Chapter 4

Uses and Limitations of Risk Assessment for Risk Management Decision-Making

Risk assessment is the systematic, scientific characterization of potential adverse effects of human exposures to hazardous agents or activities. Risk assessment as an organized activity of the federal agencies began in the 1970s. Earlier, the American Conference of Governmental Industrial Hygienists had set threshold limit values for exposures of workers, and the Food and Drug Administration (FDA) had set acceptable daily intakes of pesticide residues and food additives in the diet. In the mid-1970s, the Environmental Protection Agency (EPA) and FDA issued guidance for estimating risks associated with low-level exposures to potentially carcinogenic chemicals. Their guidance made upper-bound estimated risks of one extra cancer over the lifetime of 100,000 people (EPA) or 1 million people (FDA) action levels for regulatory attention. Estimated risks below those levels are considered negligible because they individually add so little to the background rate of about 240,000 cancer deaths per 1 million total deaths in the United States. The ultimate goal is, of course, to lower the background rate itself, a part of which can be attributed to an array of pollution-generating activities.

During 1977-1980, an interagency regulatory liaison group was actively engaged in bridging scientific, statutory, and policy responsibilities and activities of EPA, FDA, the Occupational Safety and Health Administration, the Consumer Product Safety Commission, and the Food Safety and Quality Service of the Department of Agriculture. The White House Office of Science and Technology Policy participated in the scientific discussions supporting risk assessment and risk management and published a scheme for identifying potential hazards, characterizing risks, and managing the risks, usually by reduction of use, emissions, or exposures (Calkins et al. 1980) (see Table 4.1).

That scheme makes clear that information about potential hazards can come from epidemiologic studies of workers and other people who are exposed to hazards, from direct experimental tests in animals and in cells in the laboratory, and from comparisons of chemical structures. The next stage involves the potency of the chemical (dose-response relationship), detailed understanding of exposure pathways, and the reasons for variation in responses among exposed people. Risk, then, is characterized both qualitatively (the nature of effects, the strength of evidence, and the reversibility or preventability of effects) and quantitatively (the probability of effects of various kinds and severities).

Performing full-scale risk assessments is a formidable task, requiring data, technical expertise, and peer review. Deciding to go forward with a risk assessment is a risk-management decision, and scaling the effort to the importance of the problem, with respect to scientific issues and regulatory impact, is crucial.

This section examines some of the risk assessment issues that are under debate, such as assessing toxicity and relevance to humans, accounting for variations in population exposures and susceptibility, describing uncertainties, evaluating risks of chemical mixtures, conducting ecologic risk assessments, and assessing risks associated with microorganisms and radiation.

Toxicity Assessment

Basing risk management decisions on observations and assumptions about the potential human toxicity of chemical exposures presents many challenges. The nature and magnitude of a population’s exposures to chemical contaminants generally must be extrapolated from a few data on samples obtained from the contaminated sources (see Exposure Assessment on page 72). The nature of chemical hazards and the relationships between exposures and effects often must be extrapolated to humans from toxicity tests in laboratory animals. In many cases, observations made using high doses in the laboratory or from high exposure levels in the workplace must be extrapolated to much lower environmental levels of human exposure. Extrapolating among species requires scientific information that can be used to make predictions about the relevance of a substance’s toxicity in laboratory animals to human risk. Because the results of standard toxicity tests alone often do not provide enough information to make well-informed qualitative judgments about human relevance, testing strategies that rely on mechanism-based tests to evaluate substances’ toxicity and carcinogenicity have been developed. Information about chemicals’ modes of action can make important contributions to scientifically based human health risk assessment.

This section evaluates three issues: the use of detailed toxicity information to assess the relevance of rodent bioassay results to human cancer risk, the need for more toxicity testing of chemical mixtures and ways to evaluate their risks, and the need for risk assessments to consider information about variation in susceptibility to toxic effects.

Using Rodent Tests To Predict Human Cancer Risk

Finding

Chemicals that cause cancer in rodents are appropriately considered potentially carcinogenic in humans. Investigations of chemicals’ mechanisms of action can greatly strengthen the link between findings in rodents and likely effects in humans. They can also provide biological plausibility for statistical associations in epidemiologic studies. However, some chemicals elicit tumors in rodents only through mechanisms or at doses that have been clearly demonstrated to be very different from mechanisms and exposures in humans. Regulatory agencies have been cautious in recognizing the distinctions and in issuing guidance on when such rodent responses should be discounted or disregarded.

