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Quantitative Assessment of Risks Using Docketing Approaches OVERVIEW OF TlIE SCIENCE UNDERLYING EPA'S 2001 PROPOSED REGULATION On January 22, 2001, following the publication of a proposed rule for arsenic in drinking water (EPA 2000a) and a period of public comment, EPA published a final rule for arsenic in drinking water in the Federal Register, setting a maximum contaminant level goal (MECG) of zero for arsenic in drinking water and a maximum contaminant level (MCL) for arsenic of 10 fig in drinking water (EPA 2001~. Typically, when developing an MUNG and an MCL, a risk assessment is conducted. Two important components of a risk assessment are hazard identification and dose-response assessment ARC 1983~. Exposure assessment end risk characterization are also important steps in a risk assessment, but they are beyond the scope of this subcommittee's charge and, therefore, will not be discussed here. The purpose of hazard identification is to determine whether the agent in question causes adverse effects. Deciding which end point is the most sensitive and which studies or data sets are most appropriate for assessing the risks from a chemical are major conclusions from a hazard identification. In the case of EPA's assess- ment of arsenic, the risk being assessed is the risk to the U.S. population from consumption of arsenic in drinking water. The purpose of dose-response assessment is to determine the relationship between the dose and the incidence of an adverse effect in humans. Major conclusions from the dose-response 169
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~ 70 ARSENIC IN DRINKING WA TER: 2001 UPDA TE assessment include the model or models that can be used to best determine the risks to the U.S. population from arsenic in drinking water and understanding ofthe impacts of different model choices on the risk estimates from that analy- sis. The details of EPA's hazard identification (choice of endpoint and choice of study) and dose-response assessment (choice of model, selection of a com- parison group, and adjustments for water intake, diet, and mortality versus incidence) are discussed below. Hazard Identification Choice of End Point EPA' s hazard analysis is included in Section m ofthe proposed rule (EPA 2000a) and in Section m.D. ~ of the final rule (EPA 2001~. EPA "relied upon the NRC ~ ~ 999] report as presenting the best available, peer reviewed science as of its completion and has augmented it with more recently published, peer reviewed information" in its proposed rule. EPA (2000a) concludes that acute or short-term effects are not seen at 50,ug/L (the MCL at the time of the pro- posal) and, therefore, addresses the "Iong-term, chronic effects of exposure to Tow concentrations of inorganic arsenic in drinking water." With respect to Tong-term effects, EPA concludes that "arsenic is a multi- site human carcinogen by the drinking water route," and on the basis of epide- miological studies of Asian, Mexican, and South American populations, those "with exposures to arsenic in drinking water generally at or above several hundred micrograms per liter are reported to have increased risks of skin, bladder, and lung cancer." EPA also notes that increased risk of liver and kidney cancer have been associated with arsenic exposure and that skin cancer has been associated with inorganic arsenic contamination in Argentina (re- viewed by Neubauer 1947, as cited in EPA 2000a), in Poland (EPA 2000a), and in a dose-dependent manner following exposure to arsenic in drinking water in Taiwan (Tseng et al. 196S, Tseng 19774. Other epidemiological studies also support an association between arsenic exposure and skin cancer (Roth 1956; Albores et al. 1979; Cuzick et al. 1982; Cebrian et al. 1983~. EPA discussed data from two studies in Taiwan demonstrating a statisti- cally significant increase in mortality risks for bladder, kidney, lung, liver, and colon cancer (Chen et al.1985), and a significant dose-response relationship for death from bladder, kidney, skin, and lung cancer in both sexes and from
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QUANTITATIVE ASSESSMENT OF RISKS USING MODELING APPROACHES ~ 71 liver and prostate cancer in males (Wu et al. ~ 989~. An increase in internal cancers was also seen in Argentina (bladder, lung, and kidney cancer) (Hopenhayn-Rich et al. 1996, 1998) and in Chile (bladder, kidney, and lung cancer) (Smith et al. 1998~. Tsai et al. (1999) reported that lung, bladder, intestinal, rectal, and laryngeal cancer were associated with chronic exposure to arsenic in drinking water in Taiwan. EPA also reviewed a study by Lewis et al. ~ 1999) that reported mortality of a population in Utah exposed to lower concentrations (average, 18-191 ~g/~) of arsenic in drinking water in which there was a statistically significant increase in prostate-cancer mortality, but no increase in bladder or lung cancer mortality. EPA also discussed a study by Kurttio et al. (1999) that found a significant association in a case-control study in Finland between bladder cancer and exposure to very Tow concentra- tions of arsenic in drinking water (odds ratio of ~ .53,95% confidence interval (C~ = 0.75-3.09 at 0.1-0.5,ug/~; 2.44,95% C! = ~ . ~ l-5.37 at greater than 0.5 ,ug/~. No association was seen for kidney cancer. EPA reviewed noncancer effects that are observed following chronic exposure to arsenic including dermal effects (Yeh 1973; Tseng 1977; Cuzick et al. 1982), gastrointestinal effects (Morris et al. 1974; Nevens et al. 1990; Mazumder et al. 1997), peripheral vascular disease (Tseng 1977; Zaldivar ~ 974; Cebrian ~ 987; Lewis et al. ~ 999), and diabetes (Lad et al. ~ 994; Rahman and Axelson 1995; Rahman et al. 1998~. in the final rule, EPA again summarized the acute and chronic effects of arsenic, and added a discussion of a study from Japan (Tsuda et al. 1995~. The study found an association between exposure to arsenic in drinking water and lung and bladder cancer. in addition, EPA (2001) added a short discus- sion on the potential susceptibility of children to arsenic. EPA agreed with the conclusion of the majority of the EPA Science Advisory Board (SAB) mem- bers (EPA 2000b) that children are generally at greater risk for a toxic re- sponse to any agent in water because of their greater drinking-water consump- tion (on a unit-body-weight basis), but that the available data, including a study of infant mortality in Chile (Hopenhayn-Rich et al. 2000), do not dem- onstrate a heightened susceptibility to arsenic. After discussing all the toxic effects of arsenic, the water concentrations at which they occur, and the NRC (1999) report, EPA chose cancer as the most sensitive end point, stating that it "focused its risk assessment on the carcinogenic effects of inorganic arsenic" (EPA 2000a). EPA (2001) states that lung and bladder cancer are the internal cancers most consistently seen and best characterized in epidemiological studies, and its quantitative risk assessment is based on data for those two cancers.
