6
Chemical Health Risk Assessment–Critique of Existing Practices and Suggestions for Improvement

ABSTRACT

This chapter and the previous one should be considered as a unit. A fourfold classification system of the mechanisms underlying adverse health effects is outlined below, which forms the basis for developing quantitative risk assessment approaches for both cancer and noncancer effects.

A detailed critique is then provided of existing Food and Drug Administration (FDA) risk assessments for polychlorinated biphenyls (PCBs) and methylmercury–representing the two most extensively documented examples of analyses underlying current regulatory levels for a carcinogen and a noncarcinogen in seafood, respectively. (In both cases, the committee finds considerable opportunity for improvement.)

The difficult issue of determining human intakes for a broad (though far from comprehensive) range of chemical contaminants in seafood is subsequently addressed. Estimates are made of national average daily intakes of various inorganic and organic contaminants via commercially marketed seafood, and for several organic carcinogens, upper-confidence-limit estimates of possible cancer risk are made. However, aside from the methylmercury example,1 in the absence of better information on the population distribution of the dosage of contaminants to the U.S. population, it is impossible to make even tentative quantitative estimates of potentially significant noncancer risks. Of additional serious concern are the appreciable quantities of seafood consumed following noncommercial sport and subsistence tribal fishing.

Finally, an overview of opportunities for research on different categories of potential health impacts is presented, and conclusions are drawn from both this and the previous chapter. The principal conclusions are the following:

  • From both natural and human sources, a small proportion of seafood is contaminated with appreciable concentrations of potentially hazardous organic and inorganic chemicals. Some of the risks that may be significant include reproductive effects from PCBs and methylmercury, and carcinogenesis from selected PCB congeners, dioxins, and some chlorinated hydrocarbon pesticides.

  • Consumption of some types of contaminated seafood poses enough risk that efforts toward evaluation, education, and control of that risk must be improved.

  • Present quantitative risk assessment procedures used by government agencies can and should be improved and extended to noncancer effects.

  • Current monitoring and surveillance programs provide an inadequate representation of the presence of contaminants in edible portions of domestic and imported seafood, resulting in serious difficulties in assessing both risks and specific opportunities for control.



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Seafood Safety 6 Chemical Health Risk Assessment–Critique of Existing Practices and Suggestions for Improvement ABSTRACT This chapter and the previous one should be considered as a unit. A fourfold classification system of the mechanisms underlying adverse health effects is outlined below, which forms the basis for developing quantitative risk assessment approaches for both cancer and noncancer effects. A detailed critique is then provided of existing Food and Drug Administration (FDA) risk assessments for polychlorinated biphenyls (PCBs) and methylmercury–representing the two most extensively documented examples of analyses underlying current regulatory levels for a carcinogen and a noncarcinogen in seafood, respectively. (In both cases, the committee finds considerable opportunity for improvement.) The difficult issue of determining human intakes for a broad (though far from comprehensive) range of chemical contaminants in seafood is subsequently addressed. Estimates are made of national average daily intakes of various inorganic and organic contaminants via commercially marketed seafood, and for several organic carcinogens, upper-confidence-limit estimates of possible cancer risk are made. However, aside from the methylmercury example,1 in the absence of better information on the population distribution of the dosage of contaminants to the U.S. population, it is impossible to make even tentative quantitative estimates of potentially significant noncancer risks. Of additional serious concern are the appreciable quantities of seafood consumed following noncommercial sport and subsistence tribal fishing. Finally, an overview of opportunities for research on different categories of potential health impacts is presented, and conclusions are drawn from both this and the previous chapter. The principal conclusions are the following: From both natural and human sources, a small proportion of seafood is contaminated with appreciable concentrations of potentially hazardous organic and inorganic chemicals. Some of the risks that may be significant include reproductive effects from PCBs and methylmercury, and carcinogenesis from selected PCB congeners, dioxins, and some chlorinated hydrocarbon pesticides. Consumption of some types of contaminated seafood poses enough risk that efforts toward evaluation, education, and control of that risk must be improved. Present quantitative risk assessment procedures used by government agencies can and should be improved and extended to noncancer effects. Current monitoring and surveillance programs provide an inadequate representation of the presence of contaminants in edible portions of domestic and imported seafood, resulting in serious difficulties in assessing both risks and specific opportunities for control.

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Seafood Safety Because of the unevenness of contamination among species and geographic areas, it is feasible to narrowly target control efforts and still achieve meaningful reductions in exposures. The data base for evaluating the safety of certain chemicals that find their way into seafood via aquaculture and processing is too weak to support a conclusion that these products are being effectively controlled. The principal recommendations of the committee are as follows: Existing regulations to minimize chemical and biological contamination of the aquatic environment should be strengthened and enforced. Existing FDA and state regulations should be strengthened and enforced to reduce the human consumption of aquatic organisms with relatively high contaminant levels (e.g., certain species from the Great Lakes with high levels of PCBs, swordfish and other species with high methylmercury levels). Federal agencies should actively support further research to determine the actual risks from the consumption of contaminants associated with seafood and to develop specific approaches for decreasing these risks. Increased environmental monitoring should be initiated at the state level, as part of an overall federal exposure management system. States should continue to be responsible for site closures, and for issuing health and contamination advisories tailored to the specific consumption habits, reproductive or other special risks, and information sources of specific groups of consumers. There should be an expanded program of public education on specific chemical contaminant hazards via governmental agencies and the health professions. INTRODUCTION Part of the committee's charge was to review and summarize "the current status of regulations, guidelines, and advisory statements issued by Federal and State public health authorities on environmental contaminants in seafood." Its review was to specifically address contaminants defined by Food and Drug Administration (FDA) regulations as "avoidable or unavoidable." Then, based on this, the committee was asked to "assess how well the current regulatory framework protects the public health." The committee was also charged with the task of reviewing and summarizing, specifically, the health risk assessment procedures used by FDA, the Environmental Protection Agency (EPA), and other regulatory authorities for priority environmental pollutants, including toxic metals and synthetic organic chemicals. In addition, the committee was asked to "recommend future research directions, as appropriate." To set the stage for an examination of how current risk assessment procedures can be improved, the basic concepts underlying the mechanisms of action of toxic substances are articulated in the following section, along with quantitative ideas about dose-time-response relationships. Then an extensive critique of agency risk assessments for PCBs and methylmercury is provided. Finally, the committee addresses issues of exposure assessment and risks from other substances, and opportunities for further research on potential chemical health hazards.

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Seafood Safety BROAD CATEGORIZATION OF MECHANISMS OF DIFFERENT ADVERSE EFFECTS AND IMPLICATIONS FOR DOSE-RESPONSE RELATIONSHIPS The concern from which much of our regulatory history has resulted is the potential carcinogenic effect of some contaminants. To a certain extent, this concern is based on the mutagenic mechanisms of cancer. For carcinogens that act by primary genetic mechanisms, there are good theoretical reasons to believe that at the limit of low dosage, the risk will be a linear function of exposure (Ehrenberg et al., 1983; Hattis, 1990a). However, as more is learned about the mechanisms of some other types of toxic effects–particularly reproductive effects and chronic degenerative neurological conditions (NRC, 1989, 1990)–concern about potential low-dose effects of other types has tended to increase. It is therefore important to clarify what the expectations should be for dose-response relationships from first principles, given the full range of causal processes that can lead to impairment of health. Table 6-1 shows a categorization system for biological damage mechanisms that can be helpful in guiding basic choices in risk assessment modeling (Hattis, 1982, 1986). The system is intended to distinguish between different ways of looking at the likely mechanisms of disease causation that are encouraged by different groups of scientific disciplines.2 The focus of the scheme in Table 6-1 is to sort adverse effects according to the kinds of events that are likely to be occurring at either (1) subclinical dosage levels (doses that do not produce unusual function) or (2) preclinical stages in the development of the pathological process (i.e., the time before an overt manifestation of a latent disease, such as cancer, occurs). Under these conditions, one first asks Are the events occurring ordinarily fully reversible (or very nearly so), given a prolonged period with no further exposure to the hazard? TABLE 6-1 Types of Health Hazards Requiring Fundamentally Different Risk Assessment Approaches 1. "Traditional" toxicity resulting from overwhelming body compensatory processes: below some threshold, in individuals who are not already beyond the limits of normal function without exposure, response is reversible. • Traditional acute toxicity–Toxic action is completely reversible or proceeds to long-term damage within about three days of exposure (paralytic shellfish poisoning, puffer fish poisoning; probably many teratogenic effects). • Traditional chronic toxicity—Toxic process typically proceeds to permanent damage over a period of several days to several months, due to either (1) reversible accumulation of a toxic agent (e.g., methylmercury, lead) or (2) accumulation of a slowly reversible toxic response (e.g., cholinesterase inhibition). 2. Effects resulting from insidious processes that are irreversible or poorly reversible at low doses or early stages of causation. • Molecular biological (stochastic process) effects—Effects occur as a result of one or a small number of irreversible changes in information coded in DNA: mutagenesis, most carcinogenesis, and some teratogenesis. • Chronic cumulative effects–Effects occur as a result of a chronic accumulation of many small-scale damage events: emphysema, noise-induced hearing loss, atherosclerosis, and probably hypertension; possibly depletion of mature oocytes.