Recommendation

 In general, tumors and other adverse effects observed in properly conducted animal bioassays should be considered predictive of similar effects or risks in humans. Chemicals found to elicit such effects should be regulated accordingly. If after adequate testing a chemical is found to produce only tumors that occur as a result of mechanisms or doses that have been clearly demonstrated to be not relevant to humans, that chemical should not be regulated as a carcinogen and should not require extensive risk assessment. Regulatory agencies should distinguish between tumor responses that are predictive and those that are not (see Table 4.2), and these judgments should be updated with advances in scientific knowledge about the underlying mechanisms.

The policy of presuming that a chemical that causes cancer when tested in laboratory rodents is potentially carcinogenic in humans is justified by considerable evidence and by the precautionary principle of being protective when uncertain. Rodent bioassays have played an important role in identifying human carcinogens numerous times. All 23 recognized human carcinogens are also carcinogenic in laboratory animals; for 18 of those, cancers occured in one or more organ sites in humans that are the same as those identified in the animal studies (see Table 4.3) (Rall 1988). There are other cases, however, where rodent tumor responses have been shown to be irrelevant to humans or may occur at doses far exceeding any recognized human exposures including workplace exposure. The Delaney clause prohibits chemicals that have been identified as carcinogens in rodents from being used as food additives, reguardless of whether the effects they produce are relevant to human carcinogenicity;other statutes permit scientific judgment.

From a risk management perspective, it is wasteful to expend limited risk assessment resources, risk management time, and public and legal involvement revisiting the issue of human relevance of the specific rodent response chemical by chemical. Of course, the evidence for hazard identifecation, exposure levels, and other effects must be evaluated for each chemical. Table 4.2 lists examples of rodent mechanisms and tumor responses that are candidates for classification as "not likely" to be predictive of carcinogenicity in humans according to EPA’s Proposed Guidelines for Carcinogen Risk Assessment (EPA 1996b). That classification includes a subcategory of agents that elicit only rodent tumors that are irrelevant to human risk and another of agents that produce tumors at doses and via routes of exposure that need to be compared with known human occupational and general population exposures to determine relevance to human risk. Chemicals that produce tumors only in rodents because of striking pharmacokinetic differences can also be addressed. In general, the chemicals listed in Table 4.2 are not genotoxic; that is, they do not react directly with DNA. Instead, they cause local injury or otherwise stimulate local hyperplasia and cell division, which is associated with a low incidence of tumor formation.

For example, some chemicals are recognized to induce the accumulation of large amounts of a-2u globulin protein in the male rat kidney. Most scientists agree that this accumulation leads to damage to the kidney tubules, cell death, sustained cell proliferation, and tumor formation. Some scientists do not agree (Melnick et al. 1996). This response is not believed to occur in female rats or in other species, including humans. After 4 years of extensive study and review by EPA’s Risk Assessment Forum and Science Advisory Board, the agency decided to disregard that particular rodent response for certain chemicals (EPA 1991). If that response is disregarded, risk assessment and regulation can be directed, as appropriate, at any other adverse effects, including kidney tumors not due to this protein-mediated mechanism.

Another tumor response that is believed to be irrelevant to humans is that which occurs only in the rodent forestomach after administration of a chemical by gavage (that is, via a tube placed in the stomach). Gavage is convenient for determining whether a chemical can cause tumors in organs distant from the stomach after absorption into the bloodstream, but can result in local cytotoxicity and hyperplasia. At least three commercially important chemicals (Table 4.2) have been found to produce tumors only in the forestomach and only following gavage. For example, butylated hydroxyanisole (BHA) was reviewed for FDA by a Federation of American Societies for Experimental Biology panel, which concluded in 1994 that there is a threshold for its tumor-producing cell proliferation. There is no evidence of a similar effect in humans (who lack forestomachs) and no scenario in which similar high dose local exposure would occur.

The saccharin debate of 1978-1979 highlighted rodent bladder tumors. An International Life Sciences Institute panel on rodent bladder carcinogenesis ultimately concluded that chemicals that precipitate in urine, or that elicit effects leading to precipitation of other chemicals, should be considered carcinogens only at high doses (Neumann and Olin 1995). If human exposures to such chemicals are much lower than the doses tested, the rodent response can be disregarded. Of course, bladder tumors can arise by other mechanisms that are relevant to human cancer.

