Cancer Risk Evaluation: Methods and Trends

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For those and several other reasons, it is difficult or impossible to identify cause-effect relationships clearly with epidemiologic methods OSTP, It is rare that convincing causal relationships are identified with a single study.

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Epidemiologists usually weigh the results from several studies, ideally involving different populations and investigative methods, to determine whether there is a consistent pattern of responses among them. Some of the other factors that are often considered are the strength of the statistical association between a particular disease and exposure to the suspect chemical; whether the risk of the disease increases with increasing exposure to the suspect agent; and the degree to which other possible causative factors can be ruled out.

Needless to say, different experts will weigh such data differently, and consensus typically is not easily achieved IARC, In the case of chemicals suspected of causing cancer in humans, expert groups "working groups" are regularly convened by the International Agency for Research on Cancer IARC to consider and evaluate epidemiologic evidence. These groups have published their conclusions regarding the "degrees" of strength of the evidence on specific chemicals sometimes chemical mixtures or even industrial processes when individual causative agents cannot be identified.

No similar consensus-building procedure has been established regarding other forms of toxicity. Some epidemiologists disagree with IARC's cancer classification judgments in particular cases, and there seems to be even greater potential for scientific controversy regarding the strength of the epidemiologic evidence of non-cancer e.

There has been much less epidemiologic study of other toxic effects, in part because of lack of adequate medical documentation. When epidemiologic studies are not available or not suitable, risk assessment may be based on studies of laboratory animals. One advantage of animal studies is that they can be controlled, so establishing causation assuming that the experiments are well conducted is not in general difficult. Another advantage is that animals can be used to collect toxicity information on chemicals before their marketing, whereas epidemiologic data can be collected only after human exposure.

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Indeed, laws in many countries require that some classes of chemicals e. Other advantages of animal tests include the facts that. The quantitative relationship between exposure or dose and extent of toxic response can be established. The animals and animal tissues can be thoroughly examined by toxicologists and pathologists, so the full range of toxic effects produced by a chemical can be identified. The exposure duration and routes can be designed to match those experienced by the human population of concern.

But laboratory animals are not human beings, and this obvious fact is one clear disadvantage of animal studies. Another is the relatively high cost of animal studies containing enough animals to detect an effect of interest. There are reasons based on both biologic principles and empirical observations to support the hypothesis that many forms of biologic responses, including toxic responses, can be extrapolated across mammalian species, including Homo sapiens , but the scientific basis of such extrapolation is not established with sufficient rigor to allow broad and definitive generalizations to be made NRC, b.

One of the most important reasons for species differences in response to chemical exposures is that toxicity is very often a function of chemical metabolism. Differences among animal species, or even among strains of the same species, in metabolic handling of a chemical, are not uncommon and can account for toxicity differences NRC, Because in most cases information on a chemical's metabolic profile in humans is lacking and often unobtainable , identifying the animal species and toxic response most likely to predict the human response accurately is generally not possible.

It has become customary to assume, under these circumstances, that in the absence of clear evidence that a particular toxic response is not relevant to human beings, any observation of toxicity in an animal species is potentially predictive of response in at least some humans EPA, a. This is not unreasonable, given the great variation among humans in genetic composition, prior sensitizing events, and concurrent exposures to other agents. As in the case of epidemiologic data, IARC expert panels rank evidence of carcinogenicity from animal studies. It is generally recognized by experts that evidence of carcinogenicity is most convincing when a chemical produces excess malignancies in several species and strains of laboratory animals and in both sexes.

The observation that a much higher proportion of treated animals than untreated control animals develops malignancies adds weight to the evidence of carcinogenicity as a result of the exposure.

At the other extreme, the observation that a chemical produces only a relatively small increase in incidence of mostly benign tumors, at a single site of the body, in a single species and sex of test animal does not make a very convincing case for carcinogenicity, although any excess of tumors raises some concern. EPA combines human and animal evidence, as shown in Table , to categorize evidence of carcinogenicity; the agency's evaluations of data on individual carcinogens generally match those of IARC. For noncancer health effects, EPA uses categories like those outlined in Table Animal data on other forms of toxicity are generally evaluated in the same way as carcinogenicity data, although this classification looks at hazard identification qualitative and dose-response relationships quantitative together.