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Seafood Safety If the answer to the first question is yes, then it will generally be appropriate to treat the condition within the framework of traditional toxicology.3 Some examples of such reversible changes are the following: Buildup of a contaminant in blood or other tissues. It is rare for there to be a zero rate of excretion of any material. Given time and no further exposure, toxicant buildup should be reversible, although it can be quite prolonged. [Current estimates are that only about 9% of more persistent polychlorinated biphenyl (PCB) isomers may be metabolized or excreted per year in humans (Yakushiji et al., 1984).] Most enzyme inhibition (generally, even irreversible inactivation of enzyme molecules can be "reversed" through the synthesis of replacement molecules) Induction of short-term biological responses that act to maintain homeostasis (e.g., sweating in response to heat, tearing in response to eye irritation) If the answer to the above question is no and events are likely to be occurring at subclinical exposure levels or preclinical stages that are not ordinarily reversible, the modeling of biological risks will have to be based on concepts that are fundamentally different from the homeostatic system/threshold paradigm. Examples of such irreversible or poorly reversible events include changes in genetic information or the heritable pattern of gene expression after these are effectively "fixed" into a cell's genome expression by replication; death of nonreplicating types of cells (adult neurons); destruction of nonregenerating structures (alveolar septa); and generation and buildup of incompletely repaired lesions (atherosclerotic plaques). Appropriate modeling for conditions that are the result of irreversible or poorly reversible processes must be based fundamentally on the likely dose-response characteristics of the events that cause the basic irreversible changes. Once the primacy of such changes is established for a particular event, the analyst should then ask whether clinical manifestations are likely to be the direct result of only a few, or very many, individual irreversible damage events. If only a few events are believed to contribute directly to a particular clinical manifestation (e.g., a small number of heritable changes within a singe cell line leading to cancer), the effect can be considered a "molecular biological" disease. The risk assessment models used must follow from an understanding of the stochastic nature of the basic process. On the other hand, if thousands, millions, or billions of individual irreversible events directly contribute to a particular condition (e.g., very large numbers of individual neurons must die to cause the clinical manifestations of Alzheimer's or Parkinson's disease), the biological harm should be dealt with under the novel category of chronic cumulative conditions (see below).

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Seafood Safety Traditional Acute Toxicity Three kinds of insights for acute toxicity risk assessment follow naturally from the homeostatic system paradigm of physiology and traditional toxicology: There will be a series of toxic effects as different compensatory processes are overwhelmed and as impairment broadens from more-to less-sensitive cells and functions. For each effect in each individual who is not already beyond the limits of normal functioning in the absence of exposure, there will be some subthreshold level of exposure that will be insufficient to produce the effect. Individuals will differ in their thresholds. A caveat to the general expectation of individual thresholds is that some tasks may so tax the capabilities of a system (perhaps, during fetal life, the struggle to mobilize metabolic resources to grow and differentiate as fast as possible so as to cope with the external world at birth) that any impairment of a key limiting functional parameter required for the task could compromise function to some degree. (This would also apply to reaction time for a driving task, for example.) Of particular relevance to the committee's task in this regard is the suggestion of some studies that dietary PCB exposure may be associated with either changes in birth weight (Fein et al., 1984; Sunahara et al., 1987; Taylor et al., 1989) or indices of neurological function in infants (Jacobson et al., 1985; Rogan et al., 1986). The first job in assessing acute toxic effects is to define the series of acute responses to the disturbing influence in question. Ideally, the analysis should then attempt to determine (to whatever degree of precision is possible) the nature and magnitude of the dosage and the disturbance of physiological parameter(s) that are necessary to cause each type of acute toxic response, along with the frequency of each response in a diverse human population.4 Such mechanism-based analysis is, however, not common in the field. Rather, the current state of the art in those rare cases where acute toxic effects are treated quantitatively is to use probit equations (Finney, 1971) of the general form Prohibit of response = a + b ln (Cn T), where a, b, and n are constants, C is external concentration, and T is exposure time; n represents the basic trade-off between intensity and duration of exposure, and b defines the breadth of an assumed lognormal distribution of threshold responses. Although some of the available animal data on irritant gases appear to be well summarized by equations of this form (Appelman et al., 1982), this is basically an empirical formula that does not incorporate quantitative representations of the various processes underlying toxicity. It is therefore difficult to decide what adaptations should be made in applying the empirical relationships to diverse subsets of humans. Even more common, unfortunately, is the simple use of the no-effect level (NOEL)/"safety factor" analysis for arriving at acceptable daily intake (ADI) levels for chemical contaminants. Rather than estimate the numbers of people with specific degrees of particular effects, the general approach is to arrive at an ADI by a rule-of-thumb

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Seafood Safety procedure derived from observed NOELs [or, more recently, no-adverse effect levels (NOAEL), after often contentious discussion over what effects are considered "adverse"], or sometimes low effect levels (LOELs) in animal experiments or human studies. When projections are made from animal data, generally a 100-fold "uncertainty factor" is allowed between the NOEL and the ADI (Dourson and Starra, 1983). The 100-fold factor is often decomposed as 10-fold to account for possible differences in sensitivity between humans and the most sensitive species tested and 10-fold to account for possible interindividual differences in susceptibility among humans. This approach has a few advantages: It is "quick and dirty"—relatively straightforward to apply and does not require complicated model building or analysis. Through thousands of applications in the past, it is not yet known to have led to catastrophic adverse effects in humans (using Ozonoff's working definition of a "catastrophe" as an effect so large that even an epidemiological study can detect it (D. Ozonoff, Boston University School of Public Health, personal communication, 1990). On the other hand, for the long term, the simple uncertainty factor approach has a number of disadvantages: No one knows how protective it really is, either in general or in specific cases. What fraction of the diverse human population can be expected to experience adverse effects when exposed at the level calculated to be "acceptable" under the formula? (In general, there may be some finite fraction of individuals who, because of disease or other reasons, are marginal for biological functions affected by the chemical and who may be pushed beyond a functional threshold for an adverse effect by a small finite dose of the chemical.)5 The procedure incorporates one specific social policy standard for setting "acceptable" levels without making clear where technical analysis leaves off and policy/value analysis begins. Effects are generally scored as either present (operationally, statistically significant) or not present (not statistically significant) at a particular dose. There is usually no quantitative analysis of the effects of sample size or the dose-response relationship for the effect in question. There is no defined or obvious way to incorporate newer types of relevant data on human interindividual differences in rates of uptake/absorption for a constant environmental exposure ("exposure variability"); rates of activating or detoxifying metabolism and excretion, producing differences in the concentration x time of active metabolites per unit of absorbed dose at the site of toxic action ("pharmacokinetic variability"); and differential risk of response ("response variability") for a given concentration x time of active metabolites at the site of toxic action. In particular, the committee suspects that the inability of the uncertainty factor paradigm as usually formulated to incorporate newer types of relevant information into

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Seafood Safety a systematic procedure for updating assessments of health hazards has tended to discourage both the collection and the analysis of potentially important data. One example of this is information on human interindividual variability in parameters that could affect susceptibility. Table 6-2 outlines some idealized components of a full quantitative analysis of a noncancer health effect that is mediated by what is called a "functional intermediate" parameter. Such a parameter is generally a continuous variable that has a strong causal influence on performance of an important biological function (although it will not generally be the sole determinant of performance). It should, in turn, be affected by the toxin/exposure under study, and it should be reasonably likely that effects on the final health condition of concern are primarily mediated through effects on this functional intermediate parameter. For example, a key functional intermediate for some reproductive effects of PCBs may well be changes in birth weights (Fein et al., 1984; Jacobson et al., 1985; Sunahara et al., 1987; Taylor et al., 1989). Similarly, blood or tissue concentrations of lead constitute a useful intermediate parameter for lead toxicity. TABLE 6-2 Elements of a New Analysis for Noncancer Health Effects Mediated by a "Functional Intermediate" Parameter 1. Elucidate the quantitative relationships between internal dose/time of toxin exposure and changes in the functional intermediate parameter. 2. Assess the preexisting "background" distribution of the functional intermediate parameter in the human population. 3. Assess the relationship between the functional intermediate parameter and diminished physiological performance or adverse health effects. 4. Assess the magnitude of parameter changes likely to result from specific exposures in humans (taking into account human interindividual variability in metabolism and other determinants of pharmacokinetics) and consequent changes in the incidence and severity of health effects. 5. Do not attempt, from the biology alone, to determine "acceptable" levels of parameter change or exposure. (Let policymakers decide what changes in the incidence and severity of health effects are "acceptable" in the context of modes of exposure and in light of the feasibility of reducing or avoiding exposure.) The illustrative calculations in Table 6-3 (from Ballew and Hattis, 1989) show how modest changes in the population distribution of a key parameter such as birth weight can be reflected in serious changes in the outcome of infant mortality. It can be seen that birth weights are very strongly related to infant mortality and that the relationship is continuous. Although very low birth weight infants are at dramatically higher risk than infants in the normal weight range, even infants weighing about 3,000 grams (g) can be expected to have their risks increased somewhat by an agent that causes a marginal change in birth weight. As indicated in the table, because there are many more infants in the 2,500-3,500-g weight range, the expected population aggregate mortality increase is as large for these categories as the population aggregate mortality increase for infants in the very low birth weight range (500-1,500 g).