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72 ARSENIC IN DRINKING WA TER: 2001 UPDA TE Choice of Study An important decision in a quantitative risk assessment is the choice of critical study (or studies) to be used in the dose-response assessment. At the time of EPA's proposed rule, few animal carcinogenicity bioassays had been conducted for arsenic, and there were no positive animal models for dose- response modeling. There was, however, a "large data base on the effects of arsenic on humans" (EPA 2000a, p. 38902~. EPA concluded that questions remain about the shape of the dose-response relationship at Tow concentra- tions. The advantages of using the studies from southwestern Taiwan (Chen et al. 1985; Wu et al. 1989) for quantitative risk assessment, according to EPA, are the duration of exposure and follow-up, the size of the population (more than 40,000 individuals), the extensive pathology data, and the homoge- nous lifestyles ofthe population. Those studies are limited, however, by their design (i.e., they are ecological epidemiology studies), which makes quantita- tive evaluation of dose-response relationships more difficult. EPA also stated that the studies from Chile (Smith et al. 1998) and Argentina (Hopenhayn- Rich et al.1996; 1998) are more limited than the Taiwanese studies (Chen et al. 1985; Wu et al. 1989) and not suitable for quantitative dose-response as- sessment, but that they provide supportive evidence for the effects seen in southwestern Taiwan. EPA concluded that "tt~hese epidemiological studies provide the basis for assessing potential risk from Tower concentrations of inorganic arsenic in drinking water" (EPA 2000a, p.38902~. In its final rule, EPA also concluded that the Utah study by Lewis et al. (1999) "is not power- ful enough to estimate excess risks with enough precision to be useful for the Agency's arsenic risk analysis." Therefore, in its final rule, EPA (2001) still considered the southwestern Taiwan data to be the critical data set for conducting a quantitative risk assess- ment for exposure to arsenic in drinking water. Dose-Response Modeling Model Choice and Selection of a Comparison Group In its proposed arsenic rule, EPA concluded, on the basis of the NEC (1999) report, that there is "no basis for determining the shape of a sublinear dose-response curve for inorganic arsenic" (EPA 2000a). Therefore, EPA estimated the risks of cancer from exposure to arsenic in drinking water using
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QUANTITA TI BE ASSESSMENT OF RISKS USING MODELING APPROA CHES ~ 73 a linear extrapolation from the data observed in the southwestern Taiwanese epidemiological studies down to the origin. EPA's default to a linear extrapo- lation in the absence of adequate mode-of-action data (EPA ~ 996) is in part a policy decision. For the proposed rule, EPA used the bladder cancer risk estimates presented in the NRC (1999) report (see Table 5-1 for examples). EPA cited a lifetime risk estimate with a 95% upper confidence limit of ~ to ~ .347 per 1,000, calculated by a Poisson regression mode] not using any base- line data (i.e., no comparison group) (NRC 1999), and EPA used four distribu- tions of risk estimates (mean and 95% CT) from NRC (1999) as representative risks in a Monte Cario analysis to estimate the potential health benefits from the proposed rule. Those four distributions all come from analyses of the southwestern Taiwanese data (Chen et al. 1985; Wu et al. 1989) using a Pois- son regression model with age entered as a quadratic function and dose en- tered as a linear function, either with or without baseline data, or a Poisson regression mode] with a point-of-departure approach, with or without baseline data. On October 20, 2000, EPA published a Notice of Data Availability in the Federal Register (EPA 2000b) in which it discussed statistical modeling published by Morales et al. (2000) and indicated that those analyses would be considered in its final rule for arsenic in drinking water. Morales et al. (2000) estimated bladder, lung, and liver cancer risks for the southwestern Taiwanese population based on the same data set that was analyzed by NRC (1999~. Morales et al. (2000) calculated cancer risk estimates using 10 risk models and also considered how well those models fit the data sets. Of those models, EPA chose a single model that did not use an external comparison population either from all of Taiwan or part of southwestern Taiwan, because most of the models that incorporate a comparison population result in a dose-response curve that is supralinear at low doses. EPA indicated that there is no biologi- cal basis for a supraTinear curve. In addition, differences other than arsenic exposure between the study population and the comparison population could affect the results. The decision to use a dose-response model that does not incorporate a comparison population agreed with the SAB's recommendation that the analysis should be conducted without a comparison group (see discus- sion below). Of the models that did not incorporate a comparison population, mode] ~ from Morales et al. (2000), in which "the relative risk of mortality at any time is assumed to increase exponentially, with a linear function of dose and a quadratic function of age," was used because it best fit the data based on the Akaike information criterion (EPA 2001~. However, EPA did not
