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Reprinted from regulatory toxicology and pharmacology 18, 313-356 (1993) Scientific Principles for Evaluating the Potential for Adverse Effects from Chlorinated Organic Chemicals in the Environment Robert F. Willes, Earle R. Nestmann, Patricia A. Miller-, Joan C. Orr, and Ian C. Munro i Copyright 1993 by Academic Press, Inc. All rights of reproduction in any form reserved. R&S 143943 Reprinted from regulatory toxicology and pharmacology 18, 313-356 (1993) Scientific Principles for Evaluating the Potential for Adverse Effects from Chlorinated Organic Chemicals in the Environment Robert F. Willes, EarlE R. Nestmann, Patricia A. Miller, Joan C. Orr, and Ian C. Munro CanTox Inc.. Consultants in Toxicology. Health, and Environmental Sciences, 308-2233 Argentia Road, Mississauga, Ontario, Canada L5N 2X7 Received May 4,1993 PREFACE This report describes a number of important scientific principles that govern the evaluation of the potential for chlorinated organic chemicals to cause adverse effects on the environment and human health. The report is not intended to be a compre hensive evaluation ofavailable data pertaining to chlorinated organic chemicals. Rather, certain of the published data reporting adverse effects are examined in light of the scientific principles elaborated to underline the point that, as in any branch ofscience, rigurous criteria must be applied in assessing the significance of the scientific data on chlorinated organic chemicals. The report was reviewed by a panel of scientists chosen for their expertise in human health and environmental toxicology. John Giesy, Ph.D., Michigan State University, and Martin van den Berg, Ph.D., University of Utrecht, provided critical review in the area of environmental toxicology. Dietrich Henschler, M.D., Institut fur Pharmakologie und Toxikologie der Universitat Wurzburg, and Renate D. Kimbrough, M.D., Institute for Evaluating Health Risks, provided critical review of human health aspects of the report. ABSTRACT The term chlorinated chemicals is used to describe diverse groups of chemicals of varying chemical structure, including those used in water disinfection, as well as nu merous aliphatic, aromatic, and polycyclic chlorinated substances. This report elabo; ates a number of scientific principles that govern the evaluation of the potential for chlorinated organic chemicals to cause adverse effects on the environment and to human health. The purpose ofthe report is to demonstrate the importance ofapplying these scientific principles in the evaluation of potential adverse effects of chlorinated 313 0273-2300/93-S5.00 Copyright 1993 by Academic Press, Inc. All rights of reproduction in any form reserved R&S 143944 314 WILLES ET AL. organic chemicals. The four major principles upon which such a scientific analysis must be based are: (1) the fate and biological activity of a compound are determined by the chemical properties of the compound; (2) compounds do not show adverse effects below certain threshold concentrations, and the magnitude of response is related to dose; (3) inherent metabolic processes allow organisms to accommodate low doses of chlo rinated organic chemicals; (4) observations associated with the presence of a certain compound must be biolog ically plausible effects, based on the specificity of the compound's activity in experi mental systems. With respect to the first of these principles, there is abundant scientific evidence that the physical and chemical properties of chlorinated organic chemicals govern their bioaccumulative potential, toxicological properties, and thus their potential be havior and effects in the environment. Chemicals that have low solubility in water are highly lipophilic and have low vapor pressure, tend to accumulate in biological systems, and degrade slowly in the environment. Chlorinated organic chemicals that possess these characteristics include those having a carbon ring structure and multiple chlorine substitution. Other chlorinated organic chemicals with lesser degrees of chlorine sub stitution, such as 2,4-dichlorophenoxyacetic add (2,4-D), trichlorophenol, chloroform, and dichloroethane, do not share the physical and chemical properties of the high molecular weight, cyclic, polychlorinated compounds and, as such, do not have the same potential to bioaccumulate. These differences among chlorinated organic chem icals with respect to their physical and chemical properties and behavior in the envi ronment preclude the generalization that all organic chemicals containing chlorine behave similarly in the environment and act as persistent, bioaccumulative chemicals. The reactivity ofchlorinated organic chemicals and hence their potential to produce biological effects depends on their specific molecular features. The substitution of chlorine into an organic molecule may increase or may reduce its biological activity. One determinant ofbiological activity is whether or not chlorinated and nonchlorinated organic chemicals can form electrophilic moieties which can react with critical targets in biological systems. Other chlorinated and nonchlorinated organic chemicals may interact with cellular receptor systems, resulting in changes in homeostatic functions through a variety ofsecondary effects. These are not properties shared by all chlorinated organic substances because not all are capable of being metabolically activated to reactive substances or of interacting with cellular receptor systems critical to the maintenance of normal functions. A second important principle to be considered in assessing the significance of chlo rinated organic chemicals in the environment is that of the dose-response relationship. This relationship dictates that the magnitude of the response of biological systems to chemicals depends on their concentrations in organisms, which in turn depend on the magnitude and duration of exposure. Adverse effects are observed when the rate of damage is greater than the rate of repair of cellular components. The magnitude of exposure is proportional to concentrations in the environment. The chlorinated compounds with greatest potential for bioaccumulation, such as some polychlori nated biphenyls (PCBs), some chlorinated dioxins and furans, l,l,l-trichIoro-2-2bis(parachlorophenyl)ethane (DDT), dieldrin, and other structurally similar com- CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 315 R&S 143946 pounds, act through threshold-type mechanisms. This means they would not be ex- v pected to elicit a toxic response at environmental concentrations which result in ex posures below a threshold. National regulatory authorities have acknowledged the existence of environmental concentrations below which there would be no cause for ' concern, and this concept provides the basis for existing regulatory standards. The existence of thresholds for effects has been validated for the chlorinated organic chem icals, and adverse effects would occur only when these threshold exposures are exceeded: For example, in the Great Lakes, where bioaccumulative chlorinated organic chemicals as a group have been associated with reproductive effects and population declines of certain species of piscivorous birds, decreasing concentrations of representatives of " these groups of chemicals have been associated with recovery in most species, partic ularly herring gulls and double-crested cormorants. Currently, reported adverse effects on piscivorous bird populations in the Great Lakes Basin are limited to localized areas of greatest environmental concentrations, such as Green Bay in Lake Michigan and Saginaw Bay in Lake Huron. Further evidence of a threshold concentration for chlorinated organic chemicals below which adverse effects on the environment do not occur comes from the fact that to date approximately 1500 of these substances are known to be produced by natural processes. Chlorinated organics are released into the environment by a number oforganisms, including bacteria, fungi, algae, ferns, and higher plants, by decomposition of seaweed, and by combustion sources such as forest and grass fires and volcanic activity. Naturally produced chlorinated organic chemicals indude chlorinated alkanes, alkenes, phenols, polychlorinated biphenyls, and chlorinated dioxins and furans, en compassing a broad range of substances representing the same classes of chemicals as produced by human activities. Natural processes are known to result in significant background concentrations of some of these chemicals; however, the available infor mation is inadequate to estimate total environmental releases for most chlorinated organic chemicals. Many of the chlorinated and other organic chemicals are metab olized through the action ofthe mixed function oxygenase (MFO) system, a nonspecific enzyme system present in diverse species from different phylogenetic classes. The ubiquity of chlorinated organic chemicals in the environment and the presence in diverse groups of mammalian and nonmammalian species of a common enzyme system capable of detoxifying these compounds provide evidence that organisms have evolved in the presence of such chemicals. The existence of the naturally occurring chlorinated organics supports the conclusion that exposures to sufficiently low con centrations of similar anthropogenic chemicals would not result in the occurrence of adverse effects on ecosystems. The final principle which governs the assessment of the potential adverse effects of chlorinated chemicals in the environment is that of the specificity and biological plau sibility of associating reported adverse effects with possible causal agents. Since chlo rinated organic chemicals have diverse chemical structures, and therefore vastly dif fering bioaccumulative and toxic potentials, it is critical to use sound scientificjudgment to establish associations between adverse environmental effects and causative agent(s). The establishment of cause-effect relationships is essential to invoke the development of effective mitigation procedures for improvement and protection of the environment. The application of the principle of specificity of association is demonstrated in the following three examples. Mass mortalities of marine mammals have been alleged to os associated with exposures to chlorinated organic chemicals; however, the application 316 WILLES ET AL of criteria for the establishment of causal associations leads to the conclusion that natural disease processes are often to blame. The weight of available evidence appears to support an association between adverse effects on reproduction of fish-eating birds in the Great Lakes region and historical exposures to PCBs and DDE (the principal metabolite of DDT), although other less studied factors including mercury and other chemicals, habitat destruction, and natural diseases also are involved. In the case of pulp and paper effluents, complex mixtures of chlorinated and nonchlorinated chem icals have been implicated in cause-effect relationships between rates of chemical release and adverse effects on the aquatic environment. The results of more rigorous studies in this area indicate that the effects of pulp mill effluents cannot be attributed**solely to the release of process-related chlorinated chemicals into the environment. Similar effects on fish are observed near mills that use or do not use chlorine, and significant reductions in the use of chlorine at other mills have not resulted in corre sponding reductions of environmental effects. Similarly, the application of criteria for judging causal associations ofhuman health effects leads to the conclusion that, except for well-documented instances of toxicity related to specific incidences of accidental releases and certain defined occupational exposures, there have been no definitive , cause-effect relationships established between exposure to environmental concentra tions of chlorinated organic chemicals and adverse health consequences in the human population. It is clear, from the available scientific information, that chlorinated organic chem icals possessing the physical and chemical properties that lead to bioaccumulation have the greatest potential to produce adverse effects on the environment. It is also' clear, because environmental concentrations ofthese compounds have decreased during the past 20 years, that animal populations previously thought to be affected by such chemicals are recovering, consistent with the principle that sufficiently small environ mental concentrations of these compounds will not adversely affect the environment. This principle also is supported by the fact that naturally occurring sources are re sponsible for significant releases of chlorinated organic chemicals, including some having a high degree ofchlorine substitution. While the significance of natural sources of chlorinated organic chemicals to environmental health is not well understood, the available evidence indicates that the earth's species have evolved with metabolic ca pability to accommodate small concentrations of these substances. INTRODUCTION The purpose of this report is to provide a discussion ofscientific principles governing the evaluation of the potential for chlorinated organic chemicals to produce adverse ects on the environment and human health. Their potential to produce adverse effects has been the focus of a continuing debate among environmental groups, gov ernments, and industry. The term "chlorinated organic chemical" describes a broad range of chemicals of diverse chemical structure, ranging from simple aliphatic sub stances to complex chlorinated aromatic and polycyclic substances. Also included are many substances of medicinal and health importance such as pharmaceuticals, pes ticides, and disinfectants. Available information demonstrates that physical, chemical, and biological prop erties of chlorinated organic chemicals determine their behavior in the environment R&S 143947 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 317 R&S 143948 and their potential effects on the ecosystem. Information on specific physical/chemical properties of chemicals has proven reliable in predicting their behavior or fate in the environment (Mackay et al., 1992; Klopman et al, 1993). In addition, these properties ofchemicals influence the metabolism of chemicals by organisms (including humans) /.and this, along with the level of exposure (dose), determine their potential to produce toxicity. However, physical/chemical properties have proven to be less reliable pre dictors of adverse effects than of environmental fate (Henschler, 1990). The association of bioaccumulation with specific types of chemical structures is well documented. For example, it has been established that increased concentra tions in tissues of certain polychlorinated biphenyls (PCBs) or l,l,l-trichloro-2,2bis(parachlorophenyl)ethane (DDT) have been linked to adverse responses (Tillitt et*~~ al., 1992). The occurrence of elevated concentrations of such chemicals'in tissues is principally due to their potential to biomagnify in aquatic food chains (Environment Canada, 1991a). The bioaccumulation of