No risk or hazard ranking schemes similar to those used for carcinogens have been adopted. The hazard-identification step of a risk assessment generally concludes with a qualitative narrative of the types of toxic responses, if any, that can be caused. Limited evidence from epidemiologic studies and sufficient evidence from animal studies B1 ; or inadequate evidence from epidemiologic studies or no data and sufficient evidence from animal studies B2.

In addition to the epidemiologic and animal data, information on metabolism and on the behavior of the chemical in tissues and cells i. Identifying the potential of a chemical to cause particular forms of toxicity in humans does not reveal whether the substance poses a risk in specific exposed populations. The latter determination requires three further analytic steps: emission characterization and exposure assessment discussed in Chapter 3 , dose-response assessment discussed next , and risk characterization discussed in Chapter 5.

In the United States and many other countries, two forms of dose-response assessment involving extrapolation to low doses are used, depending on the nature of the toxic effect under consideration. One form is used for cancer, the other for toxic effects other than cancer.

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For all types of toxic effects other than cancer, the standard procedure used by regulatory agencies for evaluating the dose-response aspects of toxicity involves identifying the highest exposure among all the available experimental. The sufficient-evidence category includes data that collectively provide enough information to judge whether a human developmental hazard could exist within the context of dose, duration, timing, and route of exposure.

This category includes both human and experimental-animal evidence. Sufficient Human Evidence : This category includes data from epidemiologic studies e. A case series in conjunction with strong supporting evidence may also be used. Supporting animal data might or might not be available. Sufficient Experimental Animal Evidence or Limited Human Data : This category includes data from experimental-animal studies or limited human data that provide convincing evidence for the scientific community to judge whether the potential for developmental toxicity exists.

The minimal evidence necessary to judge that a potential hazard exists generally would be data demonstrating an adverse developmental effect in a single appropriate, well-conducted study in a single experimental-animal species. The minimal evidence needed to judge that a potential hazard does not exist would include data from appropriate, well-conducted laboratory-animal studies in several species at least two that evaluated a variety of the potential manifestations of developmental toxicity and showed no developmental effects at doses that were minimally toxic to adults.

This category includes situations for which there is less than the minimal sufficient evidence necessary for assessing the potential for developmental toxicity, such as when no data are available on developmental toxicity, when the available data are from studies in animals or humans that have a limited design e. The difference between the two values is related to the definition of adverse effect. The NOAEL is the highest exposure at which there is no statistically or biologically significant increase in the frequency of an adverse effect when compared with a control group.

A similar value used is the lowest-observed-adverse-effect level LOAEL , which is the lowest exposure at which there is a significant increase in an observable effect. All are used in a similar fashion relative to the regulatory need. For human risk assessment, the ratio of the NOAEL to the estimated human dose gives an indication of the margin of safety for the potential risk. In general, the smaller the ratio, the greater the likelihood that some people will be adversely affected by the exposure.

The uncertainty-factor approach is used to set exposure limits for a chemical when there is reason to believe that a safe exposure exists; that is, that its toxic effects are likely to be expressed in a person only if that person's exposure is above some minimum, or threshold. At exposures below the threshold, toxic effects are unlikely. To establish limits for human exposure, the experimental NOAEL is divided by one or more uncertainty factors, which are intended to account for the uncertainty associated with interspecies and intraspecies extrapolation and other factors.

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Depending on how close the experimental threshold is thought to be to the exposure of a human population, perhaps modified by the particular conditions of exposure, a larger or smaller uncertainty factor might be required to ensure adequate protection. For example, if the NOAEL is derived from high-quality data in necessarily limited groups of humans, even a small safety factor 10 or less might ensure safety, provided that the NOAEL was derived under conditions of exposure similar to those in the exposed population of interest and the study is otherwise sound.

If, however, the NOAEL was derived from a less similar or less reliable laboratory-animal study, a larger uncertainty factor would be required NRC, There is no strong scientific basis for using the same constant uncertainty factor for all situations, but there are strong precedents for the use of some values NRC, The regulatory agencies usually require values of 10,, or 1, in different situations.

For example, a factor of is usually applied when the NOAEL is derived from chronic toxicity studies typically 2-year studies that are considered to be of high quality and when the purpose is to protect members of the general population who could be exposed daily for a full lifetime 10 to account for interspecies differences and 10 to account for intraspecies differences. The requirement for uncertainty factors stems in part from the belief that humans could be more sensitive to the toxic effects of a chemical than laboratory animals and the belief that variations in sensitivity are likely to exist within the human population NRC, a.