Grossly overloading the rat lung’s clearance mechanisms by administering particles directly to the lung has also been considered irrelevant to humans (Oberdörster 1995). EPA delisted titanium dioxide from the Toxic Release Inventory in 1988 for this reason (Fed Reg 53:23107-23202, 1988). The phenomenon may be applicable to particles in general, not only to titanium dioxide, but it has been declared irrelevant to humans only in the case of titanium dioxide. Declaring responses to other particles as not likely to predict human cancer risk would require criteria to determine what are "gross" particle overloads. Ultrafine particles (<0.1 microns) may well present a risk at much lower concentrations. Particles may also be carriers of hazardous chemicals that have adsorbed to them.

High doses of several pesticides and fungicides induce liver enzymes or thyroid enzymes that affect thyroid hormone levels, leading to hyperplasia and ultimately to thyroid tumor formation in rodents. Because the feedback and transport systems for rodent thyroid hormones are very different from those in humans (McClain 1994), many believe that humans are far less sensitive to this response. EPA still assesses rat thyroid data on a case by case basis.

Finally, there have been many challenges to the interpretation of mouse liver tumor formation (not listed in Table 4.2). At least six potential mechanisms have been described, some of which occur in humans. Mouse liver tumors are among the most common seen in bioassays and pose particularly vexing problems for interpreting effects of chlorinated organic solvents.

Judgments about the likelihood of a chemical’s or a tumor’s human relevance should include careful evaluations of the weight of the scientific evidence. Some considerations include:

• Adequacy of experimental design and conduct.

• Occurrence of common versus rare tumors.

• Progression, or lack thereof, from a benign to a malignant tumor.

• Latency until tumor induction.

• Dose-response relationships.

• Genetic toxicity.

Toxicity testing protocols used to evaluate a chemical’s carcinogenicity are a subject of intense debate. Leading toxicologists are eager to substitute newer tests for at least one of the two rodent species generally used in standard lifetime cancer bioassays. These newer tests employ newborn mice, which are quite sensitive and yield results in a few months, and specially developed transgenic mice with mutant p53 genes or other cancer-predisposing genes to make the mice more sensitive and provide mechanistic information. The goals are to apply scientific advances, get more information, and hopefully do so at lower cost and in less time.

Bringing a risk management perspective to the scientific review process might galvanize action. EPA reviews of the male rat kidney and rat thyroid tumor responses have required many years. The Commission recognizes that time is required to investigate chemicals’ modes of action and endorses EPA’s current plans to identify tumor responses in rodents that are not likely to be relevant to humans. We encourage EPA to apply those distinctions as early as possible in the risk assessment process, before time and resources are wasted. Other agencies should follow similar practices.

Evaluating Chemical Mixtures

Finding

Humans are exposed to many chemicals and other potentially toxic agents in the environment, but toxicity testing and regulations generally focus on one chemical at a time, often just in air, water, or food. Most risk assessments evaluate individual chemicals and then combine them by simple addition to estimate risk related to chemical mixtures. This method ignores potential synergistic or antagonistic interactions that could lead to under or overestimation of total risk, respectively. Knowledge of mechanisms of action can guide judgments of whether risks related to combinations of particular chemicals will be additive or independent.

Recommendation

 Toxicity testing of complex environmental mixtures of regulatory importance should be performed for hazard identification and to generate comparative potency estimates of human risk. For risk assessments involving multiple chemical exposures at low concentrations, without information on mechanisms, risks should be added. If the chemicals act through separate mechanisms, their dose-response relationships should be considered separately.

As commonly practiced today, risk assessment and risk management consider exposures and risks in isolation from one another, typically chemical-by-chemical. For example, risks associated with air pollution are not put into the context of concurrent risks associated with contaminated drinking water or foodborne pesticide contamination. That fragmented approach to risk characterization is mostly a result of the fragmentation of responsibilities of different regulatory agencies and programs, but it can also be attributed to the limitations in our knowledge of the interdependence of different risks.

Failure to account for multiple and cumulative exposures is one of the primary flaws of current risk assessment and risk management, according to testimony received from Michael McCloskey, chairman of the Sierra Club, and others. Many people are surprised to learn that scientists usually do not test mixtures and that risk assessors and managers do not even try to account for the full array of exposures and health (or ecologic) risks. If the Framework is implemented and experience with testing and evaluating multiple chemical risks increases, it should be feasible to move beyond fragmentation. A promising new statute, the Food Quality Protection Act of 1996, requires estimates of aggregate, cumulative, and combined exposures to pesticides; some 9,000 tolerances for registered pesticides will need to be reassessed under this new mandate during the next 10 years.