these substances results from their molecular stability coupled with their greater solubility in fat compared to water (McEwen and Stephenson, 1979). These properties of stability and lipid solubility are not general properties of all chlorinated organic substances. Most chlorinated chemicals are de graded rapidly by inherent processes common to numerous organisms in the envi ronment. Furthermore, the view that chlpjinated organic chemicals are singularly a by-product of human activities is not borne out by an examination of the scientific literature which demonstrates conclusively that a broad array ofchlorinated chemicals is produced naturally in the environment (Gribble, 1992). Importantly, enzymatic processes capable of metabolizing these substances are present in organisms in the environment (Environment Canada, 1991c). An additional factor that warrants consideration in any examination of the potential adverse effects of chlorinated substances concerns their concentrations in various en vironmental media. Adverse biological effects are dose-related, and the contention that the presence of a chlorinated chemical in the environment is an adequate basis to conclude that adverse consequences will occur is not supported by the application of the scientific principles of toxicology. In this report, available data concerning re ported adverse effects of chlorinated organic chemicals are examined in light of the concept of dose-response principles. The report also makes note of the fact that, except for some continuing "hot spots," the concentrations of chlorinated organic chemicals in environmental media and organisms have declined considerably since the 1970s, thus reducing the potential for future risk. An additional important consideration in the evaluation of the potential for chlo-' rinated organic chemicals to cause adverse effects is the adequacy of data supporting . the relationships between the presence of these substances in the environment and reported adverse consequences. As with any chemical, the toxic potential ofchlorinated organic chemicals to organisms depends on their chemical structure, pharmacokinetics and metabolism, and interactions with cellular systems. Thus, the potential for adverse effects ofsubstances within the varied classes ofchlorinate^ chemicals must beevaluated taking these factors into consideration. The strength of associations between the con centrations of chlorinated organic chemicals and reported adverse effects warrants careful examination in light of the quality and quantity of available data. Evidence for adverse effects is strengthened by the availability of laboratory data demonstrating dose-response relationships for individual chlorinated substances, while field data 318 WILLES ET AL. demonstrating only statistical associations must be critically examined before scientific judgments of causal associations are reached. Keeping in mind these important concepts of physical, chemical and biological properties, dose-response relationships, natural occurrence and applicable metabolic processes, and specificity of association of reported adverse effects, what follows is a critical examination of how these principles relate to the evaluation of the potential for chlorinated chemicals to cause adverse effects in the environment, including human health. PHYSICAL AND CHEMICAL PROPERTIES OF CHLORINATED ORGANIC CHEMICALS Chlorinated chemicals comprise structurally diverse groups of chemicals with a broad spectrum of physical and chemical properties. Chlorinated compounds are rep resented in all the major chemical classifications including inorganic salts and acids, monocyclic, polycyclic, heterocyclic, simple and branched chain aromatics, and aliphatics. The physical and chemical properties of chemicals, including chlorinated organic chemicals, govern their behavior in the environment and profoundly affect their activity on biological systems. The reactivity oforganic molecules depends on the structural features oftheir carbon backbone (e.g., straight chains,, cyclic, polycyclic) and the nature, number, and position of functional groups attached to this backbone. In the basic or parent unsubstituted organic chemical, hydrogen atoms are attached to the carbon atoms. The hydrogen atoms can be replaced by various functional groups (e.g., oxy-, hydroxyl-, carboxy-, alkyl-, allyl-, alkoxy, nitro-, amino-, sulfoxy-, sulfhydryl-, chloro-, bromo-, iodo-, fluorogroups plus many others). The introduction of various functional groups changes the physical and chemical characteristics of the parent organic molecule by changing its molecular weight, size, shape, and electrical and reactive properties (Politzer and Laur ence, 1984). These changes in molecular properties by the addition ofvarious functional grbups to organic chemicals are determinants of the interactions of these chemicals with biological systems. Overall, the properties conferred by the addition of different functional groups to organic molecules are part of a continuum, with the intensity of change in properties varying with type, number, and positioning of the functional group(s) added to the molecule. Hence, significant changes may be observed in the physical/chemical pa rameters between substituted and nonsubstituted parent molecules (Verschueren, 1983). Molecular volume is a major factor affecting the physical/chemical parameters of organic molecules. For example, the effects on molecular volume of the addition of one atom of chlorine are approximately the same as the addition of one methyl (CH3) group (Mackay et al., 1992). Therefore, the substitution of a hydrogen atom by one atom of chlorine or by one methyl group would result in approximately the same change in vapor pressure, Henry's law constant, water solubility, and octanol/ water partition coefficient of an organic molecule although different substituents would impart different degrees ofreactivity (Mackay et al., 1992). The scientific understanding ofhow the addition ofvarious functional groups alters the physical/chemical properties of chemicals is now considered to be sufficient to allow judgments to be made regarding likely behavior in the environment of specifically designed molecules, thus providing R&S 143949 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 319 a process for identifying and then avoiding releases of those with undesirable char acteristics that would result in unwanted environmental distribution and deposition, with possible adverse effects (Mackay et ai, 1992). Predictions ofBehavior and Fate of Chlorinated Organic Chemicals in the Environment Predictions of fate and bioaccumulation of chemicals in the environment based on their physical/chemical properties have been quite successful. Water solubility and w. octanol-water partition coefficient (log i^w) are measures of chemical solubility in water and fat, respectively. Vapor pressure is a measure of the extent to which a chemical evaporates from soil or water into the air. For example, the physical and chemical properties of PCBs indicate that these compounds are not very soluble in water (Mackay et ai, 1992). Therefore, in an aquatic environment, PCBs partition to the organic phase (e.g., sediments, biota) to a much greater extent than to the aqueous or water phase. The greater their degree of chlorination, the greater the tendency to 1 accumulate and persist in soils and sediment that have greater organic carbon con centrations, and the greater their capacity to bioaccumulate in aquatic and terrestrial wildlife and demonstrate substantial food chain biomagnification (Environment Can ada, 1991a). The physical/chemical properties of the majority of the chlorinated organics in use by society today are substantially different from those ofthe higher chlorinated PCBs. For example, the water solubility of chlorinated alkanes and alkenes range from 1100 to 9300 mg/liter (Verschueren, 1983). Correspondingly, log Kow values for these chem icals are low and range from 1.45 to 2.73 (Hansch and Leo, 1979; Baneijee et al., 1980; Valvani et.ai, 1981). Because of their high water solubility and low solubility in lipid, there is negligible bioaccumulation of these chemicals in the environment. Furthermore, chlorinated organic chemicals possess a range of vapor pressures. For example, 1,2-dichloroethane has a much greater vapor pressure compared to the vapor pressures of chlorinated dioxins (e.g., 8132.7 Pa at 25C compared to 0.073 to 1.10 X 10-, Pa) (Moore et ai, 1991). Thus, most of the 1,2-dichloroethane released into the environment volatilizes to the atmosphere, where it is degraded by interactions with hydroxyl radicals produced by natural sunlight (Clement Associates, 1989). As a result, this chemical would be expected to volatilize rapidly from surface waters (Clement Associates, 1989; Moore et ai, 1991). This behavior is predictable based on environmental fate modeling derived from physical/chemical properties, which indicate that the majority (about 97%) of 1,2-dichloroethane would partition to air, with only very small proportions being distributed in soil, sediments, and fish (Mackay etai, 1992). Differences in physical/chemical characteristics, which are the result of the degree and pattern of chlorination, among other parameters, influence the degradation of chlorinated organic chemicals in the environment. For example, PCBs without sub stitution in the ortho positions of the carbon rings, and with chlorine atoms in both para positions and at least one meta position, form "coplanar" structures where the carbon rings lie molecularly in flat planes. Some coplanar PCBs tend to bioaccumulate to a greater extent than noncoplanar PCBs in organisms higher in the phylogenetic R&S 143950 320 WILLES ET AL. tree (Tillitt et al, 1992). The presence of a significant degree of chlorination is not, of itself, sufficient to confer bioaccuraulative potential on a compound. This is supported^' by the observation that moderately chlorinated aromatics, such as 2,4,5-trichloro- phenol, with a log w of 4.19 (Schellenberg et al, 1984), are not particularly bioad- cumulative. Chlorophenols are rapidly absorbed by aquatic invertebrates and fish; ' however, they are also rapidly metabolized and excreted and therefore generally do not bioconcentrate to the extent expected based on the log K^. Given the difference in physical/chemical characteristics, the behavior and fate, characteristics associated with bioaccumulative compounds are not applicable to other chlorinated organic.- compounds with substantially different physical/chemical characteristics. _' In applying the scientific principles that dictate the relationship between physical/-" chemical properties and biological effects and environmental fate characteristics, it'fsT possible to make judgments as to the likely behavior and potential for adverse effects of various groups of chlorinated chemicals. For example, compounds-having low water .: solubility, low vapor pressure, and high lipid solubility are generally very stable corrf- pounds under most environmental conditions. Upon introduction to the aquatic^n- vironment, PCBs and chlorinated dioxins readily partition into the sediment (Pavlou and Dexter, 1979; Phillips, 1986). Due to their hydrophobicity (low water solubility), these chemicals adsorb to dissolved organic material (DOC) and organic carbon of suspended particulates in the water column, dissolved organic material in the pore water, organic carbon of the bed sediments, and fatty tissues of aquatic biota. Lesser chlorinated PCBs and chlorinated dioxins/furans are more soluble in water (Haque et al, 1974; Haque and Schmedding, 1976) and are expected be transported further than those with greater chlorine substitution (Phillips, 1986). The greater vapor pressure and lesser binding coefficient to particulate materials make the lesser chlorinated dioxins and furans more susceptible to degradation in the atmosphere by ultraviolet light (Eitzer and Hites, 1989; Hites, 1990). Since some PCBs, chlorinated dioxins and furans, ; and other highly chlorinated aromatics are relatively stable both thermally and chem ically, environmental concentrations would be expected to change slowly with time (Chou and Griffin, 1986). This is further exacerbated by the continued release of PCBs from municipal landfills and various waste storage sites. In addition, there are releases ofchlorinated dioxins and furans from numerous uncontrolled human activities (e.g., emissions from internal combustion engines used in automobiles and other applica tions, combustion oforganic fuels for heating and other purposes) and an undetermined extent of release from naturally occurring sources (Bumb et al, 1980; Nestrick and Lamparski, 1982; Clement et al, 1985; Travis and Hattemer-Frey, 1989; Tashiro et al, 1990). Thus, these chemicals have the potential to be an environmental problem in the future, despite their restricted or banned use in certain specific instances. Many chlorinated organic chemicals can be degraded by various chemical reactions, including hydrolysis, photochemical oxidation, as well as biomediated oxidative and reductive reactions such as degradation by bacteria and fungi. A chemical may be resistant to one or more forms of degradation, but susceptible to other forms. For example, hydrolysis of chloroform has been reported to be a very slow process (Mabey and Mill, 1978; Jeffers et al, 1989), but volatilization ofchloroform from water occurs rapidly with a half-life of I to 31 days (Zoeteman et al, 1980; SRC, 1989) and it is degraded in the environment by microorganisms (Tabak et al, 1981; SRC, 1989). Conversely, chlorinated phenols have relatively low vapor pressures at environmental temperatures in the aquatic environment resulting in minimal volatilization (Pignatello R&S 143951 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 321 et al, 1983; Sugiura et al, 1984; HWC, 1988). These chemicals are removed from; the aquatic environment primarily through photolysis and through microbial trans- ' v formation (Agriculture Canada, 1987). These examples illustrate the importance of degradation pathways in the natural environment Even among chemicals with aromatic structures, the presence ofone or more chlorine atoms confers neither persistence in the environment nor a significant potential for bioaccumulation. For example, 2,4-dichlorophenoxyacetic acid, or 2,4-D, a chlorophenoxy herbicide, has an aromatic structure substituted with two chlorine atoms and. does not persist under most soil conditions because it is readily metabolized in plants, .; soil, and in the aquatic environment by bacteria and fungi that adapt to 2,4-D, using _ it as a food source (McEwen and Stephenson, 1979; Stevens and Sumner, 1991$L<f Consequently 2,4-D is rapidly degraded in the environment, with a half-life measured*" in days to weeks (McEwen and Stephenson, 1979). Although this chemical is aromatic and substituted with chlorine atoms, it is not a potent mammalian or avian toxicant (Munro et al, 1992), and it has a very low potential for bioaccumulation. Thus, although PCBs, chlorinated dioxins and furans, 1,2-dichloroethane, trichlorophenol, chloroform, and 