Chapter 5* - THE USE OF BIOLOGICAL MATERIAL

5.1. Introduction

Natural events and anthropogenic influences can affect the aquatic environment in many ways (see Chapter 2): synthetic substances may be added to the water, the hydrological regime may be altered or the physical or chemical nature of the water may be changed. Most organisms living in a water body are sensitive to any changes in their environment, whether natural (such as increased turbidity during floods) or unnatural (such as chemical contamination or decreased dissolved oxygen arising from sewage inputs). Different organisms respond in different ways. The most extreme responses include death or migration to another habitat. Less obvious responses include reduced reproductive capacity and inhibition of certain enzyme systems necessary for normal metabolism. Once the responses of particular aquatic organisms to any given changes have been identified, they may be used to determine the quality of water with respect to its suitability for aquatic life.

Organisms studied in situ can show the integrated effects of all impacts on the water body, and can be used to compare relative changes in water quality from site to site, or over a period of time. Alternatively, aquatic organisms can be studied in the laboratory (or occasionally in the field) using standardised systems and methods, together with samples of water taken from a water body or effluent. These tests, sometimes known as biotests, can be used to provide information on the intensity of adverse effects resulting from specific anthropogenic influences, or to aid in the evaluation of the potential environmental impact of substances or effluents discharged into surface or groundwater systems. Most kinds of biological analysis can be used alone or as part of an integrated assessment system where data from biological methods are considered together with data from chemical analyses and sediment studies. A full appreciation of natural changes and anthropogenic influences in a water body can only be achieved by means of a combination of ecological methods and biotests. Sometimes these studies need to be carried out over a period of many years in order to determine the normal variation in biological variables as well as whether any changes (natural or unnatural) have occurred or are occurring. An example of a continuous programme of biological assessment using a variety of methods is that carried out in Lake Baikal, Russia (Kozhova and Beim, 1993).

It is not possible to describe in this chapter, in detail, all the methods and variations that exist for biological analysis of water quality. There are several comprehensive texts and reviews which cover this subject (e.g. Ravera, 1978; OECD, 1987; Newman, 1988; Abel, 1989) and the details of many of the methods are published in appropriate reports and journals. Since many biological methods have been developed for local use and are based on specific species, an attempt has been made in this chapter to give only the basic principles behind the methods. With the help of such information it should be possible to decide whether such methods are applicable to the water quality assessment objectives in question. In many cases, when such methods are chosen, it will be necessary to adapt the basic principle to the local hydrobiological conditions, including the flora and fauna.

5.2. Factors affecting biological systems in the aquatic environment

5.2.1 Natural features of aquatic environments

The flora and fauna present in specific aquatic systems are a function of the combined effects of various hydrological, physical and chemical factors. Two of these factors specific to water bodies are:

· The density of the water, which allows organisms to live in suspension. Organisms which exploit this are called plankton, and consist of photosynthetic algae (phytoplankton), small animals (zooplankton) which feed on other planktonic organisms and some fish species which feed on other plankton and/or fish. The development of a rich planktonic community depends on the residence time (or retention time) of the water in the water body (see sections 6.4.2 and 7.2.5). Fast flowing water tends to carry away organisms before they have time to breed and to establish populations and, therefore, planktonic communities are more usually associated with standing waters such as lakes and reservoirs. As many fish are strong swimmers they are able to live in rivers, provided there are suitable breeding grounds present (see sections 6.4.1 and 6.4.2). Organisms living permanently in fast flowing waters, require specific adaptations to their body shape and behaviour (see section 6.4.1).

· The abundance of dissolved and particulate nutrients in the water. The constant supply of these often allows diverse and rich communities of planktonic and benthic (those living in or on the bottom) organisms to develop. An abundance of dissolved nutrients in shallow, slow flowing or standing waters allows the growth of larger aquatic plants (macrophytes), which in turn provide food, shelter and breeding grounds for other organisms.

The photosynthetic organisms which depend on the dissolved nutrients and sunlight for their own carbon production are termed the primary producers. These organisms are the food source of the zooplankton and small fish (secondary producers), which in turn are the food source of other fish (tertiary producers). This simplified system is known as the food chain and, together with the processes of decay and decomposition, is responsible for carbon transfer within the aquatic environment. In practice, the interactions between different groups are more complex and may be referred to as the food web. For more detailed information on the fundamentals of biological systems in water bodies see Hutchinson (1967), Hynes (1970), Wetzel (1975), Whitton (1975) and Moss (1980).

