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8 Risk Assessment The Safe Drinking Water Committee evaluated a number of approaches to assessing the risks of a variety of health effects, including cancer, reproductive and developmental impairments, and neurological diseases. Among the major considerations in any of these approaches are the variety and extent of the variables that will be encountered. For example, different sources of data may be used in the risk-assessment process, ranging from short-term tests for mutagenicity to long-term epidemiological studies of humans. Other variables that make risk assessment difficult and sometimes subject to substantial variation include metabolic differences between spe- cies and the variety of ways that exposures can be delivered, e.g., multiple, sporadic, or peak. Other considerations include route of exposure (e.g., inhalation, dermal, or ingestion) and source of exposure (e.g., drinking water, workroom air, or food). In some circumstances, risk assessment can be enhanced when there is an opportunity to evaluate pharmacokinetic mechanisms involved in responses to environmental agents. In this chapter, the committee first reviews the discussions of risk as- sessment set forth by previous Safe ~ng Water Committees. It then examines three topics common to all risk assessment: estimation of exposure from various sources and by different routes, pharmacokinetics, and inter- species extrapolation. The remainder of the chapter is devoted to a discussion of risk assessment for different categories of health effects. Cancer is ex- amined first, because risk assessment in this area has the longest history and, therefore, the methods are more highly developed than those for over adverse health effects. However, many of these methods are germane to assessing the risk for noncancer end points as well. The last two sections cover risk 250

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Risk Assessment 251 assessment for developmental and reproductive effects and for neurotoxic effects. Conclusions and recommendations drawn from all these discussions are presented at the end of the chapter. To assess risks associated with exposure to toxic chemicals, agencies have developed a systematic scientific and administrative framework (NRC, 1983~. The process commonly begins with the identification of a hazard often brought to light by data from studies on laboratory animals, from other laboratory procedures, or sometimes from case reports involving humans. The next step is determination of the dose-response relationships between specific quantities of a substance and associated physical re- sponses, such as the development of tumors, birth defects, or necrologic deficits. Then follows exposure estimation and assessment. At that time, a search is made for an answer to the question: To what dose levels or range of dose levels will human populations most likely be exposed? Finally, the dose-response model is applied to the expected exposure levels to produce a quantitative estimate of risk. To date, quantitative risk assessment (QRA) has been used largely for estimating the risk of developing or dying from cancer, and has been used little in evaluating or estimating noncancer health effects of exposure to materials in the environment. The earliest approaches to QRA for cancer are probably those of Mantel and Bryan (19611. Modifications to these approaches have been published through the years. Recently, the U.S. Environmental Protection Agency (EPA) issued proposed guidelines for assessing the risk of exposure to carcinogens, mutagens, and develop- mental toxicants (EPA, 1984a,b,c). A developing, interdisciplinary field, risk assessment is not without unresolved issues, including the extrapolation of results to doses outside the range of observations in the experiments; the choice of the appropriate dose-response model for extrapolating from nigh-dose animal data to the anticipated low levels of exposure of humans in the ambient environment; the appropriate translation of data from the laboratory animal to humans, which in turn involves questions concerning the influence of body size, life span, and possible metabolic differences; and the potential for con- founding (e.g., whether there is synergistic or antagonistic response to exposures to other materials). Problems that lie at the interface of science and policy include selection of test method, selection of animal data and bioassay results for use as the basis for extrapolation to humans, deter- mination of how one should use data on so-called benign as well as malignant tumors, selection of appropriate safety factors for developing standards, and selection of the mathematical model to be used for ex- trapolation. QRA methods have been the subject of a number of publications (An- derson and CAG, 1983; California Department of Health Services, 1982;

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252 DRINKING WATER AND H"LTH OSTP, 1985), which will not be reviewed here. They can be used to provide a broad range of estimated risks for the purpose of setting priorities for regulatory action or to identify a specific acceptable or permissible level of human exposure to a carcinogen. Most experts and policymakers agree that current QRA techniques at best indicate a range of risks rather than a precise number. Since 1977, when the first Safe Drinking Water Committee conducted QRAs on waterborne carcinogens, National Re- search Council committees have consistently highlighted limitations of the methodology (NRC, 1977, 1980, 1983~. This committee endorses attempts to develop, validate, and apply QRA techniques to the evaluation of potential noncancer toxic responses, such as neurotoxicity, reproductive toxicity, and developmental toxicity, as well as liver damage, kidney damage, and respiratory responses other than lung cancer. Dunng the last decade, cancer has been regarded as a nonthreshold phenomenon, whereas most noncancer responses are be- lieved to require a minimum (i.e., threshold) dose before any toxic man- ifestation will appear. Some recent research has suggested, however, that this distinction may not be a clear one. PREVIOUS SAFE DRINKING WATER COMMITTEES' VIEWS ON RISK ASSESSMENT Earlier Safe Drinking Water Committees considered the problems of risk assessment in the first and third volumes of Drinking Water and Health (NRC, 1977, 19801. The present committee affirms many of the views expressed by those groups. Issues Concerning Threshold After reviewing the many problems inherent in assessing the risks of carcinogenesis and mutagenesis, the 1977 committee acknowledged that many scientists distinguish between injuries produced by chemicals likely to have a threshold dose and effects for which there is likely to be either no threshold (e.g., carcinogenesis and mutagenesis) or no way known to estimate one for large, heterogeneous populations. On the basis of this observation, the 1977 committee concluded, "It is more prudent to treat some kinds of toxic effects that may be self-propagating or strictly cu- mulative, or both, as if there were no threshold and to estimate the upper limits of risk for any given exposure" (NRC, 1977, p. 254. The report included among self-propagating or strictly cumulative effects those that result from early, chemically induced alterations in cellular DNA that are transmitted by cell propagation and irreparable injuries, such as destruction

