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A Critique of Effluent Bioassays
Clyde E. Goulden
The Clean Water Act of 1977 states, ''It is the national policy that the discharge of toxic pollutants in toxic amounts be prohibited" (Peltier and Weber, 1985, p. 1). Thorough assurance that this goal is met would require complete chemical profiles of every effluent; knowledge of the sensitivity of all potentially affected organisms to all chemicals in effluents, including both direct toxic effects and indirect effects, such as the effects of toxins on forage species; and an understanding of all synergistic interactions between compounds in effluents. It is not feasible to obtain such comprehensive information.
The U.S. Environmental Protection Agency (EPA) concluded that a cost-effective alternative approach would be to measure effluent toxicity by exposing aquatic organisms to effluents in "bioassays." Bioassays measure "the potency of any stimulus, physical, chemical, or biological, physiological or psychological, by means of the reactions that it produces in living matter" (Finney, 1952a, p. 1). The rationale is that through bioassays, test organisms reveal whether an effluent is toxic. This paper describes the development of the bioassay approach and evaluates whether through its use ecosystems can be sufficiently protected from toxic materials.
The History of Bioassays
The basic design of bioassays was developed in the nineteenth century, but test species were used as an assay of exposure to some stimulus well before that time. Finney (1952b) suggests that the basic principles of bioassays are found in early texts and quotes an example:
And it came to pass at the end of forty days, that Noah opened the windows of the ark which he had made:
And he sent forth a raven, which went to and fro, until the waters were dried up from off the earth.
Also he sent forth a dove from him, to see if the waters were abated from off the face of the ground;
But the dove found no rest for the sole of her foot, and she returned unto him into the ark, for the waters were on the face of the whole earth: then he put forth his hand, and took her, and pulled her into him into the ark.
And he stayed yet another seven days; and again he sent forth the dove out of the ark;
And the dove came in to him in the evening: and lo, in her mouth was an olive leaf pluckt off; so Noah knew that the waters were abated from off the earth. (Genesis, 8, vi-xi)
The three essential components of a bioassay are present in this example: a stimulus (the depth of water); a biological test subject (the dove); and a response (the plucking of an olive leaf).
Formal bioassays were developed to study the potency of insecticides during the early twentieth century at Rothamstead Station in England. Sir R. A. Fisher and other statisticians developed experimental designs and basic statistical procedures in collaboration with toxicologists and entomologists (Bliss, 1934a,b; Finney, 1952a; Gaddum, 1933).
Since that time, the role of bioassays has been expanded considerably. Prior to the 1970s, bioassays were used only to measure the toxicity of particular chemicals in, for example, medical, pharmacological, or agricultural studies (McKee and Wolf, 1963; Sprague, 1969). During the 1970s, the EPA began to use bioassay results for particular chemicals to establish the water-quality criteria that are published in the EPA green, blue, red, and gold books (e.g., United States Environmental Protection Agency, 1973, 1986). Since then, the application of bioassays has increased to include testing the toxicity of novel chemical compounds, licensing manufactured chemicals already in use (Federal Insecticide Fungicide Rodenticide Act), measuring toxicity at superfund sites (Resource Conservation and Recovery Act), and, of most importance for this paper, testing the toxicity of effluents.
The toxicity of effluents can be assessed by exposing test organisms to a series of dilutions of an effluent and measuring the organisms' responses. There are two basic classes of effluent bioassays. Acute bioassays measure survival of the target organisms. Chronic bioassays measure their growth, reproduction, or behavior. The toxic concentration of an effluent is defined as the lowest concen-
tration of the effluent that causes a statistically significant effect on survival, growth, reproduction, or behavior, compared with a control.
Initial EPA-approved effluent bioassays were acute toxicity tests (Peltier, 1978). Beginning in 1985, EPA introduced guidelines for chronic toxicity tests. Chronic toxicity tests are important tools because they enable the detection of toxic effects that although sublethal may have important consequences for individuals exposed to the toxins. For example, a toxin could destroy an adult organism's ability to reproduce without killing the organism itself. Acute toxicity tests would fail to detect such effects. As the worst cases of environmental pollution have been detected and addressed, chronic tests have become increasingly important. The EPA has approved chronic toxicity tests that measure growth rates of algae populations, growth rates of larval fishes, and reproduction of small freshwater crustacean zooplankton (Homing and Weber, 1985). Because effluent bioassays have become the primary tool for ensuring the protection of ecosystems from toxic effluents, it is important to evaluate whether they achieve this goal.