2,4-D are all chlorinated organic chemicals, differences in their phys ical/chemical properties cause them to behave differently in the environment. Clearly, L | the behavior and fate in the environment of one type of chlorinated.organic chemical, ; such as highly chlorinated bioaccumulative compounds, cannot be used to predict the behavior and fate of all chlorinated organic chemicals. The striking differences in behavior and fate in the environment among individual chlorinated organic chemicals indicate that scientifically valid evaluations of the potential adverse effects from en vironmental concentrations ofthese chemicals must be based on specific data related to chemical and physical properties of individual compounds. R&S 143952 Predictions ofAdverse Effects ofChlorinated Organic Chemicals The predictions of the effects of adding functional groups>on the biological activity of molecules have proven more difficult than predicting their influence on environ mental behavior. The addition of a specific functional-group with electronegativity and electron-withdrawing characteristics (e.g., chlorine and other halogens, N02, SO3) to an organic molecule may increase or reduce the chemical reactivity ofthe molecule, depending on the position and number of. substitutions, and the atoms replaced by the functional groups. This means that not all chlorinated chemicals are more chem ically or biologically reactive than their nonchlorinated parent chemicals, and that the addition to the molecule of other atoms or moieties (e.g., N02, S03, NH2, etc.) can result in the production of chemicals with greater reactivity than those containing chlorine. Other relevant factors are the presence of additional electronically active functional groups on the molecule and the chemical structure of the molecule itself (e.g., steric factors). The electron-withdrawing properties of chlorine and other halogens are considered to be the main factors governing the reactivity of organohalogens (Huheey, 1965; Temay, 1979) and in this regard, there are major differences among chlorinated organic chemicals. For example, the chemical reactivity ofchloro- and haloalkenes is dependent on the nature, the number, and the position of halogen substituents, as well as the number and the position of the double bonds (Williamson and Cvetanovic, 1968; 322 WILLES ET AL. Temay, 1979; Woo et al, 1985). In the ethylene series, introduction of the electro negative halogen atoms decreases the electron density in the double bond and exerts a stabilizing effect. Indeed, the reactivity of ethylene decreases dramatically with the increase in the degree of chlorination (Williamson and Cvetanovic, 1968). In the propene series, allyl halides (3-halopropenes) are expected to be much more susceptible to nucleophilic substitution of the halogen atom than the corresponding haloalkanes (Woo et al., 1985). Vinyl halides are poor alkylators, and enzymatic activation k required, with vinyl bromide being slightly more potent than vinyl chloride (Barbir et al, 1975; Eder et al., 1980). In contrast, allyl halides are good alkylators even without metabolic activation^theii activity depending on the halogen substitute. Iodine-substituted allyl halides are^trom alkylators, with bromine and chlorine being weaker (Eder et al, 1980). The variou: factors that may affect this behavior have been summarized by Woo et al (1985). Th< addition of greater quantities of chlorine to organic molecules tends to decrease thei reactivity and increase their stability and bioaccumulative potential in the environment Larger, more complex molecules with larger quantities of chlorine (e.g., PCBs, chlo rinated dioxins and furans) are less likely to have reactive metabolic intermediates or if they do, the quantities formed are smaller than formed from lesser chlorinatechemicals. Enzyme systems involved in the initial stages of metabolism of chemicals (i.e Phase I metabolism) generally catalyze the formation of reactive nucleophilic an electrophilic metabolites ofchlorinated organic chemicals that can interact with celluk macromolecules (e.g., DNA). For example, reactive metabolites of chlorinated alky benzenes, chlorinated nitrobenzenes, chloroform, and minor amounts of metabolite of PCBs and other chlorinated polycyclic compounds form through reactions wit the cytochrome P450-containing monooxygenase enzyme system in the liver. The; reactive metabolites can covalently bind to various cellular macromolecules, but the also react during Phase II metabolism with reduced glutathione (GSH) through tl catalytic actions ofthe glutathione ^-transferase enzyme system. The covalent bindii of the reactive metabolites to GSH prevents their reaction with vital cellular macr molecules, thereby protecting cells from the toxic effects associated with these reactio: (Sipes and Gandolfi, 1991). The balance between the formation of reactive metabolites from chlorinated orgar chemicals that undergo such reactions and their subsequent inactivation by GSH critical to the potential expression of adverse effects from such chemicals. Imbalan in the glutathione S-transferase-GSH conjugation system can substantially affect t toxic potential ofchemicals that produce toxicity via reactive metabolites. In additic large quantities ofreactive metabolites (e.g., as may be observed with extreme exposu: used in laboratory tests) can deplete the available concentrations of GSH in ce Since GSH is a cofactor for various normal cellular reactions, an array of second; toxic effects can result from its depletion (e.g., GSH is a cofactor for glutathic peroxidase, and depletion of GSH can result in increased lipid peroxidation lead: to a variety of toxic effects) (Sipes and Gandolfi, 1991). These considerations sf doubt on the validity of extrapolating information from extreme exposures (e.g., th< typical of most laboratory tests, or associated with large quantity accidental relea of chemicals into the environment) to lesser exposures more typical of those rela to usual environmental pathways where GSH depletion would not occur. R&S 143953 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 323 Chemical structure also has an important bearing on the nature of toxicity. Except for narcosis, which is a nonspecific, high-dose effect of most chlorinated and non-. _ chlorinated organic substances (Franks and Lieb, 1982; 1984), molecular configuration^ ; greatly affects the specific form of toxicity. An example of this phenomenon can be ; , , drawn from knowledge of chemical carcinogenesis. Evidence to date indicates that carcinogenic chemicals can be divided into two major classes, genotoxic carcinogens and epigenetic or nongenotoxic carcinogens (Williams and Weisburger, 1991). (i) Genotoxic carcinogens chemically interact with DNA, and adduct formation is a definitive way to identify genotoxic carcinogens. Genotoxic carcinogens are subdi- ~-J vided into those where the parent chemical is active and those that require bioactivation^ where a metabolite is the active form. (ii) Epigenetic or nongenotoxic carcinogens do not interact with genetic material, but act via an indirect biological mechanism as the basis for their carcinogenicity. , Epigenetic carcinogens function as promoters, modifiers ofhormonal activity, cytotoxic agents, immunosuppressors, peroxisome proliferators, etc. Many of the carcinogens in this class increase DNA synthesis, mitosis, and cell replication rates following extreme exposures. Some produce secondary genotoxic effects through modifications in other cellular and metabolic systems. These properties are important in assessing the carcinogenicity ofchlorinated organic chemicals (Henschler, 1985), and the distinction of the two classes is critical to risk assessment. Genotoxic carcinogens represent a threat to human health because of the interactions with genetic material and because ofthe possibility that they can act after a single exposure or in a cumulative manner with multiple exposures. Thus, more dringent criteria are required to establish acceptable or allowable exposures to genotoxic carcinogens (OSTP, 1985; Williams and Weisburger, 1991). Epigenetic carcinogens, on the other hand, show carcinogenic effects only following extreme and sustained exposures that are necessary to maintain abnormal physiological functions, hormonal imbalance, tissue injury, cell proliferation, or other nongenetic mechanisms (Williams and Weisburger, 1991). Exposures that do not produce the required alterations in physiologic, homeostatic parameters that are affected by the epigenetic carcinogen would not result in an increased incidence of cancers. Therefore, for epigenetic car cinogens it is possible to establish exposure thresholds below which cancers would not develop (Williams and Weisburger, 1991). While a great deal is known about the relationship of chemistry and mechanisms of carcinogenesis, it has proven difficult to predict the carcinogenicity of chemicals based on their specific molecular structures (Henschler, 1990). This difficulty is believed to be related to the complexity of carcinogenesis, which involves many biological processes, including uptake, metabolism, distribution, excretion, and specific inter actions of chemicals with cellular macromolecules (e.g., DNA). For example, cleavage I of GSH conjugates of chemicals by -lyase in the kidney results in the formation of reactive metabolites that can lead to toxicity and cancer (Vamvakas et al., 1993). Detailed QSAR analyses oflarge carcinogenicity databases have shown that structural elements are not related consistently to carcinogenicity (Chou and Jurs, 1979; Jurs et al., 1979; Craig and Enslein, 1981; Enslein and Craig, 1982; Rose and Jurs, 1982; Klopman, 1984; Klopman et al., 1985; Frierson et al., 1986; Ashby and Tennant, 1991). Among compounds containing "structural alerts," a large proportion were not carcinogenic, and 27 compounds not containing structural alerts did result in an in- R&S 143954 324 WILLES ET AL. -'used tumor incidence in laboratory rodents (Ashby and Tennant, 1991). Despite the difficulties in making generalizations about specific molecular characteristics and the development of cancer, empirical observations demonstrate that the presence of chlorine in a molecular structure does not necessarily render the compound carci nogenic. In fact, the opposite is true in the case ofbenzene, a potent chemical carcinogen which has been associated with leukemia in humans at sufficiently great exposures but which has several chlorinated derivatives that are not carcinogenic (Ashby and Tennant, 1991). Some chlorinated compounds might be expected to show carcinogenic activity by virtue oftheir structures, while they are actually noncarcinogens. Examples are 2-chloroacetophenone, a compound containing a reactive chlorine, and chlorc-" ethane, a potent mutagen with a great degree ofalkylating activity (Ashby and TennSSt, 1991). Some chlorinated epoxides, compounds which might be expected to be carci nogenic based on their structure, are indeed not carcinogenic. For example, ethylene oxide is carcinogenic, while its di-, tri-, and tetrachlorinated derivatives are not (Politzer and Laurence, 1984). On the other hand, propylene oxide and its chlorinated deriv atives, 1-chloropropylene oxide and 1,3-dichloropropylene oxide, all cause increased cancers in test animals (Politzer and Laurence, 1984). Among a large group of elec trophilic compounds tested in the National Toxicology Program, 17/30 (56%) of the positive carcinogens contain a halogen (i.e., chlorine, bromine, iodine, or fluorine) substituent, while 4/11 (36%) of the chemicals that were not associated with increases in tumors contain a halogen. Among compounds containing a nonreactive halogen, 26 were positive and 16 were negative in cancer bioassays, and 8 are considered equiv ocal (Ashby and Tennant, 1991). These empirical data demonstrate that the presence of chlorine in a molecule does not necessarily impart carcinogenicity, and that specific JlfljljUj testing is needed to assess the carcinogenic potential of chlorinated organic chemicals. In summary, the information now available demonstrates that physical/chemical properties and chemical structures can be used to predict the environmental behavior of chemicals with reasonable confidence. On the other hand, less confidence can be placed in the predictions of specific toxicity, although various structural alerts and critical assessment procedures (e.g., adduct formation) can provide important clues about the potential toxicity of chemicals. R&S 143955 DOSE-RESPONSE RELATIONSHIPS In evaluating the significance of environmental concentrations of chlorinated or ganics, it is of critical importance that principles governing dose-response relationships be considered (Klaassen and Eaton, 1991). The presence of chlorinated organic chem icals in the environment (regardless of their concentrations or specific chemical struc tures) does not in itselfconstitute a risk ofadverse effects, in keeping with a fundamental premise of toxicology that effects are related to dose. In addition, in the case of most if not all chemical compounds, there is a dose below which no adverse effects can be observed. Known as the threshold dose, its magnitude varies for different chemicals and varies for different organisms exposed to the same chemical. Factors that determine the threshold dose include the extent to which the chemical is absorbed by the organism, the rate and mechanism of clearance from the organism, the extent of metabolic production of toxic metabolites and/or detoxification products, and the degree to which the organism can repair any cellular damage by the a CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 325 R&S 143956 chemical. The ability of the organism to repair damage is a critical consideration when . v; judging the applicability of the extrapolation information obtained from genetically inbred strains of animals to general heterogenetic populations. In establishing a safe level of exposure to chlorinated chemicals in the environment, empirical evidence of safety based on threshold doses observed in controlled studies must be considered, in addition to mechanistic evidence which helps to explain the basis for the existence of the threshold. In this section, the application ofthe threshold concept in regulatory decision making is discussed, and several examples illustrating the operation of threshold and doseresponse principles in the natural environment are presented. For acute or lethal toxic effects, the existence of an exposure threshold below which lethality would not occui^ (i.e., zero probability of a response) has long been recognized (Klaassen and Eaton, 1991). For most chronic adverse effects, there is generally a discernible threshold below which it is not possible to measure adverse effects in exposed populations-, even though the absence of a threshold cannot be proven for certain endpoints. The fundamental basis for the assessment of potential adverse effects on the envi ronment and subsequent regulatory decision making is that the magnitude ofa response of an organism to a chemical