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Risk Assessment 253 of neurons, noting that "destruction of enough neurons leads to a decrease in central nervous system function." Carcinogenic and Mutagenic Effects The same committee outlined the following principles that underlie efforts to assess the irreversible effects of long-continued exposure to carcinogenic substances at low dose rates: 1. "Effects in animals, properly qualified, are applicable to man." 2. "Methods do not now exist to establish a threshold for long-ter effects of toxic agents." 3. "The exposure of experimental animals to toxic agents in high doses is a necessary and valid method of discovering possible carcinogenic hazards in man." 4. "Material should be assessed in terms of human risk, rather than as 'safe' or 'unsafe' " (NRC, 1977, pp. 53-561. Noncarcinogenic Effects For presumably reversible noncarcinogenic and nonmutagenic effects, the committee advised, "For noncarcinogens for which it seems likely that there are thresholds for toxic effects, the acceptable dose should be below the threshold. If a threshold cannot be shown, the acceptable dose must be related to the data from animal experimentation and consideration of the seriousness of the toxic effects, as well as the likelihood and ease of reversibility, the variability of the sensitivity of the exposed population, and the economic and health-related importance of the material" (NRC, 1977, pp. 57-58~. The present committee appreciates that there is not a clear distinction between cancer and other toxic responses. Safety, or Uncertainty, Factors To this general statement the committee added specific advice on the use of safety, or uncertainty, factors (NRC, 1977, p. 8041. An uncertainty factor of 10 was recommended when there are valid results from studies on prolonged ingestion by humans and no indication of carcinogenicity. An uncertainty factor of 100 was recommended when there are few or no toxicological data on ingestion by humans, but there are valid results from long-term studies in animals and no indication of carcinogenicity. An uncertainty factor of 1,000 was recommended when there are no long- term or acute data on humans, only scanty data on animals, and no indication of carcinogenicity.

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254 DRINKING WATER AND HEALTH The Use of An emal Data to Predict Human Risk The first Safe Drinking Water Committee remarked, "Current knowl- edge of the proper principles for extrapolating toxicological data from high dose to low dose, and from one species to another, is inadequate" (NRC, 1977, p. 251. In Volume 3 of Drinking Water and Health (NRC, 1980), the reconstituted committee reexamined the prediction of risks to human health using acute and chronic toxicity data on laboratory animals. The entire third chapter of Volume 3 was devoted to this issue.The com- mittee summarized its conclusion as follows: The concentration of most potentially toxic chemicals in drinking water is usually so low that it is difficult to predict potentially adverse effects from drinking the water. In cases of noncarcinogenic toxicity, the preferred procedure would be to make a risk estimate based on extrapolation to low dose levels from ex- perimental curves obtained from much larger doses for which effects can be readily measured. In most instances, such data are not available, and the acceptable daily intake (ADI) approach should be used until better data are obtained. In the ADI approach, "safety factors" based on the quality of the data are applied to the highest no-observable-effect dose found in animal studies. The [Safe Drinking Water Committee's] Subcommittee on Risk Assessment believes that the ADI approach is not applicable to carcinogenic toxicity, and that high dose to low dose extrapolation methods should be used for known or suspected carcinogens. Six models were evaluated for low dose carcinogenic risk estimation. They were the dichotomous response model; linear, no-threshold model; tolerance distribution model; logistic model; "hitness" model; and time-to-tumor- occurrence model. Because of the uncertainties involved in the true shapes of the dose-response curves that are used for extrapolation, a multistage model was judged to be the most useful. Such a model has more biological meaning than other models, e.g., the probit or logistic model. Moreover, it tends to be con- servative in that at low doses it will give higher estimates of the unknown risk than will many others. More confidence would be placed in mathematical models for extrapolation if they incorporated biological characteristics such as pharmacokinetic data and time- to-occurrence of tumors. Until such data are available, the extrapolation from animals to humans should be done on the basis of surface area (NRC, 1980, pp. 2-34. As the discussion in that volume indicates, the computation of risk depends upon several assumptions, ranging from the choice of mathe- matical model for low-dose extrapolation to the minor operating assump- tions made within the model itself to implement a specific computer program. The present committee noted that within the multistage model one can compute the dose necessary to develop a given level of risk by using either an experimentally restricted model or a generalized model. The restricted model limits the number of possible stages in the multistage