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~ 74 ARSENIC IN DRINKING WA TER: 2001 UPDA TE TABLE 5-1 NRC's Risk Estimates for Bladder Cancer Mortality from 1999 NRC Reporta Margin of Point of Risk at 50 ~g/L Exposure at Method ofAnalysis Departure, pub (X 1,OOO) 50 ALL Poisson model, linear dose, no background data Poisson model, linear dose, background data included a Estimated points of departure at the 1% excess risk level, corresponding margin of exposure at 50 ~g/L arid corresponding excess lifetime risk estimates at 50 ,ug/L for bladder cancer in males. Figures in parentheses are 95% confidence limits (lower for the point-of-departure estimates; upper for estimated risk at 50 vigil). Risk estimates are those predicted in Taiwan using U.S. ingestion rates. b Me point of departure represents an estimate or observed level of exposure or dose associated with an increase in adverse effects in the study population. An example of a point of departure is ail EDo~ c A margin-of-exposure analysis compares the levels of arsenic to which the U.S. population is exposed with the point of departure to characterize the risk to the U.S. population. The larger the ratio, the greater degree of assumed safety for the population. Abbreviation: EDo,, 1% effective dose. Source: Modified from NRC (1999~. 404 (323) 1.237 (1.548) 8.08 (6.46) 443 (407) 1.129 (1.229) 8.86 (8.14) publish the theoretical risk estimates on which it based its analyses in the Federal Register (EPA 2000a, 200 ~ ). The risk estimates that it presents (EPA 2001) are adjusted for the occurrence of arsenic in U.S. drinking water; con- sideration of such an adjustment is beyond the charge to this subcommittee. Because EPA did not present theoretical lifetime excess bladder or lung can- cer risk estimates, the subcommittee used linear extrapolation from the EDo~s presented in Morales et al. (2000) to estimate these risks at 3, 5, 10, and 20 ~g/L (Table 5-2~. Adjustments for Water Intake To estimate cancer risks associated with a given arsenic concentration in drinking water, assumptions must be made about water consumption in both
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QUANTITATIVE ASSESSMENT OF RISKS USING MODELING APPROACHES 1 75 TABLE 5-2 Theoretical Estimates of Excess Lifetime Risk (Incidence per 10,000 people) of Lung and Bladder Cancer at Various Concentrations of Arsenic in Drinlcing Water Based on EDo, Values from Morales et al. (2000)a Arsenic Bladder Cancer Lung Cancer Concentration (,ug/L) Females Males Females Males 3 1.2 .76 1.2 0.82 5 2.0 1.3 1.9 1.4 10 4.0 2.5 3.9 2.7 20 7.9 5.1 7.8 5.5 a Excess cancer risk estimates were calculated using the EDo,s estimated by Morales et al. (2000) using a model in which the relative risk of mortality at arty time is assumed to increase exponen- tially, with a linear function of dose arid a quadratic function of age (i.e., a multiplicative Pois- son linear regression); no external comparison population was used (see Model 1 of Table 8 from Morales et al. 2000~. Risk estimates are rounded to two significant figures. The Taiwar~- ese exposure per kilogram of body weight is assumed to be 2.2 times the U.S. exposure. the U.S. population and the study population. EPA estimated mean daily average per capita consumption of water by individuals in the United States is ~ L/person/day for "community tap water" and 1.2 L/person/day for "total water" (which includes bottled water) based on data from the 1994-1996 Continuing Survey of Food Intakes by Individuals (CSF~ (EPA 20003~. The 90th percentile is 2. l L/person/day and 2.3 L/person/day for community tap water and total water, respectively. Rather than only using a point estimate for its risk assessment, EPA conducted a Monte Cario analysis using the CFSU data to incorporate water intake. Those distributions take into account age, sex, and weight. EPA assumed that the Taiwanese consumed relatively more water per unit of body weight than Americans, estimating consumption of 3.5 and 2.0 L/day for men and women in Taiwan, respectively. As dis- cussed in the following section, EPA also added water consumption to account for water used in cooking in Taiwan. It should be noted that assumptions that increase the amount of arsenic consumed (~inking water and diet) by the study population reduce the "potency factor" or estimated slope of the linear dose-response function when applied to other populations, thereby decreasing the estimated risk in other populations. Conversely, underestimation of the