increases as the dose increases, and this dose-response relationship is based on three key observations (Klaassen and Eaton, 1991): (i) A direct biological effect from a chemical can occur when the chemical causing the effect is present in the tissues of the organism where the toxic effect occurs. Sec ondary effects at other sites can occur that are triggered by the direct effect; (ii) Once the concentration of the chemical in the tissues reaches a critical value that exceeds the ability of the tissue to counteract the effects of the chemical, the magnitude of the effects on the organism then increases in proportion to the concen tration of the chemical in the tissues; (iii) The amount ofchemical at the site of toxic action in the tissues is proportional to the amount of chemical absorbed into the organism, and the amount of chemical absorbed depends, in part, on the concentration of the chemical in the environment to which the organism is exposed. The consideration of the bioavailability of the chemical is important in the evaluation of potential adverse effects, since many chlo rinated organic chemicals in complex environmental media are absorbed much less efficiently than from ideal test vehicles used for administration in laboratory studies. In keeping with the above observations, regulatory processes are designed to reduce risks to acceptable levels, recognizing that the theoretical ideal of "Zero Risk" is not practical nor is it attainable (Johannsen, 1990). The existence of threshold doses for numerous potentially toxic substances has been acknowledged by the American Con ference of Governmental Industrial Hygienists (ACGIH) in its threshold Emit values, by the FAO/WHO Joint Expert Committee on Food Additives, and the FAO/WHO Joint Meeting on Pesticides in their establishment of acceptable daily intakes (ADI) for hundreds of food additives and pesticide residues, and by regulatory agencies in various countries around the world, in the development of regulatory standards for numerous and diverse types of chemicals encountered in food, consumer products, and the environment. Chlorinated organic chemicals are generally agreed to have exposure thresholds for adverse effects, for example, PCBs (EPA, 1988), chlorinated dioxins (Kociba et al, 1978), DDT, and dieldrin (Newton et al., 1989), all of which elicit a toxic response 326 WILLES ET AL. at or above a threshold tissue concentration or dose. A threshold for adverse effects does not mean that reversible changes in various physiological and biochemical pa rameters, occurring without adverse consequences to the organism, would not be observed at small exposures (Klaassen and Eaton, 1991). For example, a proposed mechanism of liver tumor production by chlorinated dioxins is through a combination of their cytotoxic effects on liver cells and indirect effects which cause imbalances in the endocrine system due to the induction of enzyme systems involved in hormone metabolism. At doses below those resulting in liver cell toxicity and hormonal changes, no carcinogenic effect is expected. Similarly, thresholds would exist for adverse effects on reproduction, immune function, and other effects. Many of these adverse effects are secondary to changes in enzyme activities induced by chemicals (e.g., induction ofaryl hydrocarbon hydroxylase (AHH) enzyme systems by chlorinated dioxins).''fKese enzyme systems also change activities as part ofnormal physiological responses; there fore, merely inducing the enzyme does not necessarily mean an adverse effect will occur. In Canada, regulatory agencies have based their evaluation ofthe potential risk of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the fact that practical thresholds exist for tumorigenesis and reproductive effects, and have set an exposure limit of 1C pg/kg/day (CEPA, 1990) with the recognition that a safe exposure exists, below which there would be no expectation of adverse effects. Other countries, including Norway Sweden, the United Kingdom, and Germany also regulate 2,3,7,8-T4CDD as a chemica with a threshold-type mechanism of action; however, the safety factors used in estab lishing maximum allowable exposure limits vary from 100-fold, as used in Canada to 1000-fold as used in Germany. Mounting evidence supports the premise that there is a practical threshold for adverse effects. For example, as concentrations of groups of bioaccumulative chemicals in thi environment have decreased, species of wildlife previously affected by historicall; greater concentrations are returning to normal health status, even in the presence o lesser concentrations of these groups of chlorinated organic chemicals. These finding support the regulatory premise that thresholds exist below which adverse effects wouk not be expected to occur, and that the presence of a chemical in the environment doe not necessarily result in toxic effects. In the remainder of this section, examples c this dose-response phenomenon and evidence for the existence ofa threshold in report of dose responses in exposed wildlife will be presented. The importance of considering dose-response relationships in the evaluation c chlorinated organic chemicals is demonstrated by the available literature relating dos to effect for specific chlorinated organic chemicals. In the past, exposures to enviror mental concentrations of certain bioaccumulative chlorinated organic chemicals hav been sufficiently great to implicate these chemicals in the reproductive failure and bi deformities that have been observed in several species of colonial piscivorous birc inhabiting the Great Lakes Basin (Yamashita et al, 1993; Hoffman et al., 1987; Kubir. et al, 1989; Fox et al, 1991; Gilbertson et al, 1991; Ankley et al, 1992; Tillitt et a. 1991, 1992). Decreased reproductive performance and/or population declines wei reported through the 1960s and 1970s in at least nine top-predator bird species in tl Great Lakes and other regions of North America. These declines have been associate with increased tissue concentrations of several bioaccumulative chlorinated organ chemicals (e.g., chlorinated dioxins/furans, PCBs, hexachlorobenzene, lindane, DD dieldrin), although the simultaneous destruction of necessary habitat and exposur to an array of other nonchlorinated chemicals and metals are also likely involve R&S 143957 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 327 While exposure of affected species to these bioaccumulative chemicals has decreased in recent years, and several species have experienced recovery, some species have not yet fully recovered. Developmental and reproductive effects still are observed in the areas of the Great Lakes Basin (Environment Canada, 1991a; Tillitt etal, 1991, 1992) and Western Europe (van den Berg et al, 1992; Bosveld et al, 1992) where the greatest concentrations of these chemicals and loss of habitat due to various human activities occur. These observations suggest that localized exposures are above the thresholds for adverse effects. In other areas of the Great Lakes Basin, the environmental con centrations ofthese chemicals appear to have decreased such that exposures are below the threshold and no longer produce measurable adverse effects (Fox et al, 1991). j Another manifestation of the dose-response relationship observed in populations of wild birds is the decline and recovery of the bald eagle associated with the rise and fall ofenvironmental concentrations of DDT and its metabolites. As a tertiaiypredator, the bald eagle may be exposed to the greatest concentrations in its food and receive the greatest dose of such bioaccumulative chemicals of all wildlife. It is believed that exposures to DDT and its metabolites (and possibly other bioaccumulative chemicals) were associated with severe population declines of bald eagles on the shores of Lakes Ontario and Michigan (Gilbertson et al., 1991), although loss of habitat was also a likely factor. The current reproductive status of Great Lakes bald eagle has greatly improved in several regions, with estimates of breeding pairs at 2400 in 1989 compared to a low of 400 breeding pairs in 1964. Although populations of bald eagles have not yet recovered on Lakes Huron and Ontario, the Lakes Erie and Superior populations now appear to be breeding normally. The lack of suitable habitat (e.g., loss of physical habitat, excessive human populations, large urban areas) may also have an important role in the poor reestablishment of bald eagles, particularly in some areas of the lower Great Lakes (Grier, 1982). Similar situations have been observed in other piscivorous bird populations around the Great Lakes (Struger and Weseloh, 1985; Noble and Elliot, 1986; Gilbertson et al, 1991). Increases in concentrations of certain chemicals in their tissues have been associated with population declines. Subsequent reductions in concentrations of chemicals have been associated with population recoveries. By the early 1960s, the double-crested cormorant had essentially ceased reproducing in Lakes Michigan and Ontario, and populations were also declining in the other Great Lakes (Environment Canada, 1991b). In 1972,95% ofthe eggs laid in the Georgian Bay colony disappeared or were broken by the end of the incubation period (Environment Canada, 1991c). Cormorants are particularly sensitive to eggshell thinning, an effect associated with DDE (Tillitt et al, 1992), since they incubate their eggs by standing on them (Envi ronment Canada, 1991c). Between 1979 and 1987, the incidence of bill defects in cormorants in the Green Bay (52.1 per 10,000) and Beaver-Mackinac (12.3 per 10,000) areas of Lake Michigan were substantially greater than in Lake Superior, Huron, or Ontario* in Northern Ontario, or in the Canadian Prairies (Fox et al, 1991). It has been postulated, based on a weight-of-evidence assessment, that the bill deformities observed in cormorants were associated with elevated tissue concentrations of certain bioaccumulative chlorinated chemicals such as the non-orr/io-chlorine substituted dioxins, furans, and PCBs (Fox et al, 1991; Environment Canada, 1991c). This pos tulate is based on observations that some of the non-ort/zo-chlorine-substituted isomers in the tissues of these birds are known to cause abnormalities in skeletal development R&S 143958 328 WILLES ET AL. in experimental laboratory animals (Courtney et ai, 1976; Marks el ai, 1981; Warden et ai, 1982; EPA, 1985; MOE, 1985; Weber et al., 1985). In keeping with the principle that response is a function of dose, decreased concen trations of persistent chlorinated organic chemicals in the Great Lakes were associated with a 56% annual increase in double-crested cormorant populations between 1974 and 1982 (Price and Weseloh, 1986). Current populations are greater than at any other time since cormorants entered the region in about 1928 (Environment Canada, 1991a). The adverse reproductive effects now observed are confined to a few greatly contaminated areas, such as Green Bay in Lake Michigan, where the occurrence of crossed bills in cormorants remain elevated (Tillitt et al., 1991, 1992). The above discussion indicates that developmental effects, impaired reproduction* or complete reproductive failure reported in several species of piscivorous birds in the Great Lakes Basin appear to be associated with excessive concentrations of bioaccu mulated chemicals in various tissues (Mineau et al., 1984; Price and.Weseloh, 1986; Bishop and Weseloh, 1990; Gilbertson et ai, 1991). In the early 1970s, the concen trations of several chlorinated organic chemicals in herring gull eggs and other waterbirds from the Great Lakes were among the highest reported anywhere in the world. Results from wildlife monitoring programs in effect since 1971 indicate that the com centrations of PCBs and several other chlorinated chemicals in herring gull eggs have declined since the mid-1970s (Bishop and Weseloh, 1990) (Fig. 1). Results reported by Newton et al. (1989) indicate that piscivorous birds in Britain also are recovering, with concurrent reductions in their tissue concentrations of various chemicals. Newton et al. (1989) conducted a detailed analysis of organochlorine and mercury concentrations in the eggs ofBritish peregrine falcons. The authors examined possible time trends in residue levels, egg shell thickness, and analyses of chemical residues as they may relate to breeding success. The chemicals measured were DDE, l,2,3,4,10,10-hexachloro-6,7-epoxy-l,4,4a,5,6,7,8,8a-octahydro-l,4:5,8-dimethano- H Lake Ontario 9 Lake Erie ED Lake Huron 0Lake Michigan E3 Lake Superior Fig. 1. Trends in annual average total PCB concentrations in herring gull eggs collected from the Great Lakes, 1974-1989. Data from Environment Canada (1991b). Snake Island, Lake Ontario; Middle Island. Lake Erie; Double Island, Lake Huron; Big Sister Island, Lake Michigan; Granite Island, Lake Superior. R&S 143959 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 329 naphthalene (HEOD, the active ingredient ofdieldrin), PCBs, and mercury. The authors;reported that the recent breeding success and recovery of peregrine falcon populations were associated with a general decline in concentrations of DDE, HEOD, and mercury in peregrine eggs over the study period, 1963 to 1986. In summary, several studies have proposed an association between the adverse effects in wildlife populations based on field observations and various bioaccumulative chemicals, including mercury, PCBs, DDE, chlorinated dioxins, dieldrin, and some others. The weight of available evidence indicates that these,.chemicals, and in some cases, other causal factors including other chemicals, decreased quality of habitats, and various diseases unrelated to chemical releases are implicated in the adverse effects^ observed in Great Lakes piscivorous wildlife. Since the curtailment of the use of PCBSr DDT, dieldrin, and hexachlorobenzene in the early 1970s, the egg and tissue concen trations ofthese chemicals and by implication the exposures ofwildlife to these chem icals have declined in most species around the Great Lakes (Fig. 2) (Bishop and Weseloh, 1990). Coincident with the declining tissue concentrations of chemicals, most species have exhibited some degree of population recovery, providing evidence that any adverse effects ofthese chemicals are reversible and appear to be dose-related. The declining environmental concentrations of bioaccumulative chlorinated organic chemicals also are reflected in human populations (Regulatory Network Inc., 1992), although such comparisons over time may be confounded by temporal variations in the analytical procedures used and in sampling protocols. Diet is a major source of human exposure to PCBs, primarily due to the concentrations of PCBs in major components ofthe human diet (e.g., fish, dairy products, various red meats). Exposures of humans to PCBs through diet decreased by approximately 10-fold during the 1970s (from approximately 6.9 fig PCB/day in 1971 to approximately 0.7 fig PCB/day in 1980) and then continued to decline to 0.05 Mg PCB/day by 1989, a 138-fold decrease from the concentrations in 1971 (Fig. 3). R&S 143960 1974 1976 1978 1980 1982 1984 1986 1988 H Lake Ontario H Lake Erie @ Lake Huron O Lake Michigan 0 Lake Superior Fig. 2. Trends in annual average total DDE concentrations in herring gull eggs collected from the Great Lakes, 1974-1989. Data from Environment Canada (1991b). Snake Island, Lake Ontario; Middle Island, Lake Erie; Double Island, Lake Huron; Big Sister Island, Lake Michigan; and Granite Island, Lake Superior. 