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OCR for page 214
6
Hazard Assessment
After completing a hazard identification and dose-response assessment, it is
important to place the conclusions of the assessment in the context of similar
analyses and in the context of a real-worId situation to ensure that the risk
estimates are reasonable in view of available information. Therefore, this
chapter will summarize the findings ofthe subcommittee, compare the results
of the subcommittee's dose-response assessment with those of the previous
NRC ( 1999) subcommittee and EPA (2001), and, finally, examine whether the
estimated risks are plausible when considered in the context of the U.S.
population.
FINDINGS OF THE SUBCOMMITTEE
There is increasing evidence that chronic exposure to arsenic in drinking
water may be associated with an increased risk of hypertension and diabetes.
The existing studies from Taiwan and Bangladesh, discussed in Chapter 2,
have observed substantial increases in the risk of these medical conditions at
levels of arsenic exposure that are within one to two orders of magnitude of
the Tower levels of current regulatory concern in the United States. Pending
214
OCR for page 215
HAZ4RD ASSESSMENT 215
further research that characterizes the dose-response relationship for these end
points, the magnitude of possible risk that exists at low levels is nonquantifi-
able. Nevertheless, because these endpoints are common causes of morbidity
and mortality, even small increases in relative risk at Tow dose could be of
considerable public-health importance. This potential impact should tee quaTi-
tatively considered in the risk-assessment process.
A sound and sufficient database exists on the carcinogenic effects of
arsenic in humans. The main end points for a quantitative risk assessment
following exposure to arsenic in drinking water are lung and bladder cancers.
The human data from southwestern Taiwan used by EPA in its risk assessment
remain the most appropriate for determining quantitative risk estimates. Hu-
man data from more recent studies provide additional information for use in
the risk assessment. Based on some ofthese studies, the subcommittee recom-
mends using an external comparison population when analyzing the earlier
studies from southwestern Taiwan, rather than comparing high- and low-expo-
sure groups within the exposed population, because of concerns regarding
probable exposure misclassification in the low-exposure villages within the
data set and because of new data from southwestern Taiwan that suggest that
confounding is unlikely. The data on the mode of action of arsenic do not
indicate what form of-extrapolation should be used below the exposure range
of human data. The observed data should be modeled using a biologically
plausible model form that best fits the data to determine a ~ % effective dose
(EDo~. The subcommittee used an additive Poisson model with a linear term
in dose for the southwestern Taiwan cancer data. The dose-response relation-
ship should be extrapolated linearly from the EDGE to zero. Because the hu-
man data include exposures to arsenic concentrations relatively close to some
U.S. exposures, the distance of extrapolation is very small less than 1 order
of magnitude.
The subcommittee calculated EDo,s based on the southwestern Taiwanese
data (Chen et al. 1985, 1992; Wu et al. 1989), the Chilean data (Ferreccio et
al. 2000), and the northeastern Taiwanese data (Chiou et al. 2001). It caTcu-
lated cancer risk estimates for the southwestern Taiwanese data (Chen et al.
1985, 1992; Wu et al. 1989), and the Chilean data (Ferreccio et al. 2000)
discussed below. Cancer risks were not estimated for northeastern Taiwan
(Chiou et al. 2001) because of instability of the model calculated with the
small number of cases in that study.
.
.
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21 6 ARSENICINDR]NKING WATER: 2001 UPDA TE
COMPARISONS OF RESULTS OF DOSE-RESPONSE
ASSESSMENTS
Estimates of Effective Dose for a I% Response: EDGE
Doses of an agent associated with the onset of a defined rate of observable
response in a study population (often termed EDGE when referring to a re-
sponse rate of 1°/O) can be useful in several key respects in risk assessment.
The EDGE can be used as a point of departure for extrapolation to lower doses
(typically to the origin) when insufficient data exist to characterize the shape
of the dose-response curve in a region. They can also be used to assess a
margin of exposure (MOE) between a dose with observed adverse effects and
the level of exposure that exists in the general population. A MOE is calcu-
lated by dividing the dose associated with a defined level of response, such as
a 1% (EDo~) or 10°/O (EDIT) response in animal or epidemiological studies, by
the actual or projected human exposures (EPA ~ 996~eciding which level
of response to use (e.g., 1% or 10°/O) is a policy choice that depends, in part,
on the size and quality of the epidemiological or animal data sets available.