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Seafood Safety TABLE 6-3 Expected Infant Mortality Effects of a 1% (33.66-g) Reduction in Birth Weight Weight Range (g) Fraction of Births Mortality Risk per 1,000 Births in Category Fraction of Births x Mortality Risks/1,000 Original Population After 1% Weight Reduction Without Birth Weight Reduction After 1% Birth Weight Reduction Net Change White Infants Under 500 0.0006912 0.0007701 1,000 0.6912 0.7701 0.0789 500-999 0.002171 0.002335 673.31 1.4612 1.5720 0.1107 1,000-1,499 0.005249 0.005488 237.85 1.2485 1.3053 0.0568 1,500-1,999 0.009182 0.009575 76.86 0.7057 0.7360 0.0302 2,000-2,499 0.029192 0.032804 26.746 0.7808 0.8774 0.0966 2,500-2,999 0.15164 0.16568 8.3565 1.2672 1.3845 0.1174 3,000-3,499 0.36237 0.37081 4.2566 1.5424 1.5784 0.0359 3,500-3,999 0.31749 0.30337 3.0451 0.9668 0.9238 0.0430 4,000-4,499 0.10100 0.09027 3.0293 0.3060 0.2734 0.0325 4,500+ 0.021021 0.01890 4.941 0.1039 0.0934 0.0105 Total 1 1   9.0736 9.5142 0.4406 Black Infants Under 500 0.0026095 0.0028661 1,000 2.6095 2.8661 0.2566 500-999 0.006279 0.006666 645.90 4.0558 4.3058 0.2500 1,000-1,499 0.012709 0.013154 167.98 2.1348 2.2096 0.0748 1,500-1,999 0.020673 0.02165 57.72 1.1932 1.2495 0.0563 2,000-2,499 0.067052 0.074351 21.482 1.4404 1.5972 0.1568 2,500-2,999 0.24894 0.26444 9.832 2.4476 2.6000 0.1524 3,000-3,499 0.38248 0.37936 6.636 2.5381 2.5174 0.0207 3,500-3,999 0.20683 0.19100 5.581 1.1543 1.0660 0.0883 4,000-4,499 0.04418 0.04029 5.89 0.2602 0.2373 0.0229 4,500+ 0.008253 0.006226 12.33 0.1018 0.0768 0.0250 Total 1 1   17.9358 18.7257 0.7899 In principle, the use of such intermediate parameters can provide windows on the pathological processes that occur earlier in the development of toxicity, are more sensitive to the action of potential toxicants (compared with attempts to observe actual cases of illness), and are more accessible to direct comparative measurement in both animal models and humans. It is desirable, for these purposes, that the intermediate parameters chosen be as close as possible to the actual causal pathway leading to harm. However, even a parameter such as birth weight, which may not itself bear a direct causal relation to infant deaths, may be a close enough indicator of the actual causal processes to serve as a useful intermediate predictor. Because there will generally be a series of steps in the sequence between toxin uptake and ultimate manifestation of adverse effects, the analyst may often have choices of which parameter(s) to use for assessing human risk. These choices will generally be based on the availability of measurement techniques and theory for observing or estimating the parameter(s) in question.

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Seafood Safety Traditional Chronic Toxicity The basic principles that govern the analysis of acute toxic effects are by and large directly applicable to cases of chronic toxicity. Chronic toxic analyses tend to differ from analyses of acute toxic effects primarily in that considerable emphasis must be placed on the "slow step" of the process, which causes the effect to be chronic rather than acute. This slow step is generally either a long-term accumulation of a toxic agent that is poorly excreted under ordinary conditions (e.g., lead, mercury) or an accumulation of some slowly reversed residual effect (e.g., acrylamide). With lead as an example, the following statements can be made: Appreciable information is available on the pharmacokinetics of lead absorption, transport, storage, and excretion (Barry, 1975; Bernard, 1977; Campbell et al., 1984; Chamberlain, 1985; Marcus, 1985a-c; Rabinowitz et al., 1976); and much better information could be obtained with the aid of natural experiments such as strikes among lead-exposed workers, which can given information on the rate of decrease of blood lead levels after a reduction in exposure (Hattis, 1981). Inhibition of heme synthesis enzymes at essentially all dose levels is well characterized (Haeger-Aronsen et al., 1974), and the inhibition of heme synthesis may be important in producing some of the neurotoxic effects of lead (Silbergeld et al., 1982), although the short- and long-term functional significance of different degrees of inhibition in different individuals is far from clear. Effects on some measures of neurological function and kidney function are susceptible to study in reasonably straightforward ways. Effects on higher-order development of central nervous system functions are more difficult to determine because of an ignorance of basic mechanisms; however, some good studies have become available in recent years (Baker et al., 1983; Bellinger et al., 1987, 1990; HHS, 1988; Needleman et al., 1979, 1990; Waternaux et al., 1989). The impairment of very complex neurological functions by lead raises a significant issue in the application of the traditional toxicological paradigm to risk analysis. As indicated above, the usual assumption is that there is some functional reserve capacity in "normal" individuals that maintains "adequate" performance despite a "small" degree of perturbation of a biological parameter by a "low" dose of toxic material. However, if the function is already taxed to its limit in certain situations, even in the absence of exposure (perhaps for a first grader learning to read or for a developing fetus mobilizing all its available metabolic energy to grow and differentiate), and if the biological parameter being perturbed is limiting to the performance of that function, then any level of exposure may produce at least some reduction in performance. Addressing the issues of the population distribution of different functional reserve capacities, and the relationship of functional reserve capacities to specific biochemical parameters, is essential to the future research needs of risk assessment for classical chronic toxic agents. Also in the area of neurotoxicology, Silbergeld (1982) has written of the potential of new radioimmunoassay and functional measurement techniques to help shift the focus of research away from traditional morphological criteria of neurological damage toward more sensitive and sophisticated measures of performance.

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Seafood Safety Molecular Biological (Stochastic Process) Diseases In addition to most carcinogenesis, molecular biological diseases include mutagenesis and at least some teratogenesis. The subject of quantitative risk assessment for carcinogenic hazards has been discussed extensively elsewhere (American Industrial Health Council, 1987; Bishop, 1987; Crump et al., 1976, 1977; EPA, 1986a; Hattis, 1982; Hattis and Smith, 1986; Moolgavkar, 1986; Rai and Ryzin, 1981; Whittemore, 1980). However, it is worth briefly recapitulating some basic features of the carcinogenic process and their implications for cancer dose-response relationships. Science is now much closer than it was a decade ago to understanding the fundamental mechanisms involved in carcinogenic transformation. For some time it has been clear that tumors arise as a result of a series of changes or rearrangements of information coded in DNA within single cells (Cleaver and Bootsma, 1975; Fialkow, 1977; Hattis, 1982; Knudson, 1973, 1977; McCann et al., 1975; Vogel and Motulsky, 1979). These changes are often induced by electrophilic metabolites of the parent compounds to which organisms are exposed (Miller and Miller, 1981). With the identification of "oncogenes," some detailed molecular characterization is being provided of the changes resulting in DNA (Fischinger and DeVita, 1984; Hoel, 1985; Modali and Yang, 1984; Yunis, 1983). It has also been apparent for some time that further headway cannot be made in elucidating the shapes of carcinogenesis dose-response relationships at low dosages simply by increasing the numbers of animals studied in conventional bioassays. A variety of mathematical models with dramatically different consequences for low-dose risk can always be found that fit the observations about equally well (Maugh, 1978; Whittemore, 1983). Low-dose risk projections are, therefore, inevitably much more determined by the choice of model than by the available data (Guess et al., 1977; Whittemore, 1980), if what is meant by "data" is restricted to observations of the incidence of ultimate adverse effects in small groups of animals. Because of sample size limitations, animal carcinogenesis bioassays must be done within a limited range of relatively high dose levels. Typically, the difference between the minimum detectable response and a response that effectively saturates the system or causes interference through overt toxicity is only one to two orders of magnitude (often even less). Over this high dose range near levels where the agent produces overt toxic effects, enzyme saturation and other forms of pharmacokinetic nonlinearities are most likely. If in dose-response modeling for risk assessment, the nonlinearities of pharmacokinetic origin are not separated from the nonlinearities that may arise from the multiple mutation mechanism that is central to carcinogenesis, our ordinary curve-fitting procedures will implicitly attribute the pharmacokinetic nonlinearities to the fundamental carcinogenic process (Hoel et al., 1983). The resulting errors are particularly serious if one wishes to produce the best point estimates of carcinogenic risk in addition to upper confidence limits. Clearly, to make real progress in modeling carcinogenic risks, knowledge of the fundamental processes involved must be used to break open the black box between external exposure levels and ultimate production of tumors. The use of pharmacokinetic models and intermediate parameters ("markers") to characterize the dose-response characteristics of small segments of the causal pathway to carcinogenesis has considerable potential to improve dose-response modeling for the process as a

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Seafood Safety whole (Hattis, 1988). Such markers can include both those that may lie directly along the causal pathway, such as DNA adduct formation, and putative correlates, such as hemoglobin adduct formation, that can be good indicators of the concentration-time product of active intermediates in the systemic circulation. One key fact must be recognized from the beginning about pharmacokinetic modeling, however. Whatever nonlinearities may be produced at high doses by the saturation of enzymes, the saturation of active transport processes, the depletion of cellular reactants for electrophilic agents, or changes in cell division rates to make up for cell killing due to overt toxicity, all of these nonlinearities must necessarily disappear as one approaches very low dose rates (Hattis, 1990a). The slope of the line relating ultimate DNA lesions in replicating cells to external dose may well be very different at low than at high doses, but it must be linear. The basic reason for this is that at low doses the rates of the transport and transformation processes that lead to DNA damage and repair depend directly on the number of collisions between molecules of an "input" chemical (or activated intermediate or DNA adduct) and a resident cellular reactant (or hole in a membrane or repair enzyme molecule). At low doses the number of resident cellular reactant molecules does not change appreciably as a function of the concentration of the input. Therefore, the number of relevant collisions and the rates of reactions and side reactions in the causal sequence at low dosage must be direct linear functions of the amounts of input chemical and its activated derivatives. Some finite fraction of the ultimate DNA lesions must escape repair before the next cell replication as long as the cells affected have a nonzero turnover rate, there are a finite number of repair enzyme molecules, and the repair molecules operate at a finite rate. All carcinogens – in particular the PCBs and dioxins – are not thought to act primarily by causing DNA mutations (Safe, 1989). Table 6-4 lists a variety of other types of mechanisms whereby chemicals can affect carcinogenesis. There has been a tendency in some quarters to assume that if a chemical does not act via a primary genetic mechanism, one should revert to the traditional toxicological paradigm for analysis, including all of the old presumptions about thresholds and safety factors (Weisburger and Williams, 1983). As Rodericks (1989) has noted, There is disagreement about how to estimate risks from carcinogens. In the United States, regulatory agencies generally estimate risks in the same way for both genotoxic and non-genotoxic carcinogens. Regulatory agencies adopt this position because they believe that full knowledge of the mechanism of action of non-genotoxic agents is needed before they can be assumed to exhibit thresholds. In several foreign countries, non-genotoxic carcinogens are generally assumed to have thresholds below which there is expected to be no risk. The disagreement is not confined to official government positions; some scientists prefer to treat non-genotoxic carcinogens as having thresholds, some do not. In the view of the committee, the quantitative implications of the many and diverse mechanisms listed in Table 6-4 must be worked out on a case by case basis. Even for specific types of mechanisms for which some data are available, as in the receptor binding studies that provide a framework for understanding the multiple effects of PCBs and dioxins, the implications for the shape of the dose-response relationship at

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Seafood Safety substantial unexploited opportunities to reduce risks. Information on the fetal and chronic neurological risks of methylmercury further suggests the need to reevaluate the current tolerance in that case. In the development of advisories for reproductive effects, due weight must be given to the persistence of different toxicants in people. For methylmercury, with an elimination half-life averaging about 70 days, it may well be sufficient to direct advice to couples who intend to have children in the near future. For PCBs on the other hand, with half-lives measured in several years or even decades, reductions in intake only in the months prior to and during pregnancy can be expected to have little impact on effective body burdens and fetal exposure. In that case, cautionary advice may need to extend to the entire population of reproductive and prereproductive ages. NOTES 1.   In the methylmercury case, the committee was able to infer the distribution of effective internal dosage and risk to the population by using some fragmentary data on blood levels. Similar blood level population distribution data were also cited in the polychlorinated biphenyl (PCB) case study, although a translation into intake distributions or risks was not possible there because of the incompleteness of information about the pharmacokinetics of different PCB congeners. 2.   A major theme, if not the central organizing principle of traditional physiology and toxicology, is the concept of the "homeostatic system." Biological processes are seen as part of a complex interacting web, exquisitely designed so that modest changes in any parameter will automatically give rise to compensating processes to restore optimal functioning (e.g., too much heat input automatically induces sweating so that temperature is kept within a normal range). In this view, as long as a toxic material or any other disturbing stimulus does not push one or more parameters beyond a specified limit ("threshold"), adaptive processes will repair any damage that may have been temporarily produced and completely restore the system to its normal functional state. This paradigm has enjoyed great success in guiding the design and interpretation of a wide range of experimental findings on acute responses to toxic chemicals, heat, cold, and other agents in which the mechanism of damage, does, in fact, consist of grossly overwhelming a particular set of bodily defenses. Another type of damage mechanism dominates thinking in molecular biology and genetics. At the molecular level, some fundamental life processes are basically fragile, in particular, the integrity of the information coded within the deoxyribonucleic acid (DNA) of each cell. An unrepaired error ("mutation") in copying will usually be passed on to all descendants of the mutated cell, and even if the mistake is confined to a single DNA base, massive adverse consequences may result if important genetic information has been altered in a way that affects its function. For the molecular biologist it is intuitively obvious that even a single molecule of a substance that reacts with DNA has some chance of producing a biologically significant result if it happens to interact with the right DNA site. For the traditional toxicologist, basic intuition leads to the opposite expectation: for any substance there is some level of exposure that will have no significant effect on a given biological system. Clearly, application of either intuition to a particular biological response is appropriate only to the degree

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Seafood Safety     that the causal mechanism for that response resembles the paradigmatic damage-producing process that is the basis for the intuition. 3.   It should be stressed that there is no necessary association between the reversibility or irreversibility of processes causing impairment and the social significance of the impairment itself. For example, many thousands of people are killed each year in automobile accidents because of the fully reversible impairments of judgment and reaction time produced by alcohol. 4.   Quantifying human interindividual differences in pharmacokinetic or other parameters that are likely to produce different susceptibilities to adverse effects is a key enterprise if risk assessment for traditional categories of toxic effects is to move from the gross no-effect level (NOEL)/"uncertainty factor" approach to more quantitative treatments of the likely incidence and severity of adverse effects. Hattis et al. (1987a,b) have compiled some preliminary data on the interindividual variability of systemic pharmacokinetic parameters for chemicals in general and of parameters that are likely to contribute to individual susceptibility to anticholinesterase agents. 5.   For example, for healthy workers there may indeed be a functional reserve capacity for oxygen delivery to the myocardium and hence a finite tolerance for a small impairment of oxygen-delivering capacity for the blood due to carbon monoxide. However, for a worker who has just begun to experience a myocardial infarction, oxygen delivery to portions of the myocardium is known to be seriously compromised, and a small difference in oxygen-delivering capacity due to a modest blood carboxyhemoglobin concentration could prove the difference between life and death for portions of the heart muscle that are suddenly forced to rely on collateral arterial vessels for oxygen supply. 6.   Witness FDA's 2-ppm limit in fish versus extensive EPA efforts to prevent further release of PCBs by mandating collection and destruction of PCBs from used electrical equipment, cleanup of soil contaminated by PCB spills, and even serious consideration of dredging PCB-contaminated sediment from areas such as New Bedford Harbor, Massachusetts. 7.   According to R.J. Scheuplein (1988), Deputy Director of the Office of Toxicological Sciences at FDA, in 1969 the FDA identified PCB residues in milk from several dairy farms in West Virginia. Eventually, the source of contamination was traced to spent transformer fluid that was used as a vehicle for a herbicide; dairy cattle grazing nearby had become contaminated. FDA established an action level of 5.00 ppm on PCBs in milk (fat basis). This represented the first U.S. regulatory action taken because of PCB contamination of food. During the next two years, seven other major incidents of PCB contamination of food occurred in the United States. In New York State in 1970, 140,000 chickens were destroyed because testing showed PCBs in excess of an FDA 5 ppm action level. The alleged source of contamination was believed to be plastic bakery wrappers which were ground up with the bakery goods fed to the chickens. In April of 1970 FDA investigated the contamination of milk in Ohio and determined that some farmers were using a PCB-containing sealant in their silos that migrated to the silage. By late 1971, it was quite apparent the spillage or leakage of PCBs from equipment or contact with PCB-containing materials could directly contaminate food and feed. Spills and leaks are sporadic episodes of direct food contamination and it was in response to these potential "Yusho-like" incidents that FDA first intervened with action levels and with controls on distribution. But it was also becoming clear by 1971 that PCBs had become ubiquitous environmental contaminants capable of indirectly, unavoidably and persistently contaminating many types of food.

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Seafood Safety 8.   This group was defined as those consuming more than 24 pounds per year (average consumption was 38.5 pounds). If they indeed received an average of 1.75 µg/kg/day, then the average concentration of PCBs in their fish would have been approximately 2.6 ppm; 70-kg average body weights are assumed. 9.   Unfortunately, a number of other difficulties with the calculation as presented are not so easily quantified. For example, on the cost side, FDA appears to have used landed weights of fish subject to seizure to estimate the economic losses expected from various policies. The assumption is thus that (1) every violative fish caught and presented to market is detected, and yet (2) fishermen make no changes in the locations they fish, species they fish for, and sizes of fish they bring in, in response to the assumed strict enforcement system. The shifting of fishing resources (people, equipment) to locations and species with fewer contamination problems would tend to reduce the long-term economic cost of the tolerance reduction. 10.   Not all such studies are positive, it should be noted, and the human evidence of PCB carcinogenicity is still not regarded as definitive. 11.   Note that the uncertainty bounds on this estimate extend to values that are higher and lower than the estimate by at least 10-fold; thus, the FDA potency factor is by no means ruled out by current epidemiological results. 12.   The Z-score in Figure 6-1 is simply the number of standard deviations above or below the midpoint of a standard normal or lognormal distribution, inferred from the rank of a specific individual value in a data set. To create this type of plot, measurements are first arranged in order and given ranks i (1 through N). Then, a "percentage score" is calculated for each ordered value as 100 x (i – 0.5)/N. (This is simply the percentage of an infinite sample that would be expected to be less than or equal to the observed value. It differs from the usual definition of a "percentile" in which the highest observation is assigned a score of 100.) Finally, from tables of probits in Finney (1971) or areas under a cumulative normal distribution, one calculates the number of standard deviations above or below the median of a normal distribution that would be expected to be associated with each "percentage score," if the distribution of values were in fact normal (Gaussian). In the regression line calculated from this type of plot, the intercept (Z = 0) represents the expected median, and the slope represents the standard deviation; R is regression coefficient. 13.   From the two preceding sentences, it can be inferred that Tollefson and Cordle (1986) are treating the distributions of consumption rates for individual species as normal (Gaussian), rather than lognormal, which may be more accurate. 14.   Unfortunately, although the NMFS (1978) report is statistically sophisticated it appears to be biologically naive, in that is seems to focus on the distribution of daily intake, rather than periods of a month or more to be as comparable as possible to the long biological half-life of methylmercury in humans. The report is unclear enough in its methodology and end product results that the committee is unable to effectively utilize its contents, but it provides at least an illustration of the kind of distributional treatment that, if based on appropriate periods of exposure, could be toxicologically informative. 15.   In other work (not shown) these data were fit to the lognormal risk model in Clement Associates ToxRisk2 statistical package (Crump et al., 1989). This model differs slightly from the classic probit model in that it is essentially a one-hit dose-response function, incorporating a lognormal distribution of susceptibilities. Despite