330 WILLES ET AL. Fig. 3. FDA Total Diet Study: Adult intakes from 1971 to 1990. Data are from the Food and Drug Administration (FDA) Total Diet Study of 234 food items in the United States. 1989 represents an average of the years 1988, 1989, and 1990. Data from Regulatory Network, Inc. (1992). Information on concentrations of PCBs in human tissue is provided from the Na tional Human Adipose Tissue Survey (NHATS) in the United States, where adipose tissue specimens from statistically selected samples were analyzed for a series of toxic chemicals. The results were grouped as not detected, detected but less than 1 ppm, between 1 and 3 ppm, or greater than 3 ppm (Regulatory Network Inc., 1992). In 1972,62% ofthe population had PCB concentrations of greater than 1 ppm compared to 2% in 1984 (Fig. 4). In 1972, 4% of the population had greater than 3 ppm of PCBs, and the proportion increased to slightly less than 10% by 1977; however, by 1983 and 1984, less than 1% of the samples had greater than 3 ppm. Robinson et al. (1990) reported similar trends in their independent review of the NHATS (National Human Adipose Tissue Survey) data. Results from German studies of chlorinated organic chemicals in human breast milk, although subject to differences in analytical proce- G aa. S H & (/> 6e coo 4 CO <a>o CL Fig. 4. National Human Adipose Tissue Survey results, 1972-1984: PCB levels greater than 1 (13) or () ppm. Data from Regulatory Network, Inc. (1992). CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 331 dures, also indicate that concentrations have decreased over the past decade (19811991) (Beck and Heinrich-Hirsch, 1993). Since the concentrations of chlorinated organics with greatest molecular weights have decreased in human tissues, it is reasonable to suggest that the dose also has V decreased. There is no evidence that indicates conclusively that past exposure ofhumans to environmental concentrations of chlorinated organics has been above a threshold dose for adverse effects. The application of the dose-response principle to exposures ofwildlife and humans to certain bioaccumulative chlorinated organic chemicals demonstrates that, as chem ical concentrations in the environment decline, exposures to humans and wildlife also decline. Coincidentally, in species where adverse effects were observed, these effects decrease in magnitude as exposures decrease, hindering the establishment of causal relationships between specific chemicals and the adverse effects reported, since, a variety of other factors begin to play a more central role. These observations substantiate both the dose-response concept and, in cases where data are adequate, the existence of exposure thresholds below which adverse effects are not measurable. In addition, the adverse effects appear reversible as exposure diminishes, at least on a population basis. METABOLIC CAPABILITY OF ENVIRONMENTAL SPECIES--THE SIGNIFICANCE OF NATURALLY OCCURRING CHLORINATED ORGANIC CHEMICALS Recent information shows that more than 1500 different halogenated chemicals are produced and released into the environment by various plants, marine organisms, insects, bacteria, fungi, and various physical natural processes. The numbers of naturally occurring chlorinated organic chemicals that have been identified have expanded sig nificantly in the past decade (e.g., 30 naturally occurring chlorinated chemicals had been identified in 1968, compared to 1500 in 1992). According to Gribble (1992), the increased interest in the area of organohalogen natural product chemistry will result in the identification of many,more naturally occurring sources of organohalogen chemicals in the future. The naturally occurring chlorinated organic chemicals iden tified to date include chlorinated alkanes, alkenes, benzenes, phenols, PCBs, and chlo rinated dioxins and furans and encompass a broad range of substances representative of the same classes of chemicals, with the same physical/chemical properties as those produced by various human activities (Gribble, 1992). The same dose-response prin ciples govern the potential occurrence ofadverse effects from these chlorinated organic chemicals, regardless of their natural or anthropogenic origin. Furthermore, the ex istence of natural sources ofchlorinated organic chemicals demonstrates that the earth's ecosystems have been exposed to these chemicals independent of human activities and can survive normally in the presence of these types of chemicals. Naturally occurring chlorinated organic chemicals that are released into the envi ronment are produced by a number of organisms, including bacteria, fungi, algae, ferns, and higher plants (Siuda and DeBemardis, 1973; Fenical, 1981; Harper, 1985; Gribble, 1992; NCASI, 1992). For example, chloromethane is produced by marine algae, wood-rotting fungi, the evergreen cypress, and several other species (Gribble, 1992). Similarly, 2,4-dichlorophenol (a chlorinated aromatic) is produced by soil fungi, grasshoppers have been shown to secrete 2,5-dichlorophenol, and 2,6-dichlorophenol R&S 143962 332 WILLES ET AL. is secreted as a pheromone by the female lone star tick (Gribble, 1992). In addition v t many compounds that are synthesized by organisms to protect themselves from pests or competitors contain one or several chlorine atoms (Ray, 1991). Chlorinated dioxins and fufans have been produced as unwanted by-products from i a variety of processes, but .they are produced in significant amounts also through natural processes. They appear to be ubiquitous in the environment due to numerous ! chemical reactions occurring during the combustion of natural and synthetic organic materials in the presence of naturally occurring chlorine and Have been detected in a } wide range of particulate samples (e.g., ash samples from the combustion of untreated / j firewood such as birch, oak, willow, mountain ash, and applewood in fireplaces). This - ! leads to the conclusion that natural combustion during forest fires and slash burning f are contributors to the existing concentrations of these chemicals in the environment (Bumb et ai, 1980; Nestrick and Lamparski, 1982; Clement et ai, 1985; Tashiro et ai, 1990). An additional natural source of chlorinated dioxins and furans appears to be through the transformation of chlorophenols catalyzed by bacterial peroxidase en zyme systems, leading to their presence in contaminated sewage sludge, soils, and sediments as microbial transformation by-products from chlorophenols in sewage (Oberg et ai, 1991). Other natural sources of chlorinated organics include decom position of seaweed, and volcanic activity (Travis and Hattemer-Frey, 1989; Gribble, 1992). Conversely, even though chlorinated dioxins and furans are produced by natural processes, studies ofsediment cores from the Great Lakes show a correlation ofincreases in concentrations ofthese chemicals with increased human activities in the past century (Czuczwa etal., 1984; Czuczwa and Hites, 1985; Hites, 1990), indicating that historical production from human activities exceeded those produced by natural processes. While the relative contribution of natural sources of chlorinated organic chemicals to environmental concentrations would be small in areas ofthe greatest anthropogenic production, the relative contributions of natural sources to low-level background con centrations are considered quite significant, particularly in environments where largescale degradation of organic material occurs (Gribble, 1992). For example, it has been estimated that 300,000 metric tons of total absorbable organohalides (AOX) are pro duced per year through natural processes in peat bogs in Norway (Asplund et ai, 1989). The occurrence of chlorinated organic chemicals in the environment and their presence in diverse groups of mammalian and nonmammalian species demonstrate that organisms are capable of existing in the presence of certain environmental concentrations of these chemicals with no apparent adverse effects. Many chlorinated organic chemicals are metabolized by mixed function oxygenase (MFO) enzymes (Safe, 1989), an inducible, nonspecific enzyme system that metabolizes compounds into more water-soluble forms, thereby facilitating excretion (Matthews and Dedrick, 1984: Rand and Petrocelli, 1984). This enzyme system is inherent in mammals, birds, aquaticspecies, and some invertebrates (Environment Canada, 1991c) and is also involved in the metabolism and homeostasis of hormones and other endogenous substances. Other chemicals metabolized by this group of enzymes include caffeine (Environment Can ada, 1991c), /3-naphthoflavone, phytosterols found in plant tissue (Hodson et al. 1992b), naphthalene, and benzo[a]pyrene (Rand and Petrocelli, 1984). Aflatoxin B,, a toxin produced by fungal molds on peanuts, other nuts, and grains, also is metabolized by the MFO enzyme system (Schoenhard et al, 1976). In addition to metabolizing exogenous chemicals that enter the body, MFO enzymes are responsible for the me- R&S 143963 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 333 tabolism of various reproductive hormones and chemicals that regulate moulting in aquatic invertebrates (Environment Canada, 1991c). Since MFO enzymes metabolize a variety of chemicals (naturally occurring and synthetic, chlorinated and nonchlorinated) by normal physiological processes, it is clear that the MFO system does not , exist solely for the metabolism of chlorinated chemicals, and thus induction of MFO is not necessarily indicative of exposure to chlorine-containing chemicals. These observations add support to the conclusion of Mackay (1992) that there is no basis for suspecting that exposures to small quantities of humaa-made chlorinated organics would result in the perturbation ofbiological systems. Furthermore, the ubiq uity of chlorinated organic chemicals in the environment and the presence in diverse groups of mammalian and nonmammalian species of a common enzyme system, normally used for endogenous metabolic purposes but also capable of metabolizing chlorinated organic chemicals, provide evidence that organisms have evolved in the presence of these chemicals in the environment. SPECIFICITY OF ASSOCIATION AND BIOLOGICAL PLAUSIBILITY OF REPORTED ADVERSE EFFECTS There are marked differences among the various chlorinated organic chemicals with regard to potential risk from exposure. Consequently, it is important to evaluate the information carefully regarding associations between reported effects and exposures in the environment. The critical point of such an evaluation is the examination of the weight of evidence for cause-effect relationships. The establishment of reasonable cause-effect relationships based on good science is essential to invoke effective miti gation strategies. Unless this course is followed, mitigation procedures designed to resolve environmental problems may focus on the wrong factors and, consequently, be ineffective. Obtaining unequivocal or definitive proof of cause-effect relationships at a mechanistic level can be time consuming and expensive, and unwanted adverse effects could occur while waiting for "definitive proof' of causality. Therefore, a weightof-evidence approach, combined with the application of sound scientific judgment, is needed to determine when the available evidence is sufficient to establish associations between cause and effect that are definitive enough to achieve the desired results within an environmental and socioeconomic context. The establishment of cause-effect relationships usually is based on a combination ofepidemiological or ecoepidemiological information and controlled laboratory studies. Epidemiological studies focus on the possible associations between human diseases or conditions and potential causal factors (Hill, 1965). Ecoepidemiology refers to the study ofpossible associations between adverse effects on components ofthe ecosystem F (including humans) and potential causal factors (Bro-Rasumssen and Lokke, 1984). p Laboratory studies to facilitate observations of cause-effect and dose-response relaI tionships are conducted under controlled conditions where many of the variables that | can affect the responses of organisms to chemicals can be controlled. Practical conI siderations, however, relate to the applicability of laboratory information to field or m "real-world" conditions and to the relevance of field observations because ofthe myriad B of confounding variables encountered in the real world. Hill (1965) concluded that B the following features must be considered in attempting to infer causality from asso- m ciations between effects and various factors. R&S 143964 334 WILLES ET AL. (i) Strength of the association: How often is the event of concern observed in pop ulations exposed to the factors of concern, compared to populations that are not exposed to these factors? (ii) Consistency of the observed association: Has the association repeatedly been observed by different.investigators, in different places, circumstances and times? (iii) Specificity of association: Is there an association between specific populations, or to particular sites or areas and a specific disease? (iv) Temporality of association: Is the disease or effect approximately related to the time of exposure? (v) Is there a dose response between the disease or effect and the agent(s) involved? This has not been interpreted to mean that a statistically significant dose-response curve is required to establish causality; however, there should be reasonable evidence to enable a scientific judgment that the effect becomes greater as exposure increases. (vi) Plausibility of association: Is the association biologically plausible based on the biological knowledge of the day? Evidence of implausibility should have more weight than actual lack of information on the biology. (vii) Coherence of association: The cause-effect interpretation should not be seri ously at conflict with the generally known facts of natural history and disease. (viii) Is there experimental evidence supporting the association? Does removal of the associated agent cause the effect to disappear or diminish in magnitude? (ix) Analogy of association: The effects observed are known to be caused by similar agents. Fox (1991) grouped the essential features proposed by Hill (1965) for the inference of cause-effect relationships in ecoepidemiology studies. Four criteria considered to help to validate causal associations include: (i) Strength of association; (ii) Consistency of association; (iii) Predictive performance or specificity of association; . (iv) Evidence of dose-response relationships. Three criteria considered to indicate that various events are not causally associated include: (i) Incompatibility on the basis oftime order (e.g., the event