Regulatory Appeal of Effluent Bioassays
From an administrative viewpoint, regulations based on end-of-the-pipe measurements hold distinct advantages. These measurements (based on samples collected from pipes as they leave corporate or government facilities) require no information on the composition of the effluent or on when various constituents are added to the effluent. In addition, it is inherently easier to study the toxicity of effluents by working with the effluents themselves rather than with water from the receiving ecosystems after the addition of the effluents in question and any other inputs.
Acute bioassays have particular appeal because they are discrete, well-designed tests with well-established statistical procedures for characterizing dose-response effects (Bliss, 1934a,b; Finney, 1952a). If technicians follow required protocols carefully, they can perform acute tests accurately with minimal training. Chronic tests are more complicated. Because the performance of the test individuals in chronic tests is particularly sensitive to subtleties of culture conditions, food sources, and interactive effects of foods and toxins, technicians require more training and experience to perform chronic toxicity tests than acute toxicity tests. As a result, chronic tests are about three times as costly.
The National Pollution Discharge Elimination System (NPDES) administers the effluent bioassay program in individual states. The NPDES effluent toxicity standards are based on established state or federal water-quality criteria for compounds in effluents. Current recommended test species for freshwater effluents include the alga Selenastrum capricornutum , the crustacean zooplankton Daphnia magna and Ceriodaphnia dubia, and the fathead minnow, Pimephales promelas. Although other taxa can be used in effluent bioassays, most states
require strict adherence to EPA protocols and quality assurance and quality control procedures.
Are the Results of Effluent Bioassays Misused?
Bioassays are effective tools for testing the toxicity of particular chemicals dissolved in water. They can also aid efforts to identify specific toxic compounds (Toxicity Identification Evaluations) or remove toxins from effluents (Toxicity Reduction Evaluations) (Mount and Anderson-Carnahan, 1988, 1989a,b). However, two important problems limit the use of bioassays as tools for protecting ecosystems from toxic effluents. First, the approved protocols present technical problems in application. Second and most important, effluent bioassay results do not enable the prediction of the effluent's impacts on other organisms or on ecosystem processes. In other words, acceptable toxicity in a bioassay is no guarantee that an effluent will not adversely affect the receiving habitat.
The results of effluent bioassays are sensitive to multiple variables unrelated to the effluents themselves, including genetic variation in test organisms, chemical composition of the water source used to dilute the effluent, and foods used in experiments. Therefore, bioassay results may not be reproducible if these variables are not held constant. In such cases, EPA may not be able to confirm the results of tests performed by effluent generators or other laboratories. However, the EPA must be able to reproduce test results submitted by independent laboratories if they are to carry out their enforcement responsibilities. Consequently, EPA protocols specify the use of "artificial waters" (solutions that combine distilled water with various salts in an effort to mimic a pristine, natural, standardized water source), particular food sources (e.g., commercially available fish food), and even particular clones (genotypes) of the test species (Baird et al., 1989; Finney, 1952a). The detailed specification of testing protocols has forced many entities subject to EPA regulations to hire independent testing laboratories to perform their tests. The resulting costs have led to widespread dissatisfaction with effluent bioassay requirements.
Extrapolation from Test Results to Ecosystem Impacts
The second and more important shortcoming of effluent bioassays is the plethora of assumptions required to extrapolate test results to effects on receiving ecosystems. Neither a theoretical basis nor comprehensive test data validate the assumption that effluents that pass bioassays will not adversely affect receiving ecosystems. Such effluents may still harm ecosystems for either of two reasons: Test organisms may be less sensitive than other species to toxins in effluents, or
the operational definition of acute or chronic toxicity may not account for certain adverse effects.
The species now used in EPA-approved bioassays were originally chosen based on the ease of culturing them and the availability of data on their sensitivities to various compounds. The same taxa were used to develop federal and state water-quality criteria. Even though these organisms were generally the most sensitive of the species evaluated, no data support the assumption that they are more sensitive than all other species in receiving ecosystems. In fact, most studies based on laboratory cultures use "weed" species—species that are easy to work with because they are relatively insensitive to changes in physical and chemical conditions. Consequently, it is likely that species evaluated as candidates for EPA-approved protocols had below-average sensitivity to changes in environmental conditions. An added complication exists when test protocols specify particular clones. Using clones increases the reproducibility of results but sacrifices information on the genetic variation within populations. Because clones can vary in sensitivity to toxins, it is unlikely that the clones used in bioassays are the most sensitive. With time, new, more-sensitive test species have been approved, but the selection process is generally driven by a combination of convenience considerations rather than a desire to identify the most sensitive species.
To compensate for the possibility that sensitive species are not protected, effluent bioassay protocols use "application factors" to calculate acceptable effluent concentrations for compounds. Application factors are essentially safety factors that reduce permitted effluent limits below those that show toxicity in bioassays (e.g., Peltier and Weber, 1985, p. 79). For example, if the LC50 (the concentration that kills 50 percent of test organisms) of an effluent is 1 part per million, a permit may require that the effluent concentration in the receiving ecosystem remain below 0.01 part per million. However, there is little biological or theoretical basis for choosing application factors. They may be either excessively or insufficiently protective (Forbes and Forbes, 1993).
Even if bioassays used the most-sensitive species as test organisms, several features of effluent bioassays would nevertheless complicate the use of test results as predictors of community- and ecosystem-level consequences of effluent releases. Most effluent bioassays measure changes in the survival, growth rate, reproduction, or behavior of individuals. However, these measures are insufficient to predict population dynamics in a taxon such as Ceriodaphnia because they do not measure density-dependent feedbacks. For example, feedback mechanisms may modify reproductive behavior at high population densities and low food concentrations, or high infant mortality may be offset by modifications in numbers of eggs produced or in the sizes of and nutrients present in individual eggs. Such effects are not incorporated into existing bioassay protocols.
Although it is difficult to make quantitative predictions of changes in population dynamics on the basis of toxicity to individuals, it is even more difficult to
extrapolate further to community- and ecosystem-level consequences of toxicity. Suter et al. (1985, p. 400) describe the series of extrapolations involved in trying to predict community- and ecosystem-level consequences based on measures of an effluent's toxicity to individuals.
The LC50 must be extrapolated from the test species to the species of interest, to life-cycle toxicity, to long-term toxicity in the field, to changes in population size due to direct toxic effects and, finally, to the combined direct and indirect toxic effects. Similarly, the emission rate must be converted into an effective environmental concentration in an imperfectly known hydrologic, chemical, physical, and biological system.
Ecology has not reached the point where such projections can be made with confidence. As a result, we are generally not able to make comprehensive quantitative predictions about the community- and ecosystem-level consequences of changes in, for example, the feeding rate of one member of the biotic community (Golley, 1994). In fact, many would argue that we are not even able to make confident extrapolations from bioassay data to consequences for field populations of the test organisms themselves.
A 1981 report from the National Research Council (NRC), Testing for Effects of Chemicals on Ecosystems, suggested that more-effective tests might be possible if they incorporated
. . . a significant number of species representing the degree of diversity found in the ecosystem, detailed observations on physiological and behavioral responses for individual species, [and] a time period similar to the duration of expected chemical exposure in the ecosystem. (p. 7)
However, such a "multispecies microcosm" approach has its own problems, both because the test conditions are oversimplified relative to real ecosystems and because the test conditions are more complex than those of single-species toxicity tests. Because microcosms are, of necessity, simplifications of actual ecosystems, they do not allow for all of the potential pathways of toxic effects that could occur in actual ecosystems. For example, few microcosm designs are large enough to include the largest organisms that occur in the natural ecosystems that the microcosms are intended to represent. Therefore, such tests can neither detect effects on those missing organisms, nor can they detect indirect effects that involve those missing organisms. However, because the tests involve more variables than do single-species bioassays, the mechanisms of any observed effects are more difficult to determine in multispecies microcosms than in single-species toxicity tests.
All of the above criticisms of effluent bioassays were identified by the authors of that same NRC review, which concluded that
Current laboratory tests examine only the responses of individuals, which are then averaged to give a mean response for the test species. With given constraints of limited finances and number of personnel, it is not possible to identify the most sensitive species or group of species. . . . The data are too limited in scope for extrapolations to be made from them to responses of other (even closely related) species. (pp. xi-xii)
After more than 15 years, we still do not know whether effluent bioassays sufficiently protect species in the field from direct toxic effects, and we do not have well-established methods for extrapolating from single-species toxicity measurements to community- and ecosystem-level effects of effluents. Although this is perhaps not surprising given the complexity of ecosystems and the number of variables involved, it must be recognized that this limitation in our understanding severely limits our ability to reliably extrapolate from the results of single-species bioassays to effects on receiving ecosystems.
It is important to define the proper role of single-species bioassays. Single-species bioassays are suitable for developing water-quality criteria for particular chemicals, based on the assumption that these criteria protect individuals and that no synergistic effects occur with other chemicals in the environment. They are also useful as an initial screen to detect effluent toxicity. However, bioassays alone cannot ensure that effluents will not harm the ecosystems into which they are released.
Baird, D. J., I. Barber, M. Bradley, P. Calow, and R. Soares. 1989. The Daphnia bioassay: A critique. Hydrobiologia 188/189:403-406.
Bliss, C. I. 1934a. The method of probits. Science 79:38-39.
Bliss, C. I. 1934b. The method of probits—A correction. Science 79:409-410.
Finney, D. J. 1952a. Statistical Method in Biological Assay. New York: Hafner Publishing.
Finney, D. J. 1952b. Probit Analysis, 2nd ed. Cambridge, England: Cambridge University Press.
Forbes, T. L., and V. E. Forbes. 1993. A critique of the use of distribution-based extrapolation models in ecotoxicology. Functional Ecology 7:249-254.
Gaddum, J. H. 1933. Reports on Biological Standards. III. Methods of Biological Assay Depending on Quantal Response. Special Report Series of the Medical Research Council, No. 183. London: Her Majesty's Stationery Office.
Golley, F. 1994. A History of the Ecosystem Concept in Ecology. New Haven, Conn.: Yale University Press.
Horning, W. B., and C. I. Weber, eds. 1985. Short-Term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters to Freshwater Organisms. EPA/600/4-85/014. Environmental Monitoring and Support Laboratory. Cincinnati, Ohio: U.S. Environmental Protection Agency.
McKee, J. E., and H. W. Wolf. 1963. Water Quality Criteria, 2nd ed. Pub. No. 3A. Sacramento: California State Water Quality Control Board.
Mount, D. I., and L. Anderson-Carnahan. 1988. Methods for Aquatic Toxicity Identification Evaluations. Phase 1. Toxicity Characterization Procedures. EPA-600/3-88/034. Technical Report 02-88. Duluth, Minn.: National Effluent Toxicity Assessment Center.
Mount, D. I., and L. Anderson-Carnahan. 1989a. Methods for Aquatic Toxicity Identification Evaluations. Phase 2. Toxicity Identification Procedures. EPA-600/3-88/035. Technical Report 02-88. Duluth, Minn.: National Effluent Toxicity Assessment Center.
Mount, D. I., and L. Anderson-Carnahan. 1989b. Methods for Aquatic Toxicity Identification Evaluations. Phase 3. Toxicity Confirmation Procedures. EPA-600/3-88/036. Technical Report 04-88. Duluth, Minn.: National Effluent Toxicity Assessment Center.
National Research Council. 1981. Testing for Effects of Chemicals on Ecosystems. Washington, D.C.: National Academy Press.
Peltier, W. 1978. Methods for Measuring the Acute Toxicity of Effluents to Aquatic Organisms, 2nd ed. EPA/600/4-78/012. Environmental Monitoring and Support Laboratory. Cincinnati, Ohio: United States Environmental Protection Agency.
Peltier, W. H., and C. I. Weber, eds. 1985. Methods for Measuring the Acute Toxicity of Effluents to Freshwater and Marine Organisms, 3rd ed. EPA/600/4-85/013. Environmental Monitoring and Support Laboratory. Cincinnati, Ohio: U.S. Environmental Protection Agency.
Sprague, J. B. 1969. Measurement of pollutant toxicity to fish. I. Bioassay methods for acute toxicity. Water Research 3:793-821.
Suter II, G. W., L. W. Barnthouse, J. E. Breck, R. H. Gardner, and R. V. O'Neill. 1985. Extrapolating from the laboratory to the field: How uncertain are you? Pp. 400-413 in Aquatic Toxicology and Hazard Assessment: Seventh Symposium, R. D. Cardwell, R. Purdy, and R. C. Bahner, eds. ASTM STP 854. Philadelphia, Pa.: American Society for Testing and Materials.
United States Environmental Protection Agency. 1973. Water Quality Criteria 1972. EPA-R3-003. Washington, D.C.: U.S. Environmental Protection Agency.
United States Environmental Protection Agency, Office of Water Regulations and Standards. 1986. Quality Criteria for Water 1986. EPA 440/5-86-001. Washington, D.C.: U.S. Environmental Protection Agency.