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Risk Assessment 255 TABLE 8-1 Unit Riska Depending on Assumed Stages of Carcinogenesis Stages Experimentally Stages Generalized Estimate Restricted to Data Maximum likelihood estimate 1.2 x 1o-s 8.8 x 10-6 Upper 95% confidence limit estimate 2.2 x 1o-s 1.7 x 1o-s aAssurrung a daily consumption of 1 liter of water containing the compound in a concentration of 1 ,ug/liter. process to the number of doses at which the experiment was conducted minus 1. The generalized model places no such limit on the possible number of stages but, rather, permits computation of the best fit to the data without the constraint of doses used in the experiment. For example, by using data on response to acrylamide for male A/J mice with lung tumors and the two different computational assumptions concerning stages of carcinogenesis, one can estimate two slightly different risks, as shown in Table 8-1. A similar small difference was found in the risks estimated for female mice with lung tumors. Where the data are highly curvilinear, the exper- imentally restricted approach can assign less of the effect to the linear term, and may even estimate it to be zero. In contrast, the unrestricted form of the computation is more likely to identify a nonzero linear term, which may be more in keeping with current biological knowledge and assumptions. Computations in the third volume were based on both ap- proaches. The differences were only slight and, thus, not important. Fol- lowing the recommendations of the risk-assessment panel of the Consensus Workshop on Formaldehyde (1984), the computations presented in the present volume are based on the generalized form of the model. In considering the appropriate scaling for extrapolating data from the laboratory to humans, the 1980 committee remarked: The practice among cancer chemotherapists of basing dose on body surface area is useful, particularly for extrapolation from small animals to humans, and is supported by a sizeable body of experimental evidence. Since body surface area is approximately proportional to the two-thirds power of body weight, the anticancer drugs are relatively more toxic to the larger animals than to the smaller ones (NRC, 1980, p. 29~. Combined Exposures The 1980 committee considered information on the combined action of materials found in drinking water, noting first that the joint action could be additive, synergistic, or antagonistic, and that "in general, there is not

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256 DRINKING WATER AND H"LTH likely to be sufficient information on [the action of] mixtures.... Con- sequently, estimates will . . . have to be based on an . . . assumption of additivity" (NRC, 1980, p. 27~. The committee pointed out that assuming independence of action of a material in relation to background exposure instead of assuming additivity in dose could, at a low-dose level, easily lead to a 100-fold difference in estimated risk. In a recent review of the possible effects of exposure to a mixture of materials, Berenbaum (1985), arguing by analogy from the behavior of combinations of antibiotics, remarked, "It is therefore not unreasonable to assume that carcinogens will prove to behave similarly [i.e., will show synergism] . . . and it appears sensible to assume this until proved oth- erwise." Berenbaum also noted that the "effect of a marked antagonism is to produce a threshold in the curve." Reliability of Risk Estimation Considering all the variables encountered in the process of estimating risks, the 1980 committee remarked, "If the estimates of risk from low doses of carcinogens are made with reasonable data and reasonable models, [there will be] a precision of 1 or 2 orders of magnitude in the estimates" (NRC, 1980, p. 601. It has since been pointed out that maximum likelihood estimates (MLEs) are extremely sensitive to the data in that very small differences can lead to large differences in the MLEs (Cohn, 1986; T. W. Thorslund, EPA Office of Health and Environmental Assessment, Wash- ington, D.C., personal communication, 1985~. The 95% upper confidence limit estimates are much more stable. Noncarcinogens The 1980 committee noted that determination of no-effect levels for noncarcinogens depends upon both data interpretation and the number of animals in the bioassay. It stated, "The likelihood of observing a no- adverse effect at a given dose is statistically greater for experiments with few animals than for larger experiments" (NRC, 1980, p. 311. Thus, if these and other details are not described in published studies, the use of more formal (dose-response) risk-assessment procedures is impeded and it is more difficult to interpret whatever data are at hand. Among the important matters included are whether best-fit or 95% upper confidence limit curves should be used in expressing risk and whether the dose- response model, log-normal model, or log-logistic model should be applied to the dose-response curves. The 1980 committee concluded, "The po- tential utility of dose-response extrapolation methodology for noncarci-