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~ 76 ARSENIC IN DRINKING WA TER: 2001 UPDA TE actual arsenic intake in the study population increases risk estimates in other populations. Therefore, assumptions about total arsenic exposure in the study population can have a large impact on risk estimates. Adjustments for Dietary Intake of Arsenic The staple foods in the southwestern Taiwan region where the study popu- lation resided were rice and sweet potatoes. Those foods absorb a great deal of water when cooked. As part of its risk assessment of arsenic in drinking water, EPA (200 ~ ~ adjusted its lower-bound estimates to account for exposure to arsenic in food from cooking water. For that adjustment, EPA assumed that people in the study population eat ~ cup of dry rice and 2 pounds of sweet potatoes per day. To adjust for arsenic absorbed during cooking, EPA added ~ ~ of water consumption to the Taiwanese population. Therefore, in its analyses, EPA assumed that Taiwanese men and women consumed the equiva- lent of a total of 4.5 L/day and 3.0 L/day of water, respectively. Although EPA used a Monte Cario analysis to account for variability in U.S. water consumption rates, its analyses did not incorporate analogous variability in the Taiwanese water consumption rates. EPA also discussed the fact that the food in Taiwan contains more arsenic than the food in the U.S., even prior to cooking. NRC (1999) presented data indicating that individuals in Taiwan consume food containing inorganic arsenic at 50 Gay, compared with ~ O Gay for Americans. To account for the intake of arsenic from food, EPA multiplied the Tower-bound risk esti- mates by the fraction of arsenic consumed per kilogram contributed by drink- ing water (calculated by dividing the arsenic ingested from drinking water (pa/kg/day) by the total arsenic consumed from drinking water, cooking water, and food) (J. Bennett, EPA, personal commun. May 22, 2001). Adjustments for Mortality Versus Incidence EPA's dose-response assessment is derived from data on mortality from bladder and lung cancer in the Taiwanese study (Chen et al. 1985, 1992; Wu et al. 1989). Extrapolating the mortality-risk estimates calculated in the Tai- wanese population to the incident risks in the U.S. population requires an adjustment for the survival rate for bladder and lung cancer. EPA (2001)
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QUANTITATIVE ASSESSMENT OF RISKS USING MODELING APPROACHES ~ 77 noted that the Taiwanese people in the study population had Tow incomes and poor diets, and that "the availability and quality of medical care is not of high quality, by U.S. standards." Therefore, EPA assumed that the bladder cancer incidence was relatively close to the bladder cancer mortality in the Taiwanese study area. EPA calculated the survival rate for bladder cancer by considering the survival-rate data compiled by the World Health Organization (WHO) for bladder cancer in developing countries from ~ 982 through ~ 992 (lARC ~ 999) and by comparing the annual bladder cancer mortality and incidence for the general population of Taiwan in 1996. From those data, EPA concluded that "bladder cancer incidence could be no more than two-fold bladder cancer mortality; and that an 80% mortality rate would be plausible" (EPA 2001; page 7009~. Therefore, when calculating the bladder cancer cases avoided at a given MCL, EPA adjusted the upper bound by a factor of 1.25 to reflect the mortality for bladder cancer. With respect to lung cancer, EPA concluded that "because lung cancer "mortality] rates are quite high, about 88% in the U.S. [EPA 1998], the assumption was made that all Jung cancers in the Taiwan study area resulted in fatalities." OVERVIEW OF TlIE SAB'S REPORT ON THE 2001 RISK ASSESSMENT EPA charged the SAB to review the proposed arsenic rule (EPA 2000a) and to specifically review (1~ EPA's focus on inorganic arsenic as the princi- ple form of arsenic causing health effects; (2) the implications of exposure to natural arsenic through food; and (3) the appropriateness of EPA' s precaution- ary advice to use low-arsenic water in the preparation of infant formula. They also requested the SAB to address several questions related to treatment op- tions for arsenic in drinking water. On December ~ 2, 2000, the SAB issued a report on the proposed drinking- water regulation (EPA 2000c), responding to those questions on the scientific basis of EPA's health risk assessment and on the economic and engineering aspects of the final rule. The exposure assessment, costs, benefits, control technologies, and policy issues discussed by the SAB are beyond the charge to this NRC subcommittee and will not be discussed. The SAB's responses to EPA's three questions on the health effects of arsenic are discussed in this section. EPA addressed some ofthe SAB's comments in its final arsenic rule (EPA 2001~.