5.2.2 Anthropogenic influences on water bodies

In addition to natural features, biological communities are often affected directly by human activities (such as inputs of toxins, increased suspended solids, habitat modification or oxygen depletion) or indirectly by processes influenced by anthropogenic activities (e.g. chelating capacity).

The variety of effects that can be observed on different aquatic organisms as a result of anthropogenic influences can be demonstrated by the example of domestic sewage. Purely domestic sewage, without the input of modern, synthetic, harmful substances, such as chlorinated hydrocarbons, detergents etc., is characterised by high concentrations of easily biodegradable organic matter. It also contains high concentrations of bacteria, viruses and other pathogens from which water-borne diseases may arise. During the process of biodegradation of sewage in a river there is an initial rapid decline in oxygen concentration in the water resulting from microbial respiration during self-purification. However, microbial activity also leads to an increase in nutrient content and sometimes other harmful substances are formed such as hydrogen sulphide or ionised ammonia (Figure 5.1). Hydrogen sulphide is very toxic, but ionised ammonia is a nutrient which is more easily assimilated than nitrate. However, if the pH value exceeds 8.5, a rapid increase in unionised ammonia occurs (see Figure 3.2) which is very toxic to fish. Phosphate also becomes available following the biological decomposition of domestic sewage. These changes in the chemical composition of the water are followed by significant changes in the structure of the biota, some of which exploit the increased nutrients and others which can tolerate reduced oxygen concentrations (Figure 5.1). Such changes form the basis of water quality assessments using biota as indicators of the intensity of pollution.

Figure 5.1 Typical effects on water quality and the associated biota which may be observed downstream of a sewage outlet. A and B. Physical and chemical changes; C. Changes in micro-organism populations; D. Changes in invertebrate populations (After Hynes, 1960)

5.2.3 Physical alterations in the aquatic environment

The presence or absence of specific aquatic organisms depends on the physical environment and its associated habitats, such as fast flowing water with large stones or boulders or still waters with fine deposited sediments. Although these environments can easily be modified by human activities, including river damming, canalisation and drainage schemes, natural changes can occur in the physical environment due to local climatological and geographical conditions. Events such as torrential rain storms or prolonged droughts can lead to sudden or gradual modifications of a natural habitat, e.g. by increased siltation or scouring of river beds, which in turn lead to changes in the flora and fauna of the water body. These changes can be quite dramatic, including short term or permanent loss of species. It is important to understand the hydrological regime of water bodies when designing biological assessment programmes so that effects due to natural changes in the environment can be separated from those that may be caused by human activities.

5.2.4 Dissolved oxygen

Oxygen is an important factor for aquatic life and the chemical characteristics of the environment. Concentrations less than 100 per cent saturation can occur normally under certain circumstances, e.g. at the bottom of nutrient rich lakes (see Figure 7.8), or at night in slow flowing rivers (see Figures 6.19 and 6.20). In such locations, species may be found which are adapted to low concentrations of oxygen. Under normal conditions these species would be rare, but they can become widespread in association with pollution or nutrient enrichment. However, many species are able to survive a potentially harmful lack of oxygen for a short time, but rarely for days or many hours. The ability of organisms to survive different levels of oxygen depletion in water forms the basis of some biotic indices and water quality assessment methods. Tolerance of low concentrations of oxygen varies from species to species, even within the same genus and, therefore, it is more appropriate to work at the species level for some biological assessment methods. Further details are given in section 5.4.

5.2.5 The duration of exposure

The duration of exposure, or influence on the organism, is generally the period of effective concentration of the contaminant or other variable of interest in the environment, or in a laboratory test system. In a biological sense this is the duration of actual exposure of an organism to a harmful concentration, or an effective concentration of, for example, a substance which can be bioaccumulated (see section 5.8). In some field situations the actual period of influence may be longer than the measured duration of unusually high contaminant concentrations. Alternatively, it may be shorter as a result of incomplete mixing, such as occurs at the beginning of the “toxic wave” arising from a point source input to a river. Sometimes, as a result of incomplete mixing, high concentrations may occur only on one side of a river or lake, or near the bottom. Adverse effects may be felt by organisms which are free living in the water or which live in, or on, the substrate and are not able to escape from the dangerous area. In severe cases of toxic pollution or deoxygenation a “fish-kill” may occur. In many cases, long after the fish-kill is over, the continuing absence of organisms which would otherwise be present (together with the benthic fauna which colonise the surfaces of the bottom of the river, i.e. stones, gravel, sand or mud) enables the investigator to establish the severity of the event, and the length of the affected stretch of the river.

