EUROPEAN SURFACE WATERS Review Paper

Assessment of the ecological status of European surfacewaters: a work in progress

Peeter Noges Æ Wouter van de Bund ÆAna Cristina Cardoso Æ Angelo G. Solimini ÆAnna-Stiina Heiskanen

Published online: 29 July 2009

� Springer Science+Business Media B.V. 2009

Abstract The ‘normative definitions’ of ecological

water quality classes given by the Water Framework

Directive (WFD) are narrative descriptions of the

conditions present in water bodies of different

qualities relative to reference conditions found in

unimpacted sites. In order to fill these descriptions

with a more solid content, the definitions have been a

subject of intensive development of quantitative

methodologies for ecological status assessment as

well as for rules and criteria for setting of reference

conditions and ecological status boundaries for

classification of water bodies. In this article, we

recall the basic principles of the WFD that sometimes

have been overlooked and point out some gaps

remained and problems arisen during the ongoing

implementation of the directive. Defining type-spe-

cific reference conditions for water bodies and finding

biological metrics that sensitively reflect only the

anthropogenic deviations from those conditions are

the biggest challenges that the ecological status

assessment faces. So far, there is no guarantee that

reference conditions are comparable across EU

Member States due to a lack of common criteria,

which need still to be elaborated. Defining site-

specific reference conditions instead of type specific

is a novel approach that allows for minimizing

uncertainties introduced by applying broad types.

Search for new metrics has led to a real boom of

multimetric indices which ought to be the adequate

tool to measure the multiple human impairments but

which should pass a thorough check before being

included in monitoring programs. Curiously, some

biological indices constructed as surrogates for

chemistry (especially nutrients) start ‘living their

own life’ and continue indicating the disturbance

when the controlling factors change. This shows the

obvious advantage of biological indicators against

chemical ones. New challenges to WFD implemen-

tation are brought about by the need to consider the

Guest editors: P. Noges, W. van de Bund, A. C. Cardoso,

A. G. Solimini & A.-S. Heiskanen

Assessment of the Ecological Status of European Surface

Waters.

P. Noges (&) � W. van de Bund � A. C. Cardoso �A. G. Solimini � A.-S. Heiskanen

Institute for Environment and Sustainability, Joint

Research Centre, Via E. Fermi 2749, 21027 Ispra,

VA, Italy

e-mail: peeter.noges@emu.ee

P. Noges

Centre for Limnology, Institute of Agricultural and

Environmental Sciences, Estonian University of Life

Sciences, 61117, Rannu, Tartumaa, Estonia

A. G. Solimini

Department of Experimental Medicine, Sapienza

University of Rome, Viale Regina Elena 324,

00161 Rome, Italy

A.-S. Heiskanen

Marine Research Centre, Finnish Environment Institute

(SYKE), Research Programme for the Protection of the

Baltic Sea, P.O. Box 140, 00251 Helsinki, Finland

123

Hydrobiologia (2009) 633:197–211

DOI 10.1007/s10750-009-9883-9

effects of alien species and climate change in the

assessment framework, and by the nonlinear dose—

response relationships dominating in biological sys-

tems. Attempts to diminish uncertainties in quality

assessment have become a new labour-intensive field

for researchers.

Keywords Water Framework Directive �Intercalibration � Multimetric index �Biological indicators � One out/all out principle

Introduction

The launch of the Water Framework Directive (WFD;

Directive, 2000) to provide a holistic framework for

management and protection of all waters in the

European Union (EU), was parallel to the develop-

ment of the Ecosystem Approach (EA), which was

adopted as the primary framework for action under

the convention of the biological diversity (CBD;

UNESCO, 2000) for management and conservation

of nature. The WFD follows many principles of the

EA which is a holistic framework for protection of

biodiversity and for the sustainable management of

natural resources in the context of the ecological,

socio-economical and cultural aspects. The WFD

includes management at the lowest appropriate level

(river basin management) in accordance with com-

mon principles and environmental objectives, stake-

holder involvement in the definition of management

objectives, restoration goals and planning of mea-

sures. It requires initial characterization of water

bodies, and analysis of human pressures impacting

the aquatic ecosystems, assessment of ecological and

chemical status, and economic analysis of cost-

effective protection measures. At present, the

recently adopted EU Marine Strategy Framework

Directive (Directive, 2008) is following the principles

of EA, while providing less detailed definitions on the

environmental objectives and criteria for the good

environmental status than the WFD. This may be an

advantage as it allows more flexible approaches that

can be adapted for regional seas that are different in

many ecological features, but on the other hand, it

may also make diversification of approaches in the

setting of ‘good environmental status objectives’

between the regional seas.

The WFD implements the principles of the EA

requiring management and protection of the aquatic

ecosystems as whole, ensuring sustainable function-

ing and healthy structure of the ecosystems, as well as

acknowledging the critical role of human societies in

managing those systems. However, contary to that

stated in the CBD definition of EA: The objectives of

management of land, water and living resources are a

matter of societal choice (UNESCO, 2000), the WFD

requires definitions of criteria and standards for good

ecological status that is the objective for protection

and restoration of water bodies. The interpretation

and application of Article 4 in the Directive, which

describes the environmental objectives and sets out

the so-called exemptions, does not compromise the

rule that the criteria for ‘good status’ are based on the

scientific knowledge of the aquatic ecosystems and

set by the intercalibration exercise between the

Member States (CIS, 2009). However, there may be

differences in views of the level of ambition how the

criteria for good ecological status have been defined

and applied in practical implementation of the

Directive (e.g., Moss, 2008), and, of course, expert

opinions are necessary in all stages of the assessment.

These criteria (i.e. the ‘normative definitions’) are

described in narrative terms in the Annexes of the

Directive, and have been a subject of intensive

development of quantitative methodologies for eco-

logical status assessment as well as for rules and

criteria for setting of reference conditions and

ecological status boundaries for classification of

water bodies. In this article, the authors recall the

basic principles of the WFD, which sometimes have

been overlooked or misinterpreted and point out some

gaps remained and problems arisen in the ecological

status assessment during the ongoing implementation

of the directive.

Ecological status and reference conditions

The Ecological Status assessment of the WFD

combines information on several hydromorphologi-

cal, chemical and biological parameters to acquire a

comprehensive picture of the overall status on

functioning and structure of the ecosystem. Hydro-

morphological (e.g. depth, flow rate, salinity, sedi-

ment types, etc.) and some chemical parameters (e.g.

alkalinity) are used to identify water body types,

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which are sufficiently similar while chemical param-

eters (e.g. nutrient and oxygen concentrations) are

supporting parameters that are applied to identify

reference conditions and good status (CIS, 2003a, b).

Biological parameters are the basis for the eco-

logical status assessment. WFD requires that the

status assessment of the water bodies is made based

on biological ‘quality elements’, such as phytoplank-

ton, macrophytes and phytobenthos, macroalgae and

angiosperms, macroinvertebrates and fish, combina-

tion of which is depending on the surface water

category (e.g., fish is not a quality element in coastal

waters, but required for all other surface waters

including transitional waters). The initial status

assessment should be based on all biological quality

elements required for each specific water category to

obtain an overall view of the status of biological

communities. Biological quality elements express

also the sensitivity to different pressures. Commonly,

phytoplankton is reacting first to human impact and

provides a fast response to, e.g. eutrophication

pressure, while macrophytes and fish may be more

sensitive for various degrees of hydromorphological

pressures (e.g., Solimini et al., 2006).

The selection and development of indicators

representative for the biological quality elements

are crucial in the assessment process. For instance,

chlorophyll a measurements are widely used as a

proxy for phytoplankton biomass. Moreover, also the

way of calculation of the metrics is important to

provide required robustness to the indicator. For

instance, if seasonal monitoring is carried out for

phytoplankton chlorophyll, the indicator calculated

should make use of all available data instead of

reducing datasets for summer period only, unless it is

proven that a more precise indicator can be obtained

from a subset of data (Carstensen, 2007). In the case

of coastal phytoplankton, either the 90th percentile or

the summer mean are most widely used as metrics for

chlorophyll a (Carletti et al., 2009). The robustness of

the selected indicator metrics is crucial, since several

biological parameters with a large natural variability

cannot be used for ecological status assessment as it

is not possible to identify reference conditions for

them with sufficient accuracy.

One of the big challenges of the WFD is to find

common approaches for defining reference conditions

that act as the anchor point for the ecological

assessment. A very purist approach—defining

reference conditions as the total absence of any

anthropogenic influence on the ecosystem—would

leave us with no real reference sites at all. This

realization could lead to the other extreme where the

concept of reference conditions is abandoned and

replaced with a focus on ecosystem services and

human benefits (Dufour & Piegay, 2009). In practice,

however, compromises between these two extreme

approaches are usually found, e.g. if reference condi-

tions are based on measurements from reference sites

with a minimum level of anthropogenic pressure. The

weak point so far is that there is no guarantee that

reference conditions are comparable across Member

States due to a lack of common criteria; further study

will be needed to resolve this issue. Setting reference

conditions is especially difficult for those water body

types lacking credible natural reference sites, includ-

ing lowland rivers, large rivers and lakes, and certain

types of estuaries and coastal waters.

As the driver or pressure is diminishing, the

recovery of ecosystems may not follow a similar

pathway and to return linearly to the initial state that

prevailed before the pressure increase. In a recent

article, Duarte et al. (2009) describe four coastal case

studies where nutrient loading reductions did not

bring a decline in phytoplankton biomass to about the

level before nutrient enrichments. Changes in the

species composition, internal ecosystem processes

and food web interactions may have caused a shift in

the ecosystems and the restoration of the previous

stage may not be possible, but after the pressure relief

a new alternative steady state will emerge. They

raised a question whether static reference conditions

are justified or instead shifting base lines should be

accepted to provide realistic prospects and public

motivation for restoration and ecosystem recovery.

Similarly, the changing and warming climate is

affecting many characteristics of surface waters. As a

result, some water bodies, especially those located

near the type boundaries, may even change their type.

Compared to typology characteristics, water quality

parameters and biological indicators are even more

labile and may be easily affected by climate change

(Noges et al., 2007). The review of the river basin

characterization every sixth years, as required by the

WFD, needs also encompassing re-evaluation of

reference conditions according to the changes

observed using potential reference site monitoring

networks, or modelling tools that enable distinction

Hydrobiologia (2009) 633:197–211 199

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of climate change effects. As a consequence, the

restoration targets (i.e., the good ecological status)

would also need to be evaluated periodically, accept-

ing that some changes (as loss or immigration of

some species, or permanently elevated biomass levels

of phytoplankton being fed by internal sources of

nutrients) have become permanent.

Intercalibration exercise

There is a large variety of indicator metrics, both single

and multimetric indices, which are developed for

different water categories, ecoregions and surface

water types and applied by the different EU Member

States. The comparativity of such metrics and classi-

fication results based on these indicators is one of the

key questions at the EU level. The WFD intercalibra-

tion aims to provide a process for comparison of

assessment results on quality element or on single

parameter level. In many cases the assessment meth-

ods were still in development at the time of the

intercalibration exercise that had to be carried out

based on a limited amount of available data. Results of

the intercalibration exercise have been achieved for

the most widely used methods (Carletti et al., 2009;

Poikane, 2009; van de Bund, 2009), but has not been

finalized; some of the more challenging quality

elements are yet to be addressed. Because the focus

of existing and often more traditional methods with the

results that have been laid down in a legal document,

there is the risk that intercalibration will limit the

development of new and innovative approaches. This

leaves a considerable challenge for the scientific

community to develop methods that are ecologically

sound and to show how they can improve current

methodologies. On the positive side, the intercalibra-

tion exercise has catalysed an exchange of information

in the expert groups, leading to more comparable

approaches in ecological quality assessment.

Biological elements

Annex V of the WFD sets out quite clearly the

biological elements that need to be monitored for

freshwater and coastal ecosystems. However, it

allows the elements with high natural variability to

be excluded from classification schemes. This

exemption together with the reduced probabilities of

a correct classification with increasing number of

elements used, have been considered responsible for

a restriction to the range of biological elements

sampled in routine monitoring in comparison with the

Annex V shopping list (Irvine, 2004). Indeed, a

Commission Decision on Intercalibration (European

Commission, 2008) published the classification cri-

teria for only a fraction of the required biological

elements and parameters (e.g., taxonomic composi-

tion, abundance) within the elements. This may not

only reflect difficulties associated to the development

of biological assessment methods but also a cautious

approach to the inclusion of several biological

elements in classification.

Not only the number of biological elements but

also the choice of the elements included in the

Directive has been the subject to criticism. The

omission of zooplankton among the biological ele-

ments for the assessment of lakes and coastal areas

has been particularly criticized (e.g. Moss, 2007)

given the important role of this group in reflecting

food web efficiency (e.g., Calbet & Landry, 2004;

Jeppesen et al., 2004). In particular, zooplankton

plays a key role in shaping shallow lakes’ food

webs—whereby in clear water state these lakes are

dominated by submerged macrophytes—used by

zooplankton as refuge against planktivorous fish

(e.g. Schriver et al., 1995; Stansfield et al., 1997),

and high zooplankton/phytoplankton ratio suggesting

zooplankton control on phytoplankton biomass (e.g.

Schriver et al., 1995; Søndergaard & Moss, 1998).

Another important subject to criticism regards the

exclusion of fish from the coastal waters assessment.

The role of fishes in the food web structure and

ecological process in coastal water is well known.

The magnitude of the direct pressure on this element

(i.e. through fishing) and its consequent cascade

effects on food webs (e.g. Daskalov, 2002) is also

well documented. Nonetheless, the Directive allows

for some flexibility in inclusion of other biological

elements, if their importance for the ecological

assessment is justified as it allows for exemptions

to mandatory biological elements. However, the

already significant monitoring efforts may discourage

Member States from including elements other than

those directly prescribed by the WFD.

Another gap regards the assessment of ecosystem

impacts from alien species (Cardoso & Free, 2008).

Although these species are not specifically mentioned

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123

in the text of the WFD, they represent a significant

biological pressure as many of them have an invasive

behaviour and can alter the native biological structure

and ecological processes of aquatic systems (e.g.

Olenin et al., 2007). According to the WFD Annex II,

alien species can fall in the category of ‘other

pressures’. Whereas, in the Directive’s Annex V, the

definition of ‘high status for each biological quality

element’ states the taxonomic composition corre-

sponds totally or nearly totally to undisturbed

conditions, therefore the departure from naturalness

underlines the entire ‘spirit’ of the WFD, and the

presence of alien species indicates such a state. Thus,

alien species simultaneously indicate high pressure

and poor or bad ecological status and should be

considered in ecological assessment tools.

In reality, alien species are a widespread problem

affecting all biological quality elements but water

body management is based on the assumption that

ecosystems naturally improve following reductions in

human impacts. However, alien species are a self-

perpetuating pressure which is very difficult to

eradicate once established, and ecological status

declines accordingly. This problem is further aggra-

vated because the impacts of alien species are not

reliably measured by the classification tools that have

been developed for the WFD. For example, inverte-

brate kick sampling does not adequately collect

species such as crayfish and family level identifica-

tion may fail to detect some alien species.

Metrics

(i) Type-specific or site-specific approach?: The

efficiency of water body assessment depends most

crucially on two aspects: (1) the appropriate typology

that adequately takes into account regional geo-

graphic and climatic differences and guarantees the

correct application of reference conditions, and (2)

the selection of right metrics that change quantita-

tively and consistently across a range or gradient of

human influence.

As shown by Borja et al. (2009), WFD typology

schemes have often been criticized for being over-

simplified and not reflecting the high heterogeneity of

habitats within the different types and water bodies.

Transitional waters represent especially big chal-

lenges for typifying and intercalibration on a Euro-

pean scale, due to the heterogeneity of water bodies

including wide size spectra of both estuaries and

coastal lagoons with a large variety of connections to

the open sea and to continental waters. Defining site-

specific reference conditions is a novel approach that

allows minimizing uncertainties introduced by apply-

ing broad types. In large unique water bodies for

which long-term historical data exist, this approach

has been proved useful (Lyche Solcheim, 2005). With

the help of modelling, however, site-specific refer-

ence conditions can be determined also for a larger

variety of water bodies (e.g., within broad types) as

functions of watershed attributes. Carvalho et al.

(2009) recommend site-specific reference conditions

for chlorophyll in lakes in preference to type-specific

reference conditions, as they use the individual

characteristics of a site known to influence phyto-

plankton biomass, rather than adopt standards set to

generally represent a large population of lakes of a

particular type.

(ii) Difficulties finding appropriate metrics: In

1997, in an article that has became a classic in the

field of biomonitoring of surface waters, Karr & Chu

formulated most of the leading principles that form

the basis of the nowadays WFD: the concept of good

water status (‘biotic integrity’ in their terminology),

the need for a water body type-specific approach, the

supremacy of biological indicators and the assess-

ment scheme based on comparison of present status

with reference conditions. According Karr & Chu

(1997), selection of those among the multitude of

measurable biological attributes that provide reliable

and relevant signals about the biological effects of

human activities but remain insensitive to extraneous

conditions is a step of extraordinary importance. Only

these attributes are called metrics.

The authors warn about attempts of measuring all

possible attributes and argue that some attributes are

poor candidates for monitoring metrics because of their

underlying biology. In particular, they indicate abun-

dance, population size, density and production as

attributes that vary too much even in pristine areas to be

reliable indicators of human influence. According to

Karr & Chu (1997), the belief that biology is too

variable to be of practical use in monitoring comes not

from a lack of good indicators but from the fact that

studies of naturally variable attributes such as popu-

lation size, density and abundance have dominated

ecology for[50 years. They show also that ratios (e.g.,

of the abundances of two trophic groups), although

Hydrobiologia (2009) 633:197–211 201

123

they may intuitively seem useful, are inherently flawed

as they mix independent parameters in ways that make

it hard to discern their relative influence. Further, two

large numbers and two small numbers may yield the

same ratio, although the biological meaning of small

and large numbers may be very different. As one of the

rare exceptions, relative abundance (number of indi-

viduals in a taxonomic or other target group divided by

the total number of individuals in the sample) repre-

sents ratios that may reveal consistent dose–response

relationships.

Karr & Chu (1997) also show that a small number

of attributes of biological elements—such as taxa

richness and percentages of individuals belonging to

tolerant taxa—consistently emerge as reliable indi-

cators of biological condition at sites influenced by

different human activities in different geographic

areas.

Moss (2008) goes further from such a simple and

empirical division of ecological attributes into good

and bad metrics for measuring human impact and

formulates four primary characteristics of ecological

quality which intactness should be measured to

decide about impactedness of sites. These primary

characteristics include nutrient parsimony (efficiency

in recycling scarce materials), characteristic of phys-

ical settings and food web structure that ensure this

parsimony and maintenance of the intactness of the

system as a whole structure in its several aspects,

connectedness and size. Moss argues that secondary

features such as concentrations of substances or lists

of species do not adequately measure these funda-

mental characteristics, unless these are used in ways

to diagnose the state of the more fundamental

characteristics.

In the light of the above mentioned arguments by

Karr & Chu (1997) and Moss (2008), a question

arises as regards how to evaluate the common

practice of using chlorophyll concentration as one

of the main biological metrics in WFD implementa-

tion (Carvalho et al., 2009; Carstensen & Henriksen,

2009; Wolfram et al., 2009). On the one hand,

chlorophyll clearly belongs to the abundance/popu-

lation size or density attributes criticized for their

large variability, but, on the other hand, increase of

chlorophyll and algal biomass are symptomatic signs

of alteration in nutrient parsimony and food chain

structure related to eutrophication and as such are

included also in the normative definitions of

ecological status classes for all the surface water

categories. Obviously, Karr & Chu (1997) in their

division of biological attributes to suitable and non-

suitable metrics for measuring human impact had in

mind macroinvertebrates or fish rather than phyto-

plankton which, as a primary producer, has its own

specific features.

New concepts in ecology stemming from the

theory of ecosystem resilience and alternative stable

states affect also the selection of metrics and their

calibration against pressure gradients. As Borja et al.

(2009) note, the selection of ecologically relevant

attributes to be used within any methodology, should

emphasize metrics that respond to both degradative

and restorative anthropogenic action. One aspect that

complicates this approach is the different lags of

biotic metrics to changes in pressure. A long-term

study carried out in Lake Mondsee (Dokulil &

Teubner, 2005) showed that reductions in phyto-

plankton biovolume were delayed by about 5 years

relative to reduction in total phosphorus (TP) con-

centration while phytoplankton species differed in

timing of their responses to changing nutrient condi-

tions. While Planktothrix rubescens declined con-

comitantly with the decline in TP concentration, other

species indicative of higher phosphorus concentra-

tions, such as Tabellaria flocculosa var. asterionello-

ides, tended to increase further. In such situations, the

efficiency of restoration measures can be evaluated

differently depending on how closely the selected

metric follows the changes in pressure.

(iii) Are biological indicators surrogates for

chemical analyses?: For standing waters (lakes,

transitional and coastal waters), eutrophication is a

major pan-European problem and the biggest threat

for biodiversity (Watt et al., 2007) and reaching good

ecological status of surface waters (CIS, 2005).

Phytoplankton is the first group of organisms to

respond to nutrient enrichment with excessive growth

and shifts in community composition resulting in

increased turbidity, subsequent loss of submerged

aquatic vegetation and oxygen deficiency in bottom

waters. A number of trophic state indices have been

created based on simple parameters such as TP,

Secchi depth and chlorophyll concentration (Carlson,

1977) or scores of algal (LAWA, 1999) or macro-

phyte species (Schneider & Melzer, 2003) given to

them according to phosphorus ranges in which they

occur. Even the taxon-specific optima of littoral

202 Hydrobiologia (2009) 633:197–211

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invertebrates along a TP gradient have been used to

develop an ecological classification model (Donohue

et al., 2009). The latter type of indices has been

criticized for being surrogates for driving pressures

(nutrients), against which they have been calibrated

(Moss, 2008). Moss points out that such pressures can

be independently chemically measured with greater

accuracy, but admits also that the biological commu-

nity has an integrating component that spot analyses

do not have. It seems to be rather a rule rather than an

exception that biological communities affected by

eutrophication show a delayed recovery in the

oligotrophication phase (Dokulil & Teubner, 2005;

Hajnal & Padisak, 2008; Padisak & Reynolds, 1998)

or do not reflect at all the decrease in phosphorus

(Kaiblinger et al., 2009; Noges et al., submitted) even

if indices initially calibrated against phosphorus are

used. Obviously, these phytoplankton indices are able

to capture ecological status in circumstances where

factors other than nutrients (e.g. changes in light

conditions, food chain or salinity; Borics et al., 2000;

Padisak et al., 2006) take over the control over

community structure creating a resistance to restora-

tion efforts.

A trophic index although relating community

composition to eutrophication pressure, cannot be

used for a WFD-compliant assessment of ecological

quality unless reference conditions and class bound-

aries are defined. The species list and scoring system

for species may remain the same for all lake types, for

example, in the Swedish (Willen, 2007) or Austrian

system (Dokulil et al., 2005) or different species lists

and sensitivities are applied for different types, for

example in German system (Mischke et al., 2008).

WFD represents a transition from chemically to

ecologically based assessment of water quality. In

principle, the ecological quality assessment must be

based on the structure and functioning of aquatic

ecosystems. Thus, there is a need to assess the impact

of toxic substances on aquatic populations and

communities. A probabilistic, community-based

approach for defining environmental quality and

assessing risk for ecosystems is to calculate the

number of species potentially affected by a given

concentration of a toxic chemical assuming that the

sensitivity of different species toward a given stress

factor (e.g. a toxic chemical) follows a normal

distribution (see Vighi et al., 2006 for summary).

The benefit of this approach is that it is based on the

structure of the biological communities present in a

given environment and not on toxicological data

produced on standard organisms. On the negative

side, data for such assessment of toxicity impacts on

community or population level are not yet available.

This would require deeper knowledge on structure

and functioning of natural communities as well as

toxicity data on a large number of organisms. At

present such data is available only for a very low

number of chemicals. On the other hand, toxicity

tests on the basic ecotoxicological test organisms like

algae, Daphnia, and fish, are no longer sufficient for

assessing toxicological stress on the communities

(Vighi et al., 2006).

(iv) Is there hope to discover new powerful

metrics?: Since the first attempts to use biological

indicators in water quality assessment (Kolkwitz &

Marsson, 1908, 1909), vast amounts of information on

biological indicators have been accumulated and

certain traditions established, e.g. that of using

macroinvertebrates for river assessments and phyto-

plankton for lakes. The implementation of the WFD

has induced a massive search for new biological

methods to measure the anthropogenic impact on

water bodies. We may ask: Is there any hope to

discover new powerful metrics or develop more

sensitive and selective methods? After 15 years’

research in selecting metrics for ecological status

assessment of water bodies, Karr & Chu (1997) came

to a striking conclusion that the same major biological

attributes arose recurrently and served as reliable

indicators in diverse circumstances. The recent pro-

liferation of indices is so extensive that the assessment

and comparison of their suitability for quality assess-

ment have become a new labour-intensive field for

researchers that adds an element of confusion back

into what environmental managers had hoped to

simplify (Borja et al., 2009). On the other hand, the

intensive research, especially in less traditional areas,

has already yielded and will yield in the future new

metrics, methods and analysing techniques more

specific for the widening gamma of pressures.

(v) Multimetric approaches for multivariately

determined conditions: One of such fast developing

areas is the construction of multimetric indices. The

theoretical basis of this approach has reached already

a maturity level where ‘cook books’ for designing

such indices (Hering et al., 2006) can be produced to

guide researchers to the best solutions and helping

Hydrobiologia (2009) 633:197–211 203

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them to avoid the already known pitfalls. As the

relationships between occurrence of particular spe-

cies, genera or groups and environmental conditions

are multivariately determined (Moss, 2007), a multi-

metric index attempting to represent the multitude of

impacts on different biotic levels, ought to be the

adequate tool to measure the multiple human impair-

ments. By combining different categories of metrics

(e.g. taxa richness, diversity measures, proportion of

sensitive and tolerant species, trophic structure)

reflecting different environmental conditions and

aspects of the community, the multimetric assessment

is regarded as a more reliable tool than assessment

methods based on single metrics (Hering et al., 2006).

The best multimetric indices are more than a

community-level assessment because they combine

measures of condition in individuals, populations,

communities, ecosystems and landscapes (Karr &

Chu, 1997). Since the first multimetric index, the

Index of Biotic Integrity developed by Karr (1981),

which used fish as stream quality indicators, numerous

multimetric indices have been developed. In this

issue, the Austrian Macrophyte Index (Pall & Moser,

2009) based on metrics like ‘vegetation density’,

‘vegetation limit’, ‘characteristic zonation’, and ‘spe-

cies composition’ represents a multimetric approach.

A novel approach—an extension to environmental

science currently a matter of study—is the use of

Fuzzy Logic—a way of using linguistic variables to

describe and assess complex systems. It is becoming

popular in Spain in studies related to the implemen-

tation of the WFD (Ocampo-Duque et al., 2007; see

also Borja et al., 2009). One of the main advantages

of fuzzy logic is the ability to model expert human

knowledge, a necessary feature to be considered in

the complex process of environmental management.

In a case study where the fuzzy model approach was

applied in the Ebro River Basin (Ocampo-Duque

et al., 2006), the four components surveyed corre-

sponded rather well to the four primary characteris-

tics of ecological status listed by Moss (2008): total

riparian vegetation cover (= size), cover structure

(= characteristic structure in its several aspects),

cover quality (&nutrient parsimony), and channel

alterations (= connectedness).

(vi) Dominant versus less frequent taxa and the

question of taxa sensitivity: Discussing the informa-

tion that biological samples contain, Reynolds (1998)

wrote: ‘An experienced plankton scientist, confronted

with an undefined sample of phytoplankton under a

moderate quality microscope, would have little

difficulty in gleaning sufficient information from the

assemblage to be able to deduce a lot about the water

body from which it was collected—whether oligo-

trophic or eutrophic, acidic or alkaline, shallow or

deep, mixed or stratified and, in all probability, what

month of the year it was taken.’ The same is valid not

only for phytoplankton but also for other major

biological groups (or ‘elements’ in the WFD vocab-

ulary). The question is, which components of the

biological sample contain this information: is it coded

in the dominants or in some minor indicator species?

Based on a general assumption that a healthy

environment will have the species appropriate to it, a

study (Devlin et al., 2009) was carried out in the

United Kingdom to identify whether abundance and

species richness of 20 most frequently occurring

marine phytoplankton species were predictable across

seasons and typology, and whether the deviation

away from a reference condition in different risk

conditions was sufficient to apply phytoplankton

measurements into a management tool for assessment

of environmental quality. The preliminary study

showed that it may be possible to have one or very

few generic lists of species that should occur at a

particular time of year and could be used for

ecological assessment of marine waters.

On the other hand, using the most frequent species

with respect to ecological status classification has

been criticized (Rawson, 1956), as it appears that the

dominant species are often those with rather wide

tolerance (eurybionts), and thus, they may be even

less indicative of aquatic condition than the less

frequent species. In geobotany, the fidelity concept,

one of the most important aspects of the Braun-

Blanquet (1925) approach, has been widely used to

diagnose particular types of natural plant communi-

ties. Fidelity is the degree to which a species is

concentrated in a given vegetation unit, while the

most specialized ones, often occurring with medium

or low frequencies, are called diagnostic species that

are used to delimit associations. Their diagnostic

power, however, has been tested only for natural

environments (reference conditions) and cannot tell

us anything about human-caused degradation. In

order to find out, which species or metrics merit

watching for these signs, analysis should focus on

species tolerance and community changes across the

204 Hydrobiologia (2009) 633:197–211

123

range from minimal to severe human impact. This

approach has yielded several lists of aquatic organ-

isms divided into sensitive, indifferent and tolerant

taxa. In some cases, this search for indicators has led

to in-depth studies of a single species, applying

different metrics combined into a multimetric index

(Lopez y Royo et al., 2009).

Information on taxa sensitivity or tolerance has

been used in the assessment methods in various ways.

The mere presence of very sensitive taxa is a strong

indicator of good ecological status while their relative

abundance should not necessarily be high. Presence

alone of tolerant taxa, on the other hand, says little

about ecological status since tolerant groups inhabit a

wide range of places and conditions, but as conditions

deteriorate, their relative abundance rises (Karr &

Chu, 1997). Most community composition indices,

however, do not consider this difference and use the

abundance-based approach (species score multiplied

by abundance) for both sensitive and tolerant species.

Results of such approach will be strongly biased

towards the tolerance level of the dominant species.

Karr & Chu (1997) recommended that only about

10% of taxa in a region should be identified as

sensitive or tolerant avoiding the use of the remaining

70–90% taxa with intermediate tolerances, which

would swamp the strong signal coming from presence

of the most sensitive or most tolerant taxa. A

compromise option that takes into account both the

abundance and the diagnostic power of species is to

add a weighing factor to the index that depends on the

stenoecy, i.e. the species tolerance range, as it has

been done, for instance, in the German Phytoplank-

ton-Taxa-Lake-Index (Mischke et al., 2008).

One-out–all-out principle. What does the practice

to date show?

Until now, the focus of research supporting the

implementation of the WFD has been on the devel-

opment of individual indicators at the level of the

Biological Quality Element (BQE) or single param-

eters within a BQE; hundreds of such studies

proposing such indicators have appeared in the

scientific literature in recent years, including some

articles in this volume (e.g. Kaiblinger et al., 2009;

Ptacnik et al., 2009; Pall & Moser, 2009; Donahue

et al., 2009; Lopez y Royo et al., 2009). Also, the

intercalibration exercise has been carried out at the

BQE level. For the final classification of water

bodies, it is of crucial importance how classification

results for the different BQEs are combined in an

assessment of ecological status. The WFD prescribes

the ‘one-out–all-out’ approach, where the BQE

showing the worst classification determines the

ecological status. Moss et al. (2003) showed that

the one-out–all-out principle will tend to downgrade

sites unjustifiably, depending on the number of

metrics included in the assessment. Errors within

individual quality elements and metrics tend to show

considerable variation (Johnson et al., 2006). If these

are combined using the one-out–all-out principle,

reliable, precise metrics tend to be overruled by less

reliable metrics in a large proportion of water bodies.

Alternative approaches to the one-out—all-out prin-

ciple could be considered, where pressures are more

specifically taken into account. Now that WFD

monitoring programmes have been established in

most EU member states and data will be available, it

will be possible to evaluate the strengths and

weaknesses of the one-out–all-out and alternative

approaches.

Variability in ecological status assessment

and resulting classification

Ecological indicators typically exhibit non-linear

responses to anthropogenic pressures together with

great spatial and temporal variability, which underpin

emerging key characteristics of natural systems

(Solimini et al., 2009). Predicting the causes and

consequences of human activities on freshwater

ecosystems requires a quantitative assessment of the

level of uncertainty associated with our understand-

ing of ecological structure and processes. The

assessment based on Ecological Quality Ratios

(EQRs) ultimately results in the assignment of a

given site to a certain ecological status class.

However, to what extent are we confident that the

resulting classification is the true one for that site?

Here, the problem is to estimate the level of

confidence of the class assignment for a given site

or, in other words, the risk of site misclassification

which is relevant for decision makers.

The assessment of ecological status and the result-

ing classification of water bodies are complicated by

Hydrobiologia (2009) 633:197–211 205

123

the inherent variability of biological communities. In

the normative terms, the WFD refers to uncertainty

as ‘level of confidence’ and/or level of ‘precision’

when dealing with reference conditions (Annex II)

and when dealing with monitoring (Annex V). In the

first case, the methods and the number of sites used

to set spatially derived reference conditions must

provide a sufficient level of confidence about the

values for the reference conditions to ensure that the

conditions so derived are consistent and valid for

each surface water body type. In the case of

monitoring, estimates of the level of confidence and

precision of the results provided by the monitoring

programmes must be given in the river management

plan. The frequency of monitoring should be chosen

so that an acceptable level of confidence and

precision are achieved. Accordingly, several recent

articles on WFD compliant metrics, whole assess-

ment systems or reference conditions include an

estimation of uncertainty basically linked to spatial

or temporal factors (Kelly et al., 2009; Carstensen &

Henriksen, 2009; Carvalho et al., 2009; Erba et al.,

2009; Wolfram et al., 2009). However, the total

uncertainty associated with the classification of

ecological status reflects also technical, economical

and political constraints of the assessment methods

(Fig. 1).

Abiotic factors in freshwater ecosystems are

highly variable in time and space and often account

for a significant proportion of the variation of

community patterns in terms of species diversity,

abundance, biomass and production. Therefore, most

of the variances in EQR-derived classification prob-

ably derive from the unspecified sampling variability,

resulting from this spatial heterogeneity of abiotic

factors at multiple time scale that, in turn, affect the

biotic community structure. In addition, other sources

of variation include (Fig. 1): the choice of sampling

sites within a region or catchment (how many sites?),

the sampling effort (how many samples and of what

dimension?) and the frequency of sampling (how

often?). Imposed over the natural variation of

biological metrics, the human activities produce

different pressures that interact with abiotic and

biotic factors in shaping the structure of communities

and dissecting the natural from the human-induced

variance can be extremely difficult. A possible

strategy to reduce the overall uncertainty is to include

significant site-specific features in the models used,

e.g. to derive reference conditions (Carstensen &

Henriksen, 2009). The inclusion of site-specific

features will also provide more realistic boundary

values for managing specific water bodies as well as

provide more certainty in the boundary estimates

(Carstensen & Henriksen, 2009).

A variety of errors and biases can be introduced

into the final classification result when considering

also the various field and laboratory methods

Fig. 1 Sources of

uncertainty associated with

the classification of

ecological status of water

bodies

206 Hydrobiologia (2009) 633:197–211

123

(sampling design and collection, sorting, taxonomy)

and the metric used. Often the calculation of metrics

involve several intermediate steps, all with associated

errors and all potentially adding to the resulting

overall uncertainty. It should be emphasized that the

size of the various sources of error change depending

on the water bodies considered. For example,

lowland, deep and turbid rivers require assessment

techniques and have an associated variability that is

different from those used in wadeable piedmont

rivers or streams (see, for example Solimini et al.,

2000). Similarly, assessment systems of small lakes

might differ from those used for larger systems

(Solimini et al., 2008).

A degree of uncertainty is associated with both

components of EQR: the expected and the observed

value. The reference value of a given metric comes

with an associated error because of the difficulties of

defining the reference conditions for many water

bodies. Therefore, an assessment in the uncertainty

associated with the reference condition values should

also be included in the overall risk of misclassifi-

cation of a site. It is clear that metrics should be

selected after examining their variability in natural,

undisturbed conditions. For example, Erba et al.

(2009) supposed that the highest metric variability

due to natural factors can be observed at reference

sites. Metrics too variable, even at reference sites, are

unlikely to be effective for the assessment programs.

Although, relevant typology factors of water bodies

can be included to diminish variability in reference

values for ecological measures, variability may

remain high. For example, a recent assessment of

TP values in a large dataset of reference European

lakes led to coefficient of variation ranging from 1%

to over 69% depending on the lake type and sample

size (Cardoso et al., 2007). The same authors studied

a regression model to predict reference values for TP

in European lakes from simple edaphic factors, which

gave a variance explained of 51%, similar to other

models present in literature (Cardoso et al., 2007). In

another recent article, Carvalho et al. (2009) describe

the development of a series of regression models for

predicting reference chlorophyll concentrations on a

site-specific basis. None of the regression models

developed has a particularly strong predictive ability,

and the authors speculate that uncertainty arises from

errors in the data used to develop the models

(including natural temporal and spatial variability in

data) and also from additional explanatory variables

not considered in the models, particularly nutrient

concentrations, retention time and grazing (Carvalho

et al., 2009). Also Wolfram et al. (2009) found high

variability (with r2 ranging between 0.33 and 0.40)

when regressing TP and total biovolume (e.g. algal

biomass) even on a larger TP range. Nevertheless,

site-specific reference conditions should result in

reduced error in ecological status classifications,

particularly for lakes close to typology boundaries

(Carvalho et al., 2009). Water managers may face the

problem of deciding whether the uncertainty associ-

ated with the predicted average value is small enough

to detect changes with the actual lake conditions.

For decision makers, the importance of the risk of

misclassification of a given site depends on the

context of the decision to be taken and increases with

larger spatial scales and longer time horizons. In the

case of ecological data, the complexity of error

budget and the magnitude of resulting uncertainty

estimates may be very large and so the necessary

effort (e.g. the cost) for its estimation. For this reason,

many forms of uncertainty are not acknowledged in

the models that support decisions or, in some cases,

are ignored altogether (Burgman et al., 2005; but see

Kelly et al., 2009). Recently, ecologists have become

more and more sophisticated in realizing suitable

metrics that are stable and robust enough to assess

ecological status in a reliable manner (see Ptacnik

et al., 2009; Kelly et al., 2009) but it is also necessary

that the quantification of EQR related uncertainty is

implemented in future assessment programs. Notably,

Kelly et al. (2009) found that variability (expressed as

the standard deviation of samples through time) could

be modelled using a polynomial function, from which

the risk of misclassification could be calculated.

Moreover, in general, if the EQR assessment outcome

can be incorporated into a modelling framework

resulting in an overall estimation of the risk of

misclassification, then the related uncertainty might

be assessed through careful evaluation of model

predictions. A possible approach is to account for as

many forms of uncertainty as possible (such as error

propagation and sensitivity analyses see, for example

Blukacz et al., 2005; Lo, 2005) and to use several

modelling frames and examine whether the choice

makes a difference to management decisions (Burg-

man et al., 2005). Advanced computational tech-

niques, such as fuzzy logic or neural networks, may

Hydrobiologia (2009) 633:197–211 207

123

be effective in dealing with the sources of uncertainty

that affect the process of metric development (Borja

et al., 2009).

Acknowledgements The study was supported by EU FP7

grant 226273 (WISER). Special thanks go to Dr. Judit Padisak

for her valuable comments and suggestions.

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Author Biographies

Peeter Noges (b. 1957)

graduated from Tartu Uni-

versity, Estonia (PhD:

1987), biologist–algologist,

leading researcher at Esto-

nian University of Life

Sciences. In the period from

2003 to 2009 he worked as

a national expert and con-

tractual agent at the Insti-

tute of Environment and

Sustainability, Joint

Research Centre of Euro-

pean Commission in Ispra, Italy. His main research areas are

ecology of shallow lakes and the impact of climate change on

lake ecosystems. He was one of the guest editors of special

volume of Hydrobiologia 599 dedicated to European large

lakes, reprinted in Developments in Hydrobiology 199.

210 Hydrobiologia (2009) 633:197–211

123

Wouter van de Bund(PhD: University of

Amsterdam 1994) is an

aquatic ecologist, currently

employed as a scientific

officer at the Institute of

Environment and Sustain-

ability, Joint Research Cen-

tre of European Commission

in Ispra, Italy. As the co-

leader of the Working Group

on Ecological Status under

the Common Implementa-

tion Strategy of the Water Framework Directive, his field of study

is focussed on harmonization of ecological quality assessment

systems in Europe.

Ana Cristina Cardoso (b.

1966) graduated at the

University of Algarve, Por-

tugal (1991; PhD: 1999),

aquatic ecologist. Since

then, she has been a scien-

tific officer at the Institute

of Environment and Sus-

tainability, Joint Research

Centre of European Com-

mission in Ispra, Italy. She

is currently Action leader of

the Action European Eco-

logical Water Quality assessment and Intercalibration (EE-

WAI), supporting the implementation of the Water Framework

Directive (WFD). Her main research area is ecology of lakes in

relation to the impact from nutrients.

Angelo G. Solimini (degree

in Biology, 1993; PhD in

Animal Biology, 1999) is

currently researcher in

Environmental Health at

Sapienza University of

Rome. His research and

teaching interests are the

environmental impact

assessment in aquatic eco-

systems, water-related dis-

eases, the structure and

functioning of aquatic food

webs and their relation with human pressures. He has been

involved in several European projects and working groups

supporting the WFD implementation and guest editor of a

special issue in Freshwater Biology on structure and func-

tioning of running water ecosystems. Solimini has published

about 60 scientific articles, book chapters, reviewed reports and

commentaries.

Anna-Stiina Heiskanen (b.

1961) graduated from Uni-

versity of Helsinki, Finland

(PhLic: 1991; PhD: 1998),

marine biologist. She has

worked at Finnish Environ-

ment Institute (SYKE), at

the Finnish Institute of

Marine Research, at the

universities of Helsinki and

Oslo, and at the Academy

of Finland. For eight years

(2001–2008), and also

worked as a scientific officer at the Institute of Environment

and Sustainability, Joint Research Centre of European Com-

mission in Ispra, Italy. Her field of profession was research and

coordination of ecological quality classification of surface

waters, and she led the Action European Ecological Water

Quality Assessment and Intercalibration (EEWAI). Since 2008,

she is a research manager of the SYKE Marine Research

Centre that aims at evaluating the state of the Baltic Sea and its

coastal waters and also the effectiveness and adequacy of

environmental conservation measures.

Hydrobiologia (2009) 633:197–211 211

123