Organic Pollutants in Water - ACS Publications - American Chemical

34 Risk Assessment and Control Decisions for Protecting Drinking Water Quality

Downloaded by PURDUE UNIVERSITY on March 21, 2013 | http://pubs.acs.org Publication Date: December 15, 1986 | doi: 10.1021/ba-1987-0214.ch034

Joseph A. Cotruvo Criteria and Standards Division, Office of Drinking Water, U.S. Environmental Protection Agency, Washington, DC 20460 This chapter describes risk evaluation processes as they have been applied to drinking water standards and guidelines. Traditional risk assessments and standards are based upon single chemical toxicology. They typically assume that no significant interactions occur at the low levels at which chemicals are commonly found in the environment. Newer evaluation techniques might permit development of standards based upon indications of hazard from exposure to the actual environmental mixtures. New concentration techniques and biological indicator measurements will be the keys to this possible innovation in water regulation. If improvements are expected in the ability to assess risks from the consumption of drinking water, concentrates from several sources of varying quality (including reuse systems) should be tested by these techniques to determine relative qualities of these waters and to compare the results with single toxicology predictions.

SAFETY IS THE PRACTICAL CERTAINTY that injury will not result from a substance when used in the quantity and in the manner proposed for its use (J). The goal of all public health and water authorities is to assure that public drinking water supplies are safe, pure, and wholesome in the broadest sense, that is, free from contamination by substances of possible health concern as well as free from adulterants that would detract from the water quality and reduce acceptance by consumers. Not only should public water supplies be safe, but they should be perceived to be safe. In a sense, drinking water in developed countries can be a closed and totally controllable system. It consists of source, treatment, and This chapter not subject to U.S. copyright. Published 1987 American Chemical Society

In Organic Pollutants in Water; Suffet, I., et al.; Advances in Chemistry; American Chemical Society: Washington, DC, 1986.


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transport functions, each of which is subject to contamination stresses, but virtually all of which are controllable b y feasible social, political, or technological means, given the willingness to expend the required resources. Thus, in effect, society has almost total control over the quality or composition of the water that is delivered to the consumer's tap. In simple terms, sources of public water systems can be protected f r o m contamination, water can be treated for the removal of undesirable components, and the quality of the treated water can be protected during its transport through distribution mains to consumers (2). In the extreme case, water f r o m any source can be synthesized to any desired composition and packaged for total protection. Three elements drive public policy decisions to protect health and welfare f r o m perceived or actual risks regardless of origin: (1) identification of the existence or the perception of the existence of the risk; (2) assessment of the likelihood of the risk's existence, the quantification of the magnitude of the risk, and the health significance of the risk; and (3) the feasibility, cost, and effectiveness of the means of abating or managing the risk. D r i n k i n g water quality concerns range f r o m (1) infectious disease risks that are large, obvious, and quantifiable; (2) acute or chronic chemical hazards such as those f r o m arsenic or lead that are infrequent but potentially identifiable in cause and effect when they occur; (3) postulated carcinogenic risks from radionuclides or certain organic chemicals that are largely undetectable and empirically unquantifiable and usually small in magnitude relative to overall cancer incidence rates. The public perception of the existence of any risk associated with the consumption of essential drinking water can have profound consequences both locally and nationally, including the loss of public confidence in political institutions. It can cause some consumers to shift from public water supplies to private sources, self-provided water treatment, or bottled water. T o a considerable degree, the disorientation can be traced to the nonthreshold hypothesis, that is, the theoretical existence of finite, albeit small, risks at any nonzero exposure and the inability of scientists to unequivocally attest to the absolute safety of the product, even when only a trace of a potential carcinogen has been detected. Fundamental questions exist on (1) the criteria to be used to identify those substances that have the capacity to increase the risk of human cancer, (2) their mechanisms of action, and (3) the magnitude of the risk posed b y episodic or chronic regular exposure. The answers to these questions lead to major public policy determinations based on the validity and significance of the real or postulated effects, the feasibility of the possible risk reduction measures, and the economic and social costs of those measures. In a society with finite resources at its


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disposal, ultimately the benefits accrued versus the cost burden must be weighed, as well as the trade-off be ween the value of the chosen course versus other social benefits that have been forgone. f