occurs before exposure or increases as exposure decreases); (ii) Lack of evidence of factual or biological plausibility; (iii) Lack of consistency in replication. If the application of sound scientific principles is to be followed, these criteria must be applied before a causal relationship is inferred between effects and exposures to specific chemicals in the environment. As indicated previously, laboratory information must be used, whenever available, to assist in interpreting associations between specific agents and effects. Associations Based on Information from Environmental Studies In this section, the criteria for establishing causal association, as discussed above, are applied to three reported associations between environmental factors and adverse effects: R&S 143965 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 335 (i) This evaluation shows that the alleged causal association between chemical ex- *' posure and effects on marine mammals is confounded by a number of factors. (ii) In the case of potential adverse reproductive effects on fish-eating birds on the Great Lakes, the evaluation shows a relatively strong association was evident between - historical exposures to certain bioaccumulative chlorinated chemicals and reproductive health, although there were likely other contributing factors. As environmental con centrations ofthese chemicals have decreased in the Great Lake region, the occurrence of continuing effects also has diminished. (iii) In the case of pulp mill effluents, the evaluation shows that chlorine bleaching is not exclusively responsible for the effects reported in fish populations. ^ These three examples are examined in more detail in the descriptions that follow. Adverse effects on marine mammals. Mass mortalities have occurred in marine mammals over the past decade or more in various regions, although there is consid erable debate whether possible causal factors include persistent chlorinated organic chemicals or are solely expressions of natural disease processes. Reports have been published concerning the cause of death of the common seal in the Dutch area of the Wadden Sea. Dietz et al. (1989) and others (Grachev et ai, 1989; Harwood et al., 1989; Osterhaus et ai, 1989) reported that the cause of death for these seals was most likely a virus resembling canine distemper virus. Furthermore, these authors reported a number of viruses associated with mass mortalities of the seals, including influenza, herpes, and canine distemper virus. These viruses are spread in aerosols when infected seals cough. The mass mortality of Lake Baikal seals in 1987, ofNew England harbor seals in 1979-1980, and ofcrabeater seals in the Antarctic ' in 1955 were also attributed to viral infections (Geraci et ai, 1982, 1984; Hinshaw et ai, 1984; Grachev et ai, 1989). Furthermore, Geraci et al. (1982) reported that most affected populations were showing increases in numbers prior to the outbreak of the disease, which suggests a relatively healthy population or one showing no overt signs of adverse effects. Thus, it is unclear how cumulative chemicals could be causally related to such short-term events. The New England harbor seal population had doubled since 1972 at the time of the disease outbreak, in 1979-1980. Similarly, the crabeater seal population was greater than normal in 1955, when 85% of the population died of a viral pneumonia infection. Interestingly, milder winter temperatures occurred during both these epidemics, resulting in seals spending more time out of water and congregating on land (Geraci et al., 1982). This closer contact would facilitate the spread of the diseases. These facts implicate a causative factor other than environmental concentrations of persistent chlorinated organic chemicals since tissue concentrations and environmental concentrations of such chemicals change only slowly with time. It has been suggested that viral epidemics may be a result oforganochlorine-induced impairment of the immune system (Brouwer et ai, 1989). Effects on the immune system have been observed in laboratory animals following exposure to DDT (Rehana and Rao, 1992), toxaphene (Allen et ai, 1983), dieldrin (Exon et al., 1987), lindane (Allen et al., 1979), hexachlorobenzene (HCB) (Allen et ai, 1979), chlorinated dioxins . and furans (Silkworth and Vecchi, 1985), PCBs (Silkworth and Vecchi, 1985), and polybrominated biphenyls (PBBs) (Thomas and Faith, 1985). Exposure of laboratory f animals to extreme doses of these chemicals causes suppression of antibody responses, k atrophy of lymphoid organs, and increased susceptibility and sensitivity to toxic or I' infectious agents. Other chemicals that have been demonstrated to cause immunological R&S 143966 336' WILLES ET AL. effects, also at extreme doses relative to usual environmental exposures, are polycyclic aromatic hydrocarbons, benzene, lead, inorganic mercury, and other metals (Dean and Murray, 1991). Consequently, wild animals could be exposed to several chemicals that affect the immune system. '..It has also been suggested that a mechanism possibly explaining PCB-induced im mune effects allegedly leading to disease in the common seal may be related to defi ciencies in vitamin A and thyroid hormones (Brouwer et al., 1989). Vitamin A (retinol in its various forms) is a fat-soluble vitamin, important for normal vision, reproduction, and development in marine mammals (Environment Canada, 1991c). Thyroxine is a thyroid hormone which is important for normal growth and development. Seals fed fish from the Wadden Sea, having significantly greater concentrations of DDE and -oPCBs than fish from the northeast Atlantic, had lesser concentrations ofplasma retinol and thyroxine. Since laboratory studies with experimental animals have indicated an effect of PCBs in regulation of vitamin A and thyroid hormones, Brouwer 'et al. (1989) suggested that the PCBs with non-ortfio-chlorine substitutions, or hydroxy metabolites: of non-ort/zo-substituted PCBs, impair normal vitamin A and thyroxine homeostasis, which in turn causes reduced concentrations of vitamin A and an increased suscep tibility to stress. It has been suggested that reduction of both retinol and thyroxine may be caused by the interference of PCBs, or metabolites of PCBs, or of secondary products related to effects caused by these chemicals, with the retinol-thyroxine plasma carrier-protein complex (Brouwer et al., 1989). The reduced levels of vitamin A and immunoreactive thyroid hormones have been suggested to result in reduced repro ductive success, retarded growth, and increased susceptibility to infection (Reijnders,. 1986; Brouwer etai, 1989; Manson and Wise, 1991; Schumacher etal., 1991). Bergman and Olsson (1985) also speculated that PCBs in the diet and other chlorinated organic chemicals have caused adverse effects on the endocrine system of Baltic seals, leading to immunosuppression and interrupted pregnancies. The results from these studies are difficult to interpret since, although effects observed in the Baltic seals may be the result of excessive exposure to PCBs, the existence of viral epidemics in relatively pristine environments and the fact that the Baltic Sea is contaminated with numerous other chemicals that affect the reproductive and immune systems seriously confound the identification of a single causal factor. Similarly, some attempts have been made to link the mass mortality of bottlenose dolphins along the U.S. Atlantic coast (1987-1988) to organochlorine contamination associated with compromised immune function. However, the results of a compre hensive investigation concluded that the dolphin mortalities were due to contamination with excessive levels of a natural neurotoxin produced by dinoflagellate algae (Geraci, 1989). Blooms of these algae, also known as "red tide," are known to occur on the east coast, but are particularly common on the Gulf Coast of Florida. During the summer and fall of 1987, blooms were carried north by the Gulf Stream along the east coast resulting in human poisoning from the consumption of affected fish (Bonaventura and Bonaventura, 1987) and the temporary closure of the shellfish industry (Tester et al., 1989). Geraci (1989) questioned the significance of environmental con taminants in the resilience of these animals to infection, and it was concluded that the mass mortalities of bottlenose dolphins observed on the east coast of the United States were due to a natural plant toxin (Geraci, 1989). Therefore, the weight of evidence indicates that the occurrence of mass mortalities in marine mammals is likely due to R&S 143967 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 337 natural disease factors and the effects of weather and factors related to habitat. It is doubtful that there is a causal relationship with chlorinated organic chemicals. Adverse reproductive effects on birds. A second example of the inference of causeeffect relationships involves environmental concentrations of chlorinated organic chemicals and a variety of effects observed in piscivorous birds. Based on the inter pretation ofthe results from laboratory and field studies, a causal association has been inferred in Great Lakes piscivorous birds between highly chlorinated organic chemicals and a disease known as chick edema, a syndrome characterized by embryo mortality, edema, and deformities (referred to as "GLEMEDS" by the authors) (Gilbertson et al, 1991). An increased incidence of abnormalities, including bill defects and eye and-, foot deformities, was observed in colonies of common and roseate terns at Long IslandSound in 1969 and 1970 (Hays and Risebrough, 1972). Eggs from these colonies were found to contain greater concentrations of PCBs than those from reference colonies. Gilbertson (1983) reported high mortality, edema, porphyria, liver enlargement, fatty infiltration, and necrosis, as well as growth retardation and embryo deformities (symp toms similar to chick edema disease) in Great Lakes herring gulls. Retrospective anal yses of the eggs revealed the presence of significant concentrations of the chick-edema active compounds (chlorinated dioxins and furans) and coplanar PCBs (Stalling et at., 1985; Kubiak et al, 1989). Concentrations of PCBs in eggs of Forster's tern associated with adverse reproductive effects (reduced hatchability, delayed hatch, reduced number of fledglings) were: total coplanar PCBs, 1.37 to 41 ppb (wet weight (ww); total PCBs, 6.2 to 25.9 ppm (ww) (Kubiak et al, 1989). Yamashita et al (1993) suggested a role for chlorinated organic pesticides and PCBs in the reproductive disorders observed in double-crested cormorant and Caspian tern chicks from the upper Great Lakes (Kurita et al, 1987; Kubiak et al, 1989). Eggs from these species contained concentrations of PCBs and DDE in the range where reproductive effects had been observed in Forster's tern and British peregrine falcon. Of all chemicals analyzed, concentrations of DDE and PCBs occurred at the greatest concentrations. Concentrations of PCBs ranged from 3.6 to 14 ppm (ww) while con centrations of DDE ranged from 2.2 to 6.3 ppm (ww). These PCB concentrations were in the lower range of concentrations needed for adverse reproductive effects on Forster's tern reported by Kubiak and coworkers (1989). DDE concentrations were in the range of the threshold concentration for adverse effects to peregrine falcon (3 ppm) reported by Newton et al (1989). In considering the criteria for causality as elaborated by Hill (1965) and Fox (1991) with respect to an association between PCBs and DDE and effects on bird reproduction the following points may be made: (i) laboratory studies affirm the plausibility of an association, and environmental concentrations were relevant to the effects observed; (ii) changes in environmental concentrations over time and coincident changes in response reinforce a possible dose-response and temporal relationship; and (iii) con sistency of association is evident in several studies on different species and in different areas. In the absence of information that could negate the association (as per Fox, 1991), the weight of evidence appears to indicate that the reported adverse effects on the reproductive health of fish-eating birds were likely associated with exposures to PCBs and DDE; however, since wild animals are exposed to other bioaccumulative agents that may also affect reproductive health, it cannot be concluded definitively that PCBs and DDE are the sole causal factors. R&S 143968 338 WILLES ET AL. The possibility that other factors may influence reproduction is evident from the results ofcontrolled laboratory studies as well as from biomonitoring data. For example in laboratory studies, fish exposed to PCBs at concentrations similar to those found in the environment had reduced survival, reproduction, and growth (Mayer et al 1985). However, mortality of feral Chinook salmon eggs reared in a laboratory was not correlated to PCB egg residues (Williams and Giesy, 1992). In addition, mortality of Lake Michigan lake trout fry increased from 22 to 92% between 1975-1976 (Berlin et al, 1981) and 1980-1981 (Mac et al., 1985), while the total environmental con centrations of PCBs and DDT decreased over that time frame. Williams and Giesy (1992) concluded that the lack ofcorrelation between various PCB exposure parametersand survival of eggs and fry may indicate that other factors affecting habitat such-as large amounts of suspended sediments, excessive quantities of sewage discharge, un treated sewage, and elevated ammonia concentrations were involved in the rearing mortality observed. Pulp mill effluents. A third example of possible cause and effects based on field observations involves effects noted in fish exposed to pulp mill effluents (Sodergren et al., 1988; Sodergren, 1989). Pulp mill effluent has received a great deal of attention concerning toxicity and the presence of chlorinated organic chemicals, particularly dioxins and furans. Pulp mill effluents contain complex chemical mixtures, and the toxicity of the effluents varies, depending on the pulping and bleaching technology employed, the treatment system used, and even the species of tree used to make the pulp (Salkinoja-Salanen et al., 1984; McLeay and Associates, 1987). A great deal of research has been directed toward identifying the constituents) in mill effluent re sponsible for the effects noted in fish populations. It has been suggested that the toxicity of pulp mill effluents may be regulated using the measure of total adsorbable organic halide (AOX), representing the total concen tration of all chlorinated and other halogenated materials of all types in the effluent. However, the relative chemical composition of AOX (hence, potential toxicity) in effluent varies considerably, both between mills and temporarily within a specific mill with respect to the proportions of high and low molecular weight substances. Since toxicity is directly related to exposures arising from the relative concentrations of specific chemical constituents in the receiving waters in a dose-dependent manner, AOX alone is not adequate to describe the potential toxicity of the pulp