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Seafood Safety     this difference, the results obtained by this procedure were very similar to those shown for the classical Finney (1971) procedure. 16.   If the distributions of biological half-lives and thresholds for effect in terms of blood levels are in fact lognormal, if among individuals the two distributions are independent (uncorrelated), and if the two factors each act multiplicatively in affecting individuals' thresholds for effect in terms of long-term dietary dose, then the corresponding log10 variances can simply be added. Thus, the probit slope for dietary exposure = 1/{(1/blood prohibit slope)2 [log10 (half-life geometric standard deviation)2]}0.5 The log10 (half-life geometric standard deviation) calculated from the data in Figure 6-2 is 0.147016. 17.   The FDA does not publish a list of cancer potency estimates for these compounds using its own methodology, which differs from that of EPA (discussed earlier). As a standard practice, EPA accompanies the use of these cancer potency estimates with the following: "This level is an upper estimate and the actual risk may be as low as zero." Exactly how much, as a rule, these numbers are likely to overstate actual risks is the subject of much current controversy in the regulatory and toxicological communities. For a comparison of best estimates of cancer risk and EPA upper confidence limits in three cases with the aid of physiologically based pharmacokinetic analyses, see Hattis (1990b). 18.   This is in excess of the background cancer risk, which is about 1 in 5 for the U.S. population. 19.   For example, in the Alabama River, near Claiborne, Alabama, fillets sampled from large-mouth bass average 16.1 parts per trillion dioxin. If the consumption rate is 18 g daily, which is equivalent to about one-third of a pound per week, the cancer risk would be 6.7 × 10-4. 20.   Unlike the federal regulation, MDPH uses a concentration of 0.5 ppm of mercury in fish tissue as a trigger for issuance of fish consumption advisories. This level is based on a WHO recommendation that daily consumption of mercury not exceed 35 µg. This would result in a body burden approximately 10 times lower than that observed to cause effects in humans in mercury poisoning incidents in Japan and Iraq (Michigan, 1989). At the 0.5-ppm contamination level, a person could eat nearly a pound of fish per week without exceeding the WHO recommended maximum daily intake. Larger, older fish in many inland lakes throughout Michigan may have concentrations of mercury in the 0.5- to 1.5-ppm range. This discovery of mercury in fish from inland lakes is not limited to Michigan. Wisconsin, Minnesota, and Ontario have all experienced similar findings. The EPA and the upper Midwest states are currently evaluating whether factors such as acid rain may contribute to this problem. As for warning anglers about all sources of contamination, that task appears to be impossible. Michigan alone has approximately 10,000 inland lakes within its boundaries, and the state readily concedes that it will never be feasible for fish from all lakes to be tested for contaminants.

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Seafood Safety REFERENCES AIHC (American Industrial Health Council). 1987. A Discussion Memorandum of the Vinyl Chloride Decision: Improving the Science Base for Health Assessment. Submitted to EPA for consideration in its Carcinogen Risk Assessment Guidelines: Proposed Updating. September. AIHC, Washington, D.C. 19 pp. Allen, B.C., A.M. Shipp, K.S. Crump, B. Killian, M.L. Hoge, and B.K. Tudor. 1987. Investigation of Cancer Risk Assessment Methods (four volumes). Report by Clement Associates, Inc. to the U.S. Environmental Protection Agency, EPA Report No. EPA/600/6-87/007a-d, September. Al-Shahristani, H., and K. Shihab. 1974. Variation of biological half-life of methylmercury in man. Arch. Environ. Health 28:342-344. Appelman, L.M., W.F. ten Berge, and P.G.J. Reuzel. 1982. Acute inhalation toxicity study of ammonia in rats with variable exposure periods, Am. Ind. Hyg. Assoc. J. 43:662-665. Baker, E.L., R.G. Feldman, R.F. White, and J.P. Harley. 1983. The role of occupational lead exposure in the genesis of psychiatric and behavioral disturbances. Acta Psychiat. Scand. 67(Suppl. 303):38-48. Bakir, F., S.F. Damluji, L. Amin-Zaki, M. Murtadha, A. Khalidi, N.Y. al-Rawi, S. Tikriti, H.I. Dahahir, T.W. Clarkson, J.C. Smith, and R.A. Doherty. 1973. Methylmercury poisoning in Iraq. Science 181:230-241. Ballew, M., and D. Hattis. 1989. Reproductive Effects of Glycol Ethers in Females-A Quantitative Analysis. M.I.T. Center for Technology, Policy, and Industrial Development, CTPID 89-7, July. Barry, P.S.I. 1975. A comparison of concentrations of lead in human tissues. Brit. J. Ind. Med. 32:119-139. Bellinger, D., A. Leviton, C. Waternaux, H. Needleman, and M. Rabinowitz. 1987. Longitudinal analyses of prenatal and postnatal lead exposure and early cognitive development. N. Engl. J. Med. 316:1037-1043. Bellinger, D., A. Leviton, and J. Sloman. 1990. Antecedents and correlates of improved cognitive performance in children exposed in utero to low levels of lead. Environ. Health Perspect. 89:5-11. Berglund, F., M. Berlin, G. Birke, U. von Euler, L. Friberg, B. Holmstedt, E. Jonsson, C. Ramel, S. Skerfving, A. Swensson, and S. Tejning. 1971. Methylmercury in fish: A toxicological-epidemiologic evaluation of risks. Report from an expert group. Nord. Hyg. Tidskr. 4(Suppl.):19-290. Bernard, S.R. 1977. Dosimetric data and metabolic model for lead. Health Physics 32:44-46. Bertazzi, P.A., L. Riboldi, A. Pesatori, L. Radice, and C. Zocchetti. 1987. Cancer mortality of capacitor manufacturing workers. Am. J. Ind. Med. 11:165-76. Bishop, J.M. 1987. The molecular genetics of cancer. Science 235:305-311. Brown, D.P. 1987. Mortality of workers exposed to polychlorinated biphenyls: An update. Arch. Environ. Health 42:333-339. Buchler, F., P. Schmid, and C. Schlatter. 1988. Kinetics of PCB elimination in man. Chemosphere 17:1717-1726. Campbell, B.C., P.A. Meredith, M.R. Moore, and W.S. Watson. 1984. Kinetics of lead following intravenous administration in man. Toxicol. Lett. 21:231-235. Cappuzo, J., A. McElroy, and G. Wallace. 1987. Fish and shellfish contamination in New England waters: An evaluation and review of available data on the distribution of chemical contaminants. Report submitted to Coast Alliance, Washington, D.C. Chamberlain, A.C. 1985. Prediction of response of blood lead to airborne and dietary lead from volunteer experiments with lead isotopes. Proc. Roy. Soc. London B 224:149-182. Clarkson, T.W., L. Amin-Zaki, and S.K. Al-Tikriti. 1976. An outbreak of methylmercury poisoning due to consumption of contaminated grain. Fed. Proc. 35:2395-2399. Clarkson, T.W., B. Weiss, and C. Cox. 1983. Public health consequences of heavy metals in dump sites. Environ. Health Perspect. 48:113-127. Cleaver, J.E., and D. Bootsma. 1975. Xeroderma pigmentosum: Biochemical and genetic characteristics. Ann. Rev. Genet. 9:19-38. Clement Associates, Inc. 1989. Issues in Setting Drinking Water Standards for Polychlorinated Biphenyls-Alternative Approaches to Estimates of Risk for Polychlorinated Biphenyls. Clement Associates, Inc., Ruston, La. Cordle, F. 1983. Use of epidemiology and clinical toxicology to determine human risk in regulating