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Risk Assessment 257 nogenic human risk assessment does exist but has been found to be of limited value for contaminants in drinking water" (NRC, 1980, p. 351. Crump (1984) has suggested an alternative to the no-observed-effect level (NOEL) for determining an acceptable daily intake (ADI). If it is possible to define as "acceptable" some very small increase over the response level in the control group, then a benchmark dose can be estab- lished as the lower (95% or 99%) confidence limit on an exposure that would produce small (or smaller) increases over the control level. Crump's proposal makes more effective use of all the experimental data than does the NOEL approach, taking into account the slope of the exposure-response curve and the size of the experiment. Thus, there is a greater efficiency in the use of the data, and larger experiments should, in general, lead to higher ADIs (rather than the reverse, which the 1980 committee noted is characteristic of the NOEL approach). This procedure is similar to the one used by the committee in developing a recommended range of ex- posures to aldicarb (see Chapter 91. The 1980 committee also provided a detailed discussion on the estab- lishment of suggested no-adverse-response levels (SNARLs) for acute 24- hour or 7-day exposures to noncarcinogenic materials. It based its cal- culations on assumptions that " 100% of the exposure to the chemical was supplied by drinking water during either the 24-hour or the 7-day period" (NRC, 1980, p. 681. To calculate chronic SNARLs, the committee ar- bitrarily assumed that drinking water provided 20% of the intake of the chemical of concern. Estimates of Exposure from Different Sources and Various Routes All sources of exposure must be considered when assessing risk or setting safety factors for estimating ADIs of chemicals in water, regardless of the biological end point under consideration. The population may be exposed to chemicals through air and food as well as through water. For some population subgroups and for certain toxic chemicals, the air and the workplace may be the major sources of exposure. For others, drinking water may be the primary source. The various routes of exposure from these sources must also be con- sidered. Systemic absorption resulting from simultaneous exposure via multiple routes has received little attention to date. Shehata (1984) pre- sented a modeling approach that may be useful in estimating the relative contribution of multiple exposure routes to body burdens of volatile or- ganics. However, further research is required to validate these multiple route models. The present committee finds it worthwhile to distinguish between dose response and exposure response. Dose response may be defined as the .

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258 DRINKING WATER AND H"LTH response to the dose of toxicant actually delivered to the target site. Generally, determination of dose response requires knowledge of the fate and distribution of the administered material, including pharrnacokinetic behavior and metabolic activation or deactivation. Exposure response is easier to determine, since information about exposure or nominal dose is more accessible and exposures are subject to potential regulation. Ade- quate understanding and elucidation of the mechanisms of toxicity will require knowledge of dose-response relationships. For regulatory pur- poses, an adequate description of the exposure-response relationship is more important. WATER For chemicals with a threshold dose, the ADI is based on animal tox- icology data by applying a safety factor to a NOEL. To determine the maximum exposure from water that should be allowed, contributions to total exposure made by sources other than drinking water should be sub- tracted from the permissible exposure (maximum permitted intake is usu- ally the ADI times assumed adult body weight of 70 kg). The maximum permissible level in water can then be determined from the permissible exposure through drinking water and the standard volume of daily water consumption, which is normally assumed to be 2 liters in a 70-kg adult male and 1 liter in a 10-kg child (Kelly, 1980; NRC, 1977~. Where the standard volume of consumption is set high in the range of that consumed in a population, an action level of exposure may be safely set without any increase in risk. In fact, actual water intake varies considerably with physical activity, environmental temperature, and relative humidity, which in turn vary by region of the country. At this time, however, the committee was unable to estimate the contributions of these other factors and sources other than drinking water and encourages the development of better data and models on these variables. Valid arguments may be constructed for setting the standard daily con- sumption volume for drinking at the population mean, the median, the ninth docile, or some higher standard volume. If set at the population mean or median, roughly half the population will be expected to exceed this level. This choice must be related to recommended maximum per- missible levels in water. Although ingestion is the chief route of exposure to chemical contam- inants in drinking water, inhalation and dermal exposure may also con- tribute to systemic absorption of waterborne contaminants. For example, large amounts of volatile organic chemicals may be inhaled from boiling water or hot showers. Since many organic compounds are poorly soluble