Philosophical Basis for the Determination of the Control Level (Standards) E a c h society w i l l have a different legal and procedural framework for making control decisions and determining whether these are advisory or mandatory, depending upon operative laws and tradition. The philosophical bases may be different depending upon the type of contaminant, the mechanism of action, and the significance of the adverse effects. However, the philosophical basis for determination of the control level should be articulated; the public has the right to k n o w the meaning and consequences of any control value. A m o n g the many bases for establishing control levels are the following: zero or no deliberate addition; no detection b y specified analytical methods; natural background level; safe or wholesome level; no unreasonable risk level; no k n o w n adverse effect level with a margin or safety; level consistent w i t h a specified risk or probability of harm; technologically and economically feasible level; level achievable b y using the best available technology; marginal benefits are greater than marginal control costs; and costs of achieving the level are l o w and socially acceptable. The decision-making procedure may include detailed quantitative studies subject to legally specified development and scientific and public reviews, judgmental decisions b y experts in council, or legislative determinations. When p u b l i c health is at stake, the ideal goal should be to assure against the occurrence or the potential occurrence of any of the adverse effects, with a large margin of safety. O n the other hand, all decisions ultimately must reflect economic and technological feasibility; thus, it is probable that selected control levels w i l l differ f r o m ideal goals. However, the risk assessment—the process for determining the extent of the risk and the goal—should be separate and distinct f r o m the risk management—the mechanism for evaluating the feasibility and costs of the controls. In both cases, the assumptions and uncertainties should be clearly stated along with the bases for the conclusions.

Drinking Water as a Source of Risk Risk factors from drinking water include infectious disease, acute or chronic chemical toxicity, and carcinogenicity. In 1981, the North A t lantic Treaty Organization C o m m i t t e e on Challenges to M o d e r n


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Society, in its Report on the Health Aspects of D r i n k i n g Water C o n taminants (chemicals), concluded that, in general, no adverse health effects have been observed f r o m the consumption of drinking water that has been generated in a controlled public supply and that has met drinking water standards (3). Nevertheless, the committee stated that k n o w n contamination of drinking water b y chemicals f r o m disinfection practices, industrial discharges, hazardous waste disposal, and corrosion of piping is a potential hazard. Adverse health risks have been associated with failure to protect the source, to provide adequate treatment, and to ensure the integrity of the distribution system. Source. Source waters, including rivers, lakes, and ground waters, can often be selected so as to be free f r o m significant biological contaminants or protected f r o m potentially harmful anthropogenic contaminants. Source waters can be contaminated b y a variety of synthetic organic chemicals, usually in trace amounts. Ground waters in the vicinity of improperly designed waste disposal sites have sometimes been found to be heavily contaminated b y migrating chemicals, most frequently chlorinated solvents such as trichloroethylene, tetrachloroethylene, 1,1,1-trichloroethane, or carbon tetrachloride and fuel products such as benzene and aliphatic hydrocarbons (4). However, the most universally found organic contaminants in surface waters and some ground waters are natural products including humic and fulvic acids, terpenes, tannins, amino acids, peptides, and other cellular debris. Inorganic contaminants such as c o m m o n salts or trace toxic substances such as arsenic or c a d m i u m can be present. Nitrates are common in agricultural areas. A m o n g the inorganic natural contaminants of potential significance are localized deposits of arsenic or selenium and widespread sources of radionuclides such as radium and especially radon gas f r o m ground sources. The presence or absence of inorganic ions such as calcium may play a role in reducing the postulated risks of cardiovascular diseases associated with the degree of hardness of drinking water. Treatment Processes. Technology and operating procedures are available to prevent the introduction of these contaminants, and technology is available to remove all of these contaminants f r o m drinking water; however, consumer costs can be substantial when economies of scale are absent (e.g., small communities). Many chemicals are added to water to remove contaminants such as organic matter, suspended or dissolved solids, and m i c r o b i a l pathogens. A m o n g those added are alum, iron salts, polymeric coagulant aids, chlorine, and other oxidizing agents, all of which may leave residues or byproducts in the finished water. Chlorine gas often contains


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chloroform, carbon tetrachloride, or other residues, and it reacts with organic matter in the water to produce trihalomethanes, chloramines, haloacetonitriles, haloacetic acids, halophenols, and a host of other byproducts. Normally, the major source of the synthetic chemicals in treated drinking water is the interaction of chlorine or another oxidizing agent with the natural products already there. T o further complicate the risk assessment and control decisions, a preliminary unverified report has indicated that the consumption of water containing these oxidizing agents may have increased cholesterol levels in test animals on high fat and l o w calcium diets. T h y r o i d hormone levels may have also been affected (5). Distribution Systems. A substantial amount of contamination of drinking water can occur while the water is in transit to the consumer after treatment. Pipes are made of copper, galvanized iron, asbes tos-cement, lead, or plastic, and often polymeric or coal tar coatings are used. A l l of these are capable of contributing contaminants to the water, especially if the water is corrosive. L e a d , copper, cadmium, and polynuclear aromatic hydrocarbons in finished water are primarily problems of water distribution and not source water contamination. Physical deterioration of the distribution system can also permit biologi cal contamination to occur during transit.