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78 ARSENICIN DRINKING WATER: 2001 UPDATE Inorganic Arsenic As Principal Form of Arsenic Causing Health Effects The SAB pointed out that new data released since the ~ 999 NRC report indicated that inorganic forms of arsenic are not solely responsible for the toxic effects of arsenic (see Chapter 3 for discussion of new data). Because exposures to other forms of arsenic can produce health effects, the SAB rec- ommended that future risk assessments provide quantitative information on how the intake of inorganic arsenic is related to the concentration of arsenic metabolites in the urine and to bladder cancer. However, because the princi- pal forms of arsenic in drinking water are inorganic, the SAB believed "that it is appropriate for the Agency tEPA] to make "inorganic arsenic] its regula- tory focus." Implications of Exposure to Natural Arsenic Through Food The SAB concluded that on average, for the general U.S. population, ingestion of inorganic arsenic via foodwas considerably greater then ingestion of inorganic arsenic via drinking water. The SAB agreed that data were not available to determine "a well-defined nonlinear dose-response curve." Fur- thermore, the SAB concluded that insufficient data on the distribution of food intakes existed to adequately consider them in the analysis. Therefore, the SAB concluded that EPA had no choice but to calculate marginal risk reduc- lions based solely on arsenic concentrations in drinking water. The SAB noted that "there is a limit to the benefits that can be realized by reducing arsenic in drinking water" as long as the concentrations in food remain un- changed. The SAB also reiterated the recommendation in the NRC (1999) report to obtain additional studies on the noncancer effects of arsenic and incorporate that information into a risk-assessment and cost-benefit analysis. Health Advisory on Low-Arsenic Water and Infant Formula The SAB also reviewed EPA's plan to issue a health advisory for the use of Tow-arsenic water in the preparation of infant formula. Some of SAB's responses focused on risk communication and the public's ability to follow such an advisory; those comments are beyond the scope of this subcommit- tee's charge. However, the SAB also discussed the issue of children's suscep- tibility to arsenic. Most of the SAB members agreed that special circum-
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QUANTITATIVE ASSESSMENT OF RISKS USING MODELING APPROACHES ~ 79 stances make infants unique in regard to their response to drinking-water contaminants, but the SAB did not reach consensus on endorsing EPA' s intent to issue a health advisory. An SAB consultant wrote and one SAB member endorsed a minority report on the issue of infant and children's risk. The minority report stated that differences in respiratory and circulatory flow rates, cell-proliferation rates, enzymatic pathways, developmental process, life expectancies, and the disposition of chemicals in infants and other differences in cells during devel- opment can make children more susceptible to toxic chemicals, including arsenic. It further concluded that data from Hopenhayn-Rich et al. (2000) and Concha et al. (1998) "indicate that young children are a uniquely sensitive population for adverse health effects of arsenic." The minority report "departs from the majority opinion contained in the iSAB] report in strength of its conclusions if not the general reasonableness of the need for increased con- cern for children, which is also held by the tSAB]." The basis for increased concern for children is uncertainty about pulmonary and cardiovascular risks to infants, high exposure of infants on a per kilogram basis, and the longer period of exposure and outcome relative to adults. The latter is particularly relevant if the latency period for cancer development from low arsenic expo- sure is long or if the appropriate dose metric involves a less-than-lifetime exposure as discussed in Chapter 4. Although the SAB recognized that children differ from adults in many ways that could make them more susceptible to toxic chemicals, "the majority of the tSAB] did not feel that data available to them on arsenic had demon- strated an increased sensitivity to arsenic in children." The SAB concluded that available data on U.S. drinking-water consumption indicate that infants who consume food made from drinking water could have a higher dose of arsenic per unit of body weight than adults. The majority ofthe SAB did not believe that the study by Hopenhayn-Rich et al. (2000) demonstrated "a heightened sensitivity or susceptibility to arsenic" but that the study "appears to be a hypothesis generating study that, in light of the limitations [the SAB described], merits and requires further study before drawing final conclu- sions." SAB's Comments on EPA's Interpretation of the NRC Report The SAB also commented on EPA's interpretation of the 1999 NBC report. The SAB, in general, agreed with the ~ 999 NRC report, which formed