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Seafood Safety polychlorinated biphenyls in the food supply. Regul. Toxicol. Pharmacol. 3:252-274. Cordle, F., R. Locke, and J. Springer. 1982. Risk assessment in a federal regulatory agency: An assessment of risk associated with the human consumption of some species of fish contaminated with polychlorinated biphenyls (PCBs). Environ. Health Perspect. 45:171-182. Cox, C., T.W. Clarkson, D.O. Marsh, L. Amin-Zaki, S. Tikriti, and G.G. Myers. 1989. Dose-response analysis of infants prenatally exposed to methylmercury: An application of a single compartment model to single-strand hair analysis. Environmental Research 49:318-332. Crump, K., D. Hoel, and R. Peto. 1976. Fundamental carcinogenic processes and their implications for low dose risk assessment . Cancer Research 36:2973-2979. Crump, K., H. Guess, and K. Deal. 1977. Confidence intervals and test of hypotheses concerning dose response relations inferred from animal carcinogenicity data. Biometrics 33:437-451. Crump, K.S., R.B. Howe, and C. Van Landingham. 1989. Tox-Risk–Toxicology Risk Assessment Program. Clement Associates, Inc., Ruston, La. Cumbie, P.M., and S.L. Van Horn. 1978. Selenium accumulation associated with fish mortality and reproductive failure. Proc. Am. Conf. S.E. Assoc. Fish Wildlife Agencies 32:612. DFO (Canadian Department of Fisheries and Oceans). 1989. Canadian Guidelines for Chemical Contaminants in Fish and Fish Products. Government of Canada, Ottawa. Dourson, M.L., and J.F. Stara. 1983. Regulatory history and experimental support of uncertainty (safety) factors. Reg. Toxicol. Pharmacol. 3:224-238. Ehrenberg, L., E. Moustacchi, and S. Osterman-Golkar. 1983. International Commission for Protection Against Environmental Mutagens and Carcinogens. Dosimetry of genotoxic agents and dose-response relationships of their effects. Mutat. Res. 123:121-182. EPA (Environmental Protection Agency). 1981. Chemical Hazard Information Profile Draft Report. Antimony Trioxide, CAS No. 1309-64-4. Office of Toxic Substances, U.S. Environmental Protection Agency, Washington, D.C. 41 pp. EPA (Environmental Protection Agency). 1983. Lower Meramec River Water Quality Intensive Survey. EPA Region VII, Environmental Services Division, Environmental Monitoring and Compliance Branch, Environmental Evaluation Section. EPA (Environmental Protection Agency). 1986a. Guidelines for carcinogen risk assessment, Parts II-VI. Federal Register 51(185):33992-34012. EPA (Environmental Protection Agency). 1986b. Superfund Public Health Evaluation Manual EPA/540/1-86/060. Office of Emergency and Remedial Response, U.S. Environmental Protection Agency, Washington, D.C., October. 186 pp. EPA (Environmental Protection Agency). 1987. The National Dioxin Study: Tiers 3, 5, 6, and 7. EPA 440/4-87-003. Office of Water Regulations and Standards, Monitoring and Data Support Division (WH-553), U.S. Environmental Protection Agency, Washington, D.C. EPA (Environmental Protection Agency). 1989. Health Assessment Survey Tables-Third Quarter 1989. U.S. Environmental Protection Agency. Publication OERR 9200.6/303/(89-3), July. FAO/WHO (Food and Agriculture Organization/World Health Organization). 1972. Evaluation of certain food additives and of the contaminants mercury, lead, and cadmium. Joint FAO/WHO Expert Committee on Food Additives. FAO Nutrition Meetings Report Series No. 51, Rome. 32 pp. FDA (Food and Drug Administration). 1979. Polychlorinated biphenyls (PCB's); Reduction of tolerance. Federal Register 44(June 29):38330-38340. FDA (Food and Drug Administration). 1982. Levels for poisonous or deleterious substances in human food and animals feed. Food and Drug Administration, Washington, D.C. 13 pp. FDA (Food and Drug Administration). 1988. Fish List: Guide to Acceptable Market Names for Food Fish Sold in Interstate Commerce. U.S. Government Printing Office, Washington, D.C. 50 pp. Fein, G.G., J.L. Jacobson, S.W. Jacobson, P.M. Schwartz, and J.K. Dowler. 1984. Prenatal exposure to polychlorinated biphenyls: Effects on birth size and gestational age. J. Pediatrics 105:315-320. Fialkow, P.J. 1977. Clonal origin and stem cell evolution of human tumors. Pp. 439-453 in J.J. Mulvihill, R.W. Miller, and J.F. Fraumeni, Jr., eds. Genetics of Human Cancer. Raven Press, New York. Finkel, A.M. 1990. Confronting Uncertainty in Risk Management–A Guide for Decision Makers. Resources for the Future, Washington, D.C. 68 pp. Finney, D.J. 1971. Probit Analysis, 3rd ed. Cambridge University Press, Cambridge, England. 333 pp. Fischinger, P.J., and V.T. DeVita, Jr. 1984. Governance of science at the National Cancer Institute: Perceptions and opportunities in oncogene research. Cancer Res. 44:4693-4696. Friedman, B., A.R. Frackelton, Jr., A.H. Ross, J.M. Connors, H. Fujiki, T. Sugimura, and M.R. Rosner.

OCR for page 172
Seafood Safety 1984. Tumor promoters block tyrosine-specific phosphorylation of the epidermal growth factor receptor. Proc. Natl. Acad. Sci. 81:3034-3038. GAO (General Accounting Office). 1988. Seafood Safety: Seriousness of Problems and Efforts to Protect Consumers, RCED-88-135. Report to the Chairman, Subcommittee on Commerce, Consumer and Monetary Affairs, Committee on Government Operations, House of Representatives. U.S. Government Printing Office, Washington, D.C. Gartrell, M.J., J.C. Craun, D.S. Podrebarac, and E.L. Gunderson. 1985. Pesticides, selected elements, and other chemicals in adult total diet samples, October 1978-September 1979. J. Assoc. Off. Anal. Chem. 68:862-873. Gossett, R., H. Puffer, R. Arthur, J. Alfafara, and D. Young. 1982. Levels of Trace Organic Compounds in Sportfish from Southern California. Coastal Water Research Project Biennial Report. Southern California Coastal Water Research Project, Long Beach, Calif. Green, V.A., G.W. Wise, and J.C. Callenbach. 1978. Lead poisoning. Pp. 123-141 in F. Oehme, ed. Toxicity of Heavy Metals in the Environment, Part 1. Marcel Dekker, New York. Groth, D.H., L.E. Stettler, J.R. Burg, W.M. Busey, G.C. Grant, and L. Wong. 1986. Carcinogenic effects of antimony trioxide and antimony ore concentrate in rats. J. Toxicol. Environ. Health 18:607-626. Guess, H., K. Crump, and R. Peto. 1977. Uncertainty estimates from low-dose extrapolations of animal carcinogenicity data. Cancer Res. 37:3475-3483. Gunderson, E.L. 1988. FDA Total Diet Study, April 1982-April 1984, Dietary intakes of pesticides, selected elements and other chemicals. J. Assoc. Off. Anal. Chem. 71:1200-1209. Haeger-Aronsen, B., M. Abdulla, and B.I. Fristedt. 1974. Effect of lead on delta-aminolevulinic acid dehydratase activity in red blood cells. II. Regeneration of enzyme after cessation of lead exposure. Arch. Environ. Health 29:150-153. Haines, A.T., and E. Nieboer. 1988. Chromium hypersensitivity. Pp. 497-532 in J. Nrigau and E. Nieboer, eds. Chromium in the Natural and Human Environment. John Wiley and Sons, New York. Hall, R.A., E.G. Zook, and G.M. Meaburn. 1978. National Marine Fisheries Service Survey of Trace Elements in the Fishery Resource. NOAA Technical Report NMFS SSRF-721, National Technical Information Service No. PB 283 851, March. U.S. Government Printing Office, Washington, D.C. Hattis, D. 1981. Dynamics of medical removal protection for lead–A reappraisal. Report to the National Institute for Occupational Safety and Health. M.I.T. Center for Policy Alternatives Report No. CPA-81-25. Cambridge, Mass. 47 pp. Hattis, D. 1982. From presence to health impact: Models for relating presence to exposure to damage. Pp. 1-66 in N.A. Ashford and C.T. Hill, eds. Analyzing the Benefits of Health, Safety, and Environmental Regulations. M.I.T. Center for Policy Alternatives, Report No. CPA-82-16. Cambridge, Mass. Hattis, D. 1986. The promise of molecular epidemiology for quantitative risk assessment. Risk Analysis 6:181-193. Hattis, D. 1988. The use of biological markers in risk assessment. Statistical Science 3:358-366. Hattis, D. 1990a. Pharmacokinetic principles for dose rate extrapolation of carcinogenic risk from genetically active agents. Risk Analysis 10:303-316. Hattis, D. 1990b. Use of biological markers and pharmacokinetics in human health risk assessment. Environmental Health Perspectives (in press). Hattis, D., and A. Smith. 1986. What's wrong with quantitative risk assessment? Pp. 57-79 in R. Almeder and J. Humber, eds. Biomedical Ethics Reviews 1986. The Humana Press, Clifton, N.J. Hattis, D., and H. Strauss. 1986. Potential Indirect Mechanisms of Carcinogenesis, A Preliminary Taxonomy. National Technical Information Service No. NTIS/PB89-120513. Center for Technology, Policy, and Industrial Development, Report No. CTPID 86-3, Massachusetts Institute of Technology, Cambridge, Mass. 12 pp. Hattis, D., L. Erdreich, and M. Ballew. 1987a. Human variability in susceptibility to toxic chemicals–A preliminary analysis of pharmacokinetic data from normal volunteers. Risk Anal. 7:415-426. Hattis, D., S. Bird, and L. Erdreich. 1987b. Human Variability in Susceptibility to Anticholinesterase Agents. M.I.T. Center for Technology, Policy and Industrial Development, Report No. CTPID 87-4 . Cambridge, Mass. Haxton, J., D.G. Lindsay, J.S. Hislop, L. Salmon, E.J. Dixon, W.H. Evans, J.R. Reid, C.J. Hewitt, and D.F. Jeffries. 1979. Duplicate diet study on fishing communities in the United Kingdom:

OCR for page 172
Seafood Safety Mercury exposure in a "critical group". Environ. Res. 18:351-368. HHS (Department of Health and Human Services), Agency for Toxic Substances and Disease Registry. 1988. The Nature and Extent of Lead Poisoning in Children in the United States: A Report to Congress. July. U.S. Government Printing Office, Washington, D.C. Hing, E. 1989. Nursing Home Utilization by Current Residents: United States, 1985. National Center for Health Statistics, Vital Health Stat. 13(102), DHHS Publication No. (PHS) 89-1763, October. U.S. Government Printing Office, Washington, D.C. Hoel, D.G. 1985. Epidemiology and the inference of cancer mechanisms. Natl. Cancer Inst. Monogr. 67:199-203. Hoel, D.G., N.L. Kaplan, and M.W. Anderson. 1983. Implications of non-linear kinetics on risk estimation in carcinogenesis. Science 219:1032-1037. Hogue, C.J.R, J.W. Buehler, M.A. Strauss, and J.C. Smith. 1987. Overview of the national infant mortality surveillance (NIMS) project–Design, methods, results. Public Hlth. Reports 102:126-138. Humphrey, H.E.B. 1974. Mercury concentrations in humans and consumption of fish containing methylmercury. Mercury Project Progress Report. Michigan Department of Public Health, Lansing. 6 pp. Humphrey, H. 1983a. Population studies of PCBs in Michigan residents. Pp. 299-310 in F.M. D'Itri and M.A. Kamrin, eds. PCBs: Human and Environmental Hazards. Butterworth, Boston, Mass. Humphrey, H. 1983b. Evaluation of Humans Exposed to Water-Borne Chemicals in the Great Lakes. Final Report to the Environmental Protection Agency, Cooperative Agreement CR-807192. U.S. Environmental Protection Agency, Washington, D.C. 205 pp. Humphrey, H.E.B. 1988. Chemical contaminants in the Great Lakes: The human health aspect. Pp. 153-165 in M.S. Evans, ed. Toxic Contaminants and Ecosystem Health: A Great Lakes Focus. John Wiley & Sons, New York. Jacobson, S.W., G.G. Fein, J.L. Jacobson, P.M. Schwartz, and J.K. Dowler. 1985. The effect of intrauterine PCB exposure on visual recognition memory. Child Develop. 56:853-860. Jones, K.C. 1988. Determination of polychlorinated biphenyls in human foodstuffs and tissues: Suggestions for a selective congener analytical approach. Science Total Environ. 68:141-159. Kabelitz, D. 1985. Modulation of natural killing by tumor promoters. The regulatory influence of adherent cells varies with the type of target cell. Immunobiology 169:436-446. Kampert, J.B., A.S. Whittemore, and R.S. Paffenbarger, Jr. 1988. Combined effect of childbearing, menstrual events, and body size on age-specific breast cancer risk, Am. J. Epidemiol. 128:962-979. Kansas DHE (Department of Health and Environment). 1988. A Survey of Pesticides in Tuttle Creek Lakes, Its Tributaries and the Upper Kansas River. Water Quality Assessment Section, Bureau of Water Protection, Kansas Department of Health and Environment, Topeka. Kinsella, A.R., and M. Radman. 1978. Tumor promoter induces sister chromatid exchanges. Proc. Natl. Acad. Sci. 75:6149-6153. Konz, J. 1988. Fish intake study. A memo report dated September 19, 1988 from Jim Konz of Versar, Inc. to Jacqueline Moya of the Environmental Protection Agency. Versar, Inc., Springfield, Va. Knudson, A.G. 1973. Mutation and human cancer. Adv. Cancer Res. 17:317-352. Knudson, A.G. 1977. Genetics and etiology of human cancer. Adv. Hum. Genet. 8:1-66. Landolt, M.L., F.R. Hafer, A. Nevissi, G. Van Belle, K. Van Ness, and C. Rockwell. 1985. Potential toxicant exposures among consumers of recreationally caught fish from urban embankments of Puget Sound. NOAA Tech. Memo, NOS OMA 23. Rockville, Md. 104 pp. Landolt, M.L., D. Kalman, A. Nevissi, G. van Belle, K. Van Ness, and F.R. Hafer. 1987. Potential toxicant exposures among consumers of recreationally caught fish from urban embayments of Puget Sound. NOAA Tech. Memo, NOS OMA 33. Rockville, Md. 111 pp. Layde, P.M., L.A. Webster, A.L. Baughman, P.A. Wingo, G.L. Rubin, and H.W. Ory. 1989. The independent associations of parity, age at first full term pregnancy, and duration of breastfeeding with the risk of breast cancer. Cancer and Steroid Hormone Study Group. J. Clin. Epidemiol. 42:963-973. Mackay, N.J., R.J. Williams, J.L. Kacprzac, M.N. Kazacos, A.J. Collins, and E.H. Auty. 1975. Heavy metals in cultivated oysters (Crassostrea commercialis = Saccostrea cucullata) from the estuaries of New South Wales. Aust. J. Mar. Freshwater Res. 26:31-46. Marcus, A.H. 1985a. Multicompartment kinetic models for lead. I. Bone diffusion models for long-term retention. Environ. Res. 36:441-458.

OCR for page 172
Seafood Safety Marcus, A.H. 1985b. Multicompartment kinetic models for lead. II. Linear kinetics and variable absorption in humans without excessive lead exposures. Environ. Res. 36:459-472. Marcus, A.H. 1985c. Multicompartment kinetic models for lead. III. Lead in blood plasma and erythrocytes. Environ. Res. 36:473-489. Marsh, D.O., M.D. Turner, J.C. Smith, J.W. Choe, and T.W. Clarkson. 1974. Methylmercury in human populations eating large quantities of marine fish. I. Northern Peru. Pp. 235-239 in Proceedings of the 1st International Mercury Congress, May 6-10, 1974. Barcelona, Spain. Marsh, D.O., T.W. Clarkson, C. Cox, G.J. Myers, L. Amin-Zaki, and S. Al-Tikriti. 1987. Fetal methylmercury poisoning. Relationship between concentration in single strands of maternal hair and child effects. Arch. Neurol. 44:1017-1022. Matthews, H.B., and R.L. Dedrick. 1984. Pharmacokinetics of PCBs. Rev. Pharmacol. Toxicol. 24:85-103. Maugh, T.H. 1978. Chemical carcinogens: How dangerous are low doses? Science 202:37. McCaffrey, P.G., B. Friedman, and M.R. Rosner. 1984. Diacylglycerol modulates binding and phosphorylation of the epidermal growth factor receptor. J. Biol. Chem. 259:12502-12507. McCann, J., E. Choi, E. Yamasaki, and B.N. Ames. 1975. Detection of carcinogens as mutagens in the salmonella/microsome test: Assay of 300 chemicals. Proc. Natl. Acad. Sci. 72:5135-5139. McFarland, V.A., and J.U. Clarke. 1989. Environmental occurrence, abundance, and potential toxicity of polychlorinated biphenyl congeners: Considerations for a congener-specific analysis. Environ. Health Perspect. 81:225-239. Michigan DPH (Department of Public Health). 1988. Fish Consumption Advisory Issued by State Health Department . Michigan Department of Public Health, Center for Environmental Health Sciences, December 14, 1988. Lansing. 8 pp. Miller, E.C., and J.A. Miller. 1981. Mechanisms of chemical carcinogens. Cancer 47:1055-1064. Modali, R., and S. S. Yang. 1986. Specificity of Aflatoxin B1 binding on human proto-oncogene nucleotide sequence. Pp. 147-158 in M. Sorsa and H. Norppa, eds. Monitoring of Occupational Genotoxicants. Proceedings of a Satellite Symposium to the Fourth International Conference on Environmental Mutagens, Helsinki, Finland, June 30-July 2, 1985. Alan R. Liss, New York. Moolgavkar, S. 1986. Carcinogenesis modeling: From molecular biology to epidemiology, Ann. Rev. Public Hlth. 7:151-169. Moolgavkar, S.H., and A.G. Knudson, Jr. 1981. Mutation and cancer: A model for human carcinogenesis. J. Natl. Cancer Inst. 66:1037-1052. Moolgavkar, S.H., N.E. Day, and R.G. Stevens. 1980. Two-stage model for carcinogenesis: Epidemiology of breast cancer in females, J. Natl. Cancer Inst. 65:559-569. Murphy, D.L. 1988a. Basic water monitoring program. Fish tissue analysis, 1985 . Water Management Admin. Tech. Report 59. State of Maryland, Department of the Environment, Baltimore. 38 pp. Murphy, D.L. 1988b. Trace contaminants in Chesapeake Bay bluefish. Metals and organochlorine pesticides. Water Management Admin. Tech. Report 73. State of Maryland, Department of the Environment, Divisions of Standards and Certification, Baltimore. 21 pp. Murphy, D.L. 1988c. Trace contaminants in striped bass from two Chesapeake Bay tributaries. Metals and organochlorine pesticides. Technical Report 58. State of Maryland, Department of the Environment, Divisions of Standards and Certification, Water Management Administration, Baltimore. 19 pp. Needleman, H.L., C. Gunnoe, A. Leviton, R. Reed, H. Peresie, C. Maher, and P. Barrett. 1979. Deficits in psychologic and classroom performance of children with elevated dentine lead levels. N. Engl. J. Med. 300:689-695. Needleman, H., A. Schell, D. Bellinger A. Leviton, and E. Allred. 1990. Long-term effect of childhood exposure to lead at low doses: An eleven-year follow-up report. N. Engl. J. Med. 322:83-88. Ngim, C.N., and G. Devathasan. 1989. Epidemiologic study on the association between body burden mercury level and idiopathic Parkinson's disease. Neuroepidemiol. 8:128-141. NMFS (National Marine Fisheries Service). 1978. On the Chance of U.S. Seafood Consumers Exceeding the Current Acceptable Daily Intake for Mercury: Recommendations, Regulatory Controls. Seafood Quality and Inspection Division, Office of Fisheries Development, National Marine Fisheries Service. 198 pp. NMFS (National Marine Fisheries Service). 1988. Fisheries of the United States, 1987. Current Fisheries Statistics, No. 8700. U.S. Government Printing Office, Washington D.C. 115 pp. NOAA (National Oceanic and Atmospheric Administration). 1987. Report on 1984-86 Federal Survey of PCBs in Atlantic Coast Bluefish. Interpretive Report. NOAA in cooperation with the Food and Drug Administration and Environmental Protection Agency. U.S. Government Printing Office,

OCR for page 172
Seafood Safety Washington, D.C. March. 75 pp. Norback, D.H., and R.H. Weltman. 1985. Polychlorinated biphenyl induction of hepatocellular carcinoma in the Sprague-Dawley rat. Environ. Health Perspect. 60:97-105. NRC (National Research Council). 1980. Recommended Dietary Allowances, 9th ed. Committee on Dietary Allowances, Food and Nutrition Board. National Academy Press, Washington, D.C. 185 pp. NRC (National Research Council). 1989. Biologic Markers in Reproductive Toxicology. Board on Environmental Studies and Toxicology. National Academy Press, Washington, D.C. 416 pp. NRC (National Research Council). 1990. Managing Troubled Waters. Committee on a Systems Assessment of Marine Environmental Monitoring, Marine Board. National Academy Press, Washington, D.C. 125 pp. NRC (National Research Council). 1991. Neurotoxicology and Models for Assessing Risk. Board on Environmental Studies and Epidemiology. National Academy Press, Washington, D.C. (in press). NWF (National Wildlife Federation). 1989. Abbreviated summary of quantitative health assessments for PCB, dieldrin, DDT, and chlordane. Developed for the 13 April 1989 workshop on Managing the Health Risks of Consuming Contaminated Great Lakes Sport Fish. Great Lakes Natural Resource Center, National Wildlife Federation, Ann Arbor, Mich. 16 pp. Pasquinelli, P., F. Bruschi, F. Saviozzi, and G. Malvaldi. 1985. Immunosuppressive effects and promotion of hepatic carcinogenesis by thiobenzamide. Bull. Soc. Ital. Biol. Sper. 61:61-66. Phillips, D.L., A.B. Smith, V.W. Burse, G.K. Steele, L.L. Heedham, and W.H. Hannon. 1989. Half-life of polychlorinated biphenyls in occupationally exposed workers. Arch. Environ. Hlth. 44:351-354. PTI. 1987. Guidance Manual for Assessing Human Health Risks from Chemically Contaminated Fish and Shellfish . Draft Report C737-2. Submitted to Batelle New England Marine Research Laboratory by TI Environmental Services, Inc., Bellevue, Wash., December. 64 pp. Rabinowitz, M.B., G.W. Wetherill, and J.D. Kopple. 1976. Kinetic analysis of lead metabolism in healthy humans. J. Clin. Invest. 58:260-271. Rai, K., and J. van Ryzin. 1981. A generalized multihit dose response model for low dose extrapolation. Biometrics 37:341-352. Rodericks, J. 1989. Assessing and managing risks associated with the consumption of chemically contaminated seafoods. Paper presented at the Workshop on Assessing and Controlling Health Hazards from Fishery Products conducted by the Committee on Evaluation of the Safety of Fishery Products, Food and Nutrition Board, Institute of Medicine, National Academy of Sciences Study Center, Woods Hole, Mass., July 26. 49 pp. Rogan, W.J., B.C. Gladen, J.D. McKinney, N. Carreras, P. Hardy, J. Thullen, J. Tinglestad, and M. Tully. 1986. Neonatal effects of transplacental exposure to PCBs and DDE. J. Pediatrics 109:335-341. Ross, R., and J.A. Glomset. 1976. The pathogenesis of atherosclerosis (second of two parts). N. Engl. J. Med. 295:420-425. Safe, S. 1989. Polychlorinated biphenyls (PCBs): Mutagenicity and carcinogenicity. Mutat. Res. 220:31-47. Sahl, J.D., T.T. Crocker, R.J. Gordon, and E. J. Faeder. 1985. Polychlorinated biphenyl concentrations in the blood plasma of a selected sample of non-occupationally exposed southern California (U.S.A.) working adults. Sci. Total. Environ. 46:9-18. Scheuplein, R.J. 1988. Risk assessment and risk management of environmental food contaminants by FDA. Pp. 109-122 in C.R. Cothern, M.A. Mehlman, and W.L. Marcus, eds. Advances in Modern Environmental Toxicology, Vol. 15: Risk Assessment and Risk Management of Industrial and Environmental Chemicals. Princeton Scientific Publishing Co., Princeton, N.J. Silbergeld, E.K. 1982. Current status of neurotoxicology, basic and applied. Trends Neurosci. 5:291-294. Silbergeld, E.K., R.E. Hruska, D. Bradley, J.M. Lamon, and B.C. Frykholm. 1982. Neurotoxic aspects of porphyrinopathies: Lead and succinylacetone. Environ. Res. 29:459-471. Skerfving, S. 1974. Methylmercury exposure, mercury levels in blood and hair, and health status in Swedes consuming contaminated fish. Toxicology 2:3-23. Sloan, R., E. O'Connell, and R. Diana. 1987. Toxic Substances in Fish and Wildlife-Analyses since May 1, 1982, Vol. 6. Technical Report 87-4 (BEP), September. New York State Department of Environmental Conservation, Division of Fish and Wildlife, Albany. 182 pp. Sorensen, E.M.B., P.M. Cumbie, T.L. Bauer, J.S. Bell, and C.W. Harlan. 1984. Histopathological,

OCR for page 172
Seafood Safety hematological, condition-factor, and organ weight changes associated with selenium accumulation in fish from Belews Lake, North Carolina. Arch. Environ. Contam. Toxicol. 13:153-162. Sunahara, G.I., K.G. Nelson, T.K. Wong, and G.W. Lucier. 1987. Decreased human birth weights after in utero exposure to PCBs and PCDFs are associated with decreased placental EGF-stimulated receptor autophosphorylation capacity. Molecular Pharmacol. 32:572-578. Takada, K., and H. Zur Hausen. 1984. Induction of Epstein-Barr virus antigens by tumor promoters for epidermal and nonepidermal tissues. Int. J. Cancer 33:491-496. Tanabe, S., Y. Nakagawa, and R. Tatsukawa. 1981. Absorption efficiency and biological half-life of individual chlorobiphenyls in rats treated with Kanechlor products. Agricult. Biol. Chemist. 45:717-726. Taylor, P.R., J.M. Stelma, and C.E. Lawrence. 1989. The relation of polychlorinated biphenyls to birth weight and gestational age in the offspring of occupationally exposed mothers. Am. J. Epidem. 129:395-406. Tollefson, L., and F. Cordle. 1986. Methylmercury in fish: A review of residue levels, fish consumption and regulatory action in the United States. Environ. Hlth. Perspect. 68:203-208. Tsushimoto, G., C.C. Chang, J.E. Trosko, and F. Matsumura. 1983. Cytotoxic, mutagenic, and cell-cell communication inhibitory properties of DDT, lindane, and chlordane on Chinese hamster cells in vitro. Arch. Environ. Contam. Toxicol. 12:721-729. Turner, M.D., D.O. Marsh, J.C. Smith, J.B. Inglis, T.W. Clarkson, C.E. Rubio, J. Chiriboga, and C.C. Chiriboga. 1980. Methylmercury in populations eating large quantities of marine fish. Arch. Environ. Health. 35:367-378. Vogel, F., and A.G. Motulsky. 1979. Pp. 326-329 in Human Genetics-Problems and Approaches. Springer-Verlag, New York. Waternaux, C., N.M. Laird, and J.H. Ware. 1989. Methods for analysis of longitudinal data: Blood lead concentrations and cognitive development. J. Am. Stat. Assoc. 84:33-41. Weisburger, J.H., and G.M. Williams. 1983. The distinct health risk analyses required for genotoxic carcinogens and promoting agents. Environ. Health Perspect. 50:233-245. Weiss, B., and J.M. Spyker. 1974. Behavioral implications of prenatal and early postnatal exposure to chemical pollutants. Pediatrics 53:851-859. Whitlock, J.P., Jr. 1989. The control of cytochrome P-450 gene expression by dioxin. Trends Pharmacol. Sci. 10:285-288. Whittemore, A.S. 1980. Mathematical models of cancer and their use in risk assessment, J. Environ. Pathol. Toxicol. 3:353-362. Whittemore, A.S. 1983. Facts and values in risk assessment for environmental toxicants. Risk Anal. 3:23-33. Wilber, C.F. 1983. Selenium: A Potential Environmental Poison and a Necessary Food Constituent. Charles C. Thomas, Springfield, Ill. 126 pp. Yakushiji, T., I. Watanabe, K. Kuwabara, R. Tanaka, T. Kashimoto, N. Kunita, and I. Hara. 1984. Rate of decrease and half-life of polychlorinated biphenyls (PCBs) in the blood of mothers and their children occupationally exposed to PCBs, Arch. Environ. Contam. Toxicol. 13:341-345. Yunis, J.J. 1983. The chromosomal basis of human neoplasia. Science 221:227-236. Zakour, R.A., T.A. Kunkel, and L.A. Loeb. 1981. Metal-induced infidelity of DNA synthesis. Environ. Health Perspect. 40:197-206. Zeggari, M., C. Susini, N. Viguerie, J.P. Esteve, N. Vaysse, and A. Ribet. 1985. Tumor promoter inhibition of cellular binding of somatostatin. Biochem. Biophys. Res. Commun. 128:850-857. Zeitlin, D. 1989. State-Issued Fish Consumption Advisories: A National Perspective. National Ocean Pollution Program Office, National Oceanic and Atmospheric Administration, Washington, D.C. November. 73 pp.