The body of an organism takes some time (seconds or longer) to absorb a toxin and then react. Nevertheless, many aquatic organisms react very rapidly, especially against toxic substances, and this can be an advantage for the development of biomonitoring techniques. Alternatively, toxic substances which are accumulated gradually until they reach harmful concentrations which produce sudden or very subtle effects in the organisms, present a particular problem. Nutrient absorption by aquatic organisms is usually rapid but their subsequent growth takes time. Consequently, the effect of nutrient enrichment (eutrophication) in a water body is a long-term effect. In many rivers, the effects of eutrophication may occur some distance downstream of the source of nutrients. In lakes, the effects of nutrient enrichment also occur sometime after increased nutrient levels begin (for details see section 7.3.1).

5.2.6 Concentration

The physiological or behavioural reactions of aquatic organisms depend on the concentration of natural substances and pollutants in the environment, and the time required for these substances to affect the internal systems of the organisms. The actual environmental concentration of a substance or compound which produces toxic effects in an organism can also be influenced by many other environmental factors (e.g. presence of other toxins, inadequate food supply, and physical factors such as habitat alteration, sedimentation, drought or oxygen depletion). An organism under stress will not be able to survive the same concentration of a contaminant as when its environmental conditions are optimal. Consequently, the toxic effects determined by laboratory tests may vary with different experimental circumstances. Many substances also have significant differences in their toxicity to different species. Therefore, to determine environmental effects fully, it is necessary to use a set of tests under standardised circumstances.

The reactions of organisms are often not very specific in relation to any given concentration, but can be observed mainly in relation to exceeding, or remaining below, a “no observed effect concentration” (NOEC). The NOEC plays an important role in international discussions on water quality management, as it is important in determining the toxicity of substances and setting priorities for control measures for effluents discharged to freshwaters. This specialised topic is beyond the scope of this guidebook.

5.2.7 Chelating capacity

Chelation is the ability of organic compounds to bind metal ions and maintain them in solution. Examples of chelating agents are humic and fulvic acids and compounds such as EDTA (ethylenediaminetetraacetic acid). These compounds can also slowly release bound metal ions back into the water. The chelating capacity of the water, therefore, depends on the content of humic acids and other ligands, as well as on the hardness of the water (see section 3.3.11). Hardness plays an important role in the distribution of aquatic biota and many species can be distinguished as indicators for hard or soft water. Organisms with shells which are composed of calcium carbonate need high concentrations of calcium in the water, whereas stoneflies and some triclad worms are characteristic of soft water. Different requirements can be found within the same family of organisms. The microcrustacea Gammarus pulex and G. roeseli have a preference for hard water and can survive some depletion of oxygen, whereas G. fossarum is more sensitive to organic pollution and oxygen depletion but can survive in less hard water. However, G. fossarum cannot tolerate very soft water although the closely related genus Niphargus lives in soft water wells and clean mountain rivers which are very poor in calcium.

The toxicity of trace elements to a given species may also vary according to the water hardness. For example, the toxicity of copper and zinc varies over a wide range depending on the concentration of calcium in the water. The higher the concentration of calcium, the lower the toxicity of both metals. As a result the European Union directive for the protection of water as a habitat and spawning ground for fish (CEC, 1978) gives different concentration limits for zinc and copper for different degrees of water hardness (Table 5.1). However, zinc is much more toxic to bacteria than to all other organisms, including man, and bacteria are the main organisms responsible for self-purification in freshwaters. Therefore, with respect to the total environment, it is necessary to look for the most sensitive components of the system in order to establish permissible limits for toxic compounds. The toxicity of metals may be reduced in waters high in humic acids (often known as brown waters, e.g. Rio Negro, Brazil) as a result of their chelating capacity. Examples from the temperate zone include rivers and lakes in peat bog areas, both of which have specific communities of plants and animals.

