International Journal on Algae, 2007, 9(1): 41-63

Selecting appropriate bioindicator regarding the WFD guidelines for freshwaters – a Macedonian experience∗

SVETISLAV KRSTIĆ1, ZORICA SVIRČEV2, ZLATKO LEVKOV1 , TEOFIL NAKOV1

1Faculty of Natural Sciences and Mathematics, St. Ciril and Methodius University,

Skopje Macedonia 2Faculty of Natural Sciences, University of Novi Sad,

Novi Sad, Serbia and Montenegro

skrstic@iunona.pmf.ukim.edu.mk

ABSTRACT

In light of WFD (Water Framework Directive) guidelines for freshwater monitoring this paper

addresses the suggested organisms to be used in regular biomonitoring systems. Based on critical

observations and reported obstacles in applying the various developed methods, but also on long-

term research on freshwater ecosystems in Macedonia, the best bioindicator organism is proposed.

In order to have an overall assessment of the investigated ecosystem, biomonitoring organisms (such

as fish, macrophytes) or geomorphology reveal important informations. But, if the target is rapid,

reliable and cost effective bioindication of the water quality, only algae and benthic invertebrates

have most comprehensive application history. Accumulated evidence that using benthic

invertebrates pose a problem in obtaining reliable results triggered recent CEN's revision of their

applicability in monitoring programs. Therefore, amid all critically addressed disadvantages, algae

(benthic diatoms in particular) are commented as the best solution and basic methodology given for

including in regular bioindication programs in rapid detection of water quality of lotic and lenthic

freshwater environments.

KEYWORDS: WFD, algae, diatoms, bioindicator, biomonitor, water quality, Macedonia.

INTRODUCTION

Organisms, populations, biocoenoses and ultimately whole ecosystems are naturally influenced by numerous biotic and abiotic stress factors such as fluctuations in climate, varying radiation and food supply, predator-prey relationships, parasites, diseases, and competition within and between species (Markert et al., 1999). Consequently, the ability to react to stressors is an important characteristic of all living systems and conversely no

∗ Originally published in Algologia, 2006, 16(0), pp. 000-000 ISSN 1521-9429

©Begell House Inc., 2007

2 SVETISLAV KRSTIĆ et al.

development of the species and the ecosystem as a whole is possible without such natural stressors (Schüürmann & Markert, 1998). Stress is the locomotive of evolution. But in the course of the evolutive epochs, the range of variation of the stressors is generally (amid the few cornerstones of evolution) fairly constant and allows species to adjust to changing environmental conditions.

In recent centuries, nevertheless, these changes have reached new dimensions, both in quality and quantity. Human activities have introduced totally new substances into the environment (xenobiotics, many radionucleotides) and potentially harmful substances (heavy metals) in vast quantities. In addition, these stressors usually have a synergistic effect, either with natural systems or among themselves, resulting in the decreased ability (or level of tolerance) of the organisms to adjust to them (Oehlmann & Markert, 1999).

The number of species worldwide is thought to be 13 million, although only approximately 1.6 million species have been identified (Heywood & Watson, 1995). Pollution, habitat fragmentation and loss, intensification of agriculture and population pressure are leading to dramatic changes in biodiversity (McNeelay et al., 1995). The alarming loss of biodiversity in last decades represent a major challenge to the scientific community and demands the development of appropriate strategies for land management and proper monitoring tools. Beside ecological consequences, loss of species diversity may also affect economic processes. Climate change prognosis (ICC, 2001) will eventually change the viability of populations, the number and distribution of species and the structure, composition and functioning of ecosystems (Kappelle et al., 1999).

An objective of prophylactic environmental protection must be to obtain reliable informations on the past, present and future situation of the environment. Beside the classic global observation systems, such as satellites and instrumental measuring techniques like trace gas and on-line (water, air) monitoring, there should be an increased implementation of bioindicative systems that provide integrated information permitting prophylactic care of the environment and human health. In the past 20 years, bioindicators have been proven as particularly interesting and intelligent measuring systems. In 1980, Müller considered the "bioindicative source of information" one of the pillars of modern environmental monitoring, since "bioindication is the breakdown of the information content of biosystems, make it possible to evaluate whole areas".

Theoretical background of biomonitoring

Bioindication and biomonitoring seemed very promising (and cheap) methods of observing the impact of external factors on ecosystems and their development over a long period, or of differentiating between one location (unpolluted site) and another (polluted site). Intensive developing of these methods has resulted in a still unsolved problem: definitions of both bioindication and biomonitoring, and the outcomes of these methods

SELECTING APPROPRIATE BIOINDICATOR 3

have never led to a common approach by scientists, so that now different definitions and expectations exist simultaneously (Witting, 1993). It is therefore most important to list the relevant definitions on which biomonitoring system should be based, given by Markert et al. (1999):

bioindicator – is an organism (or part of an organism or a community of organisms) that contains information on the quality of the environment (or a part of the environment). These are usually a target for presence/absence of single organisms or its populations or community of organisms in the saprobity/trophy water biomonitoring system;

biomonitor – is an organism (or part of an organism or a community of organisms) that enables information on the quantitative aspects of the quality of the environment. These are usually the target for tracing the specific pollutants (e.g. heavy metals) accumulation (bioconcentration, bioaccumulation) processes in the water biota. A biomonitor is always a bioindicator as well, but not vice versa.

In terms of newly introduced methods as additional monitoring tools to classic floristic, faunal and biocoenotic investigations that record reactions to pollutant exposure at higher organizational levels of the biological systems, there are:

biomarker – measurable biological parameter at suborganismic (genetic, enzymatic, physiological, morphological) level in which structural or functional changes indicate environmental influences in general, and the action of pollutants (eg. cytochrome P-450 induction by halogenated hydrocarbons);

biosensor – a measuring device that produces a signal in proportion to concentration of a defined group of substances through suitable combination of selective biological system, e.g. enzyme, antibody, membrane, organelle, cell or tissue, and physical transmission device (e.g. amperometric electrode, optoelectronic receiver). Main examples are bacterial toximeters, eu-cyano bacterial electrodes and biotests.

Regarding the monitoring procedures introduced for observing the vegetation changes in the ecosystems the following are basic:

survey – qualitative and quantitative observations made by standardized procedures without any regard to repetition;

surveillance – an extended programme of surveys in order to provide a time series, to ascertain the variability and/or range of states or values which might be encountered over time;

biomonitoring – regular, systematic use of organisms to determine environmental quality (Cairns, 1979).

Considering genetic and non-genetic adaptation of organisms and communities to environmental stress, the following responses could be distinguished:

4 SVETISLAV KRSTIĆ et al.

tolerance – the ability of an organism or community to withstand unfavorable abiotic (climate, radiation, pollutants) or biotic factors (parasites, pathogens), where adaptive changes (e.g. enzyme induction, immune response) can be observed;

resistance – opposite to tolerance, is genetically derived ability to withstand stress; sensitivity – susceptibility of an organism or community to biotic or abiotic

change. Sensitivity is low if the tolerance or resistance to environmental stressor is high and vice versa.

Finally (Figure 1), according to Csermeley (1998): stress – the state of biotic or abiotic system under the conditions of a "force"; strain – the response to stress, i.e. the expression before the damage occurs; damage – the result of too high a stress that no longer can be compensated for.

Figure 1. Average stress response times of biotic systems as related to size and complexity

[modified by Fränzle (2003) from Korte (1987)].

Selection of appropriate biomonitoring organism

Taking into account that the preferable candidate as a monitoring object in freshwater ecosystems should possess the following useful features:

■ to be present throughout the monitored ecosystem in fairly abundant numbers and throughout the year, especially prior and after the point source of pollution;

SELECTING APPROPRIATE BIOINDICATOR 5

■ to react swiftly, either at the individual or population level, on the detected pollution pressure; ■ to have high accumulation capacities, or to show statistically significant correlation to ambient concentration of the polluters; ■ to be sensitive to as many as possible environmental factors; ■ to be easy to collect and name at species level; ■ its collection and manipulation not to represent an ethical or biodiversity obstacle. We give brief comments on the most commonly used hydrobionts in freshwater

monitoring programs in relation to proposed EU WFD (2000) recommendations. Bacteria – useful indicators for environmental monitoring and ecological risk

assessments because they are present in high amounts in all kinds of environments and play key roles in food webs and element cycles, such as nitrogen, carbon, sulphur and phosphorous (Bloem et al., 1977). The small size and high surface to volume ratio cause a high affinity for very low concentration of substances. Due to intimate contact and interaction with the environment, microbes are very sensitive and respond quickly to contamination and other types of environmental stress (Giller et al., 1998). The microbial activity reflects the sum of all physical, chemical and biological factors regulating the decomposition and transformation of nutrients (Stenberg, 1999). Microbiological indicators can therefore serve as early warnings in monitoring programs (Jordan et al., 1995).

Microbial organisms are omnipresent in wetlands, even living within the individual submerged roots of some wetland shrubs (Fisher et al., 1991). Through interactions with wetland plants and hydrology, wetland microbial assemblages can remove inorganic nutrients, heavy metals, dissolved organic carbon, particulate organic matter, and suspended solids from the water column and sediments (Mickle, 1993), as well as play a key role in supporting food webs (Schallenberg & Kalff, 1993) and influencing global climate change through their role in methanogenesis (Kumaraswamy et al., 2000). The presence of bacteria in "biofilms" on the enormous plant and detrital surface area in wetlands is fundamental to wetland ability to degrade complex organic contaminants (Taylor et al., 1996). Iron-oxidizing bacteria in roots of wetland plants also influence plant nutrition (Emerson, 1999). Production from indigenous bacteria may surpass production from algae in forested wetlands (Hudson et al., 1992). Many naturally-occurring bacteria inhibit waterborne pathogens.

Bacterial assemblages, with generation times as little as 15 minutes are well suited for detecting short-term nutrient pulses. In an Ohio marsh, experimental dosing with phosphate stimulated an increase in bacterial density (Willis & Heath, 1993). Excessive enrichment can quickly allow anaerobic taxa to gain dominance. Microbial assemblages receiving agricultural nutrient inputs in part of the Florida Everglades were dominated by

6 SVETISLAV KRSTIĆ et al.

methanogens, sulfate reducers and acetate producers. These bacteria flourish where porewater total phosphorus concentrations and conductivities are high. Excessive nutrients from agricultural operations may reduce the normal ability of wetland microbial assemblages to detoxify particular pesticides (Entry & Emmingham, 1996). Although nitrogen additions to a riparian system briefly stimulated bacterial and fungal activity, long-term effects were perceived as negative, thus potentially compromising the ability of the system to remove nitrogen via denitrification (Ettema et al., 1999).

Figure 2. Biota relation to processes in recipient water ecosystems (Svirčev, 2005).

Algae – as early as 1853, Cohn recognized the value of algae in the biomonitoring

of freshwaters and attempted to classify them in relation to water quality (Figure 2), a scheme later modified by Mez (1898). Kolkwitz and Marsson (1902, 1908, 1909) define more clearly the relation of organisms to water quality and introduced the term 'saprobic organisms'. Later, Šrámek-Hušek (1956) introduced 'the system of saprobity' to describe the biotope, that was additionally developed and improved by Kolkwitz (1950) and Liebmann (1962). The most comprehensive effort in this field is presented by Sladeček (1973), but Friedrich (1990) finally separated photoautotrophic organisms from the system of saprobity, since it is defined as the intensity of heterotrophic activity, to avoid overlapping with the latterly introduced 'system of trophy' by Kelly and Whitton (1995). In general, all these attempts were based on presence, absence or abundance of species or their communities related to quality of the environment in which they are found (Dokulil, 2003). Based on this pattern, numerous diversity, biotic and similarity indices were developed, that were critically discussed by Washington (1984), who clearly states that biotic indices [e.g. the group of saprobity indices starting with Kolkwitz and Marsson's (1908) saprobiensystem and involving Pantle and Buck (1955), Zelinka and Marvan (1961) and Dittmar (1959)] are difficult to use in water quality estimations and monitoring of pollution, especially regarding faunal biota, and recommends diversity or similarity indices [e.g. Bray-Curtis (1957) dissimilarity index]. Ghetti and Ravera (1994) present an extensive list

SELECTING APPROPRIATE BIOINDICATOR 7

of indices used for the assessment of running waters based on plankton or periphyton algal communities, generally based on diatoms.

The exact nutrient that contributes to algal community shift is often difficult to identify due to correlations among many nutrients (McCormick & O’Dell, 1996). Nonetheless, some studies have reported that diatoms seem to dominate at lower temperatures and when phosphorus (P) but not silica (Si) is limiting, whereas green algae may dominate at higher temperatures with moderate or low N:P and Si:P ratios; Cyanobacteria (blue-green algae) typically dominate at higher temperatures and at low N:P ratios, and often characterize highly enriched waters (Biggs, 1995). In the Florida Everglades, phosphorus has the largest impact on algal assemblages, followed by nitrogen (N) and iron (Fe) (McCormick & O’Dell, 1996). In enriched areas of the Everglades, nitrogen, other nutrients, and/or light play a larger role in limiting growth (Vaithiyanathan & Richardson, 1997; McCormick & Stevenson, 1998). In nitrogen limited areas, increases in Rhopalodia gibba and blue-green algae (Nostoc) are typical (Vaithiyanathan & Richardson, 1997). Hörnström (1981) (Figure 3) has proposed the Trophic Lake Index based on

intriguing observation that algal species change towards taxa with higher biovolume (or from single

cells, via colonies, to filamentous species) as the trophy status of the lake increases. Algae have also been used as biomonitor (accumulator) organisms regarding

various polluters, most frequent being heavy metals (Whitton et al., 1981; Round, 1991), or bioassay organisms in relation to in situ (phytoplankton or attached algae – Rai et al., 1981; Levkov & Krstić, 2002) or laboratory toxicity tests (Munawar et al., 1989).

Figure 3. Trophic Lake index (Hörnström, 1981).

8 SVETISLAV KRSTIĆ et al.

If regular biomonitoring includes the following directions, algae (particularly diatoms) represent an outstanding tool for fundamental and relevant monitoring programs:

■ promoting taxonomy/ecology work on algae (diatoms) for students; ■ detecting the reference site for the area, usually mountain water source for rivers or

not influenced zones by human activities for lakes and reservoirs; ■ checking primarily the benthic forms, collecting as much substrata as possible,

prior and after the point source polluters or different stretches where diffuse pollution is expected (e.g. agriculture) for rivers. The same applies for lakes as well, where the most community changes are detected in the littoral region, but also checking the planktonic algal forms regarding possible eutrophication and occurrence of 'water blooms';

■ always looking for the live cells in the samples prior to specific check on present taxa, especially for diatoms;

■ preparing and maintaining a permanent diatom slide database for cross checking and future reference.

Invertebrate fauna – Bioassessment is now an established method for measuring human influence on aquatic ecosystems, complementing traditional physical and chemical methods. The presence or absence of benthic macroinvertebrates has been shown to be a good indicator of both chronic and episodic impacts of human disturbance to river condition (Ghetti & Ravera, 1994; Rosenberg & Resh, 1993).

Benthic invertebrates (zoobenthos) have been increasingly used in freshwater monitoring programmes. Due to its large species richness covering all types of freshwater habitats and to increased ecological knowledge on species response to environmental conditions, zoobenthos can be used for different monitoring topics, such as eutrophication, acidification, changes in habitat structure and species diversity and toxicity (EU Water ..., 2000).

The classical European approaches to biomonitoring (Kolkwitz & Marsson, 1909; Chandler, 1970) are limited in that invertebrate communities are not evaluated in the context of the site-specific habitat, which may show large spatial and temporal variation. The biological potential of a site may be limited by the quality of its habitat (Barbour, 1991). As Green (1979) emphasised: to decide whether or not a site is degraded, it is necessary to collect samples ‘both where the condition is present and the condition is absent, but everything else is the same’ (Linke et al., 2005).

The methods used for sampling benthic macroinvertebrate communities are very similar among European countries. This similarity is in part due to the international (ISO), European (CEN) and national standardisation work concerning biological and ecological assessment methods that began in the late 1980´s (EU Water ..., 2000). But, recently CEN has put all standard methods using benthic macroinvertebrates in freshwater monitoring under revision due to accumulating evidence of the difficulties in interpreting the results.

SELECTING APPROPRIATE BIOINDICATOR 9

Fish – fish populations have been monitored up until the present mostly for economic reasons. Fish catches have been estimated with the aid of different types of questionnaires, but also by test fishing important information on e.g. the structure of fish populations have been provided (EU Water ..., 2000). The focus of the previous monitoring programmes have been on commercially valuable fish species only. At present the role of fish has changed, and fish populations are considered as a part of the ecosystem.

There are several reasons why fish are widely used to describe natural characteristics of aquatic ecosystems and to assess habitat alterations (Schiemer, 2000):

■ advanced knowledge of a wide range of fish species and their ecological preferences;

■ as migratory organisms fish are suitable indicators of habitat connectivity or fragmentation (Chovanec et al., 2002);

■ fish size (and size of their organs) and their life span as well permit a variety of analytical procedures, most typically regarding bioaccumulation tests;

■ as primary and secondary consumers at different levels fish reflect trophic conditions in aquatic ecosystem;

■ the reconstruction of pristine reference communities is possible due to the existence of historical information;

■ the number of species is relatively small and species are already determinable in the field.

But, as Chovanec et al. (2003) clearly state, using fish as bioindicators may pose problems due to:

■ fishery caused alterations, such as species transfer, stocking, overfishing that interferes with other man-induced degradation of aquatic ecosystems;

■ the mobility of many species makes identification of source and duration of pollution very difficult.

It is therefore quite evident that fish are very useful indicators for overall environmental quality and ecosystem health, especially regarding pollution (bioaccumulation) aspects and environmental engineering, but the species composition and community structure can be used as biomonitoring tool with great difficulties.

Macrophytes – according to EU Water ... (2000) there are many characteristics of aquatic macrophytes, which can be used as indicator in environmental monitoring. However, due to local heterogenuity of habitats generalizations of indicator values are difficult, and variation in responses should be interpreted carefully. Main uses of aquatic macrophytes in the monitoring and assessment can be grouped in five following ways (Toivonen, 2000):

in plant tests carried out under controlled laboratory / field conditions;

10 SVETISLAV KRSTIĆ et al.

by using the chemical content of certain species as indicators of heavy metal and other toxic loads;

by using species or species groups as indicators of water or habitat quality; by studying temporal (long-term) changes in flora and vegetation; by studying or assessing biodiversity in water bodies.

As the aquatic vegetation in a water ecosystem is influenced simultaneously by several factors, macrophytes only rarely have good indicative values when evaluating some specific environmental variables. Instead, they give a good general estimation of the trophic state of the site. Their use as indicators is further motivated by their relative persistence in the site. Fluctuations in the population size of aquatic macrophytes are also usually minor in comparison to many other organisms (Toivonen, 2000). Macrophytes are especially suitable for longer-term (5-20 years) studies of changes in the littoral and open water areas.

The best solution

It is quite evident that WFD intends to have the overall monitoring of water ecosystems done both in spatial and temporal ways regarding the permanent detection of the complex ecosystem's health. This approach is most appropriate, having in mind the ever expanding numerous impacts on natural habitats and related biota due to human activities and increased necessity for ecosystems management and restoration. Appreciating that biomonitoring has been acknowledged as primary over chemical or hydromorphological features, proposed target organisms can be separated in two main groups: bioaccumulators and biomonitors. It is also very clear that macrophytes and fish, and under certain circumstances, invertebrate benthic fauna, are increasingly utilized in the detection of accumulation intensity of various compounds and their biological effect on biota, if present in a particular habitat. Therefore, establishing a reliable and cost/time-effective biomonitoring system for both lotic and lentic environments is best performed on algae as primary producers or bacteria as decomposers. But considering also the rapid bacterial turnover in numbers and specific physiological groups under variable environmental conditions and pollution fluxes, we are ultimately left with algae as the best possible choice.

Being a ubiquitous (but polyphyletic) group of organisms, algae are widely distributed in all marine and freshwater ecosystems, dominating plankton or benthic communities, some causing the so-called 'water blooms' in case of increased eutrophication. Among them, the diatoms (Bacillariophyta, Diatomeae, Figure 4.) are typically present in all, except the hottest and most hypersaline habitats, and during all seasons throughout the globe (Round et al., 1990). Having silica walls with very distinctive ornamentation, diatoms were recognized very early on (Leeuwenhoek, 1703) gaining one of the most comprehensive data records among all microorganisms. Being easy to collect and preserve

SELECTING APPROPRIATE BIOINDICATOR 11

in permanent slides, diatoms have gained much attention regarding pollution/eutrophication monitoring, or various others indications such as salinity, acidity, pH value, Al concent-rations, dissolved organic carbon and humic substances (Schönfelder, 2000). In paleo-limnology, water temperatures, pH values or eutrophication patterns can be reconstructed from diatom frustules in sediments (Dokulil, 2003). Figure 4. Benthic diatoms community from Lake Ohrid – larger cells: 1 – Gomphoneis ohridana Levkov; 2 – Navicula tripunctata (O.F. Müll.) Bory; 3 – Cymbella lange-bertalotii Krammer; 4 – Cocconeis placentula var. lineata (Ehr.) Cleve.

12 SVETISLAV KRSTIĆ et al.

Figure 5. Similar diatom species found in ecologically different water habitats in Macedonia: Gomphonema parvulum Kützing, mesotrophic habitat River Vardar; Gmphonema micropus Kützing, oligotrophic habitat rivers on Mountain Baba

The importance of using algae in general for monitoring freshwaters triggered a

specific conference on the Use of Algae for Monitoring Rivers held for the first time in Germany in 1991 and continuing today (Whitton et al., 1991). Subsequent publications arising from these conferences (Whitton & Rott, 1996; Prygiel et al., 1996) have many times pointed to the practical benefits of using algae over other (mostly macroinvertebrates) organisms in the regular monitoring of running waters. But they have also stressed the major disadvantages that monitoring programs are facing today regarding diatoms. These can be summed up as: a) taxonomic problems (Figure 5) and b) the incompatibility of indices developed in different European countries (Table 1).

With respect to problems in taxonomy, there is an ongoing scientific debate as to whether the recent introduction of many new species is justified by sound biological evidence, to support the discovery of very similar 'species flocks' or taxa that are discrete biological entities. Introduction of new techniques, primarily the use of the electron microscope, has hitherto revealed many rarely observed morphological features of the diatom frustules, thus widening the data available for taxonomical interpretation. There have been several attempts (Cox, 2002) to include more biological features, such as cytological or reproductive characters, for describing new species. More recently, molecular (Scala & Bowler, 2001) or genetic (Kooistra & Medlin, 1996; Medlin et al., 2000; Edgar & Theriot, 2004) data is helping to reveal the phylogeny relations and evolution patterns. Yet these efforts have not gained appropriate attention by taxonomists.

Other major taxonomic problem, noted by Kociolek and Spaulding (2000), is 'fitting the species into known species' relative to the apparent cosmopolitan distribution of diatoms. Thus "when one uses a guide to the diatoms in Europe in Africa, one tends to fit the taxa of Africa into descriptions given, with the result that similar species lists are produced". This fact represents both taxonomic (Levkov et al., 2005) and ecological problems, influencing the diatom indices based monitoring systems. Beside the longer history of taxonomic research in Europe, Kociolek and Spaulding point out the lack of 'rigorous taxonomic work' as a result of 'only few workers able to carry out research or proficiently train students' left in the field of diatom taxonomy. As much we would like to abandon the basic Darwinian approach to taxonomy, in this case, amid overall biological discrepancy, the diatom taxonomy should be promoted at least regarding the evident ecological necessity [as Kociolek and Stoermer (2001) profoundly state].

Development and application of different indices based on species-specific response of diatom taxa (number of counted frustules or taxa abundance) on different

SELECTING APPROPRIATE BIOINDICATOR 13

environmental conditions have finally bought closer the taxonomic and ecological efforts in monitoring, mostly the rivers. But, the problems have been reported in implementation of these indices in various environments and habitats (Table 1), basically regarding the incompatibility of detected environmental conditions and obtained index for water quality classes, although reported diatom communities were quite similar. This situation has resulted with various inapplicable indices for different regions even for Europe, most likely as a consequence of: a) Mathematical approximations – most of the indices are based on Zelinka and

Marvan's (1961) equation that includes taxa abundances and related s (saprobity value) and g (indicative weight value) of the specific taxa and thus involving more or less individual criteria in developing of the final score for the water quality class. Beside excellent review by Washington (1984), who accepts only basic similarity (or dissimilarity) indices as relevant for aquatic ecosystems, there is also Fryer's (1987) profound opinion that "the mania (on the part of some it is nothing less) to express biological events in numerical terms and to support the simplest facts with a statement of their statistical significance has become so widespread as to obscure the fact that a true understanding of many biological phenomena (even in ecology where numbers are so easily generated) often demands qualitative rather than quantitative knowledge" and also "mathematics may be synonymous with the ordered structure of the physical world: it can not explain everything in biology". We can only add that living organisms tend to evolve and thrive to survive in environments that are under human pressure, changing very rapidly. Therefore, any particular assigning of a specific number to an organism and then to use it in formulae leads to erroneous neglect of genetic plasticity and evolution;

b) Complexity of environmental influences – no matter how broad the physico-chemical analysis of a specific habitat are, and their subsequent correlation to observed diatom communities (especially for rivers), it is almost impossible to detect all parameters in time/scale fashion and their true impact on the biota. That is the main reason why we find the whole range of correlations of the same taxa to a variety of measured parameters in different ecosystems. An attempt that is ultimate misleading is the correlation between the communities with as little as one (or a few) environmental variables [like Kelly and Whitton's (1995) TDI for soluble phosphorus] or to assign the s and g values for genera. Difficulties to obtain reliable results are quite evident in Table. The only solution is developing specific indices for every ecosystem (even for two neighboring rivers) what is at least scientifically unsound, beside time and money consuming;

c) Methodology – apart of a very broad range of methods used by researchers or in regular monitoring programs stated in many references, at this point we would only

14 SVETISLAV KRSTIĆ et al.

like to stress the importance of observing live diatom cells, since it has been repeatedly stressed [e.g. Cox (2002)] that it is important to known what is actually living at the specific site and that contamination with dead diatom cells could account for as much as 50% or more of the observed community.

TABLE 1. Comparison among several mostly used diatom indices for river Vardar, Macedonia and newly developed DIS index for that particular, heavily polluted,

ecosystem (Krstić et al., 2002)

Descy (1979) Watanabe et al.

(1988) Kelly & Whitton

(1995) Rott et al.

(1997)

DIS (Levkov&

Krstić, 2002)

Riv

er V

arda

r sa

mpl

ing

site

s

Value Categ. Value Categ. Value Categ. Value Categ. Value Categ.

T1 2.31 III 62.35 III 2.67 III 1.72 I-II 0.51 II

T2 1.94 IV 67.75 III 3.41 III-IV 1.90 II 0.88 III

T3 2.05 III 69.05 III 4.25 IV 2.04 II 0.97 III

T4 2.05 III 65.05 III 4.37 IV 2.11 II 1.03 III

T5 2.01 III 69.10 III 4.26 IV 2.01 II 1.00 III

T6 2.03 III 72.05 II 3.88 IV 1.98 II 1.09 III

T7 2.03 III 76.35 II 3.85 IV 1.97 II 1.04 III

T8 2.00 III 61.55 III 4.13 IV 2.30 II-III 1.42 IV

T9 2.00 III 59.30 III 4.20 IV 2.35 II-III 1.35 IV

T10 1.94 IV 64.80 III 4.08 IV 2.12 II-III 1.18 IV

T11 1.96 IV 60.10 III 4.19 IV 2.16 II-III 1.20 IV

T12 1.94 IV 58.60 III 4.37 IV 2.16 II-III 1.23 IV

T13 1.89 IV 61.35 III 4.22 IV 2.08 II 1.28 IV

T14 2.01 III 63.90 III 4.27 IV 2.13 II 1.17 IV

T15 1.83 IV 51.20 III 4.29 IV 2.12 II 1.16 IV

T16 1.81 IV 53.60 III 4.26 IV 2.21 II-III 1.25 IV

T17 1.89 IV 53.60 III 4.08 IV 2.18 II 1.22 IV

T18 1.90 IV 49.55 IV 4.14 IV 2.23 II-III 1.25 IV

T19 1.76 IV 53.75 III 4.19 IV 2.23 II-III 1.22 IV

T20 1.74 IV 53.65 III 4.10 IV 2.21 II-III 1.21 IV

SELECTING APPROPRIATE BIOINDICATOR 15

Amid all the stated difficulties for using diatoms in long-term monitoring programs, the necessary shift from species-specific to community-specific response [or "differentiating species" versus "leading indicators" (Lange-Bertalot, 1978)] to environmental changes has been recognized quite early and it is obvious today that this principle represents the best solution. Numerous publications (Lange-Bertalot, 1978; Round, 1991; Sabater et al., 1991; Schiefele & Schreiner, 1991; van Dam et al., 1994; Krstić, 1995; Cazaubon, 1996; Coring, 1996; Dell'Uomo, 1996; Sabater et al., 1996; Bukhtiyarova, 1999; Loez & Topalián, 1999; Guasch et al., 1999; Hall & Small, 1999; Kwandrans et al., 1999; Rasiga, 1999; Battegazzore et al., 2004 and many others) clearly emphasize the same dominant diatom communities (Figure 6) present in specific water quality habitats throughout Europe and some even worldwide (e.g. Watanabe et al., 1988 for Japan or Fore & Grafe, 2002 for U.S.A.). This approach overcomes almost all stated obstacles for using diatoms in rapid and precise monitoring systems, based on following advantages:

it is based on a small number of taxonomically well known species belonging to communities that have precise responses to environmental conditions. In this way, long lists of taxa and taxonomical problems in relation to out-ecological characters of numerous taxa are excluded from the regular monitoring;

no need for various indices and complex statistics, since only the basic similarity/dissimilarity indices are useful in comparison to established reference site;

well trained monitoring personnel, educated in diatom (algal) taxonomy, can very easily and rapidly evaluate the water quality class just by preparing a slide from the epilithic diatom assemblages even at the field, in a matter of minutes;

classical methodology for diatom analysis (like Round, 1993) and live diatom material (Cox, 1996), enable wide application of relatively simple and reliable standard methods that produce comparable results. Permanent diatom slides also represent the history of established monitoring system as excellent data base record.

CONCLUSIONS

By comparing various organisms in freshwater monitoring system proposed in WFD (EU Water ..., 2000), one can confidently conclude that algae, particularly diatoms, react swiftly on changes of environmental parameters by creating easily detectable specific communities. Other algae, most of all 'blooming' cyanobacteria, can also provide very useful informations on overall environmental conditions and therefore should be included in regular monitoring programs. Monitoring of other hydrobionts, as well as the changes in morphology features of a specific ecosystem, is more useful regarding the bioaccumulation processes, overall

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state of health, intensity of prolonged impact, but also bioremediation and water management activities.

SELECTING APPROPRIATE BIOINDICATOR 17

Figure 6. Representative diatom communities for ecologically different water habitats in Macedonia – A. 1 – Diatoma hyemalis (Roth) Heib. – springs on Kozuf Mountain. B. 1 – Tabellaria floculosa (Roth) Kütz.; 2 – Diatoma mesodon (Ehr.) Kütz.; 3 – Meridion circulare var. constrictum (Ralphs) Van Heur.– Lake Crno, Shar Mountain. C. 1 – Navicula lanceolata (Ag.) Ehr.; 2 – N. tripunctata (Müll.) Bory; 3 – N. trivialis Lange-Bertalot; 4 – Nitzschia linearis (Ag.) W. Smith; 5 – Nitzschia dissipata (Kütz.) Grun. – River Vardar. D. 1 – N. amphibia Grun.; 2 – Melosira varians Ag. – Arachinovo Wetland. E. 1 – Nitzschia dissipata (Kütz.) Grun.; 2 – Cyclotella meneghiniana Kütz.; 3 – Gomphonema parvulum Kütz.; 4 – Nitzschia amphibia Grun.; 5 – N. palea (Kütz.) W. Smith – River Vardar. F. 1 – Cymbella excisa var. aff. procera Kram.; 2. Nitzschia capitellata Hust.; 3 – Melosira varians Ag. – Doiran Lake.

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