Microbial Risfa from Drinking Water The principal risk factors in drinking water (except possibly natural radioactivity) are biological in origin as indicated b y the reported and projected evidence of waterborne disease. In the 12-year period f r o m 1971 to 1982 in the U n i t e d States, there were 392 outbreaks involving almost 86,000 reported cases (Table I) (6). M a n y outbreaks, probably the great majority, go unreported because of the difficulty of detecting the event and identifying the etiology of the occurrences. In one pilot study, only about one-third to one-fifth of the actual outbreaks were being recognized and reported. The diseases include uncharacterized acute gastroenteritis, giardiasis, shigellosis, hepatitis-Α, typhoid, and salmonellosis. Fortunately, massive numbers of deaths associated with waterborne cholera and t y p h o i d no longer occur in developed countries as they d i d in the last century prior to the widespread introduction of filtration and disinfection. Retrospective identification of risk f r o m waterborne infectious dis ease is a relatively simpler task compared with carcinogenic risks. M a n y acute effects can be identified with proper population surveillance, related to probable origin, and quantified. Assessments of microbial risks f r o m theoretical projections w o u l d be extremely complex. They


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Table I. U.S. Outbreaks and Cases: 1971-82 Year


1971-80 Community PWS Noncommunity PWS Individual 1981 1982 Total 1971-82 1971-80 Community PWS Noncommunity PWS Individual 1981 1982 Total 1971-82

Giardia

Bac

Virus

Chem

Unknown

Total

Outbreaks 22 12 5 9 12 60

17 16 7 3 3 46

7 15 4 1 7 34

22 7 9 5 2 45

55 110 12 14 16 207

123 160 37 32 40 392

1820 1762 28 1761 853 6224

2886 645 63 1893 18 3740

28928 10783 134 128 1836 43574

59387 18263 339 4430 3456 85875

Cases 17090 2390 72 297 561 20410

8663 2683 42 351 188 11927

N O T E : PWS denotes public water supply, Bac means bacterial, and Chem means chemical. S O U R C E : Adapted from reference 6.

w o u l d have to b e based upon individual identifications of all of the many pathogenic microorganisms potentially in water; determination of infective doses to each segment of the population; considerations of complex interactions including age, physical state, and other health stresses; immunity states; determinations of exposed populations at various levels; variability of diagnoses; variability of sources of microorganisms; secondary infection rates; and differentiation between water and food and other sources of similar infections. A l l of these intellectually challenging and intriguing theoretical exercises have been obviated, although not without conflicts, b y the introduction of two operationally simple and practical treatment techniques: disinfection and filtration. Numerous studies have shown that these conventional processes can typically remove or inactivate six to eight orders of magnitude of virus (7). B y using the larger figure (10 ), it has been calculated that treatment of a water containing 300 viral units per 380 L w o u l d result in a finished water that contained one infectious virus unit per 120 million L . B y application of the same treatment conditions uniformly applied to a source water containing as much as 95,000 viral units per 380 L (a grossly contaminated source), one infectious unit w o u l d be present i n about 400,000 L of finished drinking water (8). Only about 1% of the treated water is ultimately ingested; thus, the probability of consumption of that infectious unit b y an individual could be commensurately smaller. 8


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Figure 1 illustrates several additional examples of calculated expected infections per year from a community water supply that contains one polio or Echo 12 virus per 1000 L (9). For example, one Echo 12 virus in 1000 L is projected to produce about four infections per year per 1000 persons. The technical objective would be to assure that a consumer would not be exposed to an infectious dose of a pathogen from the drinking water. Such a dose can range from a few or perhaps only one virulent organism (PFU) for polio virus or certain protozoa such as Giardia lamblia, to hundreds of Shigella or perhaps millions of opportunistic

2000-,

10

100

1000

LITERS

Figure 1. Examples of calculated expected infections per year from a community water supply that contains one polio or Echo 12 virus per 1000 L.



700


pathogens such as Pseudomonas (10). Water production systems using available technologies can be sited, built, and operated to reduce the probability of consumer exposure to an infective dose to an extremely l o w and insignificant risk level. Thus, in the case of biological contaminants of drinking water, the goal of virtually no risk for all practical purposes is within reach in the case of a well-designed and properly operated water production and distribution facility. In the case of biological contamination, the identification of risk became obvious b y experience, the risk assessment was made unambiguous by epidemiology, and the immediate and obvious effectiveness of the risk management decisions demonstrated their w i s d o m in the absence of elegant quantitative risk extrapolation models and projections of costs per case averted. Costs of water treatment and distribution became trivial relative to almost all other essential commodities, and in the public expectation the biological safety of drinking water became axiomatic. The fundamental and unresolvable element concerning adverse health effects and trace chemical contamination of drinking water is that in all but a few exceptional cases, three elements—risk identification, risk assessment by epidemiological data, and demonstrable risk management results—may never be available.

