SCOPE 53 - Methods to Assess the Effects of Chemicals On Ecosystems

18.1 INTRODUCTION

Assessment of the environmental impact of chemicals at the ecosystem level raises theoretical and methodological problems of substantial magnitude.

Historically, the assessment of the potential environmental impact of a chemical has been achieved through laboratory testing¾overwhelmingly on single species; recently, multispecies testing has been attempted to increase understanding about the toxicity of chemical (Cairns and Orvos, 1989; Levin et al., 1990).

The prediction of toxicity has also been attempted by using quantitative structure-activity relationships (QSR) (Calamari, 1990), and in some special instances by using dispersion models (MacKay and Paterson, 1981) to compare the potential impact on sensitive species, serving as bioindicators, with the expected lethal concentration for a specified environment. The predictive value of microcosm toxicity tests to predict hazards has been hotly debated, but the testing of new chemicals in mesocosms, especially pesticides, has been more recently developed in several countries, and even regulated by the USEPA in the late 1980s (Vosheel, 1989).

For studies in natural habitats, biomarkers have been developed to improve the estimation of sublethal exposure on populations of critical species in the wild (Peakall, 1992). They provide increased accuracy in the estimation of impacts from chronic exposures to defined chemicals in an environment. The occurrence in nature of tolerant strains of exposed species has also been proposed as an indication that some threshold of toxicity for a substance has been reached (Blanck et al., 1988).

Whatever their usefulness, these methods are deficient in assessment¾at the ecosystem level¾of the impacts of chemical mixtures such as may be found in industrial effluents (Cairns and Orvos, 1989).

Several environmental toxicologists believe that the properties of an ecosystem could be equated to the sum of the properties of its individual components¾a widespread and misleading concept in ecotoxicology. Were this concept valid, the study of bioindicator organisms from among those estimated to be the most sensitive would be adequate for an overall assessment of potential damage. This hypothesis is ecologically unsound for several reasons. First, many detrimental effects (particularly impairment of reproductive performance, reductions in development or growth) may occur at concentrations well below those causing lethality. Second, even if perfectly understood, the toxicity of a chemical on a specific population is of little value in characterizing the toxicity that may be manifest in the whole ecosystem.

Therefore, the current approach of ecological risk assessment must be replaced with a paradigm that assures that the toxicity of chemicals are studied throughout entire ecosystems. Therefore, an urgent need exists to develop and validate methods to more accurately assess impacts of chemicals at the ecosystem level. Such methods will require a strong emphasis on basic ecological research. Some ecological catastrophes (like that of the oil spill from the foundering of the Exxon Valdez) have emphasized this need at a practical level, namely, the appraisal of ecological damage to determine fair levels of compensation to those who suffered economic loss.

18.2 ASSESSMENT OF EFFECTS ON COMMUNITY STRUCTURES AND DYNAMICS

Natural communities consist of complex assemblies of hundreds to hundreds of thousands of species that are in dynamic equilibrium and that interact with the complex physico-chemical components of their ecosystems.

The biota specific to an ecosystem play a major role in fundamental ecological processes, and modify its physical and chemical environment. Conversely, the ecosystem influences the composition and diversity of the community. Thus, it is most important for the ecological risk assessment to appraise the effects of chemical stresses on the community structure and dynamics. Such studies require specific methodologies to estimate the potential effects of substances on living resources.

18.3 PREDICTING RISKS IN POPULATIONS

18.3.1 REDUCTIONS IN POPULATION SIZE AND DENSITY

The most obvious and impressive effect of a chemical spill is the elimination of most organisms poisoned in the impacted area. Several such dramatic consequences have been observed recently. For example, on 30 October 1986, in Switzerland, the Sandoz accident killed most fish over several hundred kilometres of the Rhine river downstream from Easel. Even when contamination is less dramatic, ample evidence indicates that the introduction of toxic chemicals in terrestrial or aquatic habitats results in reduction of density, and even complete extinction, of populations of the most sensitive species.