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QUANTITATIVE ASSESSMENT OF RISKS USING MODELING APPROACHES 203 TABLE 5-8 Theoretical Maxi Likelihood Estimates of Excess Lifetime Risk (Incidence per 10,000 People) of Bladder Cancer for U.S. Populations Exposed at Vanous Concentrations of Arsenic In Drinking Water Using Different Ratios for Taiwanese-to-U.S. Drinking-Water Ingestion on a Per-Body-Weight Basisa Taiwanese to U.S. Drinking Taiwanese to U.S. Drinking Arsenic Water Ratio = 1.4b Water Ratio = 3c Concentration (pg/L) Females Males Females Males 3 7.7 15 3.6 6.8 5 13 25 6.0 11 10 26 50 12 23 20 51 100 24 45 a These risks are estimated using assumptions considered to be reasonable by the subcommittee; it is possible to get higher and lower estimates using other assumptions. Estimates were calcu- lated using data from individuals in the arsenic-endemic region of southwestern Taiwan (Chen et al. 1985, 1992; Wu et al. 1989), and data from an external comparison group from the overall southwestern Taiwan area. Risk estimates are rounded to two significant figures. All 95% confidence limits are less than +/- 12% of the maximum-likelihood estimate and are not pre- sented. Those confidence limits reflect statistical variability only, reflecting primarily the sample size. As such, they are not indicative of the true uncertainty associated with the estimates. b Risks were estimated assuming that the typical U.S. resident weighs 70 kg, compared with 50 kg for the typical Taiwanese, and that the typical Taiwanese drinks just over 1 L of water per day, the same as the 1 L per day in the United States; thus, the Taiwanese exposure per kilogram of body weight is 1.4 times the U.S. exposure, calculated using an additive Poisson model with linear dose. c Risks were estimated assuming that the typical U.S. resident weighs 70 kg, compared with 50 kg for the typical Taiwanese, and that the typical Taiwanese drinks just over 2 L of water per day, compared with 1 L per day in the United States; thus, the Taiwanese exposure per kilogram of body weight is 3 times the U.S. exposure, calculated using an additive Poisson model with linear dose. ing water are presented. Those values, however, should not be considered bounds on the possible risk estimates, because other assumptions could be made that would result in higher or lower values. When estimating the risks to the U.S. population, assumptions must be made about body weights and water consumption in both the United States and the Taiwanese populations. Then, when comparing cancer risk estimates, it is important to be aware of how those assumptions affect the estimates. For example, the higher the ratio of water ingestion in Taiwan relative to the United States in terms of liters per body weight per day, the smaller the U.S. cancer risk estimate will be. The
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204 ARSENIC IN DRINKING WA TER: 2001 UPDA TE assumptions that the subcommittee used to calculate the various estimates are discussed below and presented in the footnotes of the tables. Table 5-7 presents estimates for the excess lifetime risk (incidence) of lung cancer in females and mates in the U.S. population from exposure to arsenic at 3, 5, ~ 0, and 20 ~g/L of drinking water. The estimates based on the Chilean data (Ferreccio et al. 2000) were calculated assuming that the typical U.S. and Chilean resident both weigh 70 kg and that the drinking-water ingestion rates in both countries are the same that is, that exposures in Chile on a per-body- weight basis, given the same concentration of arsenic in the drinking water, are equal to those in the United States (i.e. a factor of 1~. The estimates based on the southwestern Taiwanese data set (Chen et al. 198S, 1992; Wu et al. 1989), however, were calculated assuming that the typical U.S. resident weighs 70 kg and drinks ~ ~ of water per day, and the typical Taiwanese resident weighs 50 kg and drinks just over 2 ~ of water per day. Therefore, given the same concentration of arsenic in drinking water, exposures in Tai- wan on a per-body-weight basis would be three times that of a U.S. resident. The data in Table 5-7 utilize the background rate of lung cancer in the United States. To illustrate the importance of the background rate, Table S-9 shows the same risk projections using the Taiwanese data set and the background TABLE 5-9 Theoretical Maximum-Likelihood Estimates of Excess Lifetime Risk (Incidence per 10,000 people) of Lung and Bladder Cancer for Populations Exposed at Various Concentrations of Arsenic in Droning Water, Using the Background Cancer Incidence Rate for Taiwana Arsenic Bladder Cancer Lung Cancer Concentration (,ug/L) Females Males Females Males 3 2.3 2.0 1.8 1.7 5 3.8 3.2 3.0 3.0 10 7.5 6.8 6.2 6.1 20 15 13 12 12 . a These risks are estimated using the assumptions noted in footnotes (a) and (c) of Table 5-8, and assuming a Taiwanese to U.S. drinking water ratio of 3. The estimated background incidence rates for bladder cancer in Taiwan (derived from the subcommittee's cancer risk estimates presented in Tables 5-7 and 5-8 and adjusted by the ratio of Taiwanese incidence data (You et al.2001) to U.S. incidence data (Ferlay et al. 2001~) are 6.9 (males) and 3.4 (females) per 100,000 and for lung cancer are 25.8 (males) and 11.9 (females) per 100,000.