Safety and Risk Determination for Chemical Agents Paracelsus (11) observed that " A U things are poisons, for there is nothing without poisonous qualities. It is only the dose that makes a thing a poison." Toxicity has been defined as the intrinsic quality of a chemical to produce an adverse effect (I). The toxicology of chemical substances found in drinking water is commonly divided into two broad classes: (1) acute or chronic toxicity and (2) carcinogenicity. H o w e v e r , teratogenic and mutagenic risks could also be considered. The same substance may be capable of causing classic toxic effects and imparting risks of carcinogenicity. The distinguishing characteristic between these categories of effects lies (1) in the probably unverifiable assumption that dose thresholds exist for chronic toxicity effects and (2) in the also unverifiable assumption that dose thresholds do not exist (or have not been demonstrated) for carcinogenic effects. In the case that dose thresholds exist for chronic toxicity effects, the nominal basis for standard setting is to achieve a total daily dose of the substance that is with practical certainty below the level at w h i c h any injury w o u l d result to any individual in the population. F o r toxicants assumed to be acting by nonthreshold mechanisms, it follows that some finite risk may exist at



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any nonzero dose level. Thus, standard setting objectives range f r o m zero, which is not quantifiable and often not practically achievable, to a daily dose level that contributes only a negligible theoretical incremental increase in the lifetime risk of the effect to individuals and/or the population exposed. The determination of a permissible exposure to a toxic substance requires evaluation of qualitative and quantitative factors including the identification and health significance of the adverse effect; the sensitive members of and the size of the exposed population, biological absorption, distribution, metabolism, and excretion; and the possible additivity, synergism, or antagonism with coexposed substances. The U.S. National A c a d e m y of Sciences (NAS) in the series Drinking Water and Health and other writings has described the theory and practice of toxicology and risk assessment and related them to drinking water. These w i l l be used liberally in the following discussions.

Noncarcinogenic Effects: Safety Factors Numerous substances detected in drinking waters are k n o w n to induce toxicity but usually at dose levels much higher than those found in water. Nitrates or nitrites can cause infant methemoglobinemia, lead can affect the hematopoetic or nervous system, c a d m i u m can cause renal damage, and some organohalogens may cause liver toxicity (12). When appropriate data are available f r o m human epidemiology or animal studies, the use of the acceptable daily intake (ADI) concept is a well-accepted procedure for determining concentration levels for standard setting. The A D I of a chemical is defined as the dose that is anticipated to be without lifetime risk to humans when taken daily. The A D I does not guarantee absolute safety, however, and it is not an estimate of risk. The assumption of one threshold for each individual in a large population is simplistic; the population is genetically heterogeneous with a varied history of exposure, prior disease states, nutritional status, and stresses. Thus, it is likely that each individual has a unique threshold, and certain individuals in the population w i l l be at inordinately high risk, whereas others may be at very l o w risk. The A D I concept is probably not applicable to heavy metals and lipophilic substances, which tend to bioaccumulate (13). The A D I is usually derived f r o m a detailed analysis of the toxicology of the chemical being examined. The no observed adverse effect level ( N O A E L ) is determined for the most sensitive adverse effect in the test system (usually animals but occasionally humans), and a safety or uncertainty factor is applied to the N O A E L dose to derive the safe level for the general human population.


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The A D I is computed b y multiplying the experimental N O A E L (in milligrams per kilogram per day) b y the weight of a typical adult (70 kg) and dividing b y the safety (uncertainty) factor.


A D I (mg/person/day) = N O A E L (mg/kg/day) X 70 (kg/person)/ safety (uncertainty) factor Because an A D I is intended to account for total daily intake of the toxicant f r o m all sources, inhalation and food intake as well as water should be accounted for when attempting to arrive at the maximum drinking water level or the adjusted A D I for drinking water at the maximum drinking water level considering only health factors. Thus, in the optimum case when such information is available, the daily uptake f r o m air and the daily intake f r o m food (if 100ï uptake is assumed) should be subtracted f r o m the A D I . Finally, for the determination of the acceptable drinking water concentration value, the assumption in the United States is that adults consume 2 L of water per person per day; thus, the final value should be divided b y a factor of 2. drinking water target (mg/L) = A D I (mg/day) — inhalation (mg/day) - food (mg/day)/(2 L/day) The foregoing calculation is commonly used for determining acceptable lifetime exposures f r o m drinking water to chronic toxicants. In some cases in w h i c h the concern is for shorter exposures to young children who may be at higher risk because of a higher water consumption to b o d y weight ratio, the U . S . Environmental Protection Agency ( U S E P A ) has used the 10-kg child and an assumed consumption of 1 L of water per day as the standard exposed individual for calculation purposes. In effect, such calculation introduces an additional safety factor of 3.5. Some procedures use a dose conversion method involving b o d y surface area (milligrams per square meter) as opposed to weight (milligrams per kilogram). B o d y surface area is approximately proportional to the two-thirds power of b o d y weight. This relationship may be particularly appropriate for extrapolations from small animals (rats and mice) to humans rather than for data obtained f r o m dogs or monkeys because on this basis, chemicals w o u l d be relatively more toxic to larger animals than to smaller ones (13). Figure 2 is a general illustration of a process for the calculation of an A D I for a particular substance (14). The solid line to point A is the dose-response curve determined b y the multiple dosing experiment. Point A is the highest no observed effect level in milligrams per kilogram per day for the most sensitive adverse end point that was determined f r o m the animal multiple-dose chronic study. Points B, D ,