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Risk Assessment 259 in water and are quite volatile, a substantial portion of the contaminants in water may evaporate under certain usage conditions in the home, e.g., from washing machines, dishwashers, sinks, bathtubs, and showers. Dur- ing showers with chemically contaminated water, the combined action of the spray and high temperature could result in the generation of relatively high vapor levels in confined areas. Seasonal changes in household ven- tilation may also greatly influence the extent to which volatilized chemicals are retained in the household air. Brown et al. (1984) proposed that absorption through the skin may contribute significantly to the total dose of volatile organic compounds received during normal daily use of contaminated water. Their calculations were based on skin absorption rates for toluene, xylene, styrene, and ethylbenzene, which were measured in human exposure studies by Dut- kiewicz and Tyras (1967, 19681. The reliability of the models and the accuracy of such predictions must, of course, be verified through exper- imentation. For less volatile water contaminants, there are no definitive experimental data on inhalation and dermal exposures. Thus, doses from these routes of exposure cannot be reliably estimated at this time. It may be prudent to assume that in addition to the standard 2 liters of water consumed orally each day, daily exposure to another 2 liters results from inhalation and skin absorption during bathing, showering, cooking, washing, and other activities involving water usage. The adoption of this concept would, of course, have the effect of lowering to one-half the permissible concen- ~ation of a chemical in water. FOOD Traces of pesticides and other chemicals that contaminate drinking water also contaminate foods. Exposure through food (not including any con- tribution from the water used in cooking) varies considerably, depending on diet. The estimated tolerances to a pesticide or recommended action levels for other chemicals in foods may be used to calculate a theoretical maximum residue contribution (TMRC) through food (Ariens and Si- monis, 1982; Hathcock et al., 19834. These TMRC values constitute an upper limit on the exposure through food. They are difficult to use, how- ever, because they usually exceed the measured intakes by one to several orders of magnitude. Moreover, estimates of the proportion of the diet accounted for by specific foods are often not current and there is little information on variation in diets by region, ethnicity, age, sex, or other factors (NRC, 19821.

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260 DRINKING WATER AND HEALTH AIR The background level of exposure to many toxic chemicals in the am- bient air is usually even more variable and difficult to determine than the exposure through food. The extent of both indoor and outdoor exposures through air depends strongly on occupation, region (urban levels exceeding rural levels for many substances), climate, and life-style (e.g., whether sedentary or active, length of time spent indoors and outdoors) (NRC, 1985~. WORKPLACE In occupational settings, workers may be exposed to toxic chemicals by routes other than ingestion of water, including inhalation, direct contact with skin, or ingestion after contamination of hands, cigarettes, or food. Contaminants in ambient air and in settled dust may account for much of the total workplace exposure. Contaminants may come into contact with skin from clothing contaminated by vapors, dust, or spills. HOUSEHOLD The Total Exposure Assessment Measurement (TEAM) study conducted by the EPA (Pellizzari et al., 1984) has shown tremendous variation in personal indoor exposures to toxic chemicals. Until there is a better un- derstanding of the doses provided by all the different indoor sources of pollutants, estimates of average national exposures to drinking water con- taminants in the home will not be useful in determining total exposures to the chemical of interest. Upper limits of likely ranges of exposure may be more useful than the boundaries selected. PHARMACOKINETICS An understanding of pharmacokinetic principles is necessary to the successful extrapolation of data from high to low doses. (See Chapter 6 for a detailed discussion of models based on pharmacokinetics.) The Office of Science and Technology Policy (OSTP) (1985), elaborating on a con- clusion reached by Hoel (1980), noted, "Even if only a small portion of the background incidence [of the toxic responses is associated with the same mechanistic process as the study chemical, linearity will tend to prevail at sufficiently low doses" (OSTP, 1985, p. 81~. Some arguments to the contrary have been raised. An unresolved issue that may have a bearing on the possible existence of a threshold is the dissimilarity of the kinetics of toxification-detoxification at high in com-

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Risk Assessment 283 act via separate, distinctive, and well-defined mechanisms. However, the lines are increasingly blurred as substances previously characterized as nongenotoxic are being shown to alter the structure or sequence of the genetic material. For example, carbon tetrachloride is a weak hepatocar- cinogen (IARC, 1979), but it is extremely hepatotoxic and produces a dramatic increase in cell turnover as a result of regenerative growth (Mir- salis et al., 1982; Recknagel, 19671. Carbon tetrachloride has also been reported to be nonmutagenic in bacteria (McCann et al., 1975) and in mammalian cells (Dean and Hodson-Walker, 1979; Stewart, 1981), and it fails to induce DNA repair in the hepatocytes of treated animals (Mirsalis et al., 19821. Yet, other studies have shown that carbon tetrachloride binds to DNA, RNA (Cunningham et al., 1981; DiRenzo et al., 1982; Rocchi et al., 1973), and protein (Bolt and Filser, 1977) and induces mutation in yeast (Caller et al., 19801. Other substances that were at one time believed to be acting by so-called epigenetic mechanisms, including TPA, DDT, DES, asbestos, saccharin, and phenobarbital, have since been re- ported to exhibit genetic toxicity in some assays. A distinction has been made between primary genotoxicity and secondary- genotoxicity resulting from another activity. This distinction may only be academic, however, because the latter also damages the genome (Becker et al., 1981; Birnboim, 1982~. Furthermore, new information suggests that the promoter TPA has a memory effect lasting for at least 8 weeks (Furstenberger et al., 1983~. Thus, a lack of adequate testing or a limited knowledge of the mechanisms involved could lead to the misclassification of a substance (see Chapter 51. 2. Somatic cell mutation is not necessarily the most important mechanism in carcinogenesis, nor are genetic and nongenetic mechanisms mutually ex- clusive. Many concurrent factors (both genetic and epigenetic) may take part in He process of tumorigenesis. These include chromosome abnormalities, gene rearrangements, oncogene activation, disorders of differentiation, DNA damage, and disruption of DNA repair (see Chapter 51. 3. The threshold issue concerns the shape of He dose-response curve at increasingly small doses where little or no information is available. At present, we do not know He shapes of the low-dose curves or if Here is or is not a true threshold for an animal or human population for any carcinogen (Ehling et al., 1983~. There are no convincing data to support a nonlinear dose- response curve or threshold at very low doses for any carcinogen (Hoer et al., 1983; Weinstein, 19831. This information is not available even for DES and 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), two of He better-studied compounds on He conventional lists of epigenetic agents (Hertz, 1977; Wein- stein, 1983~. 4. Experimental data demonstrating that agents such as TCDD can be carcinogenic by themselves and may have greater carcinogenic potency