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QUANTITATIVE ASSESSMENT OF RISKS USING MODELING APPROACHES 205 incidence rates for lung cancer in Taiwan, which are approximately 2- and 3- fold lower than those in the United States for males and females, respectively. For lung cancer, the MLEs of risk for mates based on the Chilean data (Fernccio et al. 2000) range from 20 per 10,000 at 3 ~g/L to 130 per 10,000 at 20 ~g/L using the average arsenic concentration during the peak exposure period of 1958 to 1970. The corresponding risk estimates in females are 14 per 10,000 to 95 per 10,000. Using the average arsenic concentration from 1930 to 1994, the risk estimates range from 75 to 500 per 10,000 males and from 15 to 330 per 10,000 females. Using the southwestern Taiwanese data, the risk estimates for arsenic at 3,ug/L of drinking water range from 1.7 to 4.0 per 10,000 for mares and from 1.8 to 5.4 per 10,000 for females, depending on which assumptions are used (Tables 5-8 and 5-9~. Different studies have estimated the risks of lung cancer following expo- sure to arsenic, and it is possible and useful to compare the risk estimates generated in the different analyses at a given arsenic concentration. Because the case mortality for lung cancer is close to 100% (SEER 2001), lung cancer incidence approximates lung cancer mortality, and the risk estimates for lung cancer mortality can be compared with risk estimates for lung cancer inci- dence. As can be seen in Table 5-7, the subcommittee's analysis ofthe south- western Taiwanese data yields an estimate for lifetime lung cancer incidence in the United States (using the U.S. background rate) at an arsenic concentra- tion in drinking water of 10 ~g/L of approximately 14 per 10,000 and 18 per 10,000 in males and females, respectively. It is noteworthy that nearly 10 years ago, Smith et al. (1992) published a risk assessment based on the same ecological southwestern Taiwanese data analyzed by the subcommittee. In that risk assessment, lifetime lung cancer mortality risks at 10 ~g/L were estimated to be 11 per 10,000 and 17 per 10,000 for males and females, re- spectively. Therefore, lung cancer risk estimates generated by this subcom- mittee and those published by Smith et al. (1992) are very consistent. The lung cane. er risk estimates derived from the Taiwanese data in Table 5-7 can be compared with those derived from the data in the recent case-control study in northern Chile by Ferreccio et al. (2000~. When the average arsenic con- centration in northern Chile Tom 1930 to 1994 is used as the dose-metric, the risk estimates for lung cancer incidence in the United States calculated by the subcommittee for an arsenic concentration of 10 ,ug/L are 250 per 10,000 mates and 170 per 10,000 females. Those estimates are approximately an order of magnitude higher than the estimates the subcommittee derived from the Taiwanese data. However, when the dose metric selected for the Chilean data is the peak years of arsenic exposure from 1958 to 1970, the correspond-
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206 ARSENIC IN DRINKING WA TER: 2001 UPDA TE ing U.S. lung cancer risk estimates are 67 per 10,000 mates and 48 per 10,000 females. Those estimates are approximately 3 to 4 times higher than the sub- committee's estimates derived from the Taiwanese data. For comparative purposes, the subcommittee also derived cancer risk estimates at 10 ,ug/L using the relative risks (see Table 2-~) for lung cancer associated with the peak period of arsenic exposure in northern Chile in the ecological study by Smith et al. (1998~. If the same formula is applied to those relative risks, as was used to estimate the subcommittee's other cancer estimates, the U.S. lung cancer estimates at 10 ,ug/L are 38 per 10,000 and 21 per 10,000 in mates and females, respectively. Overall the peak period exposure data in northern Chile and the data from southwestern Taiwan yield coherent lifetime excess risk estimates ranging from 1.4 to 6.7 per 1,000 for lung cancer in the United States at a drinking- water arsenic concentration of 10 Vigil. The finding that risk estimates de- rived from studies of individuals exposed to arsenic in Chile are similar to those estimated from Taiwanese data provides confidence in the validity ofthe risk estimates. Tables 5-8 and 5-9 present estimates for the excess lifetime risk (incidence) for bladder cancer in females and mates in the U.S. population from exposure to arsenic at 3, 5, ~ 0, and 20 ~g/L of drinking water based on the southwestern Taiwanese study (Chen et al. 1985, 1992; Wu et al. 1989), using either the U.S. background rate for bladder cancer (Table 5-~) or the Taiwanese back- ground rate (Table 5-9~. In one water-intake scenario, the subcommittee assumed that the typical U.S. resident weighs 70 kg and drinks ~ ~ of water per day, and the typical Taiwanese resident weighs 50 kg and also drinks 1 L of water per day. Therefore, given the same concentration of arsenic in drink- ing water, exposures in Taiwan on a per-body-weight basis are 1.4 times that of a U.S. resident. in a second water-intake scenario, calculated using the same data set, it was assumed that the typical U.S. resident weighs 70 kg and drinks ~ ~ of water per day, and the typical Taiwanese resident weighs 50 kg and drinks just over 2.2 ~ of water per day. Therefore, as indicated in the lung cancer estimates earlier at the same concentration of arsenic in drinking water, exposures in Taiwan on a per-body-weight basis are 3 times that of a U.S. resident. For bladder cancer, the maximum-likelihood estimate of risk using a ratio of 1.4 for Taiwanese-to-U.S. drinking-water rates on a per-body-weight basis for males range from approximately 15 per 10,000 at an arsenic concentration of 3 ,ug/L of drinking water to 100 per 10,000 at 20 ~g/L. The corresponding estimates in females range from ~ to 5 ~ per 10,000. Using a drinking water