By contrast, chronic exposure to chemicals occurring in terrestrial or aquatic habitats may lead to extensive decline of the exposed populations of animal or plant species. The author's investigations have demonstrated that the permanent exposure of a macroinvertebrate community to low concentrations of organochlorine insecticides occurring in ponds located among field areas intensively sprayed resulted in a sharp decrease in density and in the disappearance of several families of benthic invertebrates, especially those most pollutant-sensitive (Ramade et al., 1983, 1985; Ramade, 1987a).

Mortality from either acute or chronic exposures is the most obvious toxic manifestation and, thereby, the most frequently used in the assessment of impacts on populations. However, a chemical can cause non-lethal damage, such as a decline in fecundity and reproductive success, the impairment of mating in vertebrates, the slowing of growth, a delay in, or an impairment of, metamorphosis in invertebrates, inefficient pollination or gamete proliferation in algae, and partial inhibition of photosynthesis in plants. Any of these changes can have a deleterious effect on populations.

18.3.2 PREDICTING EFFECTS ON POPULATIONS

To minimize the impact of chemical pollution on wild populations and on the whole biota, safety standards for environmental protection have been promulgated. The most widely recognized environmental safety standard for any potentially dangerous substance is the maximum acceptable toxicant concentration (MATC).

The MATC is a measure of the lethal dose of a substance in a test of a single species with a vast array of organisms. An estimate of the concentration that will protect 95 percent of the taxa occurring in a given community exposed to a test chemical is calculated; therefore, only the most sensitive 5 percent are affected. The MATC is the geometric mean of either the no-observed-effect concentration (NOEC) or of the lowest-observed-effect concentration (LOEC) in chronic exposures.

Other similar methodologies have been devised by Van Leeuwen (1990). In the Netherlands, the assessment of risks to ecosystems from toxic chemicals relies on the HCp method of Van Straalen and Denneman (1989), where HCp is the hazardous concentration for a percentage p of the exposed species:

where Xm is the sample mean of log NOEC for m test species, S_m is the sample standard deviation of log NOEC values for m test species, d₁ is the fraction of the ecosystem that is not protected (recommended value d₁ = 0.05), d₂ is the probability of overestimating the HCp (recommended value 0.05), d_m is the value such that the probability of (S_m > d_m) = d₂, and T is the application factor between HCp and e^xm.

Several limitations accompany these methods, including the accuracy which decreases substantially from the population to the community level and even further from the community to the whole ecosystem. A major limitation is the lack of concordance between findings from laboratory tests and those from field investigations. As Cairns and Orvos (1989) pointed out, "this is probably responsible for the rather curious evolutionary process of laboratory tests becoming more and more sophisticated and the methodology more complex and, most important, more isolated from mainstream ecology."

Apart from the admonition that the distribution of sensitivities of the species studied is log-normal, a major criticism of these methods is that they rely on the assumption that a community is protected if the MATC is not exceeded, implying that the NOEC would be exceeded for only 5 percent of the species. This assumption is invalid if any critical species are among the 5 percent affected. Moreover, these predictive tests almost exclusively measure lethality, ignoring other toxic reactions. In many countries, this shortcoming stems from the fact that the so-called predictive tests "have been driven more by regulatory convenience than by sound ecological practices" (Cairns and Orvos, 1989).

The current methods to predict effects at the population level do not identify discrete changes in population size of any critical species; thus, a dominant or key species can trigger major changes in the overall structure of an entire community. The effects of chemicals on key species has been poorly investigated, particularly in some temperate ecosystems (Levin et al. , 1990). For example, many plant pollinators, although discrete and minute, may prove to be species essential to a community, so that an undetected decline in their densities could generate detrimental changes jeopardizing the future of the whole ecosystem (Sheehan et al. , 1984).