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284 DRINKING WATER AND H"LTH than many initiating agents contradict the implied assumption that epi- genetic agents carry lower risks. The striking reductions in human cancer risk following decreases in exposure to promoters, e.g., by cessation of smoking and by reduction or elimination of some estrogen therapy (Day and Brown, 1980), strongly support rigorous control of epigenetic agents. In general, quantitative biochemical information is not sufficient for low-dose extrapolation. For carcinogenic risk assessment, the data suggest that a multistage model is consistent with certain qualitative aspects of cancer biology. This model is attractive because for most experimental data, the curve becomes linear at low doses. However, the biochemistry also suggests that regulatory agencies should not be complacent about such a dose-response model, despite its simplicity and its apparent con- servative approach to extrapolation to low doses. The dose response may be fundamentally nonlinear at low doses, and a linear extrapolation may underestimate risk for certain individuals, species, or tissues. Even more importantly, basic biochemical rate concepts suggest that when experi- mental dose-response data are accumulated at doses near the maximal tolerated dose, carcinogen activation, detoxification, and repair pathways may become saturated. Under those circumstances, measured dose- response data might not contain the information required to make a low- dose extrapolation, even if the precise mathematical relationship between dose and response were known independently. In light of such considerations, a generalized multiparameter-fitting protocol may be a reasonable mechanism for generating a low-dose ex- trapolation. However, it is impossible to determine if this is a consistently conservative procedure, which is the type most generally favored by reg- ulatory agencies. The agencies prefer the risk-assessment approach with the greatest potential for protecting human health, i.e., treating all car- cinogens in a similar manner (EPA, 1976; NTP, 1984; Perera, 1984; Weinstein, 1983~. Although it may not be possible to use mechanistic data by extrapolation at this time, the committee hopes that as our un- derstanding of the carcinogenic process increases so will our ability to make better risk assessments. For now, any information on mechanisms of cancer induction that bears on the risk-assessment process should at least be noted by those doing the evaluation. Developmental and Reproductive Toxicity Assessing developmental and reproductive toxicity is especially com- plex due to the great variety of possible toxic end points and the likely involvement of threshold doses. At present there is limited agreement about how to apply the results of animal reproductive and developmental

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Risk Assessment 285 toxicity studies to assess the risk of exposure to a compound. This is partly because an understanding of the underlying events leading to re- productive toxicity is usually missing. There are no agreed-upon standards or quantitative methods for cross-species extrapolation. Nonetheless, there are some conditions under which the committee believes that reproductive and developmental toxicity data could be used to estimate risk in humans. When there are sufficient data, the NOEL approach is the most reasonable approach at this time (EPA-ORNL, 19821. It is usually easier to identify a LOEL from experimental data because this value can be observed di- rectly, whereas the NOEL can be orders of magnitude below the lowest experimental exposure level observed to induce a toxic effect. LOELs are most accurately identified under conditions where the response is minimal and the end point involves reversible effects, indicating that a NOEL is being approached. There are several considerations in the selection of an appropriate safety factor to be used with NOELs or LOELs for reproductive toxicity. The committee recommends that humans should be considered to be at least 50 times more sensitive than laboratory animals to agents causing well- defined developmental and reproductive toxicity. Substances that lead to developmental toxicity at levels lower than those causing maternal toxicity constitute a greater potential hazard than substances that cause develop- mental toxicity only at maternally toxic doses. The size of the safety factor should reflect the potential hazard and should consider not only the severity of the response but also the time and route of exposure. The greatest potential hazards are presented by substances causing serious effects under conditions of exposure that may be encountered by humans. Existing models for quantitative risk assessment have not been suffi- ciently well developed to be applied to reproductive toxicity data. How- ever, some encouraging work in progress may result in the production of acceptable models for some types of data. The advantage of modeling reproductive toxicity data over the use of NOELs is that the modeling approach takes into account the slope of the exposure-response curve and the size of the experiment. An example of modeling reproductive data is shown in Chapter 9, where tolerance distribution models were applied to the developmental toxicity data for nitrofen. When insufficient data are available for the NOEL approach or possibly the modeling approach, a ranking system or quantitative index may be used. Underlying this approach is the need to distinguish between sub- stances that are uniquely toxic to the embryo and those that induce de- velopmental toxicity at exposure levels that are also toxic to the mother. Agents in the latter category should be regulated on the basis of their adult toxicity, whereas those in the former would be regulated on the basis of their unique toxicity to the embryo.

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286 DRINKING WATER AND H"LTH Neurotoxicity In the four-step process concluding with a risk assessment described by a National Research Council committee (NRC, 1983), the work eval- uating neurotoxicity appears to present serious difficulties that are a natural consequence of the complexities surrounding neurotoxicity. The many neurotoxic effects that can be induced by different exposures range from barely perceptible sensory deficits to gross behavioral or functional ab- normalities. With increasing dose and duration of exposure, one specific effect may be manifested in a larger and larger proportion of the popu- lation, or the number of affected people may not increase but the effects may become more and more serious and incapacitating. Intense short- term effects may occur, but no residual effects may be detectable after exposure ends. At the other extreme, an exposure may produce no ob- servable consequences yet may leave the exposed person highly vulnerable to a subsequent exposure to the same, or even an unrelated, neurotoxicant. Nutritional status has been found to strongly affect responses. A very large number of substances are known to produce neurotoxic effects in humans, in contrast to cancer, for which 30 causative materials or industrial pro- cesses have been implicated (IARC, 19821. Because of the many neurotoxic end points, implying many different mechanisms of action, there are essentially no general mathematical mod- els of neurotoxicity leading to quantitative risk assessment. Models may have to be constructed on a material-by-material basis. In some evaluations for lead, levels of exposure were related to necrologic effects, many of which can be associated with hemolytic markers (Table 4-21. The difficulty in measuring exposures is illustrated by the need to use secondary markers of exposure such as blood lead levels. Regulatory actions are usually based on what have been identified as nominal exposure levels that are objective, measurable external quantities such as parts per million in an air sample. Exposure to these nominal levels, of course, is not the same as the dose, at least not in the biological sense. Lead found in the blood can be regarded as an internal dose (the amount of the substance or its active metabolites in body tissues) or as the biologically effective dose (the amount of the active material that interacts with the tissue or organ). These biological doses are rarely measured; very likely vary with age, sex, and genetic background; and at times may even be unmeasurable. Because quantitative dose-response models or even adequate measures of exposures do not exist, safety factors must be used. A NOEL or a LOEL must be identified and the observed doses divided by an appropriate safety factor. The intensity or seriousness of the response, age-sex vari-

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Risk Assessment 287 ation, experiment size, and similar factors need to be taken into account in setting the appropriate safety factor. Use of these safety factors is intended to lead to exposure levels that will be safe for the most sensitive individuals. In general, however, they are usually determined on a maters by-mater~al basis and do not take into account aspects of bioac- cumulation, sensitization, or multiple exposures all factors that need specific consideration. The committee's approach to a quantitative estimate of a safe exposure level through a modified LOEL approach is described for the cholinesterase inhibitor aldicarb in Chapter 9. The approach is based on the argument that if no clinical manifestation of cholinesterase inhibition is observed unless cholinesterase levels are reduced by at least 20% or 30%, then the lower 95% confidence limit on the dose that produced such an inhibition could be looked upon as the acceptable, or maximum permissible, ex- posure level. To assess risk for neurotoxic end points, research must be conducted to develop measures of exposure, including biological markers; better laboratory techniques, including short-term tests for identifying neurotox- icants; and quantitative models for low-dose extrapolation, reflecting the different types of effects and species-to-species variability. Studies must also be undertaken to learn whether or not s~ucture-function relationships can be predictive. The effects of interactions and modified (or sporadic) exposures must also be examined. Furthermore, the relationship of mor- phological changes to neurological or neurotoxic responses should be explored. REFERENCES Anderson, E. L., and CAG (Carcinogen Assessment Group of the U.S. Environmental Protection Agency). 1983. Quantitative approaches in use to assess cancer risk. Risk Anal. 3:277-295. Ariens, E. J., and A. M. Simonis. 1982. General principles of nutritional toxicology. Pp. 17-80 in J. N. Hathcock, ed. Nutritional Toxicology. Vol. 1. Academic Press, New York. Armitage, P., and R. Doll. 1954. The age distribution of cancer and a multi-stage theory of carcinogenesis. Br. J. Cancer 8:1-12. Becker, R. A., L. R. Barrows, and R. C. Shank. 1981. Methylation of liver DNA guanine in hydrazine hepatotoxicity: Dose-response and kinetic characteristics of 7-methylguanine and O6-methylguanine formation and persistence in rats. Carcinogenesis 2:1181-1188. Berenbaum, M. C. 1985. Consequences of synergy between environmental carcinogens. Environ. Res. 38:310-318. Birnboim, H. C. 1982. DNA strand breakage in human leukocytes exposed to a tumor promoter, phorbol myristate acetate. Science 215: 1247-1249.

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288 DRINKING WATER AND H"LTH Bolt, H. M., and J. G. Filser. 1977. Irreversible binding of chlorinated ethylenes to macromolecules. Environ. Health Perspect. 21: 107-112. Bresnick, E., M. Brosseau, W. Levin, L. Reik, D. E. Ryan, and P. E. Thomas. 1981. Administration of 3-methylcholanthrene to rats increases the specific hybridizable mRNA coding for cytochrome P-450c. Proc. Natl. Acad. Sci. USA 78:4083-4087. Brown, H. S., D. R. Bishop, and C. A. Rowan. 1984. The role of skin absorption as a route of exposure for volatile organic compounds (VOCs) in drinking water. Am. J. Public Health 74:479-484. Calabrese, E. J. 1983. Principles of Animal Extrapolation. John Wiley & Sons, New York. California Department of Health Services. 1982. Carcinogen Identification Policy: A State- ment of Science as a Basis of Policy. Section 2: Methods for Estimating Cancer Risks from Exposures to Carcinogens. State of California Department of Health Services, Sacramento, Calif. [88 pp.] Callen, D. F., C. R. Wolf, and R. M. Philpot. 1980. Cytochrome P-450 mediated genetic activity and cytotoxicity of seven halogenated aliphatic hydrocarbons in Saccharomyces cerevisiae. Mutat. Res. 77:55-63. Carlbort, F. W. 1982. Speculations on an extended dose-response model for carcinogenesis. Food Chem. Toxicol. 20:319-323. Cohn, M. S. 1986. Estimated carcinogenic risks due to exposure to formaldehyde released from pressed wood products. Tab E in February 1986 Briefing Package to the Com- missioners of the Consumer Product Safety Commission on Formaldehyde Emissions from Urea-Formaldehyde Pressed Wood Products. [53 pp.] (Available from Consumer Product Safety Commission, Washington, D.C. 20207.) Consensus Workshop on Formaldehyde. 1984. Report on the Consensus Workshop on Formaldehyde. Environ. Health Perspect. 58:323-381. Crouch, E., and R. Wilson. 1979. Interspecies comparison of carcinogenic potency. J. Toxicol. Environ. Health 5:1095-1118. Crump, K. S. 1984. A new method for determining allowable daily intakes. Fund. Appl. Toxicol. 4:854-871. Crump, K. S. 1985. Mechanisms leading to dose-response models. Pp. 235-277 in P. Ricci, ed. Principles of Health Risk Assessment. Prentice-Hall, Englewood Cliffs, N.J. Crump, K. S., D. G. Hoel, C. H. Langley, and R. Peto. 1976. Fundamental carcinogenic processes and their implications for low dose risk assessment. Cancer Res. 36:2973-2979. Cunningham, M. L., A. J. Gandolfi, K. Brendel, and I. G. Sipes. 1981. Covalent binding of halogenated volatile solvents to subcellular macromolecules in hepatocytes. Life Sci. 29: 1207-1212. Day, N. E., and C. C. Brown. 1980. Multistage models and primary prevention of cancer. J. Natl. Cancer Inst. 64:977-989. Dean, B. J., and G. Hodson-Walker. 1979. An in vitro chromosome assay using cultured rat-liver cells. Mutat. Res. 64:329-337. DiRenzo, A. B., A. J. Gandolf~, and I. G. Sipes. 1982. Microsomal bioactivation and covalent binding of aliphatic halides to DNA. Toxicol. Lett. 11:243-252. Doull, J., B. A. Bridges, R. Kroes, L. Golberg, I. C. Munro, O. E. Paynter, H. C. Pitot, R. Squire, G. M. Williams, and W. J. Darby. 1983. The Relevance of Mouse Liver Hepatoma to Human Carcinogenic Risk. A Report of the International Expert Advisory Committee to the Nutrition Foundation. The Nutrition Foundation,~Inc., Washington, D.C. 34pp.

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