Toxicity testing

Many complex mixtures—such as automobile exhaust, cigarette smoke, and other combustion products—have hundreds or thousands of chemical components. Attempting to identify and characterize each component and then adding their risks is clearly impractical. In those cases, the mixtures themselves can be tested for toxicity and their risks can be characterized on the same basis. For example, toxicity studies of diesel exhaust and other emissions have been conducted by the Health Effects Institute, jointly supported by EPA and motor vehicle manufacturers. The valuable results of those studies and others, such as tests of smoggy air from the Los Angeles basin, encourage us to recommend the testing of other important chemical mixtures.

Predicting a complex mixture’s toxicity or risk can be assisted by testing it in bioassay systems and comparing the results with those from similar mixtures of known toxicity or risk. Bioassays that might be useful for testing mixtures could range from mutation tests in microorganisms to evaluation of effects on organs in culture or short-term tests of rodent respiratory function. A validated database of methods, bioassays, and biologic markers of effect and knowledge of the behavior of known mixtures in those bioassays will be needed to facilitate risk predictions for environmental mixtures. Such whole mixture testing could be considerably less expensive to perform than routine monitoring by chemical analysis for over 100 drinking water contaminants, for example, and might provide results that can be more easily extrapolated to human toxicity and discussed with stakeholders. The index of biotic integrity (see Ecological Risk Assessment on page 77) is another example of the use of a bioassay to integrate effects of numerous chemical exposures.

The experimental and epidemiologic database available for generating estimates of comparative potency of mixtures is not large. Most work has been applied to predicting lung cancer risks; for example, epidemiologic data are available on the carcinogenic potencies of coke oven emissions, coal roofing tar, coal smoke, aluminum smelters, and cigarette smoke. The human cancer risks of those emissions have been characterized and compared with their potencies in experimental systems to estimate the risks associated with mixtures that lack epidemiologic data, including automotive emissions (diesel and gasoline), woodstove emissions, residential oil furnace emissions, and ambient air particles; it is assumed that the relative carcinogenic potencies observed in experiments would be similar for humans (Harris 1983, Lewtas 1993).

Enlarging the toxicity database for complex mixtures would be facilitated by coordinated research programs among epidemiologists, toxicologists, and clinical investigators (Mauderly 1993). For example, epidemiologists could provide information on the types of mixtures to which humans are exposed, patterns of exposure, populations of concern, health effects of concern, and the level of effects observed (or observable). Clinical studies could provide information on short-term responses and dose-response relationships, biological markers revealing short-term exposures and effects, and the likelihood of sensitive subpopulations. And toxicologists could provide judgments about the biological plausibility of the suspected exposure-response relationsip, the potential for chronic disease resulting from repeated exposures, causal and predictive relationships betweeen acute and chronic effects, identity of active constituents of mixtures, and effects of the exposure patterns.

Complex mixtures seemingly from the same source can vary considerably. For example, neither automobile engines nor gasolines are identical, so automobile exhaust is likely to vary substantially among sources and over time. The composition of air pollution varies with time of day and time of year, not to mention geographic location and source, so the toxicity of such mixtures is likely to vary considerably. Probabilistic approaches to describing the variability of composition within a class of mixtures and the relationship between that variability and toxicity should be explored. Coupling mathematical/statistical modeling (e.g., Monte Carlo techniques and physiologically based pharmacokinetic/pharmacodynamic dosimetry) with mechanistically based short-term toxicology studies may prove useful (Yang et al. 1995).

Assessing risks from multiple chemicals

Most of the information that is available on interactions among chemicals comes from human occupational studies and from rodent bioassays. Those studies generally evaluate doses that are much higher than the low, environmental doses commonly encountered. Interactive effects (either synergistic or antagonistic) depend heavily on dose; therefore, characterizing interactions that occur at one set of doses (such as those used in a rodent bioassay) is likely to provide very little information about interactions at very different doses (such as those generally encountered in the environment). "High" doses for combined effects are defined as those at which statistically significant increases in detrimental outcomes are observed in either laboratory or occupational studies. For the most part, exposure to chemical mixtures in the environment occurs at "low" doses—typically, one thousandth (or less) of the doses at which toxicity is observable in rodent bioassays or in epidemiologic studies of highly exposed workers. The ratio of exposures observed to cause adverse effects and actual human exposures is called the margin of exposure (EPA 1996b) (see Need for a Common Metric on page43).