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QUANTITA TIVE ASSESSMENT OF RISKS USING MODELING APPROA CHES 20 7 ratio of 3.0, the corresponding ranges are approximately 7 to 45 per 10,000 in males and 4 to 24 per 10,000 in females, using the U.S. background rate. Risk estimates are 3-fold Tower for mates and 2-fold lower for females if the Tai- wanese background rate is used (Table 5-9~. Further discussion of the subcommittee's quantitative risk estimates and comparison of them with the results of other analyses are presented in Chapter 6. SUMMARY AND CONCLUSIONS · Since EPA issued a pending standard of 10 ,ug/L, based on lung and bladder cancer data from the southwestern Taiwanese study in 2000, two additional studies have appeared in the literature that are of sufficient size and quality and with adequate quantification of dose to be considered in comput- ing EDs for arsenic in drinking water. One is a study that examined urinary tract cancer, and TCC in particular, in northeastern Taiwan (Chiou et al. 2001~; the second is a case-control study of lung cancer in Chile (Ferreccio et al. 2000). · Although it can be argued that an external comparison group for dose- response analysis of the original Taiwanese data should not be used, the sub- committee believes that such arguments are outweighed by evidence in favor of using a comparison population. A recent paper by Tsai et al. (1999) de- creases concerns about the potential role of confounding in using either the southwestern Taiwanese population or the entire Taiwanese population as an external comparison group. · Poisson models provide a flexible and useful framework for analysis of cohort data of the form seen in both the southwestern and northeastern Taiwanese studies. Although dose can be entered into the model in a variety of ways, the additive model with a linear dose effect is consistent with other analyses that have been applied to cancer cohort data. · The BEIR IV formula provides a useful approach to computing an EDGE for the United States based on relative risks obtained from a different population. The BEIR IV formula allows the incorporation of relative risk estimates obtained from case-control studies. The simple formula presented by Smith et al. (1992) is a close approximation to the BEIR IV approach. · There is insufficient knowledge on the mode of action of arsenic to justify the choice of any specific dose-response mode] the subcommittee explored a variety of models and ultimately used the additive mode] with
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208 ARSENICIN DRINKING WATER: 2001 UPDATE linear dose effect. Although different results might be obtained from other reasonable model choices, the estimates do not differ by more than an order of magnitude. · Accounting for individual exposure rate variability causes the uncer- tainty in the estimate to increase. Therefore, as shown in Table 5-S, the cen- tral tendency estimates of the EDIT increase when individual variability in drinking-water rates is considered. The corresponding lower bound estimates for each EDIT decreases, however, because the variance about the mean be- comes larger. · Assumptions regarding the differences between mean drinking-water intakes in Taiwan and the United States can have substantial impacts on the estimated EDo,. There is evidence that drinking-water intakes in Taiwan might be closer to U.S. intakes than previously suggested, and that is an im- portant source of uncertainty. increasing the assumed drinking-water intakes for Taiwan provides a suitable approach to adjusting for arsenic exposure via cooking water, although the addition of ~ it, especially for women, seems excessive and warrants further investigation. · Adjusting for background arsenic exposure through food is likely to have only a modest effect on the estimated EDIT. · Measurement error in assigning the village-specific arsenic exposure concentrations is also likely to have only a modest impact on the estimated EDIT, compared with the variability associated with model uncertainty. Fur- ther exploration of this issue, however, would be useful. · Although there are insufficient data available for a formal combined analysis, it is helpful to compare the results across studies and also across various feasible assumptions to obtain a sense of the magnitude of likely effects. Table 5-3 presents such an analysis. · The northeastern Taiwanese study has several strengths, including exposure assessment and data on potential confounders, which could inform the dose-response assessment. At present, however, the follow-up time is insufficient to provide the precision necessary for quantitative dose-response assessment. ~ Analysis of the data from the period of peak arsenic exposure in northern Chile and the data from southwestern Taiwan results in similar esti- mates of lifetime lung cancer incidence in the United States. The consistency of the results adds to the confidence in the validity of the risk estimates.
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QUANTITATIVE ASSESSMENT OF RISKS USING MODELING APPROACHES 209 RECOMMENDATIONS · Quantitative risk assessment for arsenic in drinking water should consider the results from the following two studies: The data set from the southwestern Taiwanese study reported by Chen et al. (1988, 1992;) and Wu et al. (1989), which was considered by the NRC in the 1999 report and by EPA (lung and bladder cancer in males and fe- males). The Chilean case-control study (lung cancer) reported by Ferreccio et al. (2000~. · Dose-response analysis of the southwestern Taiwanese data should incorporate an unexposed comparison group; the southwestern Taiwanese region is the recommended comparison group. · The current mode-of-action data are insufficient to guide the selection of a specific dose-response model. The additive Poisson model with a linear term in dose is a biologically plausible model that provides a satisfactory fit to the epidemiological data and represents a reasonable model choice for use in the arsenic risk assessment. Research should be conducted on techniques for integrating the results from many epidemiological studies into a risk assessment. · Information on remaining uncertainties should be incorporated into future analyses when it is acquired. REFERENCES Albores, A., M.E. Cebrian, I. Tellez, and B. Valdez. 1979. Comparative study of chronic hydroarsenicism in two rural communities in the Laguna region of Mex- ico. tin Spanish]. Boll Of~cina Sanit. Panam. 86~3~:196-205. Breslow, N.E., and N.E. Day. 1988. Statistical Methods in Cancer Research: Vol. 2. The Design and Analysis of Cohort Studies. New York: Oxford University Press. Carlin, B.P., and T.A. Louis. 1996. Bayes and Empirical Bayes Methods for Data Analysis. New York: Chapman & Hall. Carroll, R.J., D. Ruppert, and L.A. Stefanski. 1995. Measurement Error in Nonlinear Models. New York: Chapman & Hall. CDC (Centers for Disease Control and Prevention). 2001. GMWK I Total Deaths for Each Cause by 5-Year Age Groups, United States, 1993, 1994, 1995, 1996, 1997,
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Representative terms from entire chapter: