 1980s. In fact, the widespread anoxia which
facilitates the release of phosphorus from the sea floor
sediments fuels the growth and blooms of certain
planktonic algae that are capable of utilizing dissolved
nitrogen (N2) gas. These algae, termed nitrogen (N2)
fixing blue-green algae or cyanobacteria, are capable
of fixing nitrogen dissolved in the surface layers, thus
transforming it into a form that can be used by other
organisms. Large quantities of nitrogen compounds
available for the growth of other planktonic algae are
introduced to the ecosystem by cyanobacteria especially during their bloom period in the late summer. This
state is sometimes called a state of repressed recovery
(Vahtera et al. 2007).
Panel A
Panel B
Bothnian Bay (11)
Bothnian Sea (22)
Archipelago and Åland Seas (6)
Northern Baltic Proper (41)
Gulf of Finland (20)
Western Gotland Basin (9)
Eastern Gotland Basin (6)
Gulf of Riga (20)
Gulf of Gdansk (5)
Bornholm Basin (14)
Arkona Basin (4)
Kiel Bight and Mecklenburg Bight (5)
Belt Sea (10)
Kattegat (16)
0%
good
20 %
moderate
40 %
poor
60 %
80 %
100 %
bad
Panel C
Bothnian Bay (11)
Bothnian Sea (22)
Archipelago and Åland Seas (6)
Northern Baltic Proper (41)
Gulf of Finland (20)
Western Gotland Basin (9)
Eastern Gotland Basin (6)
Gulf of Riga (20)
Gulf of Gdansk (5)
Bornholm Basin (14)
Arkona Basin (4)
Kiel Bight and Mecklenburg Bight (5)
Belt Sea (10)
Kattegat (16)
0%
high
Figure 2.4 Panel A: Integrated classification of
eutrophication status in the Baltic Sea (see Fig. 2.2 for
an explanation of the interpolation method). Areas
in green represent ‘areas unaffected by eutrophication’, while areas in yellow, orange and red represent
‘areas affected by eutrophication’, from Andersen et
al. (2010a), based on HELCOM (2009a). Large circles
represent assessment sites in open basins and small
circles represent coastal assessment sites. Panel B:
Summary of the integrated classifications of ‘eutrophication status’ presented as the proportion of assessment units per sub-basin, from Andersen et al. (2010a),
based on HELCOM (2009a). The colour key is same as
in Panel A. Panel C: Interim confidence ratings of the
20 %
acceptable
40 %
60 %
80 %
100 %
low
eutrophication classifications presented as the proportion of assessment units per sub-basin. Colours: blue
represents high confidence, green represents acceptable confidence and red represents a low and hence
unacceptable confidence, from Andersen et al. (2010b),
based on HELCOM (2009a).
17
Biogeochemistry (2011) 106:137–156
DOI 10.1007/s10533-010-9508-4
Getting the measure of eutrophication in the Baltic Sea:
towards improved assessment principles and methods
Jesper H. Andersen • Philip Axe • Hermanni Backer • Jacob Carstensen •
Ulrich Claussen • Vivi Fleming-Lehtinen • Marko Järvinen • Hermanni Kaartokallio •
Seppo Knuuttila • Samuli Korpinen • Aiste Kubiliute • Maria Laamanen •
Elzbieta Lysiak-Pastuszak • Georg Martin • Ciarán Murray • Flemming Møhlenberg •
Günther Nausch • Alf Norkko • Anna Villnäs
Received: 2 July 2009 / Accepted: 2 July 2010 / Published online: 21 July 2010
The Author(s) 2010. This article is published with open access at Springerlink.com
Abstract The eutrophication status of the entire
Baltic Sea is classified using a multi-metric indicatorbased assessment tool. A total of 189 areas are
assessed using indicators where information on
reference conditions (RefCon), and acceptable deviation (AcDev) from reference condition could be
combined with national monitoring data from the
period 2001–2006. Most areas (176) are classified as
‘affected by eutrophication’ and only two open water
areas and 11 coastal areas are classified as ‘unaffected by eutrophication’. The classification is made
by application of the recently developed HELCOM
Eutrophication Assessment Tool (HEAT), which is
described in this paper. The use of harmonized
J. H. Andersen (&) C. Murray F. Møhlenberg
DHI, Agern Allé 5, 2970 Hørsholm, Denmark
e-mail: [email protected]
A. Kubiliute
Center of Marine Research, Taikos Av. 26,
91149 Klaipeda, Lithuania
P. Axe
SMHI, Nya Varvet 31, 42671 Västra Frölunda, Sweden
E. Lysiak-Pastuszak
IMGW, Maritime Branch, Waszyngtona 42,
81-342 Gdynia, Poland
H. Backer S. Korpinen M. Laamanen
HELCOM, Katajanokanlaituri 6B, 00160 Helsinki,
Finland
J. Carstensen
National Environmental Research Institute (NERI),
Frederiksborgvej 399, 4000 Roskilde, Denmark
U. Claussen
Federal Environment Agency (UBA), Wörlitzer Platz 1,
06844 Dessau-Roßlau, Germany
M. Järvinen
Finnish Environment Institute (SYKE), Jyväskyla Office,
P.O. Box 35, 40014 Jyväskyla, Finland
G. Martin
Estonian Marine Institute, University of Tartu, Mäealuse
10a, 12618 Tallinn, Estonia
G. Nausch
Leibniz Institute for Baltic Sea Research, Seestr. 15,
18119 Rostock, Germany
A. Norkko
Department of Marine Ecology – Kristineberg, University
of Gothenburg, Kristineberg 566, 45034 Fiskebackskil,
Sweden
V. Fleming-Lehtinen H. Kaartokallio S. Knuuttila A. Norkko A. Villnäs
Marine Research Centre, Finnish Environment Institute
(SYKE), P.O. Box 140, 00251 Helsinki, Finland
123
138
assessment principles and the HEAT tool allows for
direct comparisons between different parts of the
Baltic Sea despite variations in monitoring activities.
The impaired status of 176 areas is directly related to
nutrient enrichment and elevated loads from
upstream catchments. Baltic Sea States have implemented nutrient management strategies since years
which have reduced nutrient inputs. However, eutrophication is still a major problem for large parts of the
Baltic Sea. The 2007 Baltic Sea Action Plan is
projected to further reduce nutrient inputs aiming for
a Baltic Sea unaffected by eutrophication by 2021.
Keywords Eutrophication Baltic Sea Assessment HEAT Nutrients Ecological status Nutrient management strategies
Introduction
Nutrient enrichment, leading to large scale eutrophication problems in the Baltic Sea, is perhaps the single
greatest threat to the Baltic Sea environment (HELCOM
2009). Nutrient enrichment results in an increase in
productivity and undesirable changes in ecosystem
structure and function (Ryther and Dunstan 1971; Nixon
1995; Cloern 2001). The Baltic Sea ecosystem can
cope with moderate increases in eutrophication pressure, but when the limits of ‘normal’ ecosystem structure
and function are exceeded, eutrophication becomes a
problem (Ærtebjerg et al. 2003; Rönnberg and
Bonsdorff 2004; Feistel et al. 2008; HELCOM 2009).
The 2007 Baltic Sea Action Plan (BSAP),
prepared under the Convention for the Protection of
the Baltic Sea Environment, identifies eutrophication
as one of the four main issues to address in order to
improve the environmental health of the Baltic Sea
(HELCOM 2007a). The BSAP sets a strategic goal
related to eutrophication: ‘a Baltic Sea unaffected by
eutrophication’. This is linked to a set of Ecological
Objectives, which correspond to good ecological/
environmental status sensu the European Water
Framework Directive (WFD) and Marine Strategy
Framework Directive (MSFD) (Anon. 2000, 2008a,
b). The ecological objectives associated with eutrophication are: (i) concentrations of nutrients close to
natural levels, (ii) natural levels of algal blooms, (iii)
clear water, (iv) natural distribution and occurrence
of plants and animals, and (v) natural oxygen levels.
123
Biogeochemistry (2011) 106:137–156
In the BSAP, the Baltic Sea states acknowledge
that a harmonized approach to assessing the eutrophication status of the Baltic Sea is required.
Therefore, the Baltic Sea states performed a Baltic
Sea-wide thematic assessment of eutrophication status including development of a tool for integrated
assessment, the HELCOM Eutrophication Assessment Tool (HEAT). Hence, this article describes the
principles and methods of the HEAT tool.
HEAT builds on the OSPAR Common Procedure
developed for assessment and identification of ‘eutrophication problem areas’ in the OSPAR convention
area, in particular the North Sea, the Channel, the
Skagerrak and the Kattegat (see OSPAR 2003, 2008).
It also makes use of some of the key assessment
principles of the WFD, e.g. the calculation of an
Ecological Quality Ratio (EQR) and the ‘one out, all
out’ principle (Anon. 2000; Borja et al. 2009). HEAT
arrives at a primary classification of ‘areas affected by
eutrophication’. In addition, HEAT results in a
secondary assessment of the confidence of the primary
assessment, a feature missing in other eutrophication
assessment tools (Andersen et al. 2010). This study
presents the principles and mechanics of the assessment tool and its results when applied to the Baltic Sea.
Methodology
Study area
The Baltic Sea is an inland sea with a surface area of
415,200 km2 and is one of the largest brackish-water
basins in the world. It is commonly divided into several
sub-basins separated by sills, including a transition
zone to the North Sea consisting of the Kattegat and the
Belt Sea (#11–17 in Fig. 1). These sub-areas differ
considerably in several physical characteristics including ice cover, temperature, salinity, and residence time
of the water (Leppäranta and Myrberg 2009). Surface
salinity provides an illustrative example: while it is
normally 20–25 in the Kattegat area, it is only 6–8 in
the central Baltic Sea and drops below 2 in the northern
and eastern extremities of the Bothnian Bay and the
Gulf of Finland. As a result the composition of the biota
changes considerably along these gradients (HELCOM 2007b; Feistel et al. 2008).
The human population in the catchment is 85
million, and human activities display a similar,
Biogeochemistry (2011) 106:137–156
139
Fig. 1 The Baltic Sea with
location of ‘assessment
units’ in coastal waters (172
units marked with open
circles) and open basins (17
units shown with numbered
circles). Numbers refer to
Table 1. Reproduced with
permission from HELCOM
distinctive north–south, east–west pattern. Population
density outside main cities varies from more than 100
persons per km2 in the southern and south-western
parts to less than 1 person per km2 in the northern and
north-eastern parts of the catchment area (CIESN &
CIAT 2005). In terms of land use there is a high
proportion of agricultural land in the south-eastern
and south-western parts, while boreal forest, wetlands
and barren areas dominate in the north (Anon. 2001).
The long residence times (Leppäranta and Myrberg
2009) and the strong saline stratification of the water
column, including natural hypoxia in the deep basins
(Conley et al. 2009a), make large parts of the Baltic Sea
sensitive to nutrient enrichment and eutrophication.
Human activities and settlement, including e.g.
agriculture, urban and industrial waste water, energy
production and transport result in greatly increased
loads of nutrients (nitrogen and phosphorus) from the
(relatively large) 1,700,000 km2 catchment area entering the Baltic Sea (HELCOM 2004; Schernewski and
Neumann 2005; Savchuk et al. 2008; HELCOM 2009).
Data sources
Three types of data are used in this study: (1)
monitoring data for 2001–2006 (in some cases only
2001–2005 or 2001–2004), (2) information on reference conditions (RefCon), and (3) ‘target levels’
defined as acceptable deviation (AcDev) from
RefCon.
123
140
Most of the monitoring data representing actual
status (AcStat) originate from the HELCOM Cooperative Monitoring in the Baltic Marine Environment
Programme (HELCOM COMBINE, see HELCOM
(2008) for details and note that the Kattegat is
included under both HELCOM and OSPAR) carried
out in cooperation between the Baltic countries, and
partly from national monitoring and assessment
activities (e.g. Svendsen et al. 2005; OSPAR 2008).
Data representing long-term trends in inputs of
nutrients (nitrogen and phosphorus) to the Baltic Sea
are derived from the HELCOM Fifth Pollution Load
Compilation (HELCOM 2010). All measurements
and analytical methods used as well as quality
assurance procedures are described in details in the
HELCOM COMBINE Manual, Parts A, B and C
(HELCOM 2008).
In this study, specific focus has been placed
on indicators relevant to HELCOM objectives
(HELCOM 2007b; Backer and Leppänen 2008), in
particular nutrients (objective i), chlorophyll-a (objective ii), water transparency (objective iii), benthic
invertebrates and submerged aquatic vegetation
(SAV) (objective iv). For the description of the AcStat
all Baltic Sea states have used the 2001–2006 period,
except Denmark, which used the period 2001–2005 for
the Kattegat and Great Belt and 2001–2004 for all
other areas.
RefCon
RefCon, which are ‘‘… a description of the biological
quality elements that exist, or would exist, at high
status, that is, with no, or very minor disturbance
from human activities’’ (Anon. 2000) are used to
quantify the degree of disturbance observed in the
environment. Furthermore, they should represent part
of nature0 s continuum and must reflect variability.
Three principles for making the concept of RefCon
operational are (1) reference sites, (2) historical data,
and (3) modelling. Expert judgement can be used as a
supplement when spatially based (option 1 and 2),
modelled (option 3) or combinations of 1, 2 and 3 are
not possible. In this study, the RefCon are mostly
based on historical data and modelling, since reference sites no longer exist in the Baltic Sea and the use
of expert judgement is occasionally less transparent.
The RefCon’s for nutrients (dissolved inorganic
nitrogen (DIN) and dissolved inorganic phosphorus
123
Biogeochemistry (2011) 106:137–156
Fig. 2 Reference conditions (RefCon) for open areas of the c
Baltic Sea. Numbers refer to Fig. 1. For DIN and DIP grey
bars are winter mean RefCon’s and black bars are winter
maximum RefCon’s. Please note that no data on DIP are
available for area #4, no data on Secchi depth are available for
areas #12, 13, and 16 and no data on benthic invertebrates are
available for areas #4, 6, and 11–17
(DIP)), chlorophyll-a, water transparency (Secchi
depth) and benthic invertebrates in the open parts of
the Baltic Sea, obtained from various sources
described below, are shown in Fig. 2.
For nutrients, chlorophyll-a and water transparency, RefCon’s are basin specific and mostly based
on historical data (HELCOM 2006; Fleming-Lehtinen 2007; Fleming-Lehtinen et al. 2008; Henriksen
2009). Modelled and site-specific RefCon’s have
been used for parts of the Danish Straits (OSPAR
2008). The reference values used are largely in line
with those presented by other sources, e.g. Sanden
and Håkansson (1996), Aarup (2002) and Schernewski and Neumann (2005).
RefCon’s for benthic invertebrate diversity in open
water basins, measured as gamma diversity, i.e. the
average number of species in a sub-basin per year, were
calculated based upon data from 1965 to 2006
(HELCOM 2009). RefCon’s varied by an order of magnitude between the Arkona Basin and the Bothnian Bay
due to the salinity gradient, which constrains species
distributions (Bonsdorff and Pearsson 1999). For the
coastal water assessments, different national indices
have been used; see HELCOM (2009) for details.
For SAV in coastal waters, namely depth distribution of Fucus vesiculosus and Zostera marina,
which constitute monitoring species in coastal waters
only, RefCon’s are based on historical records, e.g.
Reinke (1889), Waern (1952) and von Wachenfeldt
(1975) as well as Boström et al. (2003), Martin
(1999), and Krause-Jensen et al. (2003).
AcDev
For the open basins of the Baltic Sea, AcDev values are
set basin-wise for each indicator. Two different
principles are used for setting the AcDev, according
to whether indicators show a positive response
(increasing in value) to increases in nutrient inputs or
a negative response (decreasing in value). For an
indicator showing positive response (e.g. nutrient
concentrations and chlorophyll-a), AcDev has an
Biogeochemistry (2011) 106:137–156
141
123
142
Biogeochemistry (2011) 106:137–156
The methodology used in this study to assess
eutrophication status of a water body, the HEAT, is
based on indicators, grouped according to a predefined manner. The grouping method used follows the
WFD (Anon. 2000, 2005) quality elements (physical–
chemical features, phytoplankton, SAV, benthic
invertebrates) corresponding to HELCOM eutrophication objectives i, ii, iii (physical–chemical features), iv (phytoplankton) and v (SAV & benthic
invertebrates); subsequently combined into a final
classification of ‘eutrophication status’.
Using the described RefCon, AcDev and AcStat
concepts, the basic assessment principle becomes:
RefCon ± AcDev = EutroQO, where the latter is a
‘‘eutrophication quality objective’’ (or target) corresponding to the boundary between good and moderate ecological status. When the AcStat data exceed
the EutroQO or target, the areas in question is
regarded as ‘affected by eutrophication’’ cf. the
BSAP.
Thus, following the basis assessment principle
described above, a selection of indicators with
RefCon and AcDev values turns qualitative goals
like HELCOM’s five eutrophication objectives into
operational targets, on which objective and transparent assessments of eutrophication status can be based.
While the RefCon’s can be considered the ‘‘anchors’’
of the assessment, AcDev’s from RefCon’s are the
necessary ‘‘yardsticks’’ while AcStat is actual indicator status. The assessment principles used by
HEAT are summarised in Fig. 3.
The HEAT tool integrates all the elements
described above and is based on: (1) Indicators
representing well documented eutrophication effects
with synoptic information on RefCon, AcDevs, and
AcStat, (2) Quality Elements sensu the WFD, (3)
Fig. 3 Illustration of the key assessment principles used in the
HEAT tool. Please note that HEAT combines the principles of
the HELCOM Baltic Sea Action Plan (right side of the figure
representing open waters) with principles from the EU Water
Framework Directive (left side of the figure representing
coastal waters). Fish by courtesy of Peter Pollard, Scottish EPA
upper limit of ?50% deviation from RefCon
(HELCOM 2009). Setting AcDev to 50% implies that
low levels of disturbance (defined as less than ?50%
deviation) resulting from human activity are considered acceptable while moderate (i.e. greater than
?50%) deviations are not considered acceptable for
the body of water in question. However, in exceptional
cases the ?50% AcDev can be exceeded if scientifically justified. For indicators responding negatively to
increases in nutrient input (e.g. Secchi depth and depth
limit of SAV) the AcDev’s have in principle a limit of
-25% (HELCOM 2009), although AcDev’s used for
benthic invertebrates are slightly greater in magnitude,
ranging from -27 to -40% (HELCOM 2009).
Whereas an indicator with positive response can
theoretically show unlimited deviation, indicators
showing negative response have a maximum deviation
of -100% and a deviation of -25% is, in most cases,
interpreted as the boundary between low and moderate
levels of disturbance. These ?50% and -25% ‘‘principles’’ are under discussion, but these initial and
pragmatic values are in accordance with the WFD
(Anon. 2000, 2005) and other eutrophication assessment approaches (Bricker et al. 2003; HELCOM 2006;
NOAA 2007; OSPAR 2008; Bricker et al. 2008;
Claussen et al. 2009). The AcDev’s used for the coastal
waters are largely defined by the WFD implementation
process, in particular the WFD intercalibration activity
in the Baltic Sea (Anon. 2008b).
Assessment principles and methods
123
Biogeochemistry (2011) 106:137–156
143
HELCOM Ecological Objectives, (4) weighting of
indicators within quality elements, and (5) integration
of the Quality Elements used into a final assessment
based on the ‘One out—all out’ principle sensu the
WFD.
Step 1: Indicators and boundary setting
The EQR is a dimensionless measure of the observed
value (AcStat) of an indicator compared with the
reference value (RefCon). The ratio is equal to 1.00 if
AcStat is better than or equal to RefCon and
approaches 0.00 as deviation from RefCon becomes
large.
Step 1A: Indicators with a positive numerical
relationship to nutrient input For an indicator
showing positive response to nutrient input, the
EQR is defined by:
EQR ¼ RefCon=AcStat
ð1Þ
0 EQR 1
ð2Þ
where the observed value of the indicator (AcStat) is
equal to or less than the reference value, then the
EQR is equal to the maximum achievable, 1.00. For a
given reference value, increasing values of AcStat
give lower EQR, with EQR approaching zero as the
status value becomes infinitely large (Fig. 4a).
The value of EQR is used to assign a quality class
to the observed status. The classes in descending
order of quality are RefCon, High, Good, Moderate,
Poor, Bad. The central definition of the quality
classes is given by the value of AcDev. The boundary
between Good and Moderate status is defined as
being where the deviation from RefCon is equal to
the AcDev. That is:
AcStat ¼ ð1 þ AcDevÞ RefCon
ð3Þ
Substituting for AcStat in (1) gives:
EQRGood=Moderate ¼ 1=ð1 þ AcDevÞ
ð4Þ
The EQR boundary between High and Reference
status is always set equal to 0.95. If EQR is above
0.95, it is implicitly assumed that the indicator has a
status equal to RefCon. This deviation is allowed in
order to take into account a degree of uncertainty in
the observations of RefCon and present status as well.
Thus, this permissible deviation from RefCon (5%)
represents a generic estimate of the uncertainty
margin for all indicators. The quality class of
‘‘Reference’’ will rarely be used and quality class
‘‘High’’ therefore, in practice, represents the highest
Fig. 4 Illustration of the boundary (target) setting, when the indicator responds numerically positive to nutrient loads and enrichment
(a) and when the indicator responds numerically negative (b)
123
144
Biogeochemistry (2011) 106:137–156
achievable status. However, the High/Ref boundary is
employed in determining boundaries between the
other classes.
The values for the boundary between Reference/
High status and the boundary between Good/Moderate status constitute fixed points from which the
remaining boundary values are calculated. For
practical reasons the span of the two highest classes
and the next two classes have equal width, i.e.:
as an EQR cannot be negative, irrespective of the
extent to which the observed status exceeds RefCon.
EQRRefCon=High EQRGood=Moderate
¼ EQRGood=Moderate EQRPoor=Bad
0 EQR 1
ð5Þ
That is, the difference between the values of EQR
defining the Reference/High and Good/Moderate
boundaries is equal to the difference between the
Good/Moderate and Poor/Bad boundary values. This
Eq. 10 can be rearranged to give the value for the
boundary between Poor and Bad status:
EQRPoor=Bad ¼ 2EQRGood=Moderate EQRRefCon=High
ð6Þ
For example, consider a case where the AcDev
from RefCon is 50%. The boundary between Good
and Moderate status is 1/(1 ? 0.5) = 0.667. And
according to (6), the boundary between Poor and Bad
status lies at 0.383 (Fig. 3a).
This leaves two remaining boundaries to be
defined, the boundary between Good and High status
and the boundary between Poor and Moderate Status.
These boundaries are defined as the midpoints
between the two adjacent boundaries:
EQRHigh=Good ¼ 0:5EQRRefCon=High
þ 0:5EQRGood=Moderate
ð7Þ
EQRModerate=Poor ¼ 0:5EQRGood=Moderate
þ 0:5EQRPoor=Bad
ð8Þ
For the example of AcDev equal to 50% the values
for the High/Good and Moderate/Poor boundaries
equal 0.808 and 0.525, respectively. Figure 3a shows
how the value of EQR for the boundary between the
classes varies with the AcDev from RefCon.
The method used for calculating class boundaries
does not allow for use of AcDev greater than 110%
for indicators with a positive response to nutrient
input, as the Poor/Bad boundary would otherwise
become negative (Fig. 3a). Consequently, it would
therefore become impossible to obtain a ‘‘Bad’’ status
123
Step 1B: Indicators with a numerical negative
relationship to nutrient input For an indicator
showing a negative response to nutrient input, e.g.
depth limit of SAV or Secchi depth, the EQR is
defined as:
EQR ¼ AcStat=RefCon
ð9Þ
ð10Þ
Here, for a given reference value, the EQR is
directly proportional to the observed value, and is
equal to the maximum value of 1.00 if the AcStat
equals or exceeds the reference value.
As for the case of positive response, the AcDev
from RefCon is used to define class boundaries for
classification according to EQR value. Again, the
Good/Moderate boundary lies where the deviation
from RefCon is equal to the AcDev (3).
Using (3) to substitute for AcStat in (9), and
remembering that AcDev is negative, gives:
EQRGood=Moderate ¼ ð1 AcDevÞ
ð11Þ
For an AcDev of 50%, the boundary for Good/
Moderate status is 0.5. Figure 4b shows how the class
boundaries vary with the AcDev. Given the value for
the Good/Moderate boundary and the Ref/High
boundary (0.95), the values for the remaining
boundaries are calculated in the same manner as
described above for indicators with a positive
response to nutrient input. Figure 4b is useful in
illustrating the limit on allowable AcDev for an
indicator with negative response. Choosing an AcDev
greater than 52.5% would mean that according to the
previously described method of calculating class
boundaries, the Bad/Poor boundary becomes negative
(Fig. 4b) and it is therefore impossible to arrive at a
classification of Bad, no matter how far from RefCon
the observed status is.
Step 2: Quality elements and final classification
An EQR value and a set of class boundaries are
calculated for each indicator, but the overall status
classification depends on a combination of indicators.
First, indicator EQR values are combined to give an
EQR value for a specific Quality Element (QE), and
Biogeochemistry (2011) 106:137–156
145
b Fig. 5 Ecological Quality Ratios (EQRs) calculated for open
water bodies for a Dissolved Inorganic Nitrogen (DIN),
b Dissolved Inorganic Phosphorus (DIP), c Chlorophyll-a,
d Water transparency (as Secchi depth), and e gamma diversity
for benthic invertebrates. Numbers refer to Fig. 1. Please note
that no data on DIP are available for area #4, no data on Secchi
depth are available for areas #12, 13, and 16 and no data on
benthic invertebrates are available for areas #4, 6, and 11–17
equal weights, the EQR for the QE is the average of
the indicators’ EQRs within the QE and each QE
class boundary (e.g. Moderate/Good boundary) is
found as the average of the class boundary values for
all indicators representing that specific QE.
Within a QE, it is also possible to assign weighting
factors to indicators according to expert judgement.
The classification of the QE is then given by
comparison of the weighted averages of the EQRs
with the weighted averages of the individual class
boundaries. Thus, the same weighting is applied both
in calculation of the EQR for the specific QE as well
as QE class boundary values.
The lowest rated of the QEs will because of the
‘One out—all out’ principle determine to final status
classification. This principle is employed for two
reasons: (1) all five HELCOM objectives for the open
basins are required to be met independently, and (2)
this principle is stated in the WFD (Anon. 2000) for
assessing ecological status of coastal waters.
Results
similarly the indicator class boundaries are combined
to give the class boundaries for the QE. In the
simplest case, where all indicators within a QE have
Eutrophication status in the Baltic Sea has been
calculated for 189 assessment units: 172 coastal areas
and 17 open water bodies. In the open water areas,
monitoring data was combined into larger areas by
calculation of mean values to give a common status
for an entire sub-basin, whereas the coastal areas
were assessed in smaller scale (Fig. 1). The EQR
values for nutrients, chlorophyll-a, water transparency and the gamma diversity of benthic invertebrates are presented in Fig. 5.
For the open water bodies, 15 out of 17 are
classified as ‘areas affected by eutrophication’. The
results are summarised in Table 1. Only the Bothnian
Bay and the north-eastern part of the Kattegat are
regarded as ‘unaffected by eutrophication’. The
results of the open water body classifications for
nutrients, chlorophyll-a, water transparency, and
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Biogeochemistry (2011) 106:137–156
Table 1 Classification of eutrophication status for 17 open water areas in the Baltic Sea region
No.
Area
Ecological quality ratio
PC
PP
Eutrophication status
BIC
1
Bothnian Bay
0.729 (H)
0.668 (H)
0.830 (G)
Good
2
Bothnian Sea
0.724 (G)
0.508 (P)
0.834 (H)
Poor
3
Gulf of Finland
0.468 (P)
0.220 (B)
0.394 (B)
Bad
4
Gulf of Riga
0.543 (M)
0.340 (B)
–
Bad
5
Northern Baltic Proper
0.523 (P)
0.231 (B)
0.000 (B)
Bad
6
Western Gotland Basin
0.660 (M)
0.432 (P)
–
Poor
7
Eastern Gotland Basin
0.610 (M)
0.486 (P)
0.116 (B)
Bad
8
SE Gotland Basin, open parts
0.745 (G)
0.400 (P)
0.222 (B)
Bad
Bad
9
Bornholm Basin
0.602 (M)
0.553 (M)
0.239 (B)
10
Arkona Basin
0.616 (M)
0.535 (M)
0.764 (G)
Moderate
11
Great Belt
0.356 (B)
0.295 (B)
–
Bad
12
Kattegat, south-western
0.716 (H)
0.460 (P)
0.584 (B)
Bad
13
Kattegat, south open parts
0.561 (M)
0.351 (B)
–
Bad
14
Kattegat, south-eastern
0.821 (G)
0.588 (M)
–
Moderate
15
16
Kattegat, central
Kattegat, north-eastern
0.691 (M)
0.787 (G)
0.440 (P)
0.813 (H)
0.549 (M)
–
Poor
Good
17
Kattegat, north-western
0.845 (H)
0.603 (M)
–
Moderate
The eutrophication status is based on the ‘One out—all out’ principle. See Fig. 1 for location of the areas. Detailed HEAT
calculations are available as Electronic Supplementary Material in Andersen et al. (2010). Please note that all values are EQR values
Please note that the EQR values in bold are decisive for the final classification of eutrophication status
PC physical–chemical indicators, PP phytoplankton, and BIC benthic invertebrate communities, H High, G Good, M Moderate,
P Poor, B Bad
benthic invertebrates are presented in the following
sections. The detailed HEAT classifications for are
available as electronic supplementary material in
Andersen et al. (2010).
Nutrients
The highest DIN concentrations are found in the
Bothnian Bay, which is predominantly P-limited
(Tamminen and Andersen 2007) and therefore DIN
may accumulate to reach levels above those in other
basins (for actual data, see electronically supplementary material in Andersen et al. 2010). DIN concentrations in the Gulf of Finland are also high due to
large fluvial input of nutrients mainly from the Neva
River. For the other basins, DIN winter means vary
between 3 and 4 lmol l-l. The Gulf of Riga and the
Gulf of Finland have the highest TN annual means
(26 and 24 lmol l-l, respectively), which are due to
large riverine discharges to both basins (Fig. 5a). The
other basins have TN levels between 18 and
123
21 lmol l-l, with the lowest concentrations in the
Danish Straits. From the Baltic Proper to the Danish
Straits, there is a natural decreasing spatial gradient
owing to the mixing with Skagerrak surface water
that generally has lower TN levels.
High DIP winter means are found in the Gulf of Riga
and the Gulf of Finland (0.78 and 0.84 lmol l-l,
respectively) owing to the large influence from riverine
discharges and the upwelling of bottom waters rich in
phosphorus deriving from the Baltic Proper (Pitkänen
et al. 2001). DIP levels in the Bothnian Sea, the
Baltic Proper and the Danish Straits are similar
(0.35–0.47 lmol l-l), whereas DIP concentrations in
the Bothnian Bay are very low (0.06 lmol l-l). These
spatial differences are unaltered for TP, with high
levels in the Gulf of Riga and the Gulf of Finland
(0.70 and 0.85 lmol l-l, respectively), moderate TP
levels in the Baltic Proper and the Danish Straits
(*0.58 lmol l-l) with slightly lower levels in the
Bothnian Sea (0.42 lmol l-l) and substantially lower
in the Bothnian Bay (0.16 lmol l-l).
Biogeochemistry (2011) 106:137–156
The EQR values for DIN vary between 0.22 and
0.81 (see Fig. 5a). For DIP, EQR values vary
between 0.33 and 1.00, the latter being an indication
of almost pristine conditions in the Bothnian Bay and
the Bothnian Sea (Fig. 5b). As expected, nutrient
status is acceptable in the Bothnian Bay (area 1). The
only other areas where nutrient status is acceptable
are the northern parts of the Kattegat (areas 16 and
17), areas 2 (Bothnian Sea), 8 (south-eastern Baltic
Proper), and 14 (south-eastern Kattegat).
Phytoplankton and water transparency
Mean summer (June–September) chlorophyll-a concentrations are highest for the open water bodies in
the Gulf of Finland, the Northern Baltic Proper and
the Gulf of Riga (5.4, 4.8 and 5.3 lg l-l, respectively). In other open water bodies, average chlorophyll-a concentrations range from 1.9 to 2.7 lg l-l.
The variability in summer (June–September) chlorophyll-a observations in 2001–2006 is high, with
individual values ranging from 0.1 to [50 lg l-l.
In most of the open Baltic Sea areas, chlorophyll-a
concentrations indicate eutrophication. In other
words, EQR values derived for chlorophyll-a show
a clear deviation from RefCon (Fig. 5c). In the open
sea, the chlorophyll-a derived status is the highest in
the Bothnian Bay and the Kattegat (0.67 and 0.63,
respectively) and lowest in the Gulf of Finland, the
Northern Baltic Proper, and the Gulf of Riga (0.22,
0.23 and 0.34, respectively).
Reduced water transparency is partly an effect of
increased nutrient loads, mediated through increased
phytoplankton growth. In comparison to RefCon
(Fig. 5d), water transparency status has decreased in
all Baltic Sea sub-areas at both at coastal and open
sea sites reflecting visible eutrophication effects in
the entire Baltic Sea.
Water transparency status in open sea areas
expressed as EQR values vary markedly in different
sub-basins of the Baltic Sea. Status expressed as EQR
values varies from 0.75 to 1.0 for the southern and
central sub-basins, indicating a 0–25% decrease in
water transparency from near-pristine RefCon. However, sub-basins north of the Northern Baltic Proper
have a significantly lower status with EQR values
ranging from 0.50 to 0.61, representing a reduction of
147
39–50% in water transparency compared to RefCon.
The mean EQR value for all open sub-basins assessed
is 0.72. In the south-eastern Gotland Basin and
Arkona Basin water transparency status is highest of
all open sub-basins, with EQR values of 1.0 and 0.94
respectively. In the Kattegat water transparency
status exceeds the mean status (mean EQR for
Kattegat sites 0.75). In the Bornholm Basin, the
Western and Eastern Gotland Basin, the EQR values
are nearly equal to the Kattegat (0.75–0.81). In Gulf
of Riga, the two indicators used for Secchi depth have
variable RefCon (4.0 m for the Finnish indicator and
6.0 m for the Latvian indicator) and result in different
EQR values of 0.75 and 0.57, respectively.
The Northern Baltic Proper and Gulf of Finland
represent a distinctly lower status compared to
RefCon, with EQR values of 0.61 in the open
Northern Baltic Proper and 0.50 in the Gulf of
Finland. In the open sea areas of the Gulf of Bothnia
water transparency EQR is 0.61 in the Bothnian Sea
and 0.56 in the Bothnian Bay.
Benthic invertebrates
No benthic invertebrates survive in areas with
prolonged or permanent oxygen depletion such as in
the deep parts of the Baltic Proper. In areas with
periodic oxygen depletion every late summer and
autumn, the number of benthic species is reduced
significantly and mature communities cannot develop.
In marine areas with temporary oxygen depletion,
intermittent recovery will occur whenever conditions
improve. Oxygen depletion, if rare enough, may be
viewed as a temporal and spatial mosaic of disturbance that results in the loss of habitats, reductions in
biodiversity, and a loss of functionally important
species. Macrobenthic communities are severely
degraded throughout the open sea areas of the Baltic
Proper and the Gulf of Finland, whereas conditions in
the Arkona Basin, the Bothnian Sea and Bothnian Bay
are classified as being good (Fig. 5e).
For the open waters, the EQR values vary between
0.00 and 0.83. The highest EQR values are as
indicated above found in the Arkona Basin (0.77), the
Bothnian Sea (0.83) and the Bothnian Bay (0.83). For
the Baltic Proper and the Gulf of Finland, EQR
values range from 0.00 to 0.39 indicating impaired
environmental conditions.
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Biogeochemistry (2011) 106:137–156
Table 2 Summary of eutrophication status classifications of 172 coastal water bodies in the Baltic Sea region
Basins and sub-basins
Eutrophication status classification
High
Good
Moderate
Total
Poor
Bad
Bothnian Bay
0
1
3
2
2
8
The Quark
0
1
1
0
0
2
21
Bothnian Sea
0
9
6
2
4
The Archipelago and Åland Seas
0
0
2
1
3
6
Gulf of Finland
0
0
4
6
9
19
Gulf of Riga
0
0
0
3
2
5
Baltic Proper, northern parts
0
0
3
7
30
40
Eastern Gotland Basin
0
0
0
0
7
7
19
Western Gotland Basin
0
0
0
5
14
Gulf of Gdansk
0
0
0
1
4
5
Bornholm Basin
0
0
1
7
5
13
Arkona Basin
0
0
1
1
1
3
Kiel Bight and Mecklenburg Bight
0
0
0
2
3
5
Danish Straits including the Sound
0
0
1
4
5
10
Kattegat
Total
0
0
0
11
3
25
1
42
5
94
9
172
High and Good represent ‘areas unaffected by eutrophication’, while Moderate, Poor, and Bad represent ‘areas affected by
eutrophication’
Coastal waters
Of the 172 coastal waters assessed, 161 are classified
as ‘affected by eutrophication’ (Table 2). Coastal
waters are in general more vulnerable to nutrient
inputs than open waters—important causes being the
lower retention times as well as closer benthicpelagic interactions (Borum 1996; Wasmund et al.
2001). Seasonal variations in supply, removal, and
transformation processes give rise to distinct seasonal
patterns for nutrient concentrations in Baltic Sea
coastal areas. Distinct spatial gradients are also
found, with elevated nutrient concentrations in estuaries and coastal waters compared to open waters.
This gradient is most pronounced in the Danish
Straits and Baltic Proper. Nutrient concentrations in
coastal areas of the Gulf of Finland are similar to
those in the open sea because of upwelling of
offshore bottom water. Detailed information on
nutrient status of the coastal waters can be found in
HELCOM (2009) and Lysiak-Pastuszak et al. (2009).
In a majority of coastal Baltic areas, chlorophyll-a
concentrations and water transparency measurements
123
indicate the prevalence of eutrophication (data not
shown). In other words, EQR values derived from
chlorophyll-a and water transparency measurements
show a clear deviation from RefCon. Detailed
information about the status of planktonic communities and water transparency in various coastal waters
of the Baltic Sea can be found in Feistel et al. (2008)
and HELCOM (2009).
Extensive seagrass meadows and perennial macroalgal communities harbour the highest biodiversity
in coastal, shallow-water ecosystems. Eutrophication
has complex effects on SAV causing shifting of the
distribution depth limit towards the surface, preventing the settlement of new specimens on the seafloor
due to increased sedimentation, and favouring opportunistic species with a short life cycle and rapid
development over the perennial species, thus causing
a shift in community composition. Generally, the
level of eutrophication has caused serious changes in
the Baltic Sea SAV communities, although in many
cases the gaps in historical data do not allow us to
identify the exact timing of larger shifts in communities (Torn et al. 2006). Present-day monitoring data
Biogeochemistry (2011) 106:137–156
show that the degradation of communities is ongoing
in several areas (HELCOM 2009). At the same time,
positive signs of a slowing down or reversal of some
eutrophication effects on SAV parameters could be
observed in areas of the Northern Baltic Proper and
the Gulf of Finland, where the previous distribution
of macrophyte species has recovered in some areas
(Nilsson et al. 2004; HELCOM 2009).
In the western part of the Baltic Sea (the Kattegat
and the Danish Straits), the EQR values for the depth
distribution of Zostera marina vary between 0.89 and
0.59. With a -25% AcDev, only the Danish coastal
areas of the Kattegat have average EQR values above
0.75. For the Danish Straits, all average EQR values
are below 0.75, and hence classified as ‘affected by
eutrophication’. In the central, eastern and northern
parts of the Baltic Sea, in areas dominated by Fucus
vesiculosus, average EQR values vary between 0.84
and 0.55. EQR values above 0.75 are found in the
Gulf of Riga and Eastern Baltic Proper. In the
Bothnian Sea, Gulf of Finland, and the western parts
of the Baltic Proper, the targets for SAV are generally
not met.
Macrozoobenthic communities in coastal waters
are highly variable both between and within different
sub-basins. In general, more sheltered and enclosed
coastal water bodies are in a worse state than more
exposed open coasts. Detailed information on status
of benthic invertebrates in Baltic Sea coastal water
can be found in HELCOM (2009).
Integrated assessment
Combining indicators and applying the ‘One out—all
out’ principle in order to produce a final classification
of eutrophication status represents a step forward
from assessments based on individual indicators
towards integrated assessments applying multi-metric
indicator-based assessment tools such as HEAT. The
results can be presented in several ways, e.g.: (1)
HEAT calculations (see electronic supplementary
material in Andersen et al. (2010) for details), (2)
summarised as in Tables 1 and 2 as well as (3) in the
form of maps of eutrophication status in the Baltic
Sea.
Figure 6 presents a merger of HEAT classifications for 17 open water areas (Table 1) and 172
coastal water bodies (Table 2) into an interpolated
map of eutrophication status of the Baltic Sea. All
149
open parts of the Baltic Sea except the Bothnian Bay
and the north-eastern parts of the Kattegat are
classified as ‘affected by eutrophication’. It should
be noted that also some coastal waters situated along
the Bothnian Sea are classified as ‘unaffected by
eutrophication’.
Discussion
This assessment of eutrophication status in the Baltic
Sea compares target values (EutroQOs), derived from
combining information on RefCon (representing a
‘then’ situation) and an AcDev with recent
(2001–2006) monitoring data (representing a ‘now’
situation). According to the results of this study only
open parts of the Bothnian Bay and north-eastern
Kattegat as well as some coastal waters in Bothnian
Bay are unaffected by eutrophication.
The results of this study are generally in line with
previous indicator-based assessments (HELCOM
2002, 2006; Ærtebjerg et al. 2003; Rönnberg and
Bonsdorff 2004) and can be directly compared with
the results of national coastal assessments and the EU
processes like WFD implementation in the Baltic
(e.g. Anon. 2008b). An added value of the method
employed here over e.g. WFD is that it uses
supporting parameters, e.g. nutrients and Secchi
depth, which are significantly correlated to the
biological quality elements, on the same level of
importance as the biological quality elements
(Nielsen et al. 2002a, b; Krause-Jensen et al. 2003).
The RefCon values derived for all 17 open water
‘assessment units’ are based on the analysis of
historical data. The RefCon values used for the open
parts of the Baltic Sea represent the best available
knowledge about the eutrophication status of the
Baltic Sea 50–100 years ago before the onset of the
current large scale eutrophication process (Schernewski and Neumann 2005; Savchuk et al. 2008) and
the monitoring data used in this study represent the
best available datasets for the area. Hence, these
RefCon values are in principle ready for immediate
use in regard to any updates of the BSAP, e.g. as done
by Wulff et al. 2007.
The principles of this assessment for setting ‘target
values’ (e.g. the AcDev) are in line with the WFD: it
is the boundary between Good Ecological Status and
Moderate Ecological Status according to the WFD
123
150
Biogeochemistry (2011) 106:137–156
Fig. 6 Integrated and
interpolated five-class
classification of
eutrophication status in the
Baltic Sea region. The
interpolation was made by
inverse distance weighting
method and the gradients
among the point values
were permitted to change
over intermediate distances.
While the status of offshore
areas was pooled to a single
value from multiple point
values, the coastal
assessment units were
treated as separate and were
given a 25 km effect radius.
Reproduced with
permission from HELCOM
(Anon. 2000). For Good Ecological Status, which
together with High Ecological Status, is considered
acceptable status, the values of the biological quality
elements show low levels of disturbance from
RefCon as a result of human activity. For Moderate
Ecological Status, which together with Poor and Bad
Ecological Status, is considered an unacceptable
status, the values of the biological quality elements,
compared to RefCon, deviate moderately (or more)
from those normally associated with the water body
type under undisturbed conditions.
The nutrient concentrations overall reflect the
balance between inputs from land, atmosphere and
loss processes, and are generally in line with other
studies and assessments carried out in the Baltic Sea,
e.g. Lundberg et al. 2009. Nutrient concentrations can
be influenced also by upward mixing from deeper
123
water layers (Vahtera et al. 2007; Feistel et al. 2008;
Reissmann et al. 2009). Upwelling is an important
source of phosphorus in the Gulf of Finland, the Gulf
of Riga and also in the Baltic Proper (Nausch et al.
2009). The relatively high EQR values found in the
south-eastern Baltic Proper (0.75), the western Gotland Basin (0.81), and the south-eastern Kattegat
(0.78) are assumed to be related to imprecise setting
of RefCon. There is a need for the development of
more harmonised information on RefCon values for
nutrient concentrations.
Phytoplankton is perhaps the most important
element in any assessment of eutrophication in the
Baltic Sea, since phytoplankton primary production
and biomass are essentially coupled to nutrient
concentrations. Chlorophyll-a concentrations are
widely used as a proxy for phytoplankton biomass,
Biogeochemistry (2011) 106:137–156
but other indicators should be developed, e.g. in
regard to algal species indicative of nuisance or toxic
algal blooms. The findings presented here are generally in line with other studies and assessments, e.g.
Jaanus et al. (2007), Fleming-Lehtinen et al. (2008),
Håkansson and Lindgren (2008), Wasmund and
Siegel (2008). During recent decades, chlorophyll-a
concentrations have been increasing in most of the
Baltic Sea sub-regions, although in the 2000s chlorophyll-a levels in many open sea areas show signs of
a decreasing trend. RefCon values for chlorophyll-a
in open waters seem appropriate for the time being.
For coastal waters there seem to be a need for joint
principles and methods of setting not only RefCon
values, but also AcDev’s. This has not yet been
achieved by the WFD intercalibration activity.
The assessment of water transparency is closely
linked to the assessment of phytoplankton and SAV,
and in this study water transparency is regarded as a
proxy of eutrophication. An added value in regard to
water transparency is the length of the time series,
which extends close to 100 years back in time
(Sanden and Håkansson 1996). The findings presented here are generally in line with other studies
and assessments, e.g. Kautsky et al. (1986), and
Eriksson et al. (1998, 2002). In the Gulf of Riga, low
status is consistent with lower RefCon compared to
other areas. Low status in the Gulf of Bothnia may be
attributed mostly to changes in land use affecting
water colour (humic substances), whereas in the Gulf
of Finland the increase of phytoplankton biomass is a
more likely proximate reason for the low status.
The benthic invertebrate assessment for open
waters shows that the benthic communities are
structured by a combination of physical factors (e.g.
salinity and sediment type) and eutrophication, which
result in a higher susceptibility to hypoxia/anoxia.
The findings presented here are generally in line with
other studies and assessments, e.g. Karlson et al.
(2002), and Perus and Bonsdorff (2004). A special
challenge is the difficulty in defining historical
RefCon for macrozoobenthos—this emphasizes the
importance of conducting long-term monitoring over
large spatial scales to be able to assess changes.
Assessment of SAV in coastal waters is, at least
compared to the assessment of open waters, somewhat more challenging because the status of SAV
communities depends on a variety of local environmental conditions which also affect also the
151
eutrophication processes on very limited, local scale,
e.g. changes in nutrient loading to specific river basin
or fjord or bay while open sea indicators reflect
situation on larger sea area. So it is no surprise that
especially in case of extensive archipelago areas
some SAV indicators can show development in
opposite direction than indicators of nearby open
sea areas. In our case some recovery in the depth
distribution of SAV has occurred during last decades
in the Northern Baltic Proper (extensive archipelago
areas) as well as in some areas of the Gulf of Finland,
while indicators used for open sea areas still show
declining status.
There is in our opinion no such thing as a perfect
assessment tool. More targeted monitoring and
improved understanding of the eutrophication processes will lead to better knowledge, better indicators
and subsequently better assessment tool. The strength
of HEAT compared to the OSPAR equivalent on
which it is built, is that it is modernized in the sense it
makes use of (1) the EQR and the ‘one out, all out’
principle. Hence, HEAT is directly linked to the
principles for assessment of ecological status of
coastal water sensu the WFD. An added value
of HEAT is that it enables a secondary assessment
of confidence (see Andersen et al. 2010). Compared
to OSPAR COMP, the HEAT tool has no or few
weaknesses. When using HEAT for assessment of
‘ecological status’ sensu the WFD, it can be argued
that ‘eutrophication status’ and ‘ecological status’ are
different issues. This point is for somewhat meaningless, at least for the Baltic Sea, where the major
threat to the coastal ecosystems is nutrient enrichment and eutrophication. It can also be argued the
combination of indicators per QE mixes indicators
with different boundary setting, but here it should be
eminent that the classes used by the WFD are related
to QE (cf. Annex 5), not to individual indicators or
indices.
By providing a regional overview of eutrophication status in the Baltic Sea the results of this study
provide interesting perspectives and links to the
implementation of a range of EU Directives, e.g. the
WFD, the MSFD (Anon. 2008a), the EC Urban
Wastewater Treatment Directive (Anon. 1991a) and
the EC Nitrates Directive (Anon. 1991b). The
relations in regard to boundary setting and classification are discussed and outlined in Anon. (2009) and
HELCOM (2009). If the convergence of the aims of
123
152
these directives is taken seriously, marine waters
classified as ‘affected by eutrophication’ could by no
means be accepted as having either ‘Good Ecological
Status’ or habitats with a ‘Favourable Conservation
Status’. Similarly it can be argued that waters
classified as ‘affected by eutrophication’ should be
designated as ‘sensitive’ to nutrient inputs from
industries and cities. Along the same lines waters
affected by eutrophication should be regarded as
‘polluted’ when situated downstream of catchment
dominated by agriculture, implying that the catchment should be designated as ‘vulnerable’ in regard
to losses of nitrogen from agricultural practices.
Future assessments will however be worthless if
we fail to safeguard the current spatial and temporal
resolution of HELCOM COMBINE and monitoring
for the joint HELCOM core set of eutrophication
indicators. Any weakening of these activities will
jeopardize future re-assessments of eutrophication
status of the Baltic Sea. Issues to be improved before
a re-assessment include: (1) harmonization and
evaluation of the quality of reference condition
values (RefCon), (2) improvements of the target
values (e.g. AcDev) (more research on functional
relations, natural variations etc.), (3) improvements in
spatial and temporal coverage of HELCOM COMBINE monitoring in some areas (e.g. Gulf of Riga,
eastern Baltic Proper, South-eastern Baltic proper),
(4) adequate monitoring of SAV, (5) development of
oxygen indicators, and (6) development of statistical
principles for weighting indicators.
The current impaired status of most parts of the
Baltic Sea is a consequence of a combined increase in
population density and altered agricultural practices.
This has resulted in increased discharges, emissions
(including atmospheric nitrogen emissions) and
losses of nutrients to the environment and ultimately
nutrient enrichment in the aquatic environment. Only
few data series of nutrient loading exist, e.g.
Stålnacke (1996) and Conley et al. (2007), and
hence, the long-term nutrient enrichment will have to
be documented by the temporal trends for TN and TP
concentrations as well as TN:TP ratio in surface
(0–10 m) and bottom waters ([100 m) starting from
the 1970s until 2006 (HELCOM 2009).
Nutrient concentrations increased until the 1980s,
and in all areas except for the Gulf of Finland,
phosphorus concentrations have declined during the
past two decades (HELCOM 2009). Nitrogen
123
Biogeochemistry (2011) 106:137–156
concentrations have declined in the Gulf of Riga,
the Baltic Proper and the Danish Straits. These
declines, particularly in the coastal zone, are partly
caused by lower nutrient loads from land. Furthermore, changing volumes of hypoxia in the Baltic
Proper significantly alter nutrient concentrations in
bottom waters and, through subsequently mixing,
also in surface waters. This does not affect the Baltic
Proper alone but also connecting basins through
advective exchanges. In particular, the Gulf of
Finland has been severely affected by internal
loading of phosphorus from the sediments caused
by poor oxygen conditions (Vahtera et al. 2007).
The elevated nutrient concentrations compared to
RefCon are primarily a consequence of a long-term
(100? years) increase in direct and riverine loads to
the Baltic Sea. However, management strategies
focusing mainly on direct discharges have during
the last 20 years resulted in a decrease in loads to the
Baltic Sea (Fig. 7). However, it has to be taken into
Fig. 7 Trends in inputs of total nitrogen (TN) and total
phosphorus (TP) to the Baltic Sea. Please note that the TP input
has been scaled by factor 10. The solid line indicate run off in
m3 s-1. ‘‘2003’’ = 2001–2003 and ‘‘2006’’ = 2001–2006.
2021 (grey bars) show the ultimate nutrient input targets to
be reached as agreed by the HELCOM Baltic Sea Action Plan
Biogeochemistry (2011) 106:137–156
account that decreased flow is also partly responsible
for decreasing loads (HELCOM 2009).
Improving the eutrophication status, especially of
those areas classified as affected by eutrophication,
relies on a better linking of ecosystem effects,
nutrient concentrations, loads and human activities
in upstream catchments. The key issue is to reverse
the trend of eutrophication, sometimes referred to as
oligotrophication (Nixon 2009), and to reduce inputs
of nutrients to the Baltic Sea region. Some improvement has been made in some regions (Carstensen
et al. 2006 and Fig. 7) but additional reductions are
clearly needed. Recent modelling efforts (Wulff et al.
2007; Savchuk et al. 2008) have come a long way in
providing advice on the magnitude of nutrient input
reductions required to reach identified target levels of
key parameters, such as those utilised in this study
(Fig. 2). A first round of such calculations was
actually adopted in 2007 by Baltic Sea states in the
BSAP, partly based on an ecosystem approach to
management of human activities (HELCOM 2007b;
Wulff et al. 2007). Recently, a process of revision of
these reduction figures was begun, taking into
account more assessment parameters and atmospheric
deposition to better reflect relevant ecosystem elements and all relevant pathways of nutrient input.
When developing and implementing ecosystembased nutrient management strategies, it has been
debated whether a nutrient management strategy such
as the BSAP should focus either on N, P or both
(Tamminen and Andersen 2007). Given the variations in nutrient limitation between region and
seasons—and the fact that the flow out of the Baltic
Sea passes areas which are nitrogen limited—it is
clear that alleviation of eutrophication requires a
balanced and strategic approach to control both
nitrogen and phosphorus appropriately (Conley
et al. 2009b).
What we consider in our assessment of eutrophication or ecological status being a straightforward
eutrophication signal is in reality a response not only
to nutrient enrichment, but also to many other
pressures (Jackson et al. 2001). Often the functional
relations are complicated, including issues like
thresholds, regime shifts and climate change (Duarte
et al. 2009; Duarte 2009). The implications for
management are currently being understood and
interpreted. A rational solution would be to acknowledge that other pressures (e.g. climate change) might
153
enhance eutrophication signals and that further efforts
in regard to reduction of nutrient inputs may been
needed to comply with most eutrophication related
objectives.
Conclusions
This study has introduced a multi-metric indicatorbased eutrophication assessment tool enabling a
harmonized assessment of eutrophication status in
the whole Baltic Sea. Most parts of the Baltic Sea are,
not surprisingly, judging from available scientific
literature, affected by nutrient enrichment.
The recently developed HEAT as described in this
paper provides a qualified answer to this key question
‘‘Do we have a problem or not?’’ and thus a basis for
the implementation or revision of a Baltic Sea-wide
nutrient management strategy, e.g. the BSAP.
HEAT represents a major step forward in terms of
assessing eutrophication in the Baltic Sea. Firstly,
because HEAT is based on well-established eutrophication indicators, it is in line with the principles of
the WFD, and, perhaps more importantly, it uses the
EQR approach to enable direct comparisons of all
areas assessed despite variation in monitoring activities. Secondly, HEAT classifications can be regarded
as a baseline for the reduction figures defined in the
eutrophication segment of the BSAP against which
the HELCOM vision of a Baltic Sea unaffected by
eutrophication can be judged.
HEAT has shown itself to be a good tool and
should be used for a HELCOM re-assessment of the
eutrophication status of the Baltic Sea within e.g.
6–10 years in order to follow the implementation of
the BSAP and validate the effectiveness of the
reduction measures established so far.
Future assessments should be based on the best
scientifically based indicators and assessment tools
available rather than waiting for so-called ‘perfection’. However, development of eutrophication
assessment tools and nutrient management strategies
in the Baltic and elsewhere should ideally be
adaptive: there should always be the intention to
adapt these tools when new scientific knowledge
becomes available. Similarly, nutrient management
strategies should be based on the best available
science-based functional relations between causes
and effects, using models and Decisions Support
123
154
Systems as appropriate. Eutrophication in the Baltic
Sea is a significant challenge and the absence of
faultless tools should not prevent the Baltic Sea
countries from trying to meet this challenge.
Acknowledgements The views expressed are those of the
authors and do not necessarily represent official positions of
their institutions. The authors would like to thank national and
local authorities, and colleagues for making monitoring data
available for the HELCOM integrated thematic assessment of
eutrophication in the Baltic Sea region. Special thanks are owed
to: Juris Aigars, Mats Blomqvist, Saara Bäck, Alf B. Josefson,
Henning Karup, Pirkko Kauppila, Pirjo Kuuppo, Juha-Markku
Leppänen, Bärbel Müller-Karulis, Samuli Neuvonen, Janet
Pawlak, Heikki Pitkänen, Johnny Reker and Roger Sedin. This
work has received financial support from HELCOM (HELCOM
EUTRO and HELCOM EUTRO-PRO), the Danish Ministry for
the Environment (CO-EUTRO) and DHI.
Open Access This article is distributed under the terms of the
Creative Commons Attribution Noncommercial License which
permits any noncommercial use, distribution, and reproduction
in any medium, provided the original author(s) and source are
credited.
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Ecological Indicators 15 (2012) 105–114
Contents lists available at SciVerse ScienceDirect
Ecological Indicators
journal homepage: www.elsevier.com/locate/ecolind
Human pressures and their potential impact on the Baltic Sea ecosystem
Samuli Korpinen a,∗ , Laura Meski a , Jesper H. Andersen b,c , Maria Laamanen a
a
b
c
Helsinki Commission, Katajanokanlaituri 6B, FIN-000160 Helsinki, Finland
Department of Bioscience, Aarhus University, Ny Munkegade 120, 8000 Aarhus C, Denmark
National Centre for Environment and Energy (NERI), Frederiksborgvej 399, 4000 Roskilde, Denmark
a r t i c l e
i n f o
Article history:
Received 17 March 2011
Received in revised form 15 August 2011
Accepted 19 September 2011
Keywords:
Anthropogenic pressure
Baltic Sea
Cumulative impacts
Environmental impact assessment
Marine Strategy Framework Directive
a b s t r a c t
The EU Marine Strategy Framework Directive requires Member States to estimate the level of human
impacts on their marine waters. We report the first attempt to quantify the magnitude and distribution
of cumulative impacts of anthropogenic pressures for an entire regional sea, the Baltic Sea. We used
a method which takes account of the sensitivity of different ecosystem components and gives scores
for potential impacts in 5 km × 5 km areas. Our quantification of impacts was based on data layers of
anthropogenic pressures and ecosystem components. The classification of the anthropogenic pressures
follows the MSFD and the outcome of the index was targeted to facilitate the implementation of the
directive. The study presents the cumulative impacts over the entire sea area and shows that the highest
estimated impacts were in the southern and south-western sea areas and in the Gulf of Finland. The
lowest index values were found in the Gulf of Bothnia. The results coincide with the population densities
of the adjacent catchment areas. Fishing, inputs of nutrients and organic matter and inputs of hazardous
substances comprised 25%, 30% and 30%, respectively, of the total cumulative impact. The approach used
is transparent and the results are useful in regard to ecosystem-based management, e.g. for area-based
management and assessments. Examples of uses are given together with analysis of the strengths and
weaknesses of the approach.
© 2011 Elsevier Ltd. All rights reserved.
1. Introduction
Human activities place heavy pressures on the global marine
ecosystems (Millenium Ecosystem Assessment, 2005; Halpern
et al., 2008; Crain et al., 2009). Coastal marine environments
and marginal seas in particular have experienced ecosystem
regime shifts, altered food web structures, heavily contaminated sediments and adverse effects from hazardous substances
(Jackson et al., 2001; Kappel, 2005; Casini et al., 2008; Coll et al.,
2008). A human activity may cause multiple pressures on the
marine ecosystem. For example, bottom trawling alone may produce simultaneously above-water and underwater noise, harvest
of target and non-target species, physical disturbance of the
sea bed, increased siltation, and resuspension of nutrients and
hazardous substances. Moreover, anthropogenic pressures have
varying impacts on different components of the ecosystem which
underlines the importance of including ecosystem aspects to an
assessment of impacts. Despite the long history of human activities at sea, quantitative spatial analyses of anthropogenic pressures
and their cumulative impacts on the marine ecosystems have been
∗ Corresponding author.
E-mail address: samuli.korpinen@helcom.fi (S. Korpinen).
1470-160X/$ – see front matter © 2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.ecolind.2011.09.023
conducted only recently (Lindeboom, 2005; Halpern et al., 2008,
2009; Selkoe et al., 2009; Ban et al., 2010).
Assessment of pressures and impacts is one of the key features of
the EU Marine Strategy Framework Directive (MSFD, Anon, 2008).
The directive requires Member States to make assessments not
only on pressures and impacts but also on the state of the marine
environment and then take measures towards reaching a good
environmental status (GES) by 2020. The MSFD stipulates that GES
means the environmental status of marine waters where “these
provide ecologically diverse and dynamic oceans and seas which
are clean, healthy and productive within their intrinsic conditions,
and the use of the marine environment is at a level that is sustainable, thus safeguarding the potential for uses and activities by
current and future generations.” The level of sustainable use is not
prescribed in the directive and therefore spatial presentations of
pressures and their potential impacts can be used as a starting point
to iterate such a sustainable level. Despite the tight implementation schedule of the MSFD, no assessment of cumulative impacts
has yet been made in any of the European seas.
This study estimates the distribution and magnitude of different human activities both at sea and on land, associated pressures
and their potential impacts on the marine ecosystem. We used an
assessment tool which converts pressures to potential impacts on
selected components of the ecosystem and sums up all the impacts
in predefined assessment units. The tool is based on the method and
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S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
Fig. 1. Map of the Baltic Sea showing the sub-basins, drainage basin, major rivers, cities and countries. Differences in the water depth are indicated with different shades of
blue.
global assessment by Halpern et al. (2007, 2008). To our knowledge, this is one of the first spatial visualizations of cumulative
anthropogenic pressures and impacts in a coastal sea area globally. The assessment relies on the best available compilation of
data on human activities and ecosystem in the area which has been
possible to compile due to the long standing regional cooperation of
the Baltic Sea coastal countries and the EU under the umbrella of the
Helsinki Commission (HELCOM). This assessment should be seen
as the first step towards more comprehensive impact assessments
and better validated quantification of impacts.
S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
107
2. Methods
IDENTIFICATION OF
PRESSURES:
- Relevant parameter and
unit for the pressure
- Search for
distributionsand
intensity data
The Baltic Sea is a semi-enclosed brackish sea area with a gradient in surface water salinity from 31 to 2 PSU and relatively shallow
water (average depth 52 m). The oceanic connection is maintained
through narrow and shallow Danish straits, from which saline
water flows over a series of sills, which separate the Baltic Sea into
a series of sub-basins (Fig. 1). Natural features like water residence
time of around 30 years, shallowness and large catchment area
predispose the Baltic Sea to the accumulation of pollution by nutrients and hazardous substances. Large cities such as St. Petersburg,
Stockholm and have resided on the coast for centuries, and over
time, anthropogenic pressures have heavily affected the Baltic Sea
ecosystem, which is seen in decreased health of marine top predators, over-exploited fish populations, increased areas of anoxic sea
bottoms, contaminated fish and extensive algal blooms (HELCOM,
2009a,b, 2010a).
INDEX CALCULATION
2.1. The study area
I=
m
n Pi × Ej × i,j
i=1 j=1
where Pi is the log-transformed and normalized value (scaled
between 0 and 1, and with 1 being the highest value of the pressure measured) of an anthropogenic pressure in an assessment unit
i, Ej is the presence or absence of an ecosystem component j (i.e.
populations, species, biotopes or biotope complexes; 1 or 0, respectively), and i,j is the weight score for Pi in Ej (range 0–4, cf. Halpern
et al., 2007). The impact of any P × E × combination will be zero if
a pressure is zero or an ecosystem component is absent. Thus, the
more ecosystem components an area contains and the higher is the
number of pressures in that area, the higher the index value. The
final index value was calculated for 5 km × 5 km squares, i.e. the
assessment units. A schematic presentation of the different steps
in the use of the index is given in Fig. 2.
2.3. Compilation of data on anthropogenic pressures
We used the pressures to the marine environment as determined in the MSFD (Anon, 2008); the directive recognizes 18
pressure types. The data on marine litter and ‘other substances’
WEIGHT SCORES:
- Involve experts
- Create guidelines
- Create a questionnaire and
hold a workshop
- Take medians of the results
ECOSYSTEM DATA:
- 0 / 1 for assessment
units
INDEX CALCULATION:
- Multiply the three factors and sum
them up within an assessment unit
APPLICATIONS
We define an anthropogenic pressure as a human-derived stress
factor causing either temporary or permanent disturbance or damage to or loss of one or several components of an ecosystem. Thus,
pressure may cause immediate impacts or it may also be low
enough not to cause immediate adverse impacts on biota. According to our definition, potential anthropogenic impact is the possible
negative change a pressure may cause on an ecosystem component.
The impact is only considered potential, because our estimates rely
on the current, still imperfect, expert knowledge on the relationships between pressures and impacts on the ecosystem and the
actual impact can be reduced or increased by natural variability and
other stochastic factors. By cumulative impact we mean the sum of
all potential impacts in an area, not taking into account synergistic
or antagonistic effects. By an ecosystem component we mean biological parts of the ecosystem, such as species, biotopes formed by
habitat-forming species or abiotic biotopes with a clear linkage to
certain species.
The method to calculate an impact index value (I) for the set of
anthropogenic pressures in a given area was based on the following
formula (Halpern et al., 2008):
RESOLUTION:
Size of the assessment unit
PRESSURE LAYERS:
- Log-transformation
- Normalization [0-1]
2.2. Measuring cumulative impact
IDENTIFICATION OF
ECOSYSTEM DATA:
- Relevant species, biotopes, biotope complexes distribution etc.
- Search for data
MODIFIED INDICES:
Sector or themewise
index (e.g. only nutrients,
fisheries…)
Ecosystem component wise index (e.g. only
seagrass or seals)
Index without ecosystem
factors
MANAGEMENT
PURPOSES:
Assessments (MSFD),
Protected areas,
(threats)
Marine spatial
planning, EIAs and
permitting,
Scenarios
Fig. 2. A schematic presentation of the different steps in the calculation of the index
tool and its adaptations and suggested management purposes.
being unavailable and the impacts of non-indigenous species
unknown, this study employed 15 pressure types (Table 1) divided
into 52 Baltic Sea-wide data sets of anthropogenic pressures,
which were direct pressure data, proxies for pressures, or mere
presence/absence data of an activity or pressure. Baltic Sea environmental experts considered the data sets as evenly distributed
and to include all relevant sources of each pressure. Since there
are no direct measurements for some of the pressures, they were
estimated on the basis of the causative human activities. Detailed
information on the pressure data, including data sources, is provided in Appendix A and in HELCOM (2010b).
2.4. Spatial distribution of species, biotopes and biotope
complexes
There exists relatively accurate spatial distribution data on
some biotope complexes, biotopes and species in the Baltic Sea.
Six benthic and two pelagic biotope complexes were chosen for
the index: photic sand, photic soft bottom, photic hard bottom,
non-photic sand, non-photic soft bottom, non-photic hard bottom, photic water column and non-photic water column (based
on Al-Hamdani and Reker (2007) and the EUSeaMap project). Two
benthic biotopes (mussel beds and Zostera meadows) and four
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S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
Table 1
Anthropogenic pressure types in the marine environment, included in this study.
Category
Pressure type
Physical loss of seabed
Physical damage to seabed
Other physical disturbance
Interference with hydrological processes
Contamination by hazardous substances
Smothering by dumped material, sealing of seabed
Changes in siltation, abrasion of seabed, selective extraction of non-living resources
Underwater noise
Changes in thermal regime, changes in salinity regime
Introduction of synthetic compounds, introduction of non-synthetic substances and compounds, introduction of
radio-nuclides
Inputs of nutrients, inputs of organic matter
Introduction of microbial pathogens, selective extraction of species (e.g. fishing)
Nutrient and organic matter enrichment
Biological disturbance
species-related distribution data sets (distribution of harbour porpoise [Phocoena phocoena], distribution of the three seal species
[grey seal Halichoerus grypus, ringed seal Pusa hispida botnica and
harbour seal Phoca vitulina], wintering grounds of sea birds and
spawning and nursery areas of cod Gadus morhua) were included
in the index. Overall, there were 14 ecosystem components in the
index (Figs. 3 and 4, see details in Appendix B). All the distribution
maps have been published by HELCOM (2010b).
2.5. Weighting coefficients
A weighting coefficient is a constant, which was used to transform a pressure to a potential impact (Koskela, 2004; Halpern et al.,
2007; Vörösmarty et al., 2010). The weighting coefficient is specific
to any combination of pressures and ecosystem components. All
weighting coefficients were based on a wide questionnaire among
experts from all parts of the geographical area in question and from
different fields of environmental sciences. The scores were given on
a scale from 0 to 4, reflecting no impact (0), low impact (1), moderate impact (2), strong impact (3) or massive impact (4). When filling
the questionnaire, experts gave consideration to three aspects in
setting the score: recovery time of the ecosystem component after
the pressure (<1, 1–5, 5–10 or >10 years), resilience of the ecosystem component to the pressure (very high, high, moderate, low
resilience or vulnerable) and the functional effect (i.e. whether the
pressure affects one or several species, one or few trophic levels
or the whole community). An average of the three criteria was the
final weighting coefficient.
The weighting coefficients were provided by national experts
in six countries around the Baltic Sea (Denmark, Estonia, Finland,
Lithuania, Poland, and Sweden). In addition, the scores were discussed in an expert workshop, organized by HELCOM, and six
experts at the HELCOM Secretariat gave a seventh set of scores.
%
0
20
40
60
80
100
BOB
BOS
Zostera
GOF
Mussel
Phoc sand
Phoc so
NBP
Phoc hard
Aphoc sand
WGB
Aphoc so
Aphoc hard
Cod
EGB
Seal
Porpoise
BOR
Bird
ARK
KAT
Fig. 3. Presence of ecosystem components (benthic and water column biotope complexes, benthic biotopes and species-related data layers) in 5 km × 5 km squares.
Altogether 14 data layers were used (see text), but none of the squares contained
all the ecosystem components.
Fig. 4. Proportions of ecosystem components in the main sub- basins, defined as
their presence in all assessment units of the sub-basin. Since “photic water” covers
entirely and “non-photic water” almost entirely all sub-basins, they were omitted
from the graph. Sub-basins: Bothnian Bay (BOB), Bothnian Sea (BOS), Gulf of Finland
(GOF), Northern Baltic Proper (NBP), Western Gotland Basin (WGB), Eastern Gotland
Basin (EGB), Bornholm Basin (BOR), Arkona Basin (ARK) and the Kattegat (KAT). For
full names of the ecosystem components, see Section 2.
S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
109
In order to include the greatest number of estimates, the median
value of the expert estimates was chosen as the weighting coefficient for each P × E combination. Methods and summary results of
the questionnaire are presented in Appendix C.
2.6. Data handling
The data on pressures and ecosystem components were linked
to a grid of the assessment units. Altogether 19,276 assessment
units were calculated in the Baltic Sea area. Handling of the georeferenced data sets was done by the ESRI ArcGIS software, version
9.3, with the spatial analyst extension. In addition to the calculation
of the index, spatial regression of the impact data layers (P × E × in every cell) and the index result was performed by Ordinary Least
Squares (OLS) method (see description and examples in ESRI, 2009)
by analyzing data layers first one by one and then analyzing the
effect of removing all data layers with R2 values less than 0.05 from
the full model. R2 describes the percentage of variance explained
by the regression. In addition, the influence of ecosystem components and weighting coefficients were evaluated by two separate
runs of Geographically Weighted Regression (GWR; see description
and examples in Charlton and Fotheringham, 2009) with only one
explanatory variable in each model: the sum of normalized pressures or the full index with a constant weighting coefficient (value
2) instead of the ecosystem-specific coefficients.
3. Results
3.1. Spatial distribution of cumulative impacts
The cumulative impact values in the assessment units varied
between 6.8 (the Bothnian Bay) and 456.4 (the Sound) (Fig. 5),
the theoretical maximum being 2912, assuming that all weighting
coefficients and pressure values were maximal and all ecosystem
components were present in an assessment unit. The minimum
number of anthropogenic pressures (value > 0) in an assessment
unit was 14 (in the offshore Bothnian Bay), whereas the maximum
number was 35 (in the Belt Sea, south of Zealand).
The spatial presentation of the cumulative impacts showed that
the highest potential impacts on the Baltic Sea ecosystem take place
in the south-western sea areas (the Kattegat, Belt Sea, Kiel Bay
and Mecklenburg Bay), the Gulf of Gdansk and the Gulf of Finland,
whereas the least cumulative impacts were found from the Gulf of
Bothnia in the north (Fig. 5). The data indicated that the southern
areas were under types of pressures which are rare or non-existent
in the northern parts, such as bottom-trawling, large wind farms
and large-scale extraction of seabed resources. Also other forms
of commercial fishing were heavy in the southern sub-basins and
atmospheric deposition of heavy metals and nitrogen occurred predominantly in the southern areas. In the Gulf of Gdansk and Gulf
of Finland, the high sums of impacts were due mainly to riverine
pollution.
Another feature in the cumulative impact index was the difference between the open sea and the coastal areas. In coastal areas
all over the Baltic Sea, the multitude of coastal pressures – e.g.
fish farms, municipal waste water treatment plants, river estuaries,
industries, warm-water outflows from power plants, and coastal
structures – create a heavy burden on the marine environment.
The cumulative impacts were clearly higher in the coastal than the
pelagic areas in the Bothnian Bay, the Bothnian Sea and the Northern Baltic Proper. In other areas, cumulative impacts in the pelagic
zone were more or less as high as in the coastal areas. Moreover, the
index showed also high cumulative impacts in the vicinity of large
cities, such as Copenhagen, Gdansk, St. Petersburg and Stockholm.
Fig. 5. Presentation of cumulative potential anthropogenic impacts by the Baltic Sea
Impact Index in 5 km × 5 km assessment units. The index in each assessment unit
consists of the sum of anthropogenic impacts on selected ecosystem components
present in the unit. See the index formula in Section 2.2.
The surface and midwater trawling (incl. long lines), bottom
trawling and shipping exerted pressures on the open-sea areas and
large rivers such as the River Vistula, River Neva and River Göta
seem to affect wide open-sea areas, as a result of pollution by nitrogen, phosphorus and heavy metals (see Fig. 1 for rivers and coastal
cities). In the Kattegat and Danish Straits the impacts had very high
variation among assessment units, reflecting numerous spatially
restricted activities like dredging, sand extraction and disposal of
dredged material (Fig. 6).
3.2. The effect of data determining the index result
The pressures which have the greatest contribution to the final
index value are presented in Table 2. The sum of impacts of each
pressure type shows that the inputs of nutrients and organic matter,
inputs of hazardous substances as well as fishing have the greatest overall impact on the Baltic Sea ecosystem (25%, 30% and 30%,
respectively, Table 2). These pressures hence largely determine the
outcome of the map of potential cumulative impacts. According to
R2 , the highest-ranking data layers were those that exert pressure
over the whole Baltic Sea (e.g. atmospheric deposition of metals),
while the lowest ranks were given to the point impact data layers
(e.g. dredging, construction sites and harbours). Including only 21
impact data layers, which had R2 values over 0.05, resulted in the
R2 value of 0.99 and showed the minor influence of point data for
the full index. Thus, the geographical distribution of a pressure is
more important in forming the shape of the map than its value in
each cell.
The ecosystem components have a central role in the index
tool, because each added ecosystem component per assessment
unit introduces a full set of potential impacts to the index value.
The GWR analyses showed that the pressure data layers without
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S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
Table 2
Summary statistics of the pressures. Minimum, maximum and mean impact scores and the standard deviation of them in the assessment units as well as the sum of impacts
over the sea area are shown. The contribution of each of the pressures to the final index value is presented by coefficient of determination (R2 ). Each pressure has been
converted to impacts in an assessment unit (see the index formula in Section 2). Pressures are ordered by the sum of impacts in all assessment units in the Baltic Sea.
Pressure type and pressure
Min
Max
Mean
SD
Sum
R2
Full model
Extraction of species/mid- and surface water trawling + long lines
Inputs of nutrients/atm. deposition of N
Inputs of nutrients/waterborne N
Inputs of org. Matter/Riverine organic matter
Changes in siltation/organic matter
Non-synthetic substances/atm. deposition of Pb
Inputs of nutrients/waterborne P
Extraction of species/gillnet fishery
Extraction of species/bottom trawling
Non-synthetic substances)/atm. deposition of Cd
Non-synthetic substances/atm. deposition of Hg
UW noise/all shipping
Non-synthetic substances/waterborne Zn
Non-synthetic substances/atm. depos. of dioxins
Abrasion/bottom trawling
Non-synthetic substances/waterborne Ni
Extraction of species/Bird hunting
Non-synthetic substances/waterborne Pb
Extraction of species/trap and pot fishery
Extraction of species/seal hunting
Non-synthetic substances/waterborne Cd
Changes in siltation/coastal shipping
Synthetic compounds/population density
UW noise/recreational boating
Introduct. of pathogens/passenger ships
Inputs of radioactive subst./radioactive substances
Non-synthetic substances/waterborne Hg
Changes in siltation/beaches
Synthetic compounds/harbours
Synthetic compounds/oil spills
Sealing/harbours
Sealing/coastal defence structures
Introduct. Of pathogens/Waste water treatm. plants
Changes in siltation/Dredging
Extraction of non-living resources/dredging
Abrasion/dredging
Changes in salinity/waste water treatm. plants
Smothering/Disposal of dredged matter
Changes in salinity/coastal dams
Sealing/coastal dams
Inputs of nutrients/aquaculture
Inputs of org. matter/Aquaculture
Introduct. of pathogens/Aquaculture
UW noise/construction of cables
Smothering/construction of cables
Synthetic compounds/Industries
Changes in thermal regime/Nuclear power plants
UW noise/operational windfarms
Synthetic compounds/polluting ship accidents
UW noise/oil rigs
Smothering/construction of windfarms
UW noise/construction of windfarms
6.81
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
<0.01
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
456.44
26.81
27.26
25.54
24.08
22.34
24.17
24.79
25.14
35.01
18.37
21.86
22.82
20.38
20.62
26.09
19.65
18.07
19.71
19.26
12.55
18.64
17.77
17.51
17.53
15.84
10.53
14.90
13.65
25.40
16.13
20.55
24.77
13.04
20.22
19.11
18.81
13.31
19.02
20.00
17.80
22.50
22.50
8.35
25.25
24.20
16.72
18.20
12.20
16.64
11.35
14.50
13.60
111.52
10.04
9.99
8.78
7.56
7.36
7.22
6.83
6.15
6.04
5.82
5.06
4.09
4.00
3.86
3.85
3.19
2.27
2.17
1.61
1.42
1.04
0.92
0.36
0.31
0.31
0.28
0.21
0.17
0.17
0.16
0.17
0.13
0.09
0.08
0.08
0.08
0.08
0.04
0.04
0.03
0.03
0.03
0.02
0.02
0.02
0.02
0.01
<0.01
<0.01
<0.01
<0.01
<0.01
48.86
5.11
4.64
4.23
2.98
2.95
3.81
3.12
4.68
6.58
3.05
2.96
3.55
2.97
3.03
4.42
2.97
3.60
2.86
2.86
2.66
2.44
2.17
1.49
1.38
1.01
1.17
0.83
0.88
1.33
0.88
1.08
1.33
0.65
0.93
0.87
0.87
0.58
0.64
0.56
0.47
0.49
0.48
0.28
0.54
0.44
0.33
0.27
0.20
0.20
0.14
0.11
0.11
2,149,710
193,711
192,485
169,190
145,797
141,856
139,177
131,663
118,574
116,413
110,533
96,187
77,387
75,914
73,505
73,166
60,585
43,319
41,274
30,531
26,960
19,696
17,447
6986
5941
5937
5363
4084
3242
3220
3037
2577
2488
1638
1598
1482
1478
1439
812
674
553
523
512
399
389
314
313
97
86
69
37
22
20
1.00
0.36
0.73
0.65
0.50
0.49
0.78
0.48
0.46
0.34
0.77
0.62
0.37
0.11
0.48
0.38
0.05
0.18
0.14
0.13
<0.01
0.12
0.23
0.02
0.03
0.03
0.03
0.05
0.04
0.03
0.03
0.50
0.03
0.02
0.03
0.03
0.03
0.03
0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
<0.01
ecosystem components or weighting coefficients explained 48% of
the total variation in the impact index (Figs. 7 and 8). The rest of
the variation is caused by the ecosystem data and the ecosystemspecific weighting coefficients. The cartographic presentation of
the sum of normalized pressures (Fig. 3) shows a roughly similar
result as the results in the full impact index (Fig. 7). However, differences are visible in areas where there were only very few or a
large number of ecosystem components present in the model (cf.
Fig. 3).
The weighting coefficients were based on expert scoring and
were, thus, the only non-objective part of the index, their effect
on the index result was estimated by further testing (see Section
2.6). When adding a constant coefficient (value 2) to the index
and comparing this dummy index with the real index, the dummy
index explained 99% of the real index. The test showed that the
addition of the ecosystem components – without the variance from
the weighting coefficients – explained more of the real index than
the pressures alone. The remaining variance (1%) was explained
by the weighting coefficients and their effect on the impact map
was only on details at the local level. The test results are shown in
Appendix D.
4. Discussion
This study presented an assessment of cumulative potential
anthropogenic impacts in the Baltic Sea by an assessment tool
which uses three kinds of data on assessment units: anthropogenic
pressures, ecosystem components and weighting coefficients to
transform the pressures to impacts on each of the ecosystem components. The semi-enclosed nature of the Baltic Sea with its highly
S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
0
Mean index value per assessment unit
20
40
111
Mean of the total pressure
60
BOB
0
1
2
3
4
5
BOB
BOS
BOS
GOF
Fishing
GOF
Nutrient inputs
NBP
Non-synthecs
Synthecs
EGB
NBP
Physical damage
Physical loss
Organic maer
BOR
EGB
Fishing
Nutrient inputs
Hunng
Non-synthecs
Uw noise
BOR
Synthecs
ARK
Physical damage
ARK
KAT
Physical loss
Organic maer
Hunng
Fig. 6. Mean index values of main pressure types per assessment unit in the Baltic
Sea sub-basins. The pressures have been grouped to show the impact of various
human activities (see Tables 1 and 2, only the top nine pressure types are included).
See the sub-basin names in Fig. 4.
KAT
Uw noise
Fig. 8. Total anthropogenic pressure in the Baltic Sea sub-basins. The normalized
[scale: 0–1] pressure values have been first summed within an assessment unit
and then averaged per sub-basin. The pressures have been grouped according to
Tables 1 and 2; only the top nine pressure types are included. See sub-basin names
in Fig 4.
developed and industrialised catchment area was well reflected
in this assessment of cumulative impacts: the highest cumulative
impacts on the marine ecosystem were reported from the southern
and south-western sea areas, with the highest population densities of up to 500 inhabitants per km2 in the drainage area. In
comparison, the drainages of the northern areas with the lowest
cumulative impacts are very scarcely populated (2.0 inhabitants per
km2 ). Although the coastal areas always showed high cumulative
impacts, the offshore areas were in a number of sub-basins equally
impacted, reflecting the high fishing pressure, intensive maritime
traffic and large inputs of nutrients and hazardous substances from
the atmosphere and rivers.
4.1. Reported ecosystem impacts of the dominant pressures
Fig. 7. Presentation of the sum of normalized anthropogenic pressures. Each of
the 52 pressures is on the scale from 0 to 1. The assessment units are 5 km × 5 km
squares.
The Baltic Sea marine environment has been subject to scientific
studies for decades, but this is the first assessment of cumulative pressures and potential impacts in the region. Evaluation of
the reliability of the tool requires comparison of the results with
existing knowledge on the state of the ecosystem. In this study,
nutrient inputs were estimated to result in one of the highest pressures on the entire sea area. Recent studies have shown that the
Baltic Sea suffers from a severe eutrophication problem (HELCOM,
2009a; Lundberg et al., 2009; Andersen et al., 2010). According to a
recent assessment of the eutrophication status by the Baltic Marine
Environment Protection Commission (HELCOM, 2009a), 176 of the
189 studied sites were in an impaired state. The anthropogenic
inputs of nitrogen and phosphorus are high in the entire sea area as
112
S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
natural background inputs of nitrogen and phosphorus are estimated at only 17% and 16%, respectively, of the total inputs
(HELCOM, 2004).
According to our results, fishing exerts a high pressure on the
Baltic Sea ecosystem in all areas of the Baltic Sea. In this study
we used fishing data from the year 2007, and at that time both
eastern and western cod (Gadus morhua) stocks and three of the
five assessed herring stocks (Clupea harengus) were below safe biological limits due to overfishing and climatically induced changes
in salinity and oxygen concentrations (ICES, 2007). The decline in
cod stocks has led to an increase in sprat (Sprattus sprattus) stocks,
which has been interpreted as a regime shift from a top-down regulated food web to a resource-regulated one (Österblom et al., 2007;
Casini et al., 2008). The reproductive capacity of salmon spawning
rivers was mostly (84% of rivers) below the adequate level of river
reproductive capacity (75%) (ICES, 2007).
Bottom trawling was estimated to be an activity with a high
impact on the ecosystem. This was seen in the weighting coefficient estimations by experts (Appendix C) and in the index results
(Table 2, Figs. 7 and 8). The result is in line with assessments from
other sea areas globally (Jones, 1992; Collie et al., 1997; Watling and
Norse, 1998; Jennings et al., 2001). In the Baltic Sea, bottom trawling by the Nephros fishery in the Kattegat and Skagerrak region can
amount to a by-catch which is 50% of the biomass of the Nephros
catch and can include up to 24 non-target species in one catch
(Ottosson, 2008). Harbour porpoise, grey seal, ringed seal, harbour seal and seabirds have all been found drowned throughout
the Baltic Sea in drift nets, gillnets and trawls (Lunneryd et al.,
2004; ICES, 2008; ASCOBANS-HELCOM database, 2011). Abrasion
and resuspension by bottom-trawling have been estimated as particularly destructive in the Baltic Sea (Riemann and Hoffmann,
1991; Tjensvoll et al., 2009) and globally (Watling and Norse, 1998).
Inputs of hazardous substances, particularly metals, were estimated to be among the highest pressures in all sub-basins of the
Baltic Sea. Adverse impacts of persistent organic pollutants and
heavy metals have been found at all levels of the food web. Biomarkers on molecular, cellular and tissue level have shown reproductive
and other disorders in marine invertebrates and fish (Strand et al.,
2004; Broeg and Lehtonen, 2006), contaminant levels in biota are
high and populations of predatory species have only recently recovered from reduced reproductive success (Helander et al., 2002,
reviewed in HELCOM, 2010a).
4.2. The role of weighting coefficients in the index
The weighting of pressures had two overarching aims: to estimate the degree of destructiveness of pressures and to specify the
sensitivity of ecosystem components to these pressures. As most
of these ecosystem components were described by abiotic factors,
the weighting coefficients were estimated on the basis of the community or key species, which are typical to such a biotope complex.
Substrate type and light availability are key factors for benthic fauna
and flora (e.g. Al-Hamdani and Reker, 2007).
Estimations of weighting coefficients can be sensitive to subjectivity, as experts may judge impacts of pressures by different
means. In order to avoid false scorings the expert questionnaire
included examples and several of the experts participated in a
workshop. It is however obvious that a certain degree of subjectivity remains in the weighting coefficients. The analysis of the
effect of the three factors in the index showed that the pressures
give the major shape for the index, the ecosystem components
sharpen the shape and the weighting coefficients only added detail
at a local scale. Thus, the non-objective component of the index
did not have a ruling effect on the final impact index result on
the scale of an entire sea area. It is however obvious that in a sea
area with strong environmental gradients, like the Baltic, regional
variability in impacts may be significant. Therefore regionally
adjusted weighting coefficients might improve the accuracy of the
model. The effects of the weighting coefficients are further presented and discussed in Appendix D.
4.3. Strengths and weaknesses of the method
The method we used is based on a linear response of
ecosystem components to anthropogenic pressures. This is a simplification of reality where thresholds or other non-linearities
can be expected to occur. We think that these simplifications do not however affect the main message of the end
result, which shows a clear geographical gradient of potential
cumulative impacts in the Baltic Sea. And overall, our study
emphasizes the need for further study of pressure-impact relationships and mechanisms, many of which have not been evaluated
scientifically.
The selection of the ecosystem components in the index affects
the index result. Therefore the selection of ecosystem data has
been balanced among the different types of ecosystem components. By choosing only biotopes restricted to either coastal or
pelagic areas or data on only few species one may bias the index
results to certain geographical areas or parts of the ecosystem.
We have avoided that by selecting benthic and water column
biotope complexes which can be found both in the coastal and offshore areas. Our benthic biotopes (i.e. Zostera meadows and blue
mussel beds) were mostly from the coastal areas, but that was
balanced by including pelagic species-related data layers. Thus,
the high dependence of this index method on georeferenced data,
may – in data-poor areas – lead to poor results. In the Baltic Sea,
georeferenced data was, however, rather easily accessible due to
well established regional cooperation among the coastal countries.
In this study, the coarseness of some data layers was compensated by using a rather robust grid of 5 km × 5 km assessment
units.
The method has strengths which should be considered against
the weaknesses. There exists currently no other method which
gives an estimate of cumulative anthropogenic impacts and takes
account of ecosystem sensitivity. The recent global assessment
(Halpern et al., 2008) showed that the Baltic Sea was moderately
affected by the anthropogenic pressures, but being a global assessment it lacked regional data and therefore failed to show finer scale
variation within the region. By including Baltic Sea-specific pressure and ecosystem data we have shown that the tool can be used
to describe detailed differences at least at a basin-wide scale. The
tool has also been used for assessing other areas, such as in the
Eastern Pacific to map cumulative impacts in the California Current
(Halpern et al., 2009) and Canadian waters (Ban et al., 2010) and to
assess anthropogenic impacts on coral reefs in Hawaii (Selkoe et al.,
2009).
The cumulative impact index is a tool which can be applied
to marine spatial planning, environmental impact assessments,
placement and management of marine protected areas, permitting processes and also in environmental status, pressure and
impact assessments, such as those required by the EU MSFD.
The index can be used to evaluate each ecosystem component
separately, for example, by evaluating impacts of anthropogenic
pressures on blue mussel beds. Often such practical uses require
a special selection of data layers depending on the aim of the
pursuit. At smaller scales, the weighting coefficients may also
require adjustment to local conditions. The method can also be
used to show potential impacts with predicted pressure data;
for example with climate change or ocean acidification scenarios. Together with spatial models capable of predicting spatial
drifting or extension of pressures, the index tool is a powerful
method.
S. Korpinen et al. / Ecological Indicators 15 (2012) 105–114
5. Conclusions
The tool to assess cumulative anthropogenic impacts on the
marine environment shows that the entire Baltic Sea ecosystem is under the burden caused by anthropogenic pressures
potentially resulting in high cumulative impacts on the whole
ecosystem. The approach which was first developed for global
assessments was here adapted to a regional sea, and the adaptation required modifications in data resolution, terminology and
data management. The use of this tool provides a wide array of
possibilities for management and planning purposes from local
to regional scales, thus offering also possibilities for the EU
Member States to estimate the level of anthropogenic impacts
on their marine waters under the Marine Strategy Framework
Directive. Future assessments of cumulative impacts would benefit from a more detailed and quantitative understanding of the
pressure-impact relationships and mechanisms in the marine
environment.
Acknowledgements
We want to express our gratitude to the National experts in
Denmark, Estonia, Finland, Lithuania, Poland and Sweden for taking part in estimating the weighting coefficients and attending
the HELCOM Baltic Sea Pressure Index Workshop on 11 February
2010 in Stockholm, Sweden. Thanks are also addressed to Ms.
Ulla Li Zweifel (Dimedia) for the help in coordinating the expert
survey and to Mr. Martin Isæus (AquaBiota Water Research),
Mr. Henrik Skov (DHI) and Mr. Johnny Reker (Danish Agency
for Spatial and Environmental Planning) for some of the ecosystem data layers. Special thanks to Ciarán Murray for linguistic
corrections.
Appendix A. Supplementary data
Supplementary data associated with this article can be found, in
the online version, at doi:10.1016/j.ecolind.2011.09.023.
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Watling, L., Norse, E.A., 1998. Disturbance of the seabed by mobile fishing gear: a comparison to forest clearcutting. Conservation Biology 12,
1180–1197.
Chapter 15
Ecosystem health
Maria Laamanen, Samuli Korpinen, Ulla Li Zweifel and Jesper H. Andersen
Chapter summary
Humans have inhabited the Baltic Sea drainage area for several thousand years but it has
not been until in recent times that the impacts from human activities have surpassed the
natural dynamics of the ecosystem. Human induced degradation of the ecosystem health
accelerated after the World War II. According to overview assessments of ecosystem state
in the 2000s, human pressures and impacts can be found in all areas of the Baltic Sea. Due
to the human pressures most of the Baltic Sea area is a eutrophication problem area, it is
disturbed by hazardous substances, and has impoverished biodiversity status and an
impaired overall ecosystem health. The use of the Baltic Sea ecosystem has not been at a
sustainable level. This is evidenced also by the ecosystem regime shifts in the late 20th
century caused by fishing, hunting and eutrophication, as well as changes in climatic
conditions. Although some of the ecosystem impacts from human activities such as local
sewage pollution problems, contamination by DDT and polychlorinated biphenyls (PCBs)
have been salvaged, the widespread eutrophication with bottom hypoxia and shifts in
biodiversity are likely to pose much greater challenges. There is currently international
legislation and regional cooperation in force that has explicit aims and measures for a
healthier marine ecosystem.
15.1 Ecosystems are naturally dynamic
15.1.1 Natural fluctuations occur at various time scales
Ecosystem functions and processes are dynamic by nature. Time scales of ecosystem
processes can be geological, centennial, decennial, annual, seasonal or daily depending on
the process in question.
The Baltic Sea has a long history of interchanging marine and freshwater phases and the
salinity levels and climate conditions close to the current ones have existed only for about 3
000 years. The changes between marine and freshwater phases were driven by natural
changes in the morphology and depth of the Baltic Sea basin which regulated the
connectedness of the basin to the ocean and the levels of salinity. Four geological phases of
the Baltic Sea have been distinguished since the recent glacial each with their distinct fauna
and flora. The Baltic Sea today is a young ecosystem and it has a relatively simple ecosystem
structure and few species compared to some of the stabilised old marine ecosystems such
as tropical sea areas with coral reefs or temperate seas with kelp forests.
The northernmost areas of the Baltic Sea are still subject to land uplift after having been
depressed by the masses of ice for several thousand years during the recent glacial that
ended about 11,000 years ago in the Baltic region. In the north-eastern Bothnian Bay coast,
the rising of the Earth’s crust can be 9 mm per year while in some areas of the southern rim
of the Baltic Sea coast, this change is opposite and land is sinking approximately at a rate of
1 mm/year (Schmidt-Thomé 2006). Thus, the Baltic Sea coast line is subject to a constant
directional change.
During the last one hundred years the salinity, oxygen and temperature conditions of the
Baltic Sea have varied on a decadal time-scale (Winsor et al. 2001). These changes in the
hydrography in turn cause changes in the abundance and distribution of both pelagic,
benthic and littoral species and communities in the Baltic Sea (Alheit et al. 2005).
The climatically driven inflows of saline water from the North Sea are crucial for ecosystem
structure and functions and they act through changes of the physical and chemical
conditions of the water body. The saltwater pulses flow into the Baltic along its bottoms,
from one deep sub-basin into another, slowly mixing with the brackish water and
oxygenising the deep waters and sediment. Increasing salinity of the deep water also
strengthens the vertical physical structure of the water column and the physical barrier
across the halocline, diminishing vertical mixing in the water column, and in the longer run,
making the bottom water vulnerable to hypoxia (< 2 mg l-1 oxygen) and complete anoxia.
Hypoxic conditions are physiologically intolerable to most fish and invertebrate species and
are only inhabited by procaryotes such as sulphur bacteria. The spatial extent of anoxic
bottoms has varied over time and changes in the morphology and depths of the Baltic basin
and the sills have been the main triggers for millennial scale changes in hypoxia (Zillén &
Conley 2010).
On a decadal scale, periods without saline water inflows, so called stagnation periods, have
been observed in the Baltic from the late 1970s to the early 1990s. In the Baltic Proper, the
lower frequency of saline water inflows has tended to result in freshening of both the
surface and bottom waters, weakening and deepening the positioning of the halocline, and
most importantly relieving the hypoxia of the waters lying beneath the halocline (Conley et
al. 2002).
Fluctuations over the seasonal scale are also pronounced in the Baltic Sea. They are mainly
driven by changes in temperature and to some extent by changes in precipitation. These
changes cause seasonality in the physical structuring of the water column. Probably the
most pronounced seasonal phenomenon is the formation of a thermocline in spring which
evokes the spring bloom of planktonic algae and initiates the period of higher productivity
in the pelagic zone. The spring bloom is an important phenomenon because it serves the
benthic communities with energy by sedimentation of the dead organic matter after the
bloom. The thermocline breaks down in the autumn and is followed by a full turnover and
mixing of the water column down to the halocline level. In the wintertime parts of the Baltic
are covered by ice each year but the extent of ice varies greatly from one year to another.
There are also functions that fluctuate on a diel scale, such as vertical migration of
zooplankton or primary production of phytoplankton or macroalgae. These diel variations
are largely driven by changes in solar radiation and resulting light conditions underwater.
1
15.2.1 Ecosystem dynamics have various spatial scales
There are differences in the spatial scales of various natural phenomena and the scale is
related to the drivers of these processes.
Salt water inflows from the North Sea tend to have whole Baltic Sea scale influence although
their impacts in the northernmost sub-basins of the Gulf of Bothnia are less pronounced
and take place only after a time lag.
Seasonal changes influence the whole Baltic Sea but there are spatial differences in the way
these dynamics are expressed. The growth season is the longest in the south and shortest in
the north which is manifested by an earlier induction of the thermocline and onset of spring
blooms in the south than in the north as well as the later timing of the autumn turnover. At
a local scale, a sheltered shallow bay is likely to undergo spring phenomena earlier than the
open sea environment nearby because solar radiation will heat the shallow less turbulent
water of the bay faster than waters of the deeper, more wind exposed open sea nearby and
the bay may also be subject to inflowing fresh river waters that will further strengthen the
physical layering of its water column. Big inflowing rivers such as the rivers Neva, Oder or
Vistula tend to have sub-basin scale influence. Changes or perturbations in the drainage
area, such as flooding, often influence at the sub-basin scale through the discharge of river
waters.
The coastal zone tends to host ecological phenomena that function at smaller spatial scales
than those of the open sea. This is largely due to the mosaic-like geographical structure of
the coastal zone, providing to the biota a physically much more varied environment than
that of the open sea environment. As a result, the habitat complexity as well as species
diversity tends to be greater in the coastal zone than in the open sea.
15.2 Human pressures on the marine ecosystem
have a long history
15.2.1 Human settlement and evolving human activities
The human settlement in the Baltic Sea catchment developed as humans followed the rim of
the diminishing glacial ice shelf towards the north and inhabited the Baltic Sea basin
approximately 9,000 to 13,000 years ago. A population expansion took place in the Baltic
Sea drainage area between 1000 and 1300 AD increasing the human population from 4.6 to
9.5 million people. But it was not until between 1700 and present time that a great
population boom took place increasing the population in the drainage basin from about
14.4 to 85 million (Zillén & Conley 2010).
The sea has been a resource for coastal populations. Fisheries and hunting of seals and
harbour porpoises as well as water birds has always been part of the culture of the coastal
and archipelago communities. The increased number of humans in the catchment was
naturally accompanied by an increase in the use of land for agriculture and forestry through
land reclamation, and more recently by urbanization and industrialization. The human
2
activities on the catchment area have increased the inflow of soil particles, nutrients and
more recently of pollutants to the sea.
As long as there have been humans, the Baltic has also served as a transportation route and
connected peoples living on its shores. Just like the other human activities, also shipping
activity has intensified after the 1950s and today there are about 2000 ships plying the
Baltic Sea at any time point. Shipping results in losses to the water of nutrient rich sewage,
emissions to the air of nitrogen and sulphur, and it promotes the transfer to the Baltic Sea of
alien species from other sea regions. In coastal areas and shallower off-shore areas it causes
erosion and resuspension of bottom sediments. So far, catastrophic shipping accidents with
large-scale ecosystem effects have been avoided but pollution by petrogenic substances in
smaller scale has been common throughout the time of motorized ships.
After the World War II, technological development intensified the agriculture, forestry,
shipping, as well as fisheries to an extent unthinkable in the times when the use of the sea
was more subsistence orientated. The exploitation of the sea became industrialised. All of
this has also had drastic impacts on the marine ecosystem. Signs of deterioration were first
seen near the cities due to the pollution by untreated waste waters. Contemporary inputs of
total nitrogen (N) to the Baltic Sea have been estimated to be more than twice the amount a
century ago and total phosphorus (P) about three times the quantity a century ago
(Savchuk et al. 2008).
Anthropogenic climate change puts a further pressure on the Baltic Sea by changing the
physical setting of the sea and through that changing also the chemical and biological
features of the ecosystem. The air temperature in the Baltic Sea region has warmed by
0.08°C per decade on average during the period 1871–2004 (Heino et al. 2008). This trend
is slightly steeper than the trend in the global time series where there has been a warming
of the global climate system by 0.74°C degrees over the period 1906–2005 (IPCC 2007).
During recent decades, there has been a decreased frequency of salt-water pulses from the
North Sea into the Baltic Sea. The length of the ice season has decreased by 14–44 days
during the last century. These changes have had a measurable effect on the distribution,
reproductive output and stock sizes of the Baltic biota. However, it has not been possible to
establish a distinct causal link between these changes and anthropogenically induced
climate changes, partly because of the large natural climate variability, but also owing to
possible impacts from other human pressures (Dippner et al. 2008).
15.2.2 Baltic Sea regime shifts
Ecosystems tend to exhibit stable ecosystem states but if heavily stressed by external
factors they can undergo drastic changes that reorganise the ecosystem structure and
functioning (e.g. Scheffer et al. 2001). The causative external factors can be natural, such as
climatic fluctuations, or anthropogenic stressors, or combinations thereof.
A healthy ecosystem has an inherent capacity to buffer and counteract disturbances. This
capacity of an ecosystem to absorb change and to recover from it is called ecosystem
resilience (e.g. Elliott et al 2007). Biological diversity and variability in an ecosystem tends
to build up ecosystem resilience. As a consequence of resilience, ecosystems can appear
virtually unaffected and stable while exposed to considerable stress. The apparent lack of
3
response can be explained by natural
feedback mechanisms such as
biogeochemical compensation, regulation
through trophic and competitive
interactions within the system, and, to a
certain degree, also by the functional
diversity and redundancy among species.
Regime shifts take place when the capacity
of the ecosystem to absorb and buffer
external pressures is eventually exceeded,
in other words when the resilience of the
system has been surpassed. At a certain
point even a small increment in external
pressure can cause a shift that will result in
the collapse or dramatic change of
populations or other characteristics of the
ecosystem. An idiom that reflects this type
of event is “the straw that broke the camel’s
back”.
The stress level that pushes the system into
a change is often referred to as a threshold
or tipping point (Fig. 15.1B and D). Below
the threshold level one stable state prevails
and above the threshold another stable
state. Once a new state has been reached
the system has a tendency to be selfreinforcing as feed-back mechanisms start
stabilising the new regime. Ecosystem
changes can also be gradual with no
apparent thresholds or tipping points (Fig.
15.1A and C).
Figure 15.1 Conceptual models of the possible modes
In the Baltic Sea, it has been suggested that
several regime shifts have occurred during
the past 80 years, likely driven by
variability in climate, eutrophication, as
well as seal hunting and fishing pressure
(Österblom et al. 2007, ICES 2008). Two
shifts were observed as late as in the end of
the 1980s and the mid-1990s. In the central
Baltic Sea, The period before the first shift
was characterised by relatively high cod and
herring spawner biomass and recruitment,
and high abundances of the copepod
Pseudocalanus acuspes. The period after the
of responses of ecological status in regard to different
dose-response relationships. The dashed line indicates an
environmental target for ecosystem quality, presented as
an Ecological Quality Ratio, and the red arrows indicate
the estimated reductions in pressures needed to meet the
target. The reductions increase from scenario A to D
indicating that fulfilment of the target in non-linearly
responding systems with a threshold (scenario B) or a
shifting baseline (scenario D) or combinations hereof
(scenario D), e.g. caused by climate change, call for
reductions significantly larger compared to linearly
responding systems (scenario A). Figures are based on
Duarte et al. (2009) and Kemp et al. (2009) and the figure
has been created with the assistance of Janus Larsen,
NERI, Denmark.
4
second shift has seen a sprat dominance of the system with high abundances of Acartia spp.
and Temora longicornis zooplankton species.
The decrease in the populations of top predators like seals has modified the trophic
structure of the Baltic Sea food webs. Examples from around the world show that
decimation of top predators from the food webs through hunting and fishing have made the
marine ecosystems vulnerable to other threats like pollution by hazardous substances,
eutrophication and arrival of alien species (Jackson et al. 2001). At a larger scale, the
ecosystem has changed over the past one hundred years from an oligotrophic clear water
sea with abundant populations of top predators into a heavily eutrophicated sea area
burdened by pollution of hazardous substances and with an increasing quantity of alien
species.
15.3 Present anthropogenic pressures and state of
the ecosystem
15.3.1 Human pressures on the ecosystem
Numerous pressures from various types of human activities acted on the Baltic Sea marine
ecosystem in the 2000s (Table 1). These pressures tend to result in eutrophication,
pollution effects and deterioration of biodiversity and some of the single pressures may
have multiple and even synergistic effects. For example nutrient inputs cause
eutrophication as well as changes in biodiversity through, e.g. overgrowth of certain
opportunistic species. Overfishing, on the other hand, may exacerbate the effects of
eutrophication by changing the food web structure.
The impacts may be direct or indirect, acting through other effects. For example dredging
has a direct physical impact on the benthic environment harming organisms living in the
sediment. But dredging may also have an indirect effect of releasing contaminants buried in
the sediments to the environment (Table 1).
At present, commercial fishing, nutrient inputs, smothering, and pollution by hazardous
substances are the most adverse pressures acting on the Baltic ecosystem according to
expert views compiled by HELCOM (HELCOM 2010a, b). These views are in line with
studies on top pressures from marine regions in the world. Fishing, nutrient inputs and
inputs of hazardous substances are all intense activities with wide-spread impacts. In
addition they have significant impacts on many different ecosystem components.
The actual quantity of pressures on the ecosystem, such as smothering of the bottom
habitats, is in many cases difficult to measure. Therefore, human activities that act as
drivers of the pressures are measured instead, or other proxies are used. This approach was
used to estimate the quantity and distribution of all potential pressures in the Baltic Sea,
when HELCOM created a Baltic Sea Pressure Index, BSPI (HELCOM 2010b). In the Baltic
Sea, the sum of pressures tends to be the greatest in the eastern and southern areas of the
Baltic, especially in the Gulfs of Finland, Gdansk and Riga, as well as the Sound (Figure
5
Table 15.1 Human-derived pressures on the Baltic Sea marine ecosystem with their drivers (human activity)
and potential impacts on eutrophication (E), pollution effects by hazardous substances (HS) and biodiversity
(BD). x = direct impacts; (x) = indirect impacts. From HELCOM (2010a).
Pressure
Human activity, proxy or a direct measure of the
Potential
pressure
impacts
E
HS BD
Smothering
Smothering
Smothering
Sealing
Sealing
Sealing
Changes in siltation
Changes in siltation
Changes in siltation
Changes in siltation
Abrasion
Abrasion
Selective extraction
Underwater noise
Underwater noise
Underwater noise
Underwater noise
Underwater noise
Underwater noise
Changes in thermal regime
Changes in salinity regime
Changes in salinity regime
Introduction of synthetic compounds
Introduction of synthetic compounds
Introduction of synthetic compounds
Introduction of synthetic compounds
Introduction of synthetic compounds
Introduction
of
non-synthetic
substances and compounds
Introduction
of
non-synthetic
substances and compounds
Introduction
of
non-synthetic
substances and compounds
Introduction of radionuclides
Inputs of nutrients
Inputs of nutrients
Inputs of nutrients
Inputs of nutrients
Inputs of organic matter
Inputs of organic matter
Introduction of microbial pathogens
Introduction of microbial pathogens
Selective extraction of species
Selective extraction of species
Selective extraction of species
Selective extraction of species
Selective extraction of species
Selective extraction of species
Wind farms, bridges, oil platforms (construction phase)
Cables and pipelines (construction phase)
Disposal of dredged material
Coastal defense structures
Harbours
Bridges
Shipping (coastal)
Riverine input of organic matter
Bathing sites, beaches and beach replenishment
Dredging, sand, gravel or boulder extraction
Dredging, sand, gravel or boulder extraction
Bottom trawling
Dredging, sand, gravel or boulder extraction resulting in, e.g.,
habitat loss
Shipping (coastal and offshore)
Recreational boating and sports
Cables and pipelines (construction phase)
Wind farms, bridges, oil platforms (construction phase)
Wind farms (operational)
Oil platforms
Power plants with warm-water outflow
Bridges and coastal dams
Coastal wastewater treatment plants with freshwateroutlets
to the sea
Polluting ship accidents
Coastal industry, oil terminals, refineries, oil platforms
Harbours
Atmospheric deposition of dioxins
Population density (e.g., hormones and pharmaceuticals)
Illegal oil spills
(x)
(x)
x
(x)
(x)
(x)
(x)
(x)
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
x
(x)
(x)
x
Waterborne input of Cd, Hg and Pb
x
x
Atmospheric deposition of Cd, Hg and Pb
x
(x)
Discharges of radioactive substances
Waterborne input of nitrogen
Waterborne input of phosphorus
Aquaculture
Atmospheric deposition of nitrogen
Aquaculture
Riverine input of organic matter
Coastal wastewater treatment plants with outlets to the sea
Aquaculture
Bottom trawling (landings or catches)
Surface- and mid-water trawling
Gillnet fishery
Coastal stationary gear fishery
Hunting of seals
Hunting of birds
x
x
x
x
x
x
x
(x)
(x)
(x)
15.2A). The Bothnian Bay and Bothnian Sea as well as the northern Baltic Proper are the
areas with the lowest pressures. These trends can largely be explained by the higher
6
x
x
x
x
x
x
x
x
x
x
x
x
x
x
densities of populations in the catchments in the south and east than in the north.
Altogether 52 different pressures or activities were included in this analysis.
A
B
Figure 15.2 Sum of human pressures and their potential impacts in the Baltic Sea. Panel A: the Baltic Sea
Pressure Index (BSPI). Panel B: the sum of potential impacts from these pressures on the marine ecosystem
according to the Baltic Sea Impact Index (BSII), with smaller values (blue), indicating a lower and higher values
(red) a greater sum of pressures. From Korpinen et al. (2011).
Assessment of potential impacts on the ecosystem from these pressures requires
knowledge about the distribution of ecosystem components in addition to knowledge of the
distribution of the pressures. Even more important, it also requires knowledge or estimates
of the severity of the impacts from each human pressure or activity on each ecosystem
component. HELCOM derived these relationships, so called impact factors, through expert
estimates while creating the Baltic Sea Pressure Index and the associated Baltic Sea Impact
Index (cf. HELCOM 2010b). The Impact Index provides an estimate of the sum of potential
impacts from human pressures on the marine ecosystem and offers a way to present the
spatial distribution of these impacts.
In the Baltic Sea, the coastal areas host a larger number of ecosystem components than the
open sea and consequently these areas are also more vulnerable to human pressures
(Figure 15.2B, HELCOM 2010a). Also many of the local scale pressures, such as dredging
and construction of harbours or marinas takes place in the coastal zone in addition to larger
scale pressures. In the open sea, most of the impacts are related to fishing or airborne
deposition of nitrogen or hazardous substances.
15.3.2 Nutrient inputs and eutrophication
Nutrient loading to the Baltic Sea peaked in the 1970s to 1980s and today’s loads are lower
than those two decades ago (HELCOM 2011, 2012a). Between the years 1990 and 2006
7
phosphorus loads were estimated to have declined by 45% and nitrogen by 30% (HELCOM
2009a). In 2006, 638,000 tonnes of N and 28,400 tonnes of P still entered the Baltic Sea via
waterways (HELCOM 2011). In addition, 200,000 tonnes of airborne N was deposited onto
the Baltic Sea in 2006 (Bartnicki 2011). The fraction of P entering the Baltic Sea via
atmospheric pathways is usually minor.
These input levels are still far above the so called maximum allowable nutrient input levels
of the HELCOM Baltic Sea Action Plan (BSAP, see section 15.5.4). According to the BSAP,
601,000 tonnes of N and 15,300 tonnes of P are estimated to safeguard a Baltic Sea with an
acceptable level of eutrophication (HELCOM 2007a, Backer et al 2010).
Point sources, mainly municipal and industrial waste waters used to be the main sources of
phosphorus and a large source of nitrogen to the Baltic Sea but due to the improved waste
water treatment during the recent decades agriculture has taken over the role of the single
most polluting sector for both nutrients. In 2006, losses from diffuse sources were the main
origin of the excessive inputs of both nutrients (at least 45 % for both nutrients) and only
about 12 % of the nitrogen and only 20 % of phosphorus loads originated from point
sources (HELCOM 2011). Agriculture was the major sector contributing on average 60-90%
of the diffuse inputs of both nitrogen and phosphorus. In addition, scattered dwellings and
storm waters are also significant sources.
The main countries contributing to the waterborne nutrient input in 2006 were Poland
(24% of N, 36 % of P), Sweden (19% of N, 13% of P), and Russia (17% of N, 14 % of P)
(HELCOM 2011). Between 1994 and 2008 only Denmark showed significant decreasing
riverine loads, while Estonia and Finland had significant increasing riverine loads. The
greatest area-specific nutrient losses are found in catchment areas with high population
density, many industries and high agricultural activity, such as the south-western part of
the Baltic Sea catchment.
Eutrophication means an increase in the rate of supply of organic material to the water
(Nixon 1995). In the Baltic Sea, anthropogenic inputs of N and P have resulted in elevated
levels of both nutrients in the water. Excess nutrients and the subsequent accelerated
primary production cause the well-known eutrophication effects like increased biomasses
of planktonic as well as periphytic algae, increased frequency and intensity of algal blooms,
decreased water transparency, increased secondary production and upsurge in the biomass
of e.g. cyprinid fishes. In the end of the chain of effects is hypoxia and often even complete
anoxia in the bottom waters and sediments since decomposition by bacteria of the
increased organic matter consumes the oxygen (HELCOM 2009a).
It seems that the 1950s was the turning point in the rate of increase of nutrient loading and
human-induced eutrophication of the Baltic Sea (HELCOM 2012a). Actual long term
measurements going beyond the 1950s are few but acceleration of primary production is
reflected in the organic carbon accumulation rate over the last 40 years in the sediment
records from the Bornholm Basin and Gotland Deep (Struck et al. 2000). Increased
eutrophication during the 20th century is also reflected in the water transparency where
declines of summer Secchi depths have been observed in all sub-basins of the Baltic Sea
over the last one hundred years (Fleming-Lehtinen et al. 2010). The decrease has been most
8
pronounced in the northern Baltic Proper (from 9 to 5 m) and the Gulf of Finland (from 8 to
4 m). Eutrophication is also reflected in the increase of the total hypoxic (less than 2 mg O2
l-1) area from less than 10,000 km2 in 1900-1910 to 35,000 km2 in the 1950s and to around
60 to 70,000 km2 in the recent years (HELCOM 2012a).
In 2001-2006, most of the Baltic Sea area was affected by eutrophication (HELCOM 2009a).
Only 13 of 189 areas assessed by HELCOM had an acceptable eutrophication status. Those
areas were found in the open Bothnian Bay and the north-eastern Kattegat, as well as in
coastal sites in the Gulf of Bothnia. The Gulf of Bothnia is an area with the least human
pressures and the smallest population in the catchment area, while the western sea area,
the Kattegat, is influenced by the Atlantic waters (HELCOM 2009a). The eutrophication
state was evaluated through the use of data on indicators such as nutrient concentrations,
summer water transparency, chlorophyll a levels, maximum depth penetration of
macrophytes and species diversity of macrozoobenthic invertebrates in the deep bottoms
or the use of various macrozoobenthos indices in the coastal zone (HELCOM 2009a). Each
of the indicators had an associated target level indicating the border between an
“acceptable” and “non-acceptable” eutrophication status. These target levels assisted in the
integration of the indicator information into a eutrophication assessment for each
assessment area (cf. Andersen et al. 2010).
A particular challenge for remedial actions in the Baltic Sea is that eutrophication, through
increased sedimentation of organic matter, has contributed to extension of hypoxic areas,
as well as nutrient enrichment of the sediments (HELCOM 2009a, 2012a). When sediments
become hypoxic, phosphate is released to the overlaying water mass, primarily through
dissimilatory reduction of iron oxyhydroxides by bacteria (Conley et al. 2002). Phosphate
thus becomes available for growth of primary producers that thrive in nitrogen deficient
conditions. These kind of conditions in particular fuel the blooms of nitrogen fixing
cyanobacteria which not only are self-sufficient in terms of nitrogen but also can increase
nitrogen levels in the pelagic system. Thus, there are currently processes in place that act to
maintain the new eutrophic regime (Vahtera et al. 2007). In order to revert to a less
eutrophicated state it may necessary for the nutrient loads to be reduced to levels that are
lower than those that existed before the regime shift (Figure 15.1D).
15.3.3 Pollution and effects by hazardous substances
Chemical substances can be considered hazardous if they are toxic, persistent and
bioaccumulate, or if they are very persistent and bioaccumulate. In addition, substances
with effects on hormone and immune systems are considered hazardous.
The Baltic Sea has been exposed to an extensive use of chemicals from the very beginning of
the industrialization of the region in the late 19th century. Synthetic substances such as
persistent organic pollutants (POPs) and pharmaceuticals and non-synthetic substances
like heavy metals originate from:
1) point-sources of pollution situated on the coast or inland in the catchment areas
including industries and municipal wastewater plants;
2) land-based diffuse sources such as runoff from agricultural land, forests, and other land
uses as well as urban runoff, and leaching from dump sites and landfills;
9
3) activities taking place at sea such as shipping, pipelines, dredging or operation of oil
platforms; and
4) atmospheric deposition of contaminants from all types of combustion sources as well as
volatile chemicals (e.g., pesticides).
In the Baltic Sea region, industrial and municipal point sources have been most apparent
sources of pollutants but several contaminant groups originate mainly from minor
industrial sources, agriculture (e.g. pesticides, pharmaceuticals and fertilizers), households
(e.g. various consumer products), sludge, dump sites and waste deposition in landfills
(HELCOM 2010c). Atmospheric emissions come from land traffic, shipping, energy
production, incineration of wastes and even small-scale household combustion. For some
heavy metals (e.g. mercury, lead and cadmium), atmospheric deposition is a major
component of their annual inputs to the Baltic Sea and for substances such as dioxins
atmospheric deposition may dominate over other sources. These sources can be located
outside the Baltic Sea region and it has been estimated that 60% of cadmium, 84% of lead
and 79% of mercury deposited into the Baltic Sea originate from distant sources outside the
Baltic Sea catchment area (mainly the UK, France, Belgium and Czech Republic) (Bartnicki
et al. 2008). For dioxins this figure is 60 %.
Disturbance of wildlife by hazardous substances has been documented for a number of
trophic groups and the decline of grey and ringed seal populations has been attributed to
the increase of organochloride substances. Although the populations of grey and ringed
seals had already drastically declined due to hunting by the 1950s, the increasing pollution
in the late 1960s by polychlorinated biphenyls (PCBs) from pulp and paper mills and other
industrial sources caused a second decline which turned the populations to all-time lows
(estimated at 5,000 ringed seals and 3,000 grey seals) by the 1970s (Harding et al. 1999).
Organochlorides have been associated with reproductive failures in the seals. PCBs have
been banned and since the mid-1980s the reproduction of seals has normalised, but as a
very persistent contaminant they are still found in high concentrations in marine sediments
and biota.
Similar impacts have been shown on white-tailed eagle, mainly related to DDT, and other
fish feeding birds. The adverse effects of DDT on the eggshell production of predatory birds
and the consequent decline especially of white-tailed sea eagle populations in the northern
hemisphere in the 1950s to 1970s were well demonstrated. White-tailed eagle populations
recovered after the banning of DDT and PCBs and they, as well as smaller organisms, show
today less contamination by PCBs. However, new contaminants like the estrogenic
substances from municipal waste waters are causing new problems, such as feminization of
male fish (Ferreira et al. 2009).
In 1999–2007, the entire Baltic Sea was an area with a high contamination level. An
indicator-based assessment of 144 areas in the Baltic Sea revealed that 137 areas were
contaminated by hazardous substances (HELCOM 2010c). All open-sea areas of the Baltic
Sea, except the north-western Kattegat, were classified as being “disturbed by hazardous
substances”. The waters near larger coastal cities (Rostock, St. Petersburg, Helsinki, Gdańsk,
Riga, Copenhagen and Stockholm) were generally classified as having a “moderate” or
“poor” chemical status. Substances such as PCBs, lead, 1,1-dichloro-2,2-bis(p10
chlorophenyl)ethylene (DDE), cadmium, mercury, tributyltin (TBT), and dioxins as well as
brominated substances, for example, brominated diphenylether (BDE) appeared as
contaminants with the highest concentrations in relation to the threshold levels for
disturbance.
The above integrated status assessment did not address all hazardous substances that have
been measured in Baltic fish or sediments. Such substances include, for example, all
perfluorinated substances, alkyl phenyls, bisphenol A, and pharmaceutical substances. The
concentrations of these substances in the marine environment are high and increasing.
However, the understanding of their environmental fate is poor, and there is limited
information about their spatial distribution, main sources and transport mechanisms.
15.3.4 Changes in biodiversity
Diversity of species in the Baltic Sea is lower than diversity in most true marine areas. At
present, about 6,000 species are known from the Baltic Sea and the Kattegat, including
cyanobacteria, phytoplankton, zooplankton, phytobenthos, zoobenthos, fish, marine
mammals and birds, as well as vertebrate parasites. Most species are either fresh or marine
water species, and only one species, Fucus radicans, is a species endemic to the Baltic Sea
(Ojaveer et al. 2010). Many freshwater and marine species live at their physiological limits
in the brackish Baltic Sea. In addition, a low genetic diversity has also been uncovered in the
Baltic. The low species and genetic diversity indicates that the biodiversity in the Baltic Sea
may be particularly sensitive to disturbances.
The biodiversity of the Baltic Sea has undergone major changes during the past decades.
Especially, variations in climate, whether natural or anthropogenic, have introduced
physical and chemical changes and consequently also biological shifts (HELCOM 2009b).
While the lack of comprehensive data and the natural variability in biodiversity makes it
challenging to identify the precise role of human pressures, they have no doubt contributed
to the observed changes.
HELCOM has identified species with a threatened or declining population status (HELCOM
2007b). The only species known to be extinct is the sturgeon (Acipenser sturio) but all
marine mammals, except the grey seal north of 59° N latitude, are still under a threat or
decline. The population of harbour porpoises (Phocoena phocoena), is in a perilous state
with only a few hundred animals remaining in the Baltic Proper. The harbour porpoise was
widely distributed and common until the early 20th century. Previously, the main cause of
the decline was hunting but today the most important anthropogenic threats are incidental
fisheries by-catch and prey depletion (HELCOM 2009b). In addition, factors such as
hazardous substances and noise pollution are also likely to have negative impacts on the
species.
The grey seals (Halichoerus grypus) which were nearly hunted to extirpation in the
beginning of the 20th century, are now recovering in the northern Baltic Sea, while south of
59° N the recovery is still very slow. The status of the ringed seal (Phoca hispida) is also still
unfavourable. While the impacts of hunting and hazardous substances on seals have been
reduced, fisheries by-catch and prey depletion are among the most prominent and
continuing threats to these populations (HELCOM 2009b).
11
One of the most well-established changes in the Baltic ecosystem is the shift from
dominance of demersal fish communities to pelagic clupeid fishes. The eastern Baltic cod
(Gadus morhua) stock reached high abundances in the late 1970s and early 1980s (Eero et
al. 2007) while in the 1980s, a climate-induced decrease in the cod reproductive volume,
i.e., the amount of water with good conditions for hatching of cod eggs, caused high cod egg
mortality (Köster et al. 2003). This, together with very high fishing pressure, resulted in
historically low values of the cod stock since the early 1990s, however with some signs of
improvement in the 2000s (ICES 2008). The clupeids, sprat (Sprattus sprattus) but also
herring (Clupea harengus membras), on the other hand, have increased since the 1980s.
Factors underlying the increase of especially sprat are the decreased predation by cod and
possibly also eutrophication through an increase of food resources of the pelagic fish.
In many coastal areas, fish which benefit from or tolerate eutrophication, such as percids
and cyprinids have increased but in many other areas, fish stocks have overall declined
owing to high fishing pressure. Several stocks of migratory fish species are also in a poor
condition because of damming or blocking of migratory pathways or degradation of
riverine habitat quality.
Among the bird species, long-term population declines are evident for dunlin (Calidris
alpina schinzii), and more recently for common eider (Somateria mollissima) and long-tailed
duck (HELCOM 2009b). The causes are not well understood, but climate change in the case
of the dunlin, and shipping-induced oil spills, fisheries by-catch and habitat deterioration in
the case of the ducks may have contributed to the decline. A recent threat assessment of
Baltic breeding birds (HELCOM 2012b) records the gull-billed tern (Gelochelidon nilotica)
as regionally extinct. Moreover, the kentish plover (Charadrius alexandrines) is critically
endangered and dunlin, terek sandpiper (Xenus cinereus) and black-legged kittiwake (Rissa
tridactyla) endangered. The development of bird populations also includes positive signs
such as the recovery of the white-tailed eagle (Haliaeetus albicilla) and great cormorant
(Phalacrocorax carbo sinensis).
Macrobenthic soft-sediment communities are currently severely degraded and abundances
are below a 40-year average in the entire Baltic Sea (HELCOM 2009b). The increased
prevalence of oxygen-depleted deep water is perhaps the single most central factor
influencing the structural and functional biodiversity of benthic communities in the opensea areas of the Baltic Sea.
The zooplankton community has as already mentioned also experienced significant changes
over recent decades particularly in the offshore copepod communities in the Baltic Proper
and the southern Baltic Sea (HELCOM 2009b). Changes in salinity and temperature are
likely important factors underlying observed changes. In addition, eutrophication has
contributed to the decreasing volume of oxygenated water below the halocline in offshore
areas, thereby reducing the volume of water suitable for the reproduction of zooplankton
species that require higher salinities.
Nutrient enrichment has resulted in increased phytoplankton productivity, including more
prevalent algal blooms and which themselves can be considered as a manifestation of
reduced biodiversity within the phytoplankton community (HELCOM 2009b). A number of
12
changes in the community composition have also occurred during the past thirty years, e.g.,
a shift in dominance from diatoms to dinoflagellates during spring bloom periods.
In the macrophyte communities, important habitat forming species such as bladder wrack,
eelgrass, and stoneworts have decreased in abundance and distribution in many coastal
areas, including a decrease of the depth penetration of bladder wrack (Fucus vesiculosus)
(Kautsky et al. 1986). This has been primarily attributed to reduced water transparency.
The decrease in all habitat builders is most pronounced in highly polluted and eutrophied
areas as well as areas subject to physical disturbance to the bottom (HELCOM 2009).
Decline in habitat forming species has implications beyond the local scale since they are
important living, feeding, reproduction and nursing environments for associated flora and
fauna. More recently submerged aquatic vegetation is showing signs of recovery e.g., in the
north-western and north-eastern Baltic Proper.
Alien species are both a part of and a threat to the Baltic Sea biodiversity. Since the early
1800s, about 120 alien species have been introduced to the Baltic Sea including the
Kattegat. Most of the introduced species are crustaceans (23 species) and the main
introduction pathway has been shipping (HELCOM 2009b). The benthic bristle worm
Marenzelleria spp. presents one of the best documented cases of invasive species. It took
the species roughly ten years to spread to the entire Baltic Sea and to become common in
many soft-bottom habitats. Often the ecological impacts of alien species on the Baltic Sea
ecosystem have been difficult to observe or they are poorly understood. In the most heavily
invaded coastal lagoons of the southern Baltic, several food chains and even major parts of
sea-bottom communities may be based on introduced species (Leppäkoski et al. 2002). The
increasing spread of alien species increases the risk that native species or habitats of high
conservation value will eventually be impacted.
A pilot study using indicators of biodiversity and an integrated assessment approach was
carried out in 73 open-sea and coastal areas for the time period 2003–2007 (HELCOM
2010). According to these results, 82% of the coastal areas assessed had an unfavourable
conservation status and only 18% were in a ‘good’ or ‘high’ status. In terms of ecosystem
health, the deterioration of the status of biodiversity is critical because it diminishes the
resilience or buffering capacity against large-scale shifts in the Baltic Sea ecosystem and
increases the risk for escalating degradation of the environment (HELCOM 2009b).
15.3.5 Holistic assessment of ecosystem health
An initial holistic assessment of ecosystem health of the Baltic Sea in 2003-2007 (Figure
15.4) showed that the whole Baltic Sea marine area, except one site in the Gulf of Bothnia,
was classified as having a moderate, poor or bad ecosystem health (HELCOM 2010a). The
only areas close to having “good ecosystem health status” include the Bothnian Bay, the
Bothnian Sea and parts of the northern Kattegat. The Baltic Sea ecosystem is degraded to
such an extent that its capacity to deliver ecosystem goods and services to the people living
in the nine coastal states has been reduced. These results were based on chemical,
biological and supporting indicator data from altogether 84 assessment areas (HELCOM
2010a).
The state of open waters was in some areas worse than that of the coastal waters (Figure
15.3). In the Baltic Sea Pressure Index open sea areas especially in the southern and eastern
13
areas were under a high sum of pressures from a number of sources (Figure 15.2B). The
status of the ecosystem in the open sea was heavily affected by open sea fisheries and
inputs of nutrients and hazardous substances. In addition, the poor status of sediments of
the open sea areas resulting from the accumulation of hazardous substances or hypoxia was
in many cases the underlying factor behind the poor ecosystem health status.
Figure 15.3 Ecosystem health of the Baltic Sea in 2003-2007 assessed using biological, chemical and
supporting indicators. Blue colour indicates high status, green good, yellow moderate, orange poor and red bad
status in coastal (small dot) or open sea (large dot) assessment sites. From HELCOM (2010a).
15.4 Is the Baltic Sea a healthy ecosystem?
15.4.1 Concept of ecosystem health
Ideally, a healthy ecosystem is an ecosystem with full functionality and potential, and with
the absence of distress symptoms caused by anthropogenic stressors (Rapport 2007). An
14
analogy can be drawn between humans and ecosystems: just like an increase in human
body temperature or signs of inflammation are symptoms of human illness, declines in
populations of top predators, algal blooms or bottom hypoxia can be considered symptoms
of deteriorating health of a marine ecosystem. Ecosystem vitality on the other hand is
reflected in nutrient and energy flows, biodiversity and the capacity of the ecosystems to
rebound from perturbations (Rapport 2007).
Historical data probably provides the most appropriate means to evaluate whether
symptoms of deteriorating ecosystem health are present. It enables the analysis of trends in
ecosystem state and deterioration and the comparison of current ecosystem state to a
pristine or less impacted state in the past. Alternatively, pristine sites or areas within a
marine region can be used as reference sites for a healthy ecosystem. However, such
pristine sites are no longer found in the Baltic Sea.
This kind of an ecosystem health approach to human impacts was highly relevant at the
time when ecosystems were not yet complexly impacted and impacts from human activities
only started to become increasingly apparent. More recently, the attention has turned to
estimating how much anthropogenic impact to ecosystems is acceptable and how the
ecosystems can be restored if the boundaries of acceptable levels have been surpassed.
15.4.2 Sustainable use of the marine ecosystem
Today, humans are seen as an integral part of the ecosystems and negative impacts from
human presence and resource use are considered unavoidable and to some extent
acceptable. Desirable states of marine ecosystems are increasingly defined from the point of
view of human needs, and environmental protection is implemented through the concepts
of sustainable development and sustainable use of ecosystems.
The basis for the concept of sustainable development was laid already by the 1972 United
Nations Conference on the Human Environment. It brought to the wider global political
attention the need to protect the environment while pursuing development. The concept of
sustainable development was reinvigorated by the 1992 United Nations Rio Conference on
Environment and Development which emphasized the right to development of present and
future human generations, as well as the sovereign right of states to exploit their own
resources, but in such a way that environmental protection will constitute an integral part
of the development process. Sustainable development as it is interpreted today aims at
balancing between the current economic, social and environmental interests without
compromising the well-being of future generations.
Today the ecosystem approach and ecosystem-based management of human activities form
the basis of environmental policies related to the marine environments. The twelve “Malawi
principles of the ecosystem approach to biodiversity management” presented to the Fourth
Meeting of the Parties of the UN Convention on Biological Diversity in 1998 are often
referred to as the starting point for the ecosystem approach. These principles underline for
example that conservation of ecosystem structure and functioning is important, that
appropriate balance between conservation and use of biodiversity should be pursued, and
most importantly ecosystems must be managed within the limits of their functioning. Both
the HELCOM Baltic Sea Action Plan as well as the EU Marine Strategy Framework Directive
15
(MSFD), both of which are highly relevant for the Baltic Sea, are based on the ecosystem
approach to the management of human activities.
Good understanding of the ecosystem structure and functions and of the pathways and
mechanisms through which human activities impact the ecosystem are fundamental.
Ideally, the pressure – impact relationships would be understood and presented by
response curves (cf. Figure 15.1) which would guide managers to predict a sustainable
amount of pressure. However, currently we are far from such precise understanding and
marine assessments are forced to simplify the reality.
Defining what is a “healthy ecosystem” and sustainable level of use for a particular marine
ecosystem requires not only good scientific understanding of the ecosystem’s structure and
functioning and responses to anthropogenic pressures. More importantly it requires value
judgments. The advice from science is ultimately operationalized through political decisions
related to how much degradation of the environmental quality of the ecosystem the society
is ready to accept (e.g. Mee et al. 2008).
Recently, the concepts of ecosystem thresholds, resistance or resilience, as well as carrying
capacity have become centrepieces in defining acceptable levels of pressures and impacts
resulting from human activities (e.g. Elliott et al 2007). Due to the need to keep on the right
side of the pressure thresholds for regime shifts and the incomplete knowledge on the ways
on which the pressures act on marine systems, there is a need to increase knowledge on
how these concepts materialise in reality (Mee et al. 2008).
In the Baltic Sea, it took nearly four decennia from the recognition of the DDT problem to
the recovery of the white-tailed eagle populations through political decisions to ban the
substance (Elmgren 2001). Similarly, it took some decennia from the recognition of the
mercury and PCB contamination problem to the recovery but the politicians did not wait
until scientific consensus to emerge before they took decisions to ban those substances
(Elmgren 2001). Precautionary and adaptive management approaches have been called for
by the scientists when the not-yet-so-well-known environmental problems are to be
managed (Mee et al. 2008, Elmgren 2001). This means that the society should not wait until
scientific consensus emerges and act only then but instead take action already in the
presence of uncertainty.
15.4.3 Is the present level of human use of the Baltic Sea sustainable?
The Baltic Sea has not been managed in a sustainable manner and the use of the ecosystem
goods and services has surpassed the boundaries of sustainable use.
In the assessment of the sum of human pressures on the Baltic Sea, the whole Baltic Sea was
under some level of pressure from human activities and the pressures were greatest in the
eastern and southern areas (HELCOM 2010a, 2011). The thematic assessments showed that
the levels of eutrophication, pollution by hazardous substances and degradation of
biodiversity were unacceptable in most of the Baltic Sea area (HELCOM 2009a, 2009b and
2010c). Similarly, the holistic assessment of ecosystem health demonstrated that in only
one pocket in the coastal area of the northern Bothnian Sea a good ecosystem health status
could be found (HELCOM 2010a).
16
In addition to these assessments, the prevalence of regime shifts caused partly by
overfishing and eutrophication provides first hand evidence that the levels of sustainable
use of the ecosystem have been surpassed for the Baltic Sea ecosystem (Österblom et al.
2007).
It is likely that some ecosystem impacts can be reverted, whilst others cannot. Positive
examples include those from coastal areas where the introduction or upgrading of waste
water treatment has led to the improvement of the water quality and recovery of the
ecosystem status (e.g. Finni et al. 2001). Recoveries of the northern Baltic grey seal
populations, as well as the resurge of populations of great cormorants and white-tailed
eagles, after the relaxation of hunting and banning of the use of organochlorides DDT and
PCBs are also among the recovery success stories. Reverting to an acceptable
eutrophication status and restoring the Baltic Sea biodiversity will be much more
challenging if not unachievable since there are now feedback mechanisms in play such as
internal loading of phosphorus from sediments and prevalence of nitrogen-fixing
cyanobacteria that maintain the eutrophicated state and have led to the potentially
inhibited recovery (Vahtera et al. 2007). As concerns the biodiversity of the Baltic Sea, the
communities and food webs of the Baltic Sea have changed as a result of climatic changes,
fishing and eutrophication and the ecosystem has reorganised through regime shifts. Due to
the profound changes both in the underlying physical and chemical factors as well as the
populations of species, it is highly unlikely that it will be possible to return through active
management the Baltic Sea ecosystem to a state it had before heavy human influence e.g. in
the beginning of the 20th century.
Recent reports have demonstrated that nutrient loads to the Baltic Sea have declined
(HELCOM 2011) and that there are decreasing concentrations of nutrients in the sea, but
eutrophication indicators such as chlorophyll a levels still do not shown signs of decrease
(HELCOM 2012a). Overall, the return paths to healthier states of the ecosystem are still
largely unknown for the Baltic Sea (cf. Figures 15.1A to D). When considering
eutrophication, it is clear that due to the long memory of the system, e.g. long residence
time of water, there are time lags involved in the recovery and the recovery will not result
in a similar ecosystem that was present before large scale eutrophication. This will result in
a recovery trajectory with a shifted baseline (Fig. 15.1B and D).
Even if there is no return to a previous state, there is the obligation in the current policies to
manage the ecosystem prevailing physical and climatic conditions and the requirement to
ensure that the thresholds of current regime will not be exceeded.
15.5 Marine ecosystem health objectives in
international policies
15.5.1 The aim: “Good environmental status”
At the global level, the 1982 UN Convention of the Law of the Sea (UNCLOS) determines that
states have the obligation to protect and preserve the marine environment but UNCLOS
does not define more specific goals for the desired state.
17
At the European level the EU Marine Strategy Framework Directive (MSFD) adopted in
2008 is the most explicit piece of EU environmental legislation addressing the status of
marine environments (Anon 2008a). The MSFD aims at measures that will yield “good
environmental status” of the marine waters. It gives a description of what is a “good
environmental status”, a concept that also encompasses ecosystem status. The MSFD
stipulates that "good environmental status" means the environmental status of marine
waters where these “provide ecologically diverse and dynamic oceans and seas which are
clean, healthy and productive within their intrinsic conditions, and the use of the marine
environment is at a level that is sustainable, thus safeguarding the potential for uses and
activities by current and future generations”. The MSFD further reads that good
environmental status is achieved when: “The structure, functions and processes of the
constituent marine ecosystems, together with the associated physiographic, geographic,
geological and climatic factors, allow those ecosystems to function fully and to maintain their
resilience to human-induced environmental change. Marine species and habitats are
protected, human-induced decline of biodiversity is prevented and diverse biological
components function in balance. Hydro-morphological, physical and chemical properties of
the ecosystems, including those properties which result from human activities in the area
concerned, support the ecosystems as described above. Anthropogenic inputs of substances
and energy, including noise, into the marine environment do not cause pollution effects.” This
definition has been complemented by eleven qualitative descriptors of good environmental
status (Annex 1 of the Directive), and 26 criteria and 56 sub-criteria that should be
considered when assessing the state of the sea (Anon 2010).
The Directive obliges the member states to develop and implement marine strategies which
aim at reaching the good environmental status.
15.5.2 The aim: “Good ecological status”
The EU Water Framework Directive (WFD) addresses primarily fresh water environments
but also river mouths and marine coastal waters extending one (for ecological status) or
twelve (for chemical status) nautical miles seawards from the baselines, and sets an
objective to maintain or move towards “good ecological status” (Anon 2000). The directive
stipulates what should be assessed when evaluating the status, so called “biological quality
elements”, and provides normative definitions by words for what is a good ecological
status. For coastal waters, factors such as composition and abundance of phytoplankton,
other aquatic flora and benthic invertebrate fauna need to be assessed together with
supporting hydromorphological, chemical and physico-chemical elements. In general, good
status for these quality elements is described as a status with only slight changes caused by
human pressures from the reference conditions that are defined for different water-types
and for each quality element.
15.5.3 The aim: “Favourable conservation status of species and habitats”
The UN Convention on Biological Diversity (CBD) is the most important agreement
addressing biodiversity protection at the global level. Global targets have been set for
improving the state of biodiversity, including marine biodiversity, under, e.g. the so called
Aichi biodiversity targets.
18
At the European level, Habitats and Birds Directives have set as an objective to reach
favourable conservation status of species and habitats listed in these directives (Anon 1992,
2009). Some of the species and habitats on these lists are also found in the Baltic Sea, such
as species of marine mammals, seabirds and underwater marine habitats like reefs,
estuaries and lagoons.
According to the Habitats Directive, the conservation status of a natural habitat is
‘favourable’ when its natural range and areas it covers within that range are stable or
increasing, the specific structure and functions which are necessary for its long-term
maintenance exist and are likely to continue to exist for the foreseeable future, and the
conservation status of its typical species is favourable. The conservation status of species is
in turn ‘favourable’ when population dynamics data on the species concerned indicate that
it is maintaining itself on a long-term basis as a viable component of its natural habitats, and
the natural range of the species is neither being reduced nor is likely to be reduced for the
foreseeable future, and there is, and will probably continue to be, a sufficiently large habitat
to maintain its populations on a long-term basis.
Furthermore, the Habitat Directive requires the designation of a network of protected
areas, labelled Natura 2000 that hosts the habitats and species listed in the annexes of the
Directive.
The Birds Directive is the oldest EC nature conservation directive and aims at protecting
naturally occurring bird species in Europe. It applies to birds, and their nests and eggs, and
particularly focuses on the protection of habitats for endangered as well as migratory
species. This directive also requires the designation of protected areas that nowadays are
part of the Natura 2000 network.
15.5.4 Regional cooperation for the better health of the Baltic Sea
The global or European policies and their environmental objectives do not specifically
address the Baltic Sea ecosystem. However, their implementation requires interpretation
and application at the Baltic Sea level so that we know what the jurisdictional language
means when the species, habitats and environmental conditions of the Baltic Sea are in
focus.
Governments of the Baltic Sea coastal countries have worked together to protect the marine
environment of the Baltic Sea since 1974 when the Convention for the Protection of the
Marine Environment of the Baltic Sea was first signed. Today Helsinki Commission
(HELCOM), which is the governing body of the Convention, gathers together the coastal
countries and the EU, represented by the European Commission, to discuss and agree on
what are the goals and measures needed for a healthy Baltic Sea. HELCOM also provides a
platform for regional coordination: to consider how the EU Directives’ jurisdictional
language is applied when it comes to the ecosystem of the Baltic Sea and how the
implementation of the legislation can be coordinated at the Baltic Sea level.
The Baltic Sea Action Plan (BSAP) was adopted in 2007 by the ministers and high-level
representatives of the HELCOM Contracting Parties (HELCOM 2007, Backer et al. 2010).
19
With the BSAP the coastal countries of the Baltic Sea committed themselves to reaching a
healthy Baltic Sea ecosystem by year 2021 through applying the actions outlined in the
Action Plan. The ultimate vision of HELCOM is to reach a healthy Baltic Sea environment,
with diverse biological components functioning in balance, resulting in a good
environmental/ecological status and supporting a wide range of sustainable human
economic and social activities. The plan focuses on reducing eutrophication, contamination
by hazardous substances, enhancing biodiversity and nature conservation, as well as
securing environmentally friendly maritime activities. For reducing eutrophication, the
BSAP contains a full-fledged critical load approach with maximum allowable loads of
nitrogen and phosphorus defined for each Baltic Sea sub-basin and country based on
ecological targets for eutrophication ecological modelling. The maximum allowable inputs
are accompanied by provisional nutrient load reduction requirements for each coastal
country. This scheme is the first internationally agreed targets-based nutrient load
reduction scheme for an entire regional sea (Backer et al. 2010).
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