mill effluent. A variety of nonlethal endpoints ranging from effects on growth and development to biochemical and hormonal changes have been associated with the exposure of aquatic species to pulp mill effluent. These effects must be examined critically to ascertain what constituent(s) is specifically associated with and, subsequently, what process should be controlled to minimize adverse environmental effects. Increased liver enzyme activity of mixed function oxygenases (MFO) has been used as an indicator of exposure of wild fish to aromatic hydrocarbons, although such enzyme changes do not in themselves indicate the occurrence ofadverse effects (Payne et al., 1987; Environment Canada, 199 Id). Induction of MFO activity has been ob served in fish collected downstream of pulp mill outfalls in Canada (Rogers et al., 1989; Smith and Rokosh, 1989; Munkittrick et al., 1989, 1991; Servos et al., 1992; Hodson et al, 1992a) and Scandinavia (Larsson et al, 1988; Sodergren et al, 1988; Andersson et al, 1988; Lindstrom-Seppa and Oikari, 1989, 1990a,b), compared to reference fish not exposed to pulp mill effluent. It has been suggested that increased MFO activity may lead to enhanced metabolism of circulating concentrations of sex R&S 143969 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 339 hormones thereby causing reproductive effects that have been noted in fish populations, (McMaster etal, 1991; Munkittrick etal, 1991). Although the substance(s) responsible for the observed MFO induction near pulp mills has not been identified (Munkittrick et ai, 1992a; Hewitt et al, 1992), chlorinated dioxins and furans have been proposed as causative agents (Rogers et ai, 1989; Hodson et ai, 1992a; Lockhart and Metner, 1992). More recent information, however, suggests that constituents in bleached kraft mill effluent other than chlorinated dioxins and furans may be responsible for increased MFO activity in fish (Munkittrick et al., 1992a,b). Enzyme induction by chlorinated dioxins and furans has been shown to persist for 85-180 days under laboratory con ditions (Muir et ai, 1990; Van der Weiden et ai, 1990). However, MFO activityjn three species of fish, collected downstream of secondary treated pulp mill outfall, returned to control values following a planned mill maintenance shutdown and a 2week effluent-free exposure period (Munkittrick et al., 1992a). These observations refute the contention that chlorinated dioxins and furans, or other such chlorinated organics, are a major cause of MFO induction by pulp mill effluent. Moreover, data presented by Van der Kraak et al. (1992) and Martel and Kovacs (1992) also suggest a causative factor for MFO induction other than chlorinated dioxins and furans or other chlorinated organic chemicals. These studies have shown that enzyme induction and changes in other biomarkers such as increased liver size and decreased gonad size, often used as indicators of adverse effects, are evident in fish populations in waters receiving effluents of pulp mills that do not use chlorine in their bleaching process (Martel and Kovacs, 1992). In fact, the most recent observations of Andersson (1992) show that significant reductions in the use ofchlorine between 1988 and 1990 at the Norrsundet pulp and paper mill on the Baltic Sea did not result in corresponding reductions in the biomarkers of increased liver size, blood lactate, or MFO enzyme activity in exposed fish. Thus, increasing evidence suggests that some thing other than high molecular weight chlorinated organic chemicals may be largely responsible for the effects noted in fish exposed to pulp mill effluents. Possible etiological agents may be natural plant sterols (phytosterols). Phytosterols were implicated as causal factors in the masculinized female mosquitofish observed in a northwest Florida stream receiving pulp and paper mill effluent (type not specified) (Howell et al., 1980). Female mosquitofish possessed male-like gonopodial anal fins and demonstrated elements of male reproductive behavior. Similar observations were made for three fish species (least killifish, sailfin molly, and mosquitofish) in another Florida stream receiving pulp and paper mill effluent (type not specified) (Howell et al., 1980; Bortone and Drysdale, 1981). The acquisition of male secondary sex char acteristics among poecilid female fish was observed and was reproducible in the lab oratory using exposure to microbial-degraded phytosterols. This process is considered analogous to processes occurring in effluents from kraft pulp mills. Although the mas culinized females were still able to produce viable offspring, preliminary laboratory studies demonstrated that their reproductive fitness may be impaired compared to controls (Davis, 1989). Thus, the masculinization of female fish inhabiting areas re ceiving pulp mill effluent appears to have resulted from exposure to natural plant sterols rather than process-related chlorinated organic chemicals. These studies of field populations, particularly of migratory marine mammals, are difficult to interpret. Unlike laboratory experiments in which any factor that may influence the results is controlled, confounding factors abound in field studies making 340 WILLES ET AL. it difficult to distinguish cause-effect relationships. For example, the evidence shows confounding effects of multiple chemicals, interaction with habitat, and various natural disease processes. In the pulp bleaching example, the lack of consistency in observation between chlorine bleaching and adverse effects also confound the interpretation of causal relationships. Therefore interpretation of results based on field studies should be made with caution, bearing in mind that unknown factors may influence the results. Consequently, cause-effect relationships are difficult to demonstrate and the specificity of association may be weak. The three examples presented serve to demonstrate the necessity of using the criteria for assessing causal associations (Hill, 1965; Fox, 1991) between chemicals and reported effects on organisms in the environment. ^ Associations Based on Information from Human Studies Examples of difficulties similar to those outlined above for wildlife also are seen in attempting to establish cause-effect relationships according to the criteria outlined by Hill (1965) and Fox (1991) based on information from human populations. The avail able studies often are difficult to interpret due to confounding factors such as alcohol consumption, tobacco smoking, nonprescription drug consumption, a variety ofdietary differences, and exposures to a mixture of different chemicals. Breast cancer and chlorinated organic chemicals. A causal association between several chlorinated organic chemicals (e.g., PCB, DDT, and chlorinated dioxin/furans) and the incidence of human breast cancer has been proposed (Falck et al., 1992), The associations between breast cancers and breast tissue concentrations ofbioaccumulative chlorinated organic chemicals proposed by Falck et al (1992) does not meet the re quirements for causality as outlined by Hill (1965) and Fox (1991). First, the proposed association does not appear to be biologically plausible based on information from laboratory studies where doses and confounding factors are more readily controlled. Laboratory studies on chlorinated dioxins/furans, PCBs, and DDT have been shown to increase liver tumors but not mammary tumors. The available evidence demonstrates that the increased incidence of liver tumors observed in these animal studies is related to the promotion of preexisting spontaneous lesions (Maslansky and Williams, 1981; Ito et al, 1982; Hayes, 1982; Poland, 1984; Safe, 1989; Silberhom et al, 1990). In fact, decreases in tumors of the mammary gland and various reproductive organs, as well as liver tumors at the lowest dose, were observed in some animal studies (Kociba et al, 1978). The evidence from the experimental animal data argues strongly against the biological plausibility of any proposed relationship between exposure to these chlorinated organic compounds and breast cancer (Kimbrough et al, 1975; Rossi et al, 1977; Kociba et al, 1978; Cabral et al, 1982). Second, the causal association proposed by Falck et al (1992) does not meet the requirement of observational consistency. Several publications have reported increased concentrations of bioaccumulative chlorinated organic chemicals in human tissues (Hoffman et al, 1967; Radomski et al, 1968; Casarett et al, 1968; Unger and Olsen, 1980; Unger et al, 1982, 1984; Teufel et al, 1991); however, t|iere is no consistency among the various studies in the association of the increased tissue concentrations and specific human diseases. Inconsistency or lack of replication in observations be tween studies has been identified as a major factor in detracting from cause-effect relationships based on epidemiological or ecoepidemiological information (Fox, 1991). R&S 143971 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 341 Third, the criteria for evidence of dose-response relationships required to establish causality of association (Hill, 1965; Fox, 1991) are not consistently met in proposing a causal association between human breast cancer and concentrations of persistent chlorinated organic chemicals in breast tissue. The differences in tissue concentrations of PCBs and DDE between women with breast cancer compared to those with benign breast disease were small, within the analytical error of the method, and within the range of concentrations observed in the general population. In addition, a number of factors known to affect breast cancer (e.g., smoking habits, age, diet, and family history of breast cancer) were not controlled in the study. In fact, the women with breast cancer averaged 8 years older than those with benign breast disease. This is an important variable since (i) the incidence of breast cancer is positively correlated with age, and (ii) the concentrations of PCBs and DDE in the tissues of humans are known to increase with age regardless of breast cancer status (Kreiss et ai, 1981). Based on the above analysis, it is evident that the proposed causal association of breast cancer and exposures to bioaccumulative chlorinated organic chemicals should be rejected because of (i) lack of biological plausibility based on no increased mammary tumor incidence in laboratory studies involving PCBs and DDT, (ii) lack of consistency in observation, and (iii) lack of evidence of a dose-response relationship. In addition, the lack of information on several factors known to affect breast cancer incidence (e.g., hormonal imbalance, age of first pregnancy, evidence of miscarriage, length of nursing, age of menarche, and age at menopause) weaken the strength and consistency of association. These factors, combined with the weak nature of most of the reported associations between chemical exposure and breast cancer, lead to the conclusion that the available epidemiological data do not support a causal association between exposure to chlorinated organic chemicals and increased breast cancer incidence. Human development and PCB exposure. The basic criteria forjudging the causality ofassociations (Hill, 1965; Fox, 1991) also are not met by the studies proposing causal associations between human development and exposures to PCBs. Lower birth weights and smaller head circumference have been reported for infants bom to mothers con suming greater amounts of fish from the Great Lakes than control (low fish consuming) populations (Fein et al., 1984). The strength of the proposed association between exposures to PCBs and the observed developmental effects is seriously weakened by several factors. First, the authors suggested that PCBs in the fish were the cause of these effects, although no correlation was observed between fish consumption and umbilical cord blood PCB concentrations (IDSP, 1987; Paneth, 1991). Also, there are a number of inconsistencies in the reported analyses for PCBs in maternal and cord blood sera and in breast milk. It appears that analyses for PCBs were not conducted on serum samples with lipid concentrations less than 200 mg/dl, a decision that could bias the results in an undetermined manner. Mean maternal serum concentrations of Aroclor 1260 were somewhat greater than 9 ppb in the group consuming 52 to 183 meals of fish per year, within the normal range observed in most U.S. populations in the early 1980s. The deficiency in the criteria for a dose-response relationship (e.g., the lack of consistent correlation between cord blood PCB concentrations and fish consumption), a basic requirement for establishing causality (Hill, 1965; Fox, 1991), and the lack of substantial differences between the concentrations of PCBs in the study and the general population indicate that the effects observed on human development I in the populations studied were not causally related to exposures to PCBs. R&S 143972 342 WILLES ET AL. Second, the women from the elevated fish consumption group also reported sig nificantly greater consumptions of alcohol, caffeine, tobacco, and cold remedy med ications during pregnancy than those consuming less fish, and the statistical evaluation ofthese confounding factors was not appropriate (Environment Canada, 1991a; Paneth , 1991). Many of these factors are known to result in lower birth weights of infants and lower developmental scores. Reduced head circumference and weight were not cor related with cord serum PCB levels when considering potential confounding effects of alcohol (a known teratogen), caffeine, and drug use during pregnancy. In addition', there were differences in maternal body weights (the high-fish consumers weighed on average 4.1 kg less) prior to pregnancy, and prepregnancy body weight is one of tfje major factors influencing infant birth weights. Furthermore, of the population con suming greater quantities of fish, the proportion with nonspontaneous deliveries was almost 50% higher than in the low fish-consuming group, and infants bom following nonspontaneous deliveries are more prone to apparent developmental deficits as new borns (Paneth, 1991). Other factors that were not considered in the assessment were the concentrations of other chemicals known to be present in fish from the Great Lakes. All these confounding factors seriously detract from the biological plausibility of an association between the amount of exposure to PCBs, as indicated by blood I PCB concentrations, and effects on human development. Third, the plausibility and consistency of association that are required to establish a causal relationship (Hill, 1965; Fox, 1991) are also seriously weakened by various follow-up studies of the initial pregnancies assessed by Fein et al. (1984). Greater umbilical cord serum concentrations of PCBs were marginally associated with poorer verbal and memory scores on the McCarthy Scales performance tests and with lower scores in the verbal and numerical memory subtests in the same children 4 years later (Jacobson et al., 1990a). Other components of the McCarthy Scales performance tests (perceptual performance, quantitative, motor, and general cognitive index) were un related to in view exposure to PCBs. Since no correlation was observed between fish consumption and umbilical cord concentrations of PCBs, however, no correlation can be made between contaminated fish and scores on the McCarthy Scales perfor mance tests. The arguments for plausibility of association are also weakened by observations that, even though postnatal lactational exposures to PCBs are far greater than those reported in utero (Jacobson et al., 1989), the children exposed to PCBs via lactation for longer periods had significantly higher scores on both the Memory and Verbal Scales tests (Jacobson et al., 1990a). Poorer scores were significantly associated only with the highest maternal milk concentrations of PCBs (1250 to 2600 ng PCB/ml milk). Furthermore, the results of the Jacobson et al. (1990a) study also indicated that in the 4-year-old children higher serum PCB concentrations were not associated with any cognitive deficits. The plausibility of the association proposed by Jacobson et al. (1990a) is further eroded by the observation that certain marginal deficits reported in some clusters of the McCarthy tests were associated with higher serum concentrations of PCBs in the mother, but not with greater exposures to PCBs through lactation. Furthermore, it was not apparent that the scores obtained in the tests of any of the children were outside the ranges of normal since no such ranges were given. Finally, a total of approximately 38 behavioral and neurological tests were conducted on the children, even though the results of only two tests were reported to be affected. Some association based on chance alone would be expected from this large number of tests. R&S 143973 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 343 In addition, the assessment of the test results was based on a "clustering" approach,, which is not a standard procedure in evaluating neurological tests from children. N6 information was available on the effect of the clustering on the interpretation of the results that would be expected from the general population. Therefore, it is not possible to evaluate the importance of this approach in the interpretation of the results of the studies. The plausibility of the association proposed by Fein et al (1984) is further degraded by inconsistencies in the information reported in different publications of the studies. Jacobson et al. (1990b) reported, based on Fein et al. (1984), that prenatal exposure to PCBs through mothers consuming fish was associated with a lower birth weight.,. The same children, assessed 4 years later, had the greatest serum concentration^^ PCBs (mainly as a result of lactational exposure) and were also reported by Jacobson et al. (1990b) to have reduced activity levels. This result, however, is not in accordance with findings of Jacobson et al. (1990a) that the children who had longer lactational exposure had no cognitive deficits when compared to children with shorter lactational exposure. In fact, prior to the controlling of confounding variables, the children with longer lactational exposure tended to score higher on the McCarthy Scales performance tests (Jacobson et al., 1990a). To explain this finding in the first follow-up study, Jacobson et al. (1990a) suggested that the children who were breast-fed longer had the greater intellectual stimulation from their mothers. If this argument is accepted, it is difficult to explain "reduced activity levels" in those 4-year-old children having relatively higher serum PCB concentrations since most of these children were part of the group of infants breast-fed for longer periods of time. In addition, no ranges of normal were included for any of the data sets from either the first or second fol low-up study. Fourth, the criterion of consistency of observation, that is an essential component in the establishment of causality of association based on epidemiological data (Hill, 1965; Fox, 1991), is not met by the studies reported by Fein et al. (1984) and Jacobson (1990a,b). In contrast to the results reported by these researchers, exposure to PCB (as measured by maternal serum PCB concentrations) was correlated with increased birth weight, not decreased birth weight, in a random sample of 100 participants of a 1112 women study in the Green Bay area of Wisconsin (Dar et al., 1992) (near the same area where numerous effects were observed in wildlife as discussed in the previous section). Although the concentrations of PCBs reported in the maternal serum in the Dar et al. (1992) study would appear to be lower than those observed by Fein et al. (1984), the analyses conducted by Dar et al. (1992) were congener specific rather than for total concentrations of PCBs. Greater birth weights, as opposed to lesser birth weights, as reported by Fein et al. (1984) and Jacobson et al (1990b), have also been reported in other fish-consuming populations (Olsen et al, 1990). Similarly, Fitzgerald et al. (1992) observed that the concentrations of PCBs in breast milk were the same (although the congener profiles were different) between 53 Mohawk women consuming fish from the St. Lawrence River and a population of 109 women from Warren Scho harie County in New York. The Mohawk women were expected to have higher PCB concentration in their breast milk as they were determined to consume consider ably more fish containing PCB prior to pregnancy (more than 1 year) than those women residing in Warren Schoharie County. These observations demonstrate a lack of consistency between fish consumption, concentrations of PCBs in breast R&S 143974 344 WILLES ET AL. milk and serum, and reproductive outcomes compared to the report by Fem e al. (1984). Studies designed to evaluate possible causal associations between human develop ment and exposures to PCBs, DDT, and DDE using a cohort of children from Nortl Carolina were conducted by Rogan et al. (1986a,b). Prenatal exposures of infants tc PCBs, DDT, and DDE were predicted from extrapolations across the nursing periot based on breast milk analysis of one sample at one stage of nursing. Calculated daib chemical exposures of the infants were based on these predicted concentrations of th< chemicals in breast milk, and no validation of these predictions was reported. Th< neurological development of the infants was then correlated with breast milk fart* estimate a noobservable-adverse-effect level (NOAEL) of3.4 and 1.0 ppm, respectively for Brazelton and Baileys scores (Tilson et al., 1990). The Brazelton scores were baset on 27 behavioral tests and 20 reflex tests that were summarized into clusters (Rogai et al., 1986b) following the procedures outlined by Jacobson et al. (1984). Only th. cluster scores for tonicity and reflexes were correlated with exposures to PCBs or DDE Although effects on certain of the various behavioral parameters were correlated witl estimated PCB exposures, the plausibility of the results, and therefore the strength c the association, is weakened by the fact that the women with higher milk PCB con centrations were older, had a greater rate of alcohol consumption, and greater tobaco. smoking incidence. All of these factors are known to affect scores on a wide range c neurological testing regimes for newborn infants. In addition, the effects were observe^ on a small number of infants (49 infants of a total of 856 were exposed to mil! concentrations of PCBs above 3.4 ppm, which is reported as the NOAEL, and onl 42% of these, or 21/856, showed positive tests). The small sample size of infant exposed to PCB concentrations in breast milk fat in excess of 3.4 ppm, and who scorei significantly lower on the behavioral tests, precludes the determination of a NOAEL with a high degree of confidence, for effects on behavioral testing parameters. As result of the small sample size of infants exposed to PCB concentrations of greate than 3.4 ppm in breast milk fat, and due to the uncertainties introduced into th analysis arising from the presence of confounding variables, there is little basi for establishing a causal association between the observed effects and exposure to PCBs. Based on the above analysis, and considering the marginal significance of the ot servations (most at or just below P < 0.05), the information reported by Fein et a (1984), Rogan et al. (1986a,b), and Jacobson et al. (1990a,b) does not meet the criteri as outlined by Hill (1965) and Fox (1991) for the establishment of a causal associatio for an effect of PCBs on growth and behavior in human populations. CONCLUSIONS REGARDING THE APPLICATION OF SCIENTIFIC PRINCIPLES In this brief review, certain scientific principles have been presented that must t considered in any critical evaluation ofliterature reports ofadverse effects of chlorinate organic chemicals. These key principles represent essential criteria by which repon should be examined, either to enable validation of the findings based on scientif principles or to show that the findings do not withstand objective scientific scrutin' These scientific principles, as they apply specifically to the potential effects of chlor nated organic chemicals in the environment, are: R&S 143975 CHLORINATED ORGANIC CHEMICALS IN THE ENVIRONMENT 345 (1) the fate and biological activity of a compound are determined by the chemical properties of the compound; (2) compounds do not show adverse effects below certain threshold concentrations, and the magnitude of response is related to dose; (3) inherent metabolic processes allow organisms to accommodate low doses of chlorinated organic chemicals; (4) observations associated with the presence of certain compounds must be bio logically plausible effects, based on the specificity of the compound's activity in ex perimental systems. This review has presented a critical evaluation of evidence relating specificallyHo chlorinated organic compounds by using these scientific principles. The evidence shows that different types of chlorinated organic compounds have very different types of environmental behavior and biological activity, depending upon their physical/chem ical properties. These differences suggest logical ways for controlling exposure to prevent or minimize the potential for adverse effects. For example, chemicals with a greater degree of electrophilicity, such as vinyl chloride monomer, will not bioaccumulate in environmental species, but exposed workers were historically at risk (Doll, 1988). This emphasizes the importance of a containment procedure that prevents exposure of workers and the general public to the unreacted compound. On the other hand, the high lipophilicity and only gradual breakdown of certain highly chlorinated compounds point out the potential for their bioaccumulation in food chains. Problems that have been observed in certain wildlife species, notably piscivorous birds, can be associated with elevated concentrations of certain of these bioaccumulative compounds. Such unwanted effects indicate that, in the past, environmental releases of compounds with these properties were too great. Furthermore, environmental recoveries have been noted to be associated with lower environmental concentrations of such chemi cals, except in localized hot spots. This dose-response evidence indicates a practical environmental threshold, even for highly bioaccumulative compounds. Scien tists now know the physical/chemical properties of chemicals that cause such un desired characteristics, and the future production of such chemicals can now be avoided. These types of scientific evidence should govern industrial design and regulatory decision making, with the goal of preventing potential adverse effects on the environ ment. For example, all factors that influence the rate of environmental loading of bioaccumulative compounds should be considered. Mitigating practices that would have an impact on environmental releases ofchemicals include recovery and recycling to minimize the need for disposal, with the best available technology to be used for treating waste that contains compounds that could accumulate in the environment. The physical and chemical properties determine what practices would be required to ensure that environmental release rates of specific chemicals are maintained below those that would result in adverse effects. This information, coupled with scientific evidence on dose-response relationships, provides the guidance necessary to establish no-effect levels in the receiving environment that in turn will determine the rigor of waste reduction and waste treatment practices. It is also essential to note that alternatives to any chemical product or process must be evaluated with the same scientific principles to predict potential impact on human and environmental health. The chemical characteristics of chlorobenzyltoluenes, as R&S 143976 346 WILLES ET AL. substitutes for PCBs, for example, indicate that they may have similar biological char acteristics (Mark ei al., 1990) and environmental fate (Wester and van der Valk, 1990) as PCBs. Thus, their potential quantity and rate of release into the environment should be consistent with levels that would not lead to adverse effects through their bioac cumulation. In addition, although DDT has been associated with adverse effects in wildlife species, its acute toxicity in humans and other nontarget mammals is much lower than many of the compounds introduced to take its place. Accidental.poisonings and deaths from lindane following its increased use as a replacement for DDT, and from organophosphorus insecticides that more generally have replaced the organochlorine insecticides, frequently have been observed in humans and wildlife. Scientific evaluation of alternatives may lead to important decisions not only about the overall effect of the alternatives, but also about use patterns of the original product. For example, the use of DDT may be much more important in public health for the control of malaria than its past use in agriculture. In conclusion, chlorinated organic compounds represent diverse groups ofchemicals, a limited number of which can bioaccumulate in the environment. Depending on environmental release rates, concentrations of these bioaccumulative chemicals could result in exposures sufficiently great as to be associated with adverse effects in wildlife. These types of chemicals require careful consideration and control. However, evalu ation according to the guiding scientific principles indicates that the vast majority of chemicals produced from chlorine chemistry do not have these chemical characteristics. Rather, they are subject to degradation in the environment and there is no reason to speculate that they would accumulate to concentrations that would pose a threat to ecosystem health. The analysis of dose-response relationships demonstrates that ex posures from the environmental concentrations of nonbioaccumulative chlorinated organic chemicals are far below those that would be associated with adverse effects based on laboratory studies. There is evidence that all major classes of chlorinated organic chemicals produced by human activities also are produced by natural processes. 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