5.2.8 Acidity

Some organisms are sensitive to the acidity or alkalinity of water. Aquifers, rivers and lakes situated in catchment areas consisting of acid rocks or pure quartz have waters poor in calcium and magnesium with a low buffering capacity. Such water bodies are widely distributed in North America (the Laurentian shield), Scandinavia and specific areas of the arctic, temperate and tropical zones. In these water bodies, additional acid input from “acid rain” produces a drop in the pH of the water and may result in increased concentrations of reactive aluminium species released from the soil of the surrounding watershed (Meybeck et al., 1989). Both low pH (lower than 5.5) and increased aluminium content are toxic to many invertebrates and fish. Consequently, “acid rain” has led to a reduction of fish populations and other organisms in lakes and rivers which had previously been excellent fishing grounds in North America, Scandinavia and other parts of the world (Meybeck et al., 1989). Sudden effects due to acidification of waters can also occur after heavy rain and snow melt. Although acid waters occur naturally, evidence now suggests that acidification arising from “acid rain” is also occurring in some tropical countries, e.g. Brazil and parts of South East Asia. Detailed information is available in specialised texts, e.g. Rodhe and Herrera (1988) and Bhatti et al. (1989). Other adverse effects associated with acidification arise from the mobilisation of mercury and cadmium, both of which are highly toxic and can be accumulated by fish in their body tissues (see section 5.8).

Table 5.1 Permissible concentrations of copper and zinc in freshwaters of different degrees of hardness according to the European Union directive for the support of fish life in freshwaters

Source: CEC, 1978

5.3. Uses and benefits of biological methods

5.3.1 Biological effects used for assessment of the aquatic environment

A variety of effects can be produced on aquatic organisms by the presence of harmful substances or natural substances in excess, the changes in the aquatic environment that result from them, or by physical alteration of the habitat. Some of the most common effects on aquatic organisms are:

· changes in the species composition of aquatic communities,
· changes in the dominant groups of organisms in a habitat,
· impoverishment of species,
· high mortality of sensitive life stages, e.g. eggs, larvae,
· mortality in the whole population,
· changes in behaviour of the organisms,
· changes in physiological metabolism, and
· histological changes and morphological deformities.

As all of these effects are produced by a change in the quality of the aquatic environment, they can be incorporated into biological methods of monitoring and assessment to provide information on a diverse range of water quality issues and problems, such as:

· the general effects of anthropogenic activities on ecosystems,

· the presence and effects of common pollution issues (e.g. eutrophication, toxic metals, toxic organic chemicals, industrial inputs),

· the common features of deleterious changes in aquatic communities,

· pollutant transformations in the water and in the organisms,

· the long-term effects of substances in water bodies (e.g. bioaccumulation and biomagnification, see section 5.8),

· the conditions resulting from waste disposal and of the character and dispersion of wastewaters,

· the dispersion of atmospheric pollution (e.g. acidification arising from wet and dry deposition of acid-forming compounds),

· the effects of hydrological control regimes, e.g. impoundment,

· the effectiveness of environmental protection measures, and

· the toxicity of substances under controlled, defined, laboratory conditions, i.e. acute or chronic toxicity, genotoxicity or mutagenicity (see section 5.7).

Biological methods can also be useful for:

· providing systematic information on water quality (as indicated by aquatic communities),

· managing fishery resources,

· defining clean waters by means of biological standards or standardised methods,

· providing an early warning mechanism, e.g. for detection of accidental pollution, and

· assessing water quality with respect to ecological, economic and political implications.

Examples of some of these uses are mentioned in the following sections, and in Chapters 6, 7 and 8 in relation to rivers, lakes and reservoirs.

5.3.2 Advantages of biological methods

Biological assessment is often able to indicate whether there is an effect upon an ecosystem arising from a particular use of the water body. It can also help to determine the extent of ecological damage. Some kinds of damage may be clearly visible, such as an unusual colour in the water, increased turbidity or the presence of dead fish. However, many forms of damage cannot be seen or detected without detailed examination of the aquatic biota.

Aquatic organisms integrate effects on their specific environment throughout their lifetime (or in the case of laboratory tests, during the period of exposure used in the test). Therefore, they can reflect earlier situations when conditions may have been worse. This enables the biologist to give an assessment of the past state of the environment as well as the present state. The length of past time that can be assessed depends on the lifetime of the organisms living in the water under investigation. Micro-organisms, such as ciliated protozoa, periphytic algae or bacteria, reflect the water quality of only one or two weeks prior to their sampling and analysis, whereas insect larvae, worms, snails, and other macroinvertebrate organisms reflect more than a month, and possibly several years.

When biological methods are carried out by trained personnel they can be very quick and cheap, and integrated into other studies. Compared with physico-chemical analysis, much less equipment is necessary and a large area can be surveyed very intensively in a short time, resulting in a large amount of information suitable for later assessment. Recent developments in water quality assessment, especially for the purpose of effluent control, have begun to include bioindicators and tests such as bioassays (as in Germany under the “Waste Water Levy Act”). The costs of chemical analytical equipment, trained personnel and materials, repairs and energy consumption are enormous due to the number of different polluting substances that now have to be legislated and controlled. In some situations biological methods can offer a cheaper option. The advantages of biological methods, however, do not eliminate the need for chemical analysis of water samples. Agencies and individuals responsible for establishing assessment programmes must integrate both methods to provide a system which is not too expensive and which provides the necessary information with maximum efficiency.

Acute toxicity testing (see section 5.7.1) is particularly useful in cases of emergency and accidental pollution where it can minimise the amount of chemical analysis required. When investigating a fish-kill, samples of the water are usually taken for analysis in order to determine the cause. However, if a toxicity test (using an aquatic organism in samples of the contaminated water) is conducted immediately in parallel to the chemical analyses it is possible to ascertain whether toxic concentrations are present in one, or all, of the samples taken. This initial “screening” enables the chemical laboratory to focus their efforts on the most toxic water samples and helps the water quality managers and decision-makers prepare (or stop) further action. Immediate remedial action may, therefore, be possible although the reaction of the biota in the test does not necessarily give specific information about the type of substance causing the toxicity, or an indication of the concentration present.

5.3.3 Types of biological assessment

Biological assessment of water, water bodies and effluents is based on five main approaches.

(i) Ecological methods:

· analysis of the biological communities (biocenoses) of the water body,
· analysis of the biocenoses on artificial substrates placed in a water body, and
· presence or absence of specific species.

(ii) Physiological and biochemical methods:

· oxygen production and consumption, stimulation or inhibition,
· respiration and growth of organisms suspended in the water, and
· studies of the effects on enzymes.

(iii) The use of organisms in controlled environments:

· assessment of the toxic (or even beneficial) effects of samples on organisms under defined laboratory conditions (toxicity tests or bioassays), and

· assessing the effects on defined organisms (e.g. behavioural effects) of waters and effluents in situ, or on-site, under controlled situations (continuous, field or “dynamic” tests).

(iv) Biological accumulation:

· studies of the bioaccumulation of substances by organisms living in the environment (passive monitoring), and

· studies of the bioaccumulation of substances by organisms deliberately exposed in the environment (active monitoring).

(v) Histological and morphological methods:

· observation of histological and morphological changes, and
· embryological development or early life-stage tests.

Some of these methods are widely used in freshwaters although other methods have been developed for use in specific environments, or in relation to particular environmental impacts. The principal advantages and disadvantages of the major methods for freshwater quality assessment are given in Table 5.2.

Biological assessment is of increasing importance in many places around the world, and many methods have been developed which are being used on a national or local basis. In the following sections only the basic principles, and selected methods which are in more general use, are described. More details are available in specialised texts which summarise the literature on the topic (e.g. Alabaster, 1977; Burton, 1986; Hellawell, 1986; De Kruijf et al, 1988; Yasuno and Whitton, 1988; Abel, 1989; Samiullah, 1990).

5.4. Ecological methods

Aquatic organisms have preferred habitats which are defined by physical, chemical and other biological features. Variation in one or more of these can lead to stress on individuals and possibly a reduction in the total numbers of species or organisms that are present. In extreme situations of environmental change, perhaps due to contamination, certain species will be unable to tolerate the changes in their environment and will disappear completely from the area concerned, either as a result of death or migration. Thus the presence or absence of certain species or family groups, or the total species number and abundance, have been exploited as a means of measuring environmental degradation (see the example of the sewage input into a river in Figure 5.1 and section 5.2.2). Two main approaches have been used: methods based on community structure and methods based on “indicator” organisms. An indicator organism is a species selected for its sensitivity or tolerance (more frequently sensitivity) to various kinds of pollution or its effects, e.g. metal pollution or oxygen depletion. Some groups of organisms, such as benthic invertebrates, have been exploited more than others in the development of ecological methods and this is due to a combination of the specific role of the organisms within the aquatic environment, their lifestyle and the degree of information available to hydrobiologists (Table 5.3). Principal approaches based on the commonly used groups of organisms are described below. For further details see the specialised literature quoted.

5.4.1 Indices based on selected species or groups of organisms

Biological indices are usually specific for certain types of pollution since they are based on the presence or absence of indicator organisms (bioindicators), which are unlikely to be equally sensitive to all types of pollution. Such indices often use macroinvertebrate populations because they can be more easily and reliably collected, handled and identified. In addition, there is often more ecological information available for such taxonomic groups.