The combined effects of exposure to chemicals in a mixture are determined by how individual components of the mixture affect the biological processes involved in toxicity. Components of a mixture can affect biological processes in many ways. For example, anything that affects the absorption, distribution, metabolism, or elimination of a chemical will affect the amount of that chemical that is available to react with DNA or other cellular targets. Because interactions leading to synergism or antagonism are the result of reactions of many molecules at many cellular sites, a mathematical dose-response model of a synergistic or antagonistic response that depends on such mechanisms is most likely nonlinear at low doses. Such logic strongly suggests that any disease process that depends on such interactions is only marginally important at low exposure levels. Only at high doses of one or more mixture components—such as cigarette smoke, alcohol, and some substances in occupational exposures—is the combined effect likely to be detectably greater than the sum of the individual effects. For example, occupational exposure to asbestos is associated with a mortality ratio for lung cancer of up to 5 (that is, in comparison to persons not occupationally exposed to asbestos) and smoking with a mortality ratio for lung cancer of about 10; but asbestos workers who smoke have a mortality ratio for lung cancer of 50, not 15. Similarly, the risk of liver cancer associated with aflatoxin is increased markedly by hepatitis B virus infection.

The National Academy of Sciences report Complex Mixtures (NRC 1988) also concluded that effects of exposures to agents with low response rates usually appear to be additive. The experimental evidence that can be used to infer effects at low doses appears to support the assumption that low dose additivity does not underestimate, and in most cases probably overestimates, risk (see, for example, Ikeda 1988).

When the individual components of a chemical mixture exhibit different kinds of toxicity or have different biological mechanisms of toxicity, they do not interact—they act independently at low doses. In that case, the dose-response relationships for each chemical should be considered independently. For example, if the chemicals of concern at a Superfund site are copper, a gastrointestinal toxicant; lead, a developmental toxicant; and heptachlor, a neurologic toxicant, their toxicity should be evaluated independently and not combined into a single "noncancer" risk estimate. Experiments have shown that when groups of unrelated chemicals with unrelated targets of toxicity were administered to rodents simultaneously at doses equal to their separate NOAELs, no cumulative effects were observed; each chemical acted independently (Jonker et al. 1990, Groten et al. 1994). The same is true of groups of chemicals with the same target but different mechanisms of action (Jonker et al. 1993); studies in which similar chemicals with similar mechanisms and targets were administered simultaneously indicate that antagonism, is the usual outcome (Falk and Kotin 1964, Schmähl et al. 1977).

Accounting for Differences in Susceptibility

Finding

Genetic, nutritional, metabolic, and other differences make some segments of a population more susceptible than others to the effects of a given exposure to a given chemical; however, current regulatory approaches for reducing risks associated with chemical exposures generally do not include information on differences in individual susceptibility or encourage gathering evidence to identify them. In the absence of specific information about differences in susceptibility, risk assessments rely on assumptions and safety factors that are presumed to be protective of sensitive individuals.

Recommendation

Risk assessments should include consideration of genetic and other host differences in susceptibility, recognize the spectrum of interindividual variations within normal populations, and identify subpopulations especially susceptible to specific chemical exposures. Available information on the range of a population’s susceptibility should be considered and used in place of assumptions. Where appropriate, knowledge of differences in susceptibility should be used to support additional bright lines for risk to protect especially susceptible subpopulations (see Bright Lines for Risk Management on page 54) and to tailor specific risk management actions to protect those subpopulations.

Susceptibility to the effects of chemical exposures depends on the sensitivity of a person’s response to different doses. Susceptibility is influenced by many factors, including age, sex, genetic variation in metabolism of chemicals, genetic variation in response to agents or stressors at their sites of action, ethnic origin and ethnic practices, socioeconomic status, geographic location, and lifestyle factors, such as smoking, alcoholic beverage consumption, diet, physical activity, and recreational habits. Dose-response relationships are chemical-specific and depend on a chemical’s mode of action; people are not hypersusceptible to all kinds of exposures (Omenn 1982). The influence of concurrent exposures on risk is discussed in "Identifying Highly Exposed Populations" on page 75. The following are examples of subpopulations potentially at higher risk.

There are opportunities to identify, evaluate, and reduce risks to sensitive people. Asthmatics, for example, make up 5 to 10 percent of the general population in the United States. Some air pollutants, especially sulfur oxides, particles, and ozone, are respiratory irritants that pose a greater risk to this subpopulation than to the general public. Both the number of cases of asthma and the number of deaths from asthma are increasing in the United States. Blacks have a 15% higher prevalence of asthma than whites. Likewise, susceptibility to lung cancer appears to vary among ethnic groups; in the United States, the incidence of lung cancer in black men is 1.5 times that in white men, 2.5 times that in Hispanic men, 2 to 4 times that in Asian men, and 8 times that in American Indian men (NCI 1984). One source of individual and ethnic differences in susceptibility is differences in the activity of enzymes that affect chemical toxicity. Increased risks of cancers of the bladder, skin, colon, lung, and stomach have been associated with differences in the activity of specific enzymes that can activate or deactivate carcinogens. Susceptibility to organophosphate pesticide toxicity is also markedly influenced by the activity of a specific enzyme in the blood. Metabolism however is only one of many contributors to an individual’s susceptibility .

Amendments to the Safe Drinking Water Act and to the Federal Insecticide, Fungicide and Rodenticide Act require such recognizable subpopulations as the elderly, children, and women of child-bearing age to be identified and considered more explicitly than they are currently in risk characterization and in standard-setting. The Food Quality Protection Act of 1996 requires an additional safety factor of 10 be used when pesticide risks are assessed, to allow for children’s greater intake on the basis of body weight and potentially greater susceptibility, unless data are sufficient to justify a different safety factor. Recognition of subgroup susceptibility does not necessarily result in more stringent regulation, however. For example, people allergic to particular chemicals or pet animal proteins might modify their exposures or modify their responses (with medication). Identifying the size of the population at higher risk and describing the risk peculiar to that population during risk characterization, perhaps using biologic markers of susceptibility, will make it possible to characterize risks more realistically than is possible using only estimates for the general population. Risk communication messages can then be targeted more effectively.

Exposure Assessment

Exposure assessments can be simple or complex, depending on the needs of a particular risk management question. They are based on measurements, models, and assumptions, and generally focus on individual chemicals, media, and sources. Often, unvalidated mathematical models are used to make predictions about a population’s exposure on the basis of limited information on chemical contamination and assumptions about the population. The results oversimplify actual exposure magnitudes and conditions, in part to allow for population variability. And the methods generally do not consider other sources of exposure to the same or similar chemicals and their interdependence.

This section recommends ways to generate credible and understandable exposure information for informed decisions by risk managers and the public about the need for risk reduction. The Commission recommends that agencies show a preference for actual exposure data on communities and populations at risk.

Design of Exposure Assessments To Meet Risk Management Goals

Finding

Exposure assessments vary greatly in design and content. Complex risk management decisions often are based on simplistic, deterministic estimates of exposure derived from few data, many assumptions, and inadequately validated models. In contrast, some exposure assessments are more complex than is needed for straightforward risk management decisions.

Recommendation

Exposure assessments should be designed to be commensurate with the needs of the risk management decisions at issue. The design of an appropriate exposure assessment should take place at the problem/context stage of the risk management process.

Several measurement tools, statistical methods, and other procedures and considerations can be used to design and conduct an exposure assessment. No method or group of methods should be used in all cases. Selection of appropriate methods should be discussed and evaluated during the planning stages of a risk management process (the problem in context stage of the Commission’s Risk Management Framework) to ensure that they meet the needs and expectations of risk managers and other stakeholders. The following general principles are suggested as the planning basis for an exposure assessment:

• Simple methods should be considered before more complex methods. Such a tiered assessment strategy is increasingly used in risk assessment and can be cost-effective.

• Chemicals are more biologically available in some media than in others; that is, the matrix within which chemicals occur (such as air, water, food, or soil) can greatly affect the extent of human exposure. The effect of the matrix should be considered in assessing exposure before assuming that contaminants are 100% bioavailable.

• Whenever possible, measurements should be obtained to support or validate any generic values used in exposure assessments, to check modeling results, or to provide more realistic estimates of exposure than can be obtained with models. Such measurements might include collecting data at locations where exposures are anticipated, monitoring the exposures experienced by individuals, collecting data on the physical and chemical conditions that affect the movement and bioavailability of chemicals, and providing information that relates exposure to effects, possibly using biologic markers. Measurements of exposure can be very different from estimated exposures based on source characteristics.

Using Realistic Exposure Scenarios

Finding

Because of statutory requirements and the desire not to underestimate chemical exposures, many risk assessments have estimated risks for a hypothetical, nonexistent "maximally exposed individual" (MEI) and have neglected information about the frequency, duration, and magnitude of actual population exposures. More recent assessments have used less extreme exposure scenarios. Congress specified in the 1990 amendments to the Clean Air Act that, after maximum available control technology is implemented for stationary sources, further controls must be considered if the lifetime excess cancer risk to the "individual most exposed to emissions from a source" in a category exceeds 10-6. The criteria for the "individual most exposed" were not stated; in fact, Congress mandated this Commission to advise what exposure scenarios should be used.

Recommendation

Exposure assessments should not be based on a hypothetical MEI. Screening risk assessments should rely on more representative estimates, such as EPA’s high-end exposure estimate (HEEE) or a maximally exposed actual person and estimates of the total number of potentially exposed people in the geographical areas of interest. Risk management decisions should be based on refined exposure assessments that evaluate the distribution of a population’s varied exposures and should address explicitly any segments of the population that have unusually high exposures. Exposure assessments should rely on population exposure data where possible instead of assumptions about exposure derived from source characteristics and models. The characteristics of actual or potential future populations in relation to specific sources of exposure should be emphasized and multiple sources of exposure should be reflected as appropriate in each case.

With the intention of protecting public health, past exposure assessment and health risk assessment practices have relied on exposure estimates derived from a hypothetical MEI who might spend a 70-year lifetime living at the point of greatest deposition from a plume of industrial contaminant emissions or who might spend a 70-year lifetime drinking only ground water with the highest concentrations of contaminants detected. The MEI was often so unrealistic that its use impaired the scientific credibility of health risk assessment.

Federal agencies have generally moved away from exposure assessments relying on such MEIs. For example, EPA’s exposure assessment guidelines have adopted the use of distributions of individual exposures and HEEEs chosen from values in the upper tail of those distributions (EPA 1992a). EPA’s risk characterization guidelines provide guidance on the use of exposure descriptors to characterize risk (EPA 1995a). At this time, implementation of those guidelines among EPA regional offices is uneven; some continue to use point estimates, while others use probability distributions of exposure estimates.

The Commission supports distributional approaches to exposure characterization that are based on knowledge of the characteristics of a population’s variability. Where possible, the entire distribution of the variability associated with exposure should be used in a risk characterization (see Effective Risk Characterization To Support Decision-Making on page 85). That distribution should be based on the characteristics of the entire exposed population and not solely on a highly exposed subpopulation; any highly exposed subpopulations known to exist should be considered separately. If a single value representing a population’s or subpopulation’s exposure is required, such as for priority setting, a point in the upper end of the distribution should be used, such as the 95th percentile.

Agencies should develop standard distributions to use in exposure assessments as defaults when population-specific information is unavailable. If data limitations do not permit the development of a defensible exposure distribution, a value representing a hypothetical highly exposed individual should be used. Such point exposure estimates are appropriate for screening level risk assessments. Probabilistic exposure estimates should be considered when standard default methods are expected to yield unrealistically conservative exposure estimates, when population estimates of exposure are desired, or when the exposure assessment is complex. Mark Van Putten, of the National Wildlife Federation, testified before the Commission that the environmental justice movement has provided some impetus for considering distributions instead of point estimates, on the grounds that populations with disproportionate exposures can be more explicitly identified and considered in risk assessments. We agree.

One advantage of using distributions to describe a population’s exposure is that it focuses attention on population risk, not just individual risk. Considering the size of a population in addition to the distribution of its exposures is important; for example, although emissions in a rural area might pose the same individual risk as those in an urban area, the total population risk for the latter is much greater. Another advantage is that it focuses attention on the characteristics of the population ("receptor-based" analysis) instead of basing exposure estimates primarily on the emission or other characteristics of a particular source of contamination ("source-based" analysis). A population-based approach can be source-specific but should include information on the variables that influence the mode, frequency, and duration of exposures. A complementary community-based approach would begin by determining a population’s exposures and moving from that information to identify sources of exposure. The total exposure assessment methodology (TEAM) study conducted by EPA and the Harvard Six Cities Survey, in which representative members of several urban populations wore small personal samplers to measure individual exposure to airborne chemicals (EPA 1987a, Dockery et al. 1993), are examples of a community-based approach to exposure assessment. The TEAM study also illustrates how dissimilar source-based predictions of exposures and actual exposures can be. Monitoring blood lead in a community’s children and tracing the sources of lead is another example of receptor-based analysis.

Many exposure assessments are based on source characteristics, not population characteristics. For example, air pollution sources typically have been licensed on the basis of modeled projections of their stack emissions. Few data on actual population exposures exist. (The Six Cities and TEAM studies are notable exceptions.) Such data deficiencies create problems, as emphasized by Ellen Silbergeld, representing the Environmental Defense Fund, in testimony before the Commission: there is no direct way to estimate the actual health risks experienced by an exposed population; there is no way to assess the relative contribution of multiple sources to risk; and there are no baseline data with which to evaluate the effects of new sources or of pollution reduction activities on existing sources.

Resistance to collecting data on populations’ actual exposures arises from the substantial time and expense associated with monitoring efforts, especially given the large variations in local climate and the problems associated with accurate detection of small pollutant exposures. Environmental monitoring is needed, however, to generate actual data that are consistent with a public health approach to risk assessment and with the Commission’s Risk Management Framework. In some circumstances, the costs of monitoring, such as for blood lead, are small compared to the overall costs of remediating a Superfund site, for example, and can save funds amounting to several times the cost of the study. Although multipathway modeling is not scientifically well developed, at present, exposure assessment must begin to address aggregate exposures (see also Section 2 and Evaluating Chemical Mixtures on page 68). Stimulated in part by Toxic Release Inventory reports, communities are interested not just in what they are exposed to because of a particular industrial facility, but in how that facility adds to the burden of exposures that they are already experiencing. Focusing on real populations is essential to identifying multiple exposure situations. We expect biomarkers of exposure to become useful in validating exposure estimates and in relating exposures to specific subgroups and even to individuals.

Identifying Highly Exposed Populations

Finding

Some population groups are at increased risk for toxic effects of chemical exposures because their exposures are greater than those of other population groups. Cultural practices, occupational exposures, behavior patterns, eating habits, and effects of related chemicals can be responsible. The high-risk subpopulations might be of special concern when risk assessments are conducted and risk management decisions are made. Risk assessors often have not sought information from knowledgeable citizens and consequently have not explicitly considered specific exposure conditions that might be present in minority group communities, certain occupational settings, or areas of particular socioeconomic status.

Recommendation

Risk assessments should be conducted so as to identify groups of people who are likely to have higher exposures to the chemicals of interest. Affected parties should be consulted in the early stages of an assessment to obtain information about all known sources of exposure to a particular chemical and related chemicals and to characterize exposure factors peculiar to particular subpopulations and link them with host susceptibility factors (see Accounting for Differences in Susceptibility on page 71).

Increased risks of adverse health effects from contaminant exposures can result from increased doses, as well as from increased susceptibility, which was discussed in the section Accounting for Differences in Susceptibility on page 71. Dose is a function of the concentration of a substance in the environment and the extent of exposure that a person has to the substance. Advances in the use of biologic markers will help to define relationships between exposure and dose. Below is a list of some factors that can increase risk as a result of increased exposure.

The Clinton Administration, the 103rd and 104th Congresses, interest groups, and the scientific community have attempted to address the issue of high-risk populations in several ways. For example, Executive Order 12898 on Environmental Justice requires that federal programs protect minority-group and low-income populations from disproportionately high exposures and adverse human health and environmental effects. EPA addressed the potentially greater susceptibility of children to pesticides and pesticide residues by requiring that assessments of environmental risks explicitly take health risks to children and infants into account (EPA 1995b). Congress reinforced that practice when it passed the Food Quality and Protection Act of 1996, which responded to a National Research Council report that variations in dietary exposure to pesticides related to nutritional intake, age, geographic region, and ethnicity were not addressed adequately by current regulatory practice (NRC 1993). Infants and children might be more heavily exposed to pesticides than adults because of their relatively high intake of fruit juices, for example, and they are more susceptible to the toxic effects of pesticides because of the sensitivity of their still-developing nervous systems and probably because of their greater concomitant exposures to lead and other environmental hazards.

Community assistance in characterizing exposure factors peculiar to particular segments of the population can focus a risk assessment and broaden risk management options. The Commission heard testimony from Asians and Pacific Islanders about their fish consumption patterns and about the role that education can play in risk management. Not only do they consume more fish, but they consume fish parts that are usually discarded by others and in which pollutants are often concentrated, placing themselves at higher risk than the general population for the effects of contaminants in fish. They reported that educational brochures, signs around contaminated bodies of water, and community involvement led to voluntary reduction in exposure through modest changes in fish eating in the Seattle area. Of course, education is only one risk management alternative, and other stakeholders might not consider it to be appropriate or acceptable. In contrast to the Asians and Pacific Islanders, Mark Van Putten, of the National Wildlife Federation, testified that in the Great Lakes region it was difficult to convince risk managers that subsistence fishers, such as Native Americans, should be considered in risk assessments.

Specific information gathered from the community and stakeholders could reduce the need for default assumptions and improve the quality of risk assessments in communities with multiple polluting operations, such as a municipal incinerator, a chemical plant, a dry cleaning establishment, and an abandoned hazardous waste site. Involving the community and other stakeholders in the planning stages of a risk assessment can help to engage individuals, families, schools, businesses, and municipalities in targeted pollution prevention and pollution reduction actions that reduce exposures. The Commission’s Risk Management Framework calls for stakeholders to be involved in every step of the process, including evaluation of the actions taken.

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