Those beliefs are plausible, but the magnitudes of interspecies and intraspecies differences for every chemical and toxic end point are not often known. Uncertainty factors are intended to accommodate scientific uncertainty, as well as uncertainties about dose delivered, human variations in sensitivity, and other matters Dourson and Stara, EPA's approaches to risk assessment for chemically induced reproductive and developmental end points rely on the threshold assumption.

The RfD is obtained as described above. The total adjustment or uncertainty factor referred to in the proposed guidelines for use in obtaining an RfD from toxicity data "usually ranges" from 10 to 1, The adjustment incorporates as needed uncertainty factors "often" 10 for " 1 situations in which the LOAEL must be used because a NOAEL was not established, 2 interspecies extrapolation, and 3 intraspecies adjustment for variable sensitivity among individuals. EPA's revision of its guidelines for developmental-toxicity risk assessment state that "human data are preferred for risk assessment" and that the "most relevant information" is provided by good epidemiologic studies.

When these data are not available, however, reproductive risk assessment and developmental-agent risk assessment, according to EPA, are based on four key assumptions:. An agent that causes adverse developmental effects in animals will do so in humans, with sufficient exposure during development, although the types of effects might not be the same in humans as in animals.

Although the types of effects in humans and animals might not be the same, the use of the most sensitive animal species to estimate human hazards is justified. A threshold is assumed in dose-response relationships on the basis of current knowledge, although some experts believe that current science does not fully support this position. The new guidelines state that "the existence of a NOAEL in an animal study does not prove or disprove the existence or level of a biological threshold.

The guidelines also discuss EPA's plans to move toward a more quantitative "benchmark dose" BD for risk assessment for developmental end points "when sufficient data are available"; the BD approach would be consistent with the uncertainty-factor approach now in use EPA, a. It would be derived by modeling the data in the observed range, selecting an incidence rate at a preset low observed response e. A BD thus calculated would then be divided by uncertainty factors to derive corresponding acceptable intake e.

Thus, the traditional uncertainty-factor approach is retained in the developmental-toxicity guidelines, as well as in the proposed BD approach. However, the new guidelines are unique, in that they emphasize both the possible effect of interindividual variability in the interpretation of acceptable exposures and the improvements that biologically based models could bring to developmental risk assessment EPA, a :. It has generally been assumed that there is a biological threshold for developmental toxicity; however, a threshold for a population of individuals may or may not exist because of other endogenous or exogenous factors that may increase the sensitivity of some individuals in the population.

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  • Thus, the addition of a toxicant may result in an increased risk for the population, but not necessarily for all individuals in the population. For some toxic effects, notably cancer, there are reasons to believe either that no threshold for dose-response relationships exists or that, if one does exist, it is very low and cannot be reliably identified OSTP, ; NRC, This approach is taken on the basis not of human experience with chemical-induced cancer, but rather of radiation-induced cancer in humans and radiologic theory of tissue damage. Risk estimation for carcinogens therefore follows a different procedure from that for noncarcinogens: the relationship between cancer incidence and the dose of a chemical observed in an epidemiologic or experimental study is extrapolated to the lower doses at which humans e.

    In this procedure, there is no "safe" dose with a risk of zero except at zero dose , although at sufficiently low doses the risk becomes very low and is generally regarded as without publichealth significance. The procedure used by EPA is typical of those used by the other regulatory agencies. The observed relationship between lifetime daily dose and observed tumor incidence is fitted to a mathematical model to predict the incidence at low doses.

    Several such models are in wide use. FDA uses a somewhat different procedure that nevertheless yields a similar result. An important feature of the LMS is that the dose-response curve is linear at low doses, even if it displays nonlinear behavior in the region of observation. EPA applies a statistical confidence-limit procedure to the linear multistage no-threshold model to generate what is sometimes considered an upper bound on cancer risk.

    Although the actual risk cannot be known, it is thought that it will not exceed the upper bound, might be lower, and could be zero. The result of a dose-response assessment for a carcinogen is a potency factor. EPA also uses the term unit risk factor for cancer potency. This value is the plausible upper bound on excess lifetime risk of cancer per unit of dose. In the absence of strong evidence to the contrary, it is generally assumed that such a potency factor estimated from animal data can be applied to humans to estimate an upper bound on the human cancer risk associated with lifetime exposure to a specified dosage.

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