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Figure 2. General illustration of a process for the calcuktion of an ADI for a particular substance. and Ε are the presumed thresholds for the adverse effect in the human population if the extrapolated dose-response curves A B , A D , or A E are correct. Point C w o u l d be the A D I concentration determined b y ap plication of the selected safety (uncertainty) factor to the dose at point A . Because lines A B , A D , and A E are extrapolations, the true dose-response curve in the range of concern is unknown; thus, any of the curves could be correct in a given case. The intention of the standard setter is that A B w o u l d be the true curve because the no (actual) response dose w o u l d be greater than the calculated A D I value C ; thus, the safety (uncertainty) factor was chosen appropriately. However, if A D or A E were the true dose-response curve, then the calculated A D I was too large; thus, the safety (uncertainty) factor was too small, and some members in the human population might suffer the adverse effect. A E indicates a nonthreshold dose-response. The size of the gap between C and Β is also of interest because if it were exces sively large, overregulation could result in excessive control expendi tures without any benefit. The value of an A D I is entirely dependent on the quality of the experimental data and the judicious selection of the safety (uncertainty) factor, w h i c h is entirely judgmental. A m o n g the factors influencing the quality of the experimental data, beyond the mechanics, are the selec tion of the appropriate animal model as the human surrogate, the


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number of animals at each dose and the number and range of the doses for acceptable statistical significance of results and shape of the experimental curve, the actual detection of the most sensitive adverse effect (which could be only biochemical change or frank organ damage), the length of the study (lifetime studies versus shorter term studies), and the appropriate route of exposure (inhalation, gavage, ingestion in f o o d or water, etc.). The quality of the experimental evidence determines the magnitude of the safety (uncertainty) factor to be applied.

Safety (Uncertainty) Factors The safety factor is a number that reflects the degree or amount of uncertainty that must be considered when experimental data are extrapolated to the human population. When the quality and quantity of dose-response data are high, the uncertainty factor is low; when the data are inadequate or equivocal, the uncertainty factor must be larger (15). The original use of the safety factor approach in regulation was b y Lehman and Fitzhugh (13), w h o considered that animals may be more resistant to the toxic effects of some chemicals than humans are. They proposed the use of a factor of 10 when extrapolating from animals to humans and the use of another factor of 10 to account for differential sensitivities within the human population (13). These are not, however, rigid rules, and they should be applied with a strong infusion of scientific judgment. The following general guidelines (15) have been adopted b y the N A S Safe D r i n k i n g Water Committee, and they are also used b y the U S E P A in the development of drinking water standards and guidelines and health advisories. 1. 10 Factor: V a l i d experimental results f r o m studies on p r o l o n g e d human ingestion w i t h no indication of carcinogenicity. 2. 100 Factor: Experimental results of studies of human ingestion not available or scanty. V a l i d results f r o m long-term feeding studies on experimental animals or, in the absence of human studies, on one or more species. N o indication of carcinogenicity. 3. 1000 Factor: N o long-term or acute human data. Scanty results on experimental animals. N o indication of carcinogenicity. The N A S also examined the application of quantitative models such as log probit and log logistic for human risk assessments for noncar-



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cinogenic substances but found these to be of limited value for contaminants in drinking water (16). These models could be used to estimate the risk of a toxic effect, but they require data f r o m lifetime feeding studies w i t h sufficient numbers of animals and w i t h a demonstrated dose-response. Data of this type are seldom available; thus, the N A S concluded that the A D I approach is most useful at this time. H o w e v e r , in those cases in which such data can be obtained, a risk estimate approach can be employed. A n example of an A D I calculation for p-dichlorobenzene is provided (17). A n i m a l studies at various doses have observed liver and kidney damage, porphyria, pulmonary edema, and spleenic weight changes. H u m a n exposure at high concentrations has been reported to result in pulmonary damage and hemolytic anemia. A 1-year gavage study in the rabbit contained five animals per group dosed between 0 and 1000 mg/kg/day and resulted in weight loss, tremors, and liver pathology. The highest N O A E L was 357 mg/kg/day. A subchronic study indicated a N O A E L of 150 mg/kg in the rat exposed b y gavage. Animals received doses of 37.5, 75, 150, 300 or 600 mg/kg/day in corn oil 5 days per week for 13 weeks. N o significant differences were observed in food consumption or body weight gain compared with controls for either sex at any dose. A t the two highest doses, there was a microscopically detected increase in the incidence and severity of renal cortical degeneration. B y using the experiment just described as the basis for calculations with an additional factor reflecting that the exposure in the experiment occurred for 5 out of 7 days each week, the provisional A D I could be computed as follows: A D I (mg/day) = (150 mg/kg/day X 70 kg/person X 5/7)/ (100 X 10) = 7.5 mg/person/day where 100 is the uncertainty factor appropriate for use with a N O A E L f r o m animal studies with comparable human data and 10 is an additional uncertainty factor because the exposure duration in the experiment was significantly less than lifetime. The assumed daily water intake per person was 2 L/day. M i n i m a l data were available on food and air contributions to exposure, so an arbitrary designation of 20$ was chosen as the maximum allocation from drinking water. Other factors could have been selected. drinking water target = ( A D I X water allocations)/(2 L/day) = 7.5 mg/day X 20£/2L/day = 0.75 mg/L More recent data indicate that p-dichlorobenzene is carcinogenic in test animals.


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Nonthre8hold Toxicants T w o fundamentally distinct processes are especially involved in the regulatory decision maker's role in prescribing controls for substances that may cause public health risks without a k n o w n safe exposure threshold: (1) assessment of the risk and (2) management of the risk. Risk assessment is the use of a base of scientific research to define the probability of some harm coming to an individual or population as a result of an exposure. Risk management, however, is the public process of deciding what actions to take when risk has been determined to exist. It includes integration of the risk assessment with consideration of engineering feasibility and determination of how to apply the public health official's imperatives to reduce risk in light of legal, social, economic, and political factors (18). These two functions should be formally separated within regulatory agencies. W i l l i a m Ruckelshaus, a former administrator of U S E P A , recently described the dilemma in light of all of the uncertainties as follows (18): When the action. . . has dire economic or social consequences, the person who must make the decision may be sorely tempted to ask for a "reinterpretation" of the data . . . . Risk assessment can be like a captured spy: If you torture it long enough, it will tell you anything you want to know. So it is good public policy to so structure an agency that such temptation is avoided. The assessment of human cancer risk associated with a substance is a complicated scientific endeavor requiring careful review of all pertinent information b y professionals. Such assessment involves primarily the evaluation of clinical, epidemiological, and animal studies as well as short-term tests, structure activity, comparative metabolism pharmacokinetics, and mechanism of action when possible. The U . S . Office of Science and Technology Policy (OSTP) recently prepared a very comprehensive document reviewing the science and associated principles concerning chemical carcinogens (19). It was intended to be of use to regulatory agencies in the United States as a framework for assessing cancer risks. After extensive discussions of basic principles, mechanisms, short-term tests, long-term bioasays, epidemiology, and exposure assessment, it describes a process for using scientific data in the assessment of cancer risk. The O S T P also described four steps in the risk assessment process: 1. H a z a r d identification: qualitative evaluations of the agent's ability to produce carcinogenic effects and the relevance to humans. 2. Exposure assessment: the number of individuals likely to be exposed with the types, magnitudes, and durations of the exposure.


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3. H a z a r d or dose-response assessment: the attempt to assemble the hazard and exposure information along with mathematical models to estimate an upper bound on the carcinogenic risk at a given dose. 4. Characterization of the risk associated with human exposure.


Risks from Potential Nonthreshold Toxicants In 1977, the N A S Safe Drinking Water Committee outlined four principles that it said should be useful in dealing with the assessment of hazards that involve chronic irreversible toxicity or the effects of longterm exposure (20). These principles (paraphrased as follows) were intended to apply primarily to cancer risks f r o m substances whose mechanisms involve somatic mutations and may also be applicable to mutagenesis and teratogenesis: 1. Effects in animals, properly qualified, are applicable to man. This premise underlies all of experimental biology, but it is often questioned with regard to human cancer. Virtually every f o r m of human cancer has an experimental counterpart, and every f o r m of multicellular organism is subject to cancer. There are differences in susceptibility between animal species, between different strains of the same species, and between individuals of the same strain. However, large bodies of data indicate that exposures that are carcinogenic to animals are likely to be carcinogenic to humans, and vice versa. 2. Methods do not now exist to establish a threshold for longterm effects of toxic agents. Thresholds in carcinogenesis that w o u l d be applicable to a total population cannot be established experimentally. There is no scientific basis for estimations of safe doses using classic A D I techniques for carcinogens. Experimental bioassays with even large numbers of animals are likely to detect only strong carcinogens. E v e n negative results in such bioassays do not assure that the agent is unequivocally safe for humans. Therefore, possibly fallible measures of estimating hazard to humans must be used. 3. The exposure of experimental animals to toxic agents in high doses is a necessary and valid method of discovering possible carcinogenic hazards in humans. O n l y dosages that are high in relation to expected human exposures must be given to animals under the experimental conditions that are used. There is no choice but to use numbers of animals that are small relative to exposed human populations, and then to use biologically reasonable models in extrapolating the results to estimate risk at l o w doses. A n


7m



incidence as l o w as 0.01$ w o u l d represent a risk to 20,000 people in a population of 200 million, whereas the lower limit of reproducibility in common animal studies w o u l d be an incidence of 1035. The committee concluded that the best method available today is to assume no threshold and a direct proportionality between dose and tumor incidence. The actual human risk may be greater than predicted b y the small animal study because of the longer human lifetime and exposure period. 4. Material should be assessed in terms of human risk rather than as safe or unsafe. Extrapolation techniques may permit the estimation of upper limits of risk to human populations. T o do so, data are needed to estimate population exposure; valid, accurate, precise, and reproducible animal assay procedures are required; and appropriate statistical methods are necessary. Decisions cannot involve merely risk; benefit evaluations should include the nature, extent, and recipient of the benefits. It is often necessary to accept risks when the benefits warrant the risk, but risks imposed on persons who gain no benefits are generally not acceptable. The committee concluded the following: Mankind is already exposed to many carcinogens whose presence in the environment cannot be easily controlled. In view of the nature of cancer, the long latent period of its development, and the irreversibility of chemical carcinogenesis, it would be highly improper to expose the general population to an increased risk if the benefits were small, questionable, or restricted to limited segments of the population. Such benefit-risk considerations not only must be based on scientific facts but also must be ethical, with as broad a population base as possible used in the decision-making process.

Identification of Compounds Likely To Be Carcinogenic to Humans The fundamental question of risk assessment for potential human carcinogens requires definition of substances that exceed an evidentiary threshold. Once the scientific evidence establishes a substantial basis for conclusion of k n o w n or potential human cancer, it is then in order to determine a procedure for risk quantification. Quantitative risk assessments must always be read with the qualitative evidence of the likelihood of carcinogenicity. The International Agency for Research on Cancer ( I A R C ) has provided guidelines (21) for assessing the epidemiological and animal toxicological data base leading to a conclusion of the strength of the evidence of carcinogenicity of numerous substances. USEPA has re-


34.

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709


cently proposed a similar approach with some added refinement. Three classifications are defined as follows: 1. Carcinogenic to humans. This category includes substances for w h i c h there was sufficient evidence f r o m epidemiological studies to support a causal association between exposure and cancer. 2. Probably carcinogenic to humans. This category includes substances for w h i c h the evidence ranged f r o m almost sufficient to inadequate at the other extreme. The category was subdivided as follows: (A) at least limited evidence of carcinogenicity to humans and (B) sufficient evidence in animals and inadequate data in humans. 3. Cannot be classified as to its carcinogenicity to humans. The I A R C w o r k i n g group considered that the k n o w n chemical properties of a compound and the results from short-term tests could allow its transfer to a higher ranking group. The definitions of the key terms—sufficient, limited, and inadequate—are provided for both human and animal data. D a t a f r o m Humans 1. Sufficient: causal relationship between the agent and human cancer. 2. L i m i t e d : a causal relationship is credible, but alternative explanations such as chance, bias, or confounding could not be adequately excluded. 3. Inadequate: one of three conditions including (A) few pertinent data were available; (B) available studies d i d not exclude chance, bias, or confounding; and (C) some studies d i d not show evidence of carcinogenicity. D a t a f r o m Animals 1. Sufficient: increased incidence of malignant tumors (A) in multiple species or strains; (B) in multiple experiments (preferably b y different routes and different doses); or (C) unusual i n c i d e n c e , site, tumor type, or age at onset. Dose-response, short-term tests, and chemical structure may also be factored. 2. L i m i t e d : suggestive of carcinogenicity but limited because (A) single species, strain, or experiment; (B) inadequate dosage levels, duration of exposure, follow-up period, poor survival, too few animals, or inadequate reporting; or (C) neoplasms often occurring spontaneously and difficult to classify as malignant b y histological criteria alone (e.g., lung and liver tumors in mice).


710



3. Inadequate: studies cannot be interpreted or the chemical was not carcinogenic within the limits of the test. I suggest an initial determination of at least sufficient human evidence, limited human evidence supported b y animal or short-term tests, or sufficient animal evidence for a conclusion of probable human carcinogenicity as warranting the most conservative regulatory control philosophy. L i m i t e d animal evidence without substantial support f r o m short-term tests or mechanistic data indicative of potential human risk w o u l d warrant less heroic controls. Yet, these controls w o u l d be more protective than those developed from an A D I determination and sufficient to preclude any significant risk in the event that further studies raised the classification.

Risk Extrapolation Numerous mathematical models have been developed in attempts to estimate potential risks to humans f r o m low-dose exposures to carcinogens. E a c h m o d e l incorporates numerous unverifiable assumptions. Low-dose calculations are highly model dependent, widely differing results are commonly obtained, and none of the models can be firmly justified on either statistical or biological grounds (22). Thus, the decision to use this approach and the choice of h o w to do the calculations are matters of judgment. A m o n g the choices that the decision makers must consider are w h i c h model(s) to employ, w h i c h assumptions to incorporate, and which acceptable risk to allow. Numerous bodies, including the N A S Safe D r i n k i n g Water C o m mittee, F o o d Safety C o u n c i l Scientific Committee, and U . S . Office of Science and Technology Policy, have examined risk extrapolation science and methodology in great detail. The following brief discussion from the N A S Safe D r i n k i n g Water Committee report is intended as an introduction to the description of cases in which risk extrapolation procedures have been evaluated in several regulatory decisions involving water.