Table 5.2 Principal biological approaches to water quality assessment; their uses, advantages and disadvantages

¹ e.g. biotic indices

² e.g. diversity or similarity indices

Table 5.3 Advantages and disadvantages of different groups of organisms as indicators of water quality

Source: Based on Hellawell, 1977

It is frequently argued that the indicator organisms incorporated into biotic indices should be distributed world-wide. However, few animal and plant species have true global distributions apart from ciliated protozoa which are difficult to collect, preserve and identify. Those species which do occur world-wide probably have broad ecological requirements and are, therefore, generally not suitable as indicators. Recent developments in biological monitoring methods have favoured the use of families of organisms instead of species in order to limit the time and effort required in identifying organisms to the species level. Although the ecological requirements of families of organisms are often so broad that their use in biological indices is limited (Friedrich, 1992), the basic principles can be applied to develop regionally suitable methods. Recent experience in Brazil has shown that it can be useful, when establishing a new monitoring system based on bioindicators, to use a provisional form of taxonomy such as “morphological type” which can be observed at species level (Friedrich et al., 1990). A general problem associated with establishing such systems is the lack of basic data from physico-chemical analyses. Statistical correlations of species alone are not satisfactory.

The basic principles of some indices are described below. It must be stressed that in most cases the indices only work well for the water bodies in the regions in which they were developed, and that they may give anomalous results in other types of water body, largely due to natural variations in species distributions. When applying biological indices in other regions careful selection of the appropriate method must be made to suit local conditions (Tolkamp, 1985) and caution must be applied in interpreting the results. Alternatively, the indices can be modified to suit local conditions. Reviews of some of the widely used biotic indices for water quality assessment are available in Hellawell (1978) and Newman et al. (1992).

The Saprobic system and Saprobic Indices

At the very beginning of the twentieth century the effect of point source pollution from sewage discharges on aquatic fauna and flora downstream of urbanised areas became evident. Kolkwitz and Marsson (1902, 1908, 1909) were the first to exploit these effects and present a practical system for water quality assessment using biota. Their system, known as the Saprobic system, has been used mainly in Central Europe. It is based on the observation that downstream of a major source of organic matter pollution a change in biota occurs. As self-purification takes place further ecosystem changes can be observed, principally in the components of the biotic communities. Odour and other chemical variants in the water also change. Kolkwitz and Marsson were principally concerned with an ecological approach, dealing with biological communities and not purely with indicator species. Since the taxonomy of aquatic organisms in Central Europe is well developed, it is possible to use the species level (which is the most precise) in the Saprobic system developed in that region.

The Saprobic system is based on four zones of gradual self-purification: the polysaprobic zone, the a-mesosaprobic zone, the b-mesosaprobic zone, and the oligosaprobic zone. These zones are characterised by indicator species, certain chemical conditions and the general nature of the bottom of the water body and of the water itself, as described below:

· Polysaprobic zone (extremely severe pollution): Rapid degradation processes and predominantly anaerobic conditions. Protein degradation products, peptones and peptides, present. Hydrogen sulphide (H₂S), ammonia (NH₃) and carbon dioxide (CO₂) are produced as the end products of degradation. Polysaprobic waters are usually dirty grey in colour with a faecal or rotten smell, and highly turbid due to the enormous quantities of bacteria and colloids. In many cases, the bottom of the watercourse is silty (black sludge) and the undersides of stones are coloured black by a coating of iron sulphide (FeS). Such waters are characterised by the absence of most autotrophic organisms and a dominance of bacteria, particularly thio-bacteria which are particularly well adapted to the presence of H₂S. Various blue-green algae, rhizopods, zooflagellates and ciliated protozoa are also typical organisms in polysaprobic waters. The few invertebrates that can live in the polysaprobic zone often have the blood pigment, haemoglobin, (e.g. Tubifex, Chironomus thummi) or organs for the intake of atmospheric air (e.g. Eristalis). Fish are not usually present.

· a-mesosaprobic zone (severe pollution): Amino acids and their degradation products, mainly the fatty acids, are present. The presence of free oxygen causes a decline in reduction processes. The water is usually dark grey in colour and smells rotten or unpleasant due to H₂S or the residues of protein and carbohydrate fermentation. This zone is characterised by “sewage fungus”, a mixture of organisms dominated by the bacterium Sphaerotilus natans. The mass of organisms, which form long strands, can become detached from the bottom by the gas generated during respiration and decomposition processes, and then drift in the water column as dirty-grey masses. Frequently, they form a mat over the entire surface of the stream bed. Sewage fungus is particularly common in waters containing wastes rich in carbohydrates such as sewage and effluents from sugar and wood processing factories.