A non-toxicological factor is food shortages that change susceptibility to a toxicant in a prey species, which is thought highly important ecologically, but which is also ignored in the protection of biota against chemical pollution. The decline of the population upon which a predator preys may have considerable impact on the predator population despite the absence of toxicity of the chemical for this species. Several examples demonstrate this phenomenon: (1) in Britain, the decline of the grey partridge due to the rarefaction of non-target insects in cereal fields sprayed by pesticides (Rands, 1985) and (2) the breeding success of the blue tit (Parus caeruleus) in forests sprayed by cypermethrin, due to the reduction of caterpillars, that are the basic diet of nestlings (Pascual and Peres, 1992). However, the development of multispecies predictive tests, even when accepted by regulatory authorities, still has not increased efficiency in conserving the exposed community. Accordingly, only the monitoring of polluted ecosystem and site-specific validation of the effects of a given chemical can provide the additional ecological data requested.

18.3.2.1 Reduction in diversity and species richness

The reduction in density and diversity of species in chronically polluted habitats is the primary contributor to alterations in community structure. Indicator species are also relevant to assess effects on communities. Despite several criticisms regarding their use, they provide valuable information on the early effects of chemicals on biota.

The properties required from an effective bioindicator are the ability of its population to respond to discrete changes in the environment induced by chemicals, hypersensitivity to stimuli, reliability and specificity of its responses, rapidity of response, and ease of monitoring. If these prerequisites are met, discrete changes in carefully selected indicators would measure both the rate and magnitude of change induced by a chemical in a community.

18.3.2.2 Species diversity for use as an ecological index

Conceptually, ecological diversity integrates both the number of species (richness) and their relative frequency in a given biota. Decreased diversity has been used to assess gross environmental degradation in ecosystems for several decades. Among those indices that are the most frequently applied to chemical pollution, that of Margaleff are presented:

where N = total number of individuals, S = total number of species, n_i = number of individuals of the ith species.

The major limitation of this index is its need to count all individuals from the stressed community. Thereafter, for practical applications, ecotoxicologists selected other indices that could be applied to samples, because communities cannot usually be entirely numbered. The index of Lyod, Zad, and Karr was routinely applied to the study of water pollution of the Seine river:

where C is the number of class of frequency expressed in bits (for 10 class, C =3.3219).

Shannon's index has been by far the most widely used in ecotoxicology:

Several limitations hamper the effectiveness of the diversity index to assess the effects of chemicals on community structure. Shannon's index, for example, gives the same weight to systematic units of the same abundance, whatever their taxonomic level and a fortiori affinities. To avoid these limitations, Osborne, Davis, and Linton have proposed the use of a hierarchical diversity index (HDI) devised by a formula expanded from Pielou to include three taxonomic levels (familial, generic, and specific):

where H'(F) is the familial component of the total diversity, H'(G) is the generic component, and H'(S) is the specific component of the total diversity.

The most universal criticism of the application of the Shannon diversity measure in ecotoxicology is the misleading interpretation of data from communities influenced by a large evenness component (Godfrey, 1978).

Another problem in using the diversity index is the absence of linearity of a community response to a given gradient of increasing pollution, which is never univocal. During an initial stage, at sublethal exposure, some dominant species among the most pollutant sensitive will decrease their abundance, increasing the evenness and, therefore, the diversity index value. Only the onset of lethal exposures triggers the disappearance of species less pollutant-sensitive, thereby lowering the index (Figure 18.1).

Boyle et al. (1990) have conducted a theoretical study to assess the validity of 16 indices of water quality. They started with three communities differing in species richness but presenting the same abundance-value distribution curves. Their overall conclusion was that while the Shannon index affords a good representation of changes in species richness, it generally does not provide much representativeness of changes occurring in polluted communities and can even produce misleading conclusions.

Figure 18.1. Theoretical variation of the diversity index inside a gradient of increasing concentration of a given chemical (Ramade, 1987b).

In conclusion, the diversity index reflects changes in ecosystems structure only during periods of severe stress. In moderately polluted ecosystems, changes in dominance strongly affect equatability, thereby hampering the effectiveness of diversity indices in distinguishing degraded communities from unstressed ones. The index can be improved in estimating changes in abundance in a whole community, because the changes due to chemical pollution on a taxonomic group has been identified previously, and has been carefully studied for its value as indicator. The lichens are a sound illustration of such a use to assess the level of air pollution by SO₂ and predict the ecological effects associated with the observed levels. For example, an index of atmospheric purity (IPA) was derived from biocoenotic surveys revealing the species diversity of lichen communities in relationship with the level of SO₂ pollution):

where S = number of lichen species in the area sampled,  f = frequency of each species, and Q = index of toxiphobia of each species.

18.3.2.3 Effects on frequency distribution of species

The means by which species populations are controlled is one of the most important aspects of the ecological niche. The presence of a pollutant in a habitat will either impact the area occupied by each species or the resources used by each species, depending upon the level of tolerance or sensitivity of the species. Consequently, the balance between the various parts of a community will be disturbed as pollutants force modifications in competition leading to the elimination of the most sensitive populations. Therefore, the frequency distribution of species will be distorted to varying degrees by the chronic pollution of an ecosystem. The use of importance value distribution curves can be used methodologically to assess the response of a community to a chemical contaminant.

Ramade et al. (1985) compared the importance value distribution curves computed from experimental data from two lentic communities of benthic invertebrates: one from ponds contaminated by insecticides and the other from meadow ponds used as a control (Figure 18.2). The latter fits well on a Preston distribution, whereas the community from fields ponds fits on an intermediary position between Preston's model and a log-linear one. Generally, communities strongly disturbed by any pollutant are prone to fit a log-linear model, which is applicable to communities living in constraining environments.

18.4 PRINCIPLES TO ASSESS THE EFFECTS OF CHEMICALS ON ECOSYSTEMS

An ecosystem is a structural and functional complex array of associations within a living community and with its physico-chemical environment. Consequently, a chemical may be assumed to induce changes in the community structure and function, notwithstanding possible corresponding changes in the physicochemical structure of the contaminated habitats due to changes in activity of its whole biota, particularly its decomposer biomass. Dissecting extensive studies on the impact of acid rain on lakes, the action of chronic pollution by a chemical on photosynthesis ranks among the major deleterious effects of chemicals on fundamental ecological processes.

Discrete actions of a chemical can affect the primary productivity of terrestrial or aquatic ecosystems. For example, it was demonstrated that SO₂ at only 10 ppb in the atmosphere may affect the primary productivity of the Scot pine (Pinus sylvestris) forest; no morphological or anatomical damage can be detected at this concentration (Grodzinsky et al., 1984). Phytoplanctonic algae may experience an inhibition of their photosynthetic activity at dosages lower than 1 ppb for some pesticides. Research has shown a substantial decrease in phytoplankton and filamentous algae from ponds contaminated by various herbicides (especially Chlortoluron, Triazine, and Neburon) coming from adjacent fields (Goacoulou and Echaubard, 1987).

 Figure 18.2. Comparison of importance-value curves computed for a community of benthic macroinvertebrates from field ponds chronically polluted by pesticides and from control ponds (Ramade et al., 1985)

Effects on secondary productivity, though less explored, stand as a major parameter by which to assess effects at the ecosystem level. Among the most investigated is the study of the action of acid rain and pesticides on invertebrates and on freshwater fisheries exposed to forest spraying or acidification (Ide, 1967).

For pesticides, a substantial decrease has been demonstrated in benthic macroinvertebrates productivity resulting from contamination of freshwater ponds by organochlorine insecticides run off from surrounding fields, despite the fact that concentrations in water and sediments were well below acutely toxic concentrations. In acidified lakes, a sharp decrease in abundance and biomass of zooplankton has been observed (Almer et al., 1974; Stenson and Oscarson, 1985). A sharp decline has been also observed in commercial salmon fishing as a result of organochlorine insecticide pollution as early as the 1960s. For example, the New Brunswick salmon streams were badly affected by DDT spraying intended to control the spruce budworm. The decrease in the secondary productivity was demonstrated to be not so much a consequence of the insecticide toxicity in fish, but of the collapse of the aquatic insect populations upon which young salmons normally feed (Ide, 1967). Effects of river acidification on the productivity of the salmon fishing industry has been also well documented, and careful studies have shown a strong decline directly related to the level of acidification experienced (Leivestad and Muniz, 1976; Watt et al., 1983).

Some attempts have been made to model the effects of a chemical on some major components of ecosystem structure or function to expand the risk analysis methodology for ecosystems. O'Neill et al. (1982) devised a method to extrapolate laboratory toxicity data to aquatic ecosystem effects such as decreased productivity or reduction in fish biomass. Called the standard water column model or SWACOM, this method requires translating laboratory data into changes in the parameters of an entire ecosystem. The extrapolation is accomplished with knowledge of toxicological modes of action, and by simulation of the effects of a toxic substance across different trophic levels¾accordingly on the relationship between nutrients, phytoplankton, zooplankton, and fish. Various scenarios are developed, and each models population interactions that alter both the level and the nature of the risk to ecosystem processes. Moreover, the method describes uncertainties associated with laboratory measurements and the extrapolations, and describes risks as probabilities.

Relying on this SWACOM method, O'Neill et al. (1982) simulated well the effects of phenol and quinoline in an aquatic ecosystem. However, Cairns et al. (1988) stressed that experimental simulation results must be validated to assure a high accuracy of the estimates; many ecosystem simulation units have relatively low environmental realism, making estimates highly imprecise. Consequently, field studies on natural ecosystems are essential to achieve this validation.

Since research on the effects of chemicals carried out in full-scale ecosystems may prove lengthy, complex, expensive, and inaccurate due to the vast number of variables, the mesocosm approach was developed in the 1970s to estimate the effects of chemicals on ecosystems (Mauck et al., 1976; Vosheel, 1989). The problem that still impedes assessment of the effects of chemicals by the mesocosm method stems from the expense that limits its use in ecotoxicology. Research is progressing (Caquet et al., 1992) across Europe to develop mesocosms that would be less expensive while still providing adequate representation of natural ecosystems.

18.4.1 EFFECTS ON DECOMPOSERS AND NUTRIENTS CYCLING

Another major parameter of ecosystem functioning is related to nutrient cycling. Impairment of the activity of decomposers by a chemical may lead to major alterations and even the collapse of the whole ecosystem. For example, air pollutants disrupt the nutrients cycle of saprophagous insects that feed on the dead litter in forest ecosystems, and the consumption of dead organic matter by the soil Arthropods is slowed considerably. Other studies have shown effects of the same magnitude on leaf decomposition in the waters of acidified lakes. Therefore, possible effects of pollutants must be assessed on microorganisms that decompose dead organic matter in assorted ecosystems as part of a sound ecological risk assessment.

The appraisal of possible effects on the biogeochemical cycles of major nutrients is also a major issue in risk assessments at the ecosystem level. Attempts have been made to judge impairment of the nitrogen cycle at ecosystem level: Mathes and Shultz-Berendt (1988) reported how Aldicarb alters nitrification at the agroecosystem level. From previous studies, Mathes and Weidemann (1990) proposed an integrated approach to assess changes caused by chemicals on terrestrial ecosystems, including comparisons between two parts of an ecosystem, one with the test substance and the other without to serve as a control.

This approach has several prerequisites, especially a suitable indicator system. The spatio-temporal scale of exposure and temporal variability must be considered. The selection of reference organisms should incorporate both short-lived and long-lived species.

18.5 CONCLUSIONS

Major conclusions from this review include the following:

1. The structural and functional complexity of ecosystems needs further research to identify key species and to improve the knowledge of bioindicators as a tool to identify the effects of chemicals on communities.

2. The need exists to accurately assess toxic effects and requires additional study to ascertain the impact of chemicals on fundamental ecological processes such as primary and secondary productivity at the decomposer level, on nutrient cycles, and on biogeochemical cycles.

3. Since the detrimental effects of chemicals are not usually restricted to an ecosystem but impact areas that include both terrestrial and aquatic habitats, landscape ecology must be included in any assessment of risk.

18.6 REFERENCES

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