Risk Calculation Models A l l of the mathematical models that relate dose to response rate are either dichotomous response models or time-to-response models (23). Dichotomous response models are concerned with whether or not a particular response (tumor) is present b y a particular time (e.g., the animal's normal lifetime). In time-to-response models, the relationship between initiation of exposure and the actual occurrence of the response is determined for each animal.


34.

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Dicholomous Response Models. It is often assumed that the carcinogenetic process consists of one or more stages at the cellular level beginning with a single-cell somatic mutation at which point the cancer is initiated. The Armitage and D o l l multievent theory leads to a model that relates the probability of response, P(d), to the daily dose, d, b y P(d) = 1— exp[—(Xo + kid + k d +. . . . k d )] where k represents the number of transitional events in the carcinogenic process, and λο,λι, . . .λ* are unknown nonnegative parameters. F o r very small values of d, this dose-response rate w i l l be approximately equal to k\d> assuming λο is the background rate. This model was suggested for use by the Safe D r i n k i n g Water Committee, and it is commonly used b y U S E P A in its analyses of most water exposure risks. 2


711

2

k

k

Linear, No-Threshold Model. This simplest model is based on the assumption that risk is directly proportional to the dose: P(d) = ad. When it is assumed that the true dose-response curve is convex, linear extrapolation in the low-dose region may overestimate the true risk. However, it is not k n o w n if the experimental dose is in the convex region of the curve. Tolerance Distribution Model. This model assumes that each member of the population at risk has an individual tolerance below w h i c h no response w i l l be produced and that these tolerances vary in the members of the population according to some probability distribu tion (F). The probability distribution is also assumed to involve parameters of location (a) and scale (β > 0) and can be generally denoted b y F (a + β log z ) , where ζ is the tolerance level to a particular toxic agent. The probability, P(d), that a random individual w i l l suffer a response f r o m a dose, d, is P(d) = F (a + β log d) = /-« * dF(x). Therefore, the proportion of the population expected to respond to a specific dose is indicated b y the proportion of individuals having tolerances less than this dose level. βΧο

ά

Logistic Models. This model is based upon the assumption of a logistic distribution of the logarithms of the individual tolerances. P(d) = F (a + β log d) = [1 + exp(c* + β log d]~

l

where β < 0. The Committee concluded that tolerance distribution models have little theoretical justification for carcinogenic response. Hitness Models. Models for radiation-induced carcinogenesis have been proposed on the basis of a target theory that assumes that the site of action has some number of particles (N > 1) that are hit b y k or


712



more radiation particles. The probability of a hit is assumed to be proportional to the dose. The most commonly used versions are the single hit (N = 1, k = 1), the two hit (N = 1, Jt = 2), and the two target (N = 2, k = 1). Some have suggested a multielement theory of radiation-induced carcinogenesis that involves both a linear and a quad ratic dependence upon dose. The possibility of cell killing at high doses can also be included in this model. Time-to-Tumor Models. Experiments that produce nearly 100% incidence provide little dose-response information. However, examina tion of times to response may show a monotonie relationship between means or medians and dose levels. Also, early appearing tumors may be more biologically significant and a greater hazard than later tumors. The W e i b u l l model is a generalization of the one-hit model (22): P(d) = 1 — βχρ(—βάΜ) where M and β are parameters.

^

^~Φ~ =

a constant. Thus, in the low-dose region, this model becomes linear for M = 1, concave for M < 1, and convex for M > 1. α—Ο

Case: Radium-226 and Radium-228. The concept of risk projec tions from experimental dose-response curves has been highly developed in the case of estimating risks to the population f r o m l o w doses of radiation. Such methods were later extended to estimate risks f r o m other carcinogens in drinking water and other media. Radioactivity can contribute risks f r o m teratogenic, genetic, and somatic (carcinogenic) effects. Figure 3 illustrates rough dose-response model fits with human data for ionizing radiation and leukemia incidence from atomic b o m b sur vivors (24). Data exist d o w n to about the 10" lifetime risk per person exposed. This value is close to the region of regulatory interest, and relatively small risk differences are predicted b y the three illustrated models in the dose range up to two orders of magnitude below the last observed dose value (ca. 5 rad). Carcinogenic risks are considered to be stochastic effects—those for which the probability of an effect occuring, rather than its severity, is regarded as a function of the dose without threshold. The basic assumption of the International Commission on Radiologi cal Protection (ICRP) is that for stochastic effects, a linear relationship without threshold is found between dose and the probability of an effect within the range of exposure conditions usually encountered in radiation work. However, I C R P cautions that if the dose is highly sigmoid, the risk from l o w doses could be overestimated b y linear extrapolation f r o m data obtained at high doses. Furthermore, I C R P 5


In Organic Pollutants in Water; Suffet, I., et al.; Advances in Chemistry; American Chemical Society: Washington, DC, 1986. 00

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