· b-mesosaprobic zone (moderate pollution): Aerobic conditions normally aided by photosynthetic aeration. Oxygen super-saturation may occur during the day in eutrophic waters. Reduction processes are virtually complete and protein degradation products such as amino acids, fatty acids and ammonia are found in low concentrations only. The water is usually transparent or slightly turbid, odour-free and generally not coloured. The surface waters are characterised by a rich submerged vegetation, abundant macrozoobenthos (particularly Mollusca, Insecta, Hirudinae, Entomostraca) and coarse fish (Cyprinidae).

· Oligosaprobic zone (no pollution or very slight pollution): Oxygen saturation is common. Mineralisation results in the formation of inorganic or stable organic residues (e.g. humic substances). More sensitive species such as aquatic mosses, planaria and insect larvae can be found. The predominant fish are Salmonid species.

Each of the four zones can be characterised by indicator species which live almost exclusively in those particular zones. Therefore, comparison of the species list from a specific sampling point with the list of indicator species for the four zones enables surface waters to be classified into quality categories, particularly when combined with other important and often characteristic details (e.g. generation of gas in sediments, development of froth, iron sulphide on the undersides of stones, etc.).

The above classification system has been used to design Saprobic Indices particularly for data treatment, assessment and interpretation in relation to decision making and management. The first Saprobic Index designed by Pantle and Buck (1955) has been modified by Liebmann (1962). The frequency of occurrence of each species at the sampling point, as well as the saprobic value of that indicator species are expressed numerically. The frequency ratings or abundance, a, are:

and the preferred saprobic zones of the species are indicated by the numerical values, s, as follows:

For any given species i the product of abundance a_i and saprobic zone preference s_i expresses the saprobic value S_i for that species, i.e. S_i = a_i s_i.

The sum of saprobic values for all the indicator species determined at the sampling point divided by the sum of all the frequency values for the indicator species gives the Saprobic Index (S) which can be calculated from the following formula:

The Saprobic Index S, a number between 1 and 4, is the “weighted mean” of all individual indices and indicates the saprobic zone as follows:

The Saprobic Index can be plotted directly against the distance along a river as shown in Figure 5.2. This is an example of an early use of biological data for engineers and decision makers. Note, however, that use of the index requires the organisms normally occurring in each of the river classification zones for a particular region to be known so that they can be assigned to a preferred saprobic zone during the calculation of the index. This information can only be obtained by detailed studies of the river systems, including precise identification of the individual species.

A comprehensive revision of the Saprobic system, carried out in 1973 by Sl醖ecek (1973), has been adopted for widespread use in Central and Eastern Europe (LAWA, 1976; Breitig and Von T黰pling, 1982). Based on many years of practical experience and a large amount of data (especially physico-chemical data from water analysis) the system has again been revised by a group of experts in Germany (Friedrich, 1990). The main revisions were:

· The species used must be benthic in order to reflect the situation at the place where they are found, since planktonic species reflect the situation at an unknown place upstream.

· Photoautotrophic species are no longer included to avoid interactions between indicating saprobity and indicating trophic status.

· The species selected must be able to be identified without doubt by trained biologists and not only by specialists in taxonomy.

· The species used must be distributed over most of Central Europe.

· All species included must be well known with respect to their ecological requirements.

Figure 5.2 Changes in the Saprobic Index of Pantle and Buck (1955) along the Werra river, Germany (After UNESCO/WHO, 1978)

In this revision, the organisms were assigned a saprobic value (s) between 1 and 20 for a more precise description of the ecological range of the species. The Saprobic Index is calculated using the formula of Zelinka and Marvan (1961) which takes account of the fact that very few species occur only in one saprobic zone. Using many years’ experience in Germany, species with very narrow ecological ranges have been distinguished from less sensitive ones and a weighting factor, g (a value of 1, 2, 4, 8 or 16) has been assigned to each organism and incorporated into the revised formula as follows:

Further details and the formula for calculating the uncertainty of this index are available in Friedrich (1990). This revised system has been designated as a German Standard Method (DIN 38410 T.2) and forms part of the basis of an integrated system of water quality classification, which includes biological and chemical variables.

The sampling procedure to be employed for collecting organisms for determination of the Saprobic and other biotic indices has been standardised at the national and international level (e.g. ISO, 1985).

Biotic indices

Alternative approaches to the Saprobic Index have been developed by Cairns et al. (1968), Woodiwiss (1964), Chandler (1970) and others. These methods are based on the presence or absence of certain “indicator” groups (e.g. Plecoptera, Ephemeroptera, Gammaridae), and/or “indicator” species, at the sampling point. As with the Saprobic Index, they are best suited to use in waters polluted with organic matter, particularly sewage, since the indicator organisms are usually sensitive to decreases in oxygen concentrations. However, a similar approach has recently been developed for the biological monitoring of acidification in streams and lakes using an “Acidification Index” based on the tolerance of invertebrates to acidity (Raddum et al., 1988; Fjellheim and Raddum, 1990).

The Trent Biotic Index was originally developed by Woodiwiss (1964) for assessing pollution in the River Trent in England and forms the basis for many similar types of index. The index is based on the number of defined taxa of benthic invertebrates in relation to the presence of six key organisms found in the fauna of the sample site. Depending on the number of taxonomic groups present and the key organisms found in the fauna, the index ranges from ten, for clean water, to zero for polluted water (Table 5.4). This index has been found to be rather insensitive and has been largely replaced by further developments of the basic principle which have become widely used in Europe. One variation, used in the UK, is the Chandler Biotic Index (Chandler, 1970). To derive the index for a particular river station the invertebrate fauna, collected according to a standard procedure, are identified and then counted. Each group is given a score according to its abundance as shown in Table 5.5. The total score represents the index and the higher the score the cleaner the water. Calculation of the Trent Biotic Index and Chandler Biotic Index using invertebrate samples collected from an upland stretch of a lowland river in England is illustrated in Table 5.6.

In order to limit the taxonomic requirement of earlier biotic indices to identify organisms to species level, some alternative indices have been developed which use only the family level of identification (Hellawell, 1986; Abel, 1989). An example is the Biological Monitoring Working Party-score (BMWP) which has been published as a standard method by an international panel (ISO-BMWP, 1979). This score was devised in the UK but was not specific to any single river catchment or geographical area. Invertebrates are collected from different habitats (e.g. gravel, silt, weed beds) at representative sites of river stretches. The organisms are identified to the family level and then each family is allocated a score between one and ten. The most sensitive organisms, such as mayfly nymphs score ten, molluscs score three and the least sensitive worms score one (Table 5.7). The BMWP score is calculated by summing the scores for each family represented in the sample. The number of taxa gives an indication of the diversity of the community (high diversity usually indicates a healthy environment, see next section). The average sensitivity of the families of the organisms present is known as the Average Score Per Taxon (ASPT) and can be determined by dividing the BMWP score by the number of taxa present. A BMWP score greater than 100 with an ASPT value greater than 4 generally indicates good water quality. An evaluation of the performance of the BMWP score in relation to a range of water quality variables is described by Armitage et al. (1983).

Table 5.4 Use of invertebrates to calculate the Trent Biotic Index

The bold index number represents the result of calculating the Trent Biotic Index using the species list in Table 5.6.

¹ The term “group” means any one of the species included in the list of organisms below (without requiring detailed identification)

² Baetis rhodani excluded
　

³ Baetis rhodani (Ephemeroptera) is counted in this section for the purpose of classification

Groups:

Each known species of Platyhelminthes (flat-worms)
Annelida (worms excluding genus Nais)
Genus Nais (worms)
Each known species of Hirudinae (leeches)
Each known species of Mollusca (snails)
Each known species of Crustacea (hog louse, shrimps)
Each known species of Plecoptera (stone-fly)
Each known genus of Ephemeroptera (mayfly, excluding Baetis rhodani)
Baetis rhodani (may-fly)
Each family of Trichoptera (caddis-fly)
Each species of Megaloptera larvae (alder-fly)
Family Chironomidae (midge larvae except Chironomus riparius)
Chironomus riparius (blood worms)
Family Simulidae (black-fly larvae)
Each known species of other fly larvae
Each known species of Coleoptera (Beetles and beetle larvae)
Each known species of Hydracarina (water mites)

Source: After Mason, 1981

Table 5.5 The use of invertebrates to calculate the Chandler Biotic Index

Source: After Mason, 1981

Table 5.6 Examples of the calculation of the Trent and Chandler Biotic Indices using invertebrates identified in a benthos sample taken from the upper stretch of a river in lowland England¹

¹ Combination of five two-minute kick samples collected in October