Therefore, the smaller the MOE is for a given population, the closer the popu-
lation exposures are to exposures shown to have an adverse effect. The MOE
can provide risk managers with information about the extent of apparent pro-
tection for the population. The MOE approach is complementary to more
traditional approaches for determining a safe level of exposure, each approach
providing different information to the riskmanagers (Presidential/ Congressio-
nal Commission on Risk Assessment and Risk Management 1997a,b). Be-
cause the human epidemiological data set for arsenic encompasses exposure
levels close to those for which the subcommittee calculated Edgers, the sub-
committee elected to present its EDo~s rather than Eddies used in EPA's
margin-of-exposure analyses.
Table 5-3 presents the EDo~s estimated by the subcommittee (based on
mortality or incidence data, depending on the study) for the Chilean data
(Ferreccio et al. 2000), the northeastern Taiwanese data (Chiou et al. 2001),
and the southwestern Taiwanese data (Chen et al. ~ 985, ~ 992; Wu et al. ~ 989~.
Those values were estimated using a number of statistical models to fit the
data, including additive and multiplicative models using linear or logarithmic
terms in dose. The EDo~s were estimated using the published or calculated
relative risk values and a modification of the BEIR IV (NRC 1988) formula,
as described in Chapter 5. Despite the variability, it is evident that most ofthe
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HAZARD ASSESSMENT 21 7
EDo~s are less than a factor of 10 higher than the current U.S. maximum con-
taminant level (MCL`) of 50 ,ug/~.
The subcommittee determined EDo~s (i.e., the dose at which there is a I%
response in the study population) for various studies using a number of statis-
tical models. For example, the estimated EDo~s from the Chilean study on
lung cancer ranged from 5 to 27 ~g/L, depending on the exposure data used.
The previous Subcommittee on Arsenic in Drinking Water estimated
EDo~s of 404 to 450 vigil, depending on the model used, for arsenic and male
bladder cancer mortality. Those values are approximately within the range of
EDo~s estimated by this subcommittee. However, because the EDGY values
reported by the current and prior subcommittees were derived using different
biostatistical approaches, they are not directly comparable. The EDGY values
in NRC (1999) reflect a I% increase relative to background cancer mortality
in Taiwan, whereas the current subcommittee's approach, using a modifica-
tion of the BE1:R IV analysis (NRC 1988), reports EDo~s based on a 1°/O in-
crease relative to the background cancer mortality in the United States. This
is an important difference because the background rates for lung and bladder
cancer are substantially different between Taiwan and the United States.
Background rates for lung cancer in the United States are approximately 3-
and 2.3-fold higher than in Taiwan for females and males, respectively; and
bladder cancer risks are approximately 1.4- and 3-fold higher in females and
males, respectively, in the United States when compared to Taiwan.
Cancer Risk Estimates
The subcommittee presents the theoretical lifetime excess cancer risks for
lung and bladder cancer incidence for the U.S. population (females and males
calculated separately) at fixed arsenic concentrations in drinking- water of 3,
5, 10, and 20 go/. Table 6-1 presents the maximum-likelihood estimates
(MLEs) of the risk of bladder and lung cancer combined based on the data
from southwestern Taiwan. Estimates calculated using the U.S. background
cancer incidence data or Taiwanese background cancer incidence data are
presented. The U.S. background cancer incidence data is taken from SEER
(2001~. The Taiwanese background cancer incidence data were estimated by
multiplying the subcommittee's corresponding U.S. lifetime incidence rate
(Tables 5-7 and 5-~) by the ratio of the Taiwanese annualized rate (You et al.
2001) to the U.S. annualized rate (Ferlay et al. 2001~. The relatively small
confidence limits around the MLE (+/- less than 12% ofthe MLE) reflect the
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21~ ARSENIC IN DRINKING WA TER: 2001 UPDA TE
TABLE 6-1 Theoretical Maxi Likelihood Estimates of Excess Lifetime Risk
(Incidence per 10,000 people) of Lung Cancer and Bladder Cancer for U.S. Popula-
tions Exposed at Vanous Concentrations of Arsenic in Drinking Watera
Bladder Cancer
Arsenic
Concentration
(fig/ L)
Lung Cancer
U.S. Back-
ground Rateb
T.
alwanese
Background
Ratec
Taiwanese
U.S. Background
Background Rateb Rater
-
Fernales Males Females Males Females Males Females Males
3 3.6 6.8 2.3 2.0 5.4 4.0 1.8 1.7
5 6.0 1 1 3.8 3.2 8.9 6.8 3.0 3.0
10 12 23 7.5 6.8 18 14 6.2 6.1
20 24 45 15 13 36 27 12 1 1
a Estimates were calculated using data from individuals in the arsenic-endemic region of south-
western Taiwan and data from an external comparison group from the overall (mostly unex-
posed) southwestern Taiwan area. The risks are estimated using what the subcommittee consid-
ered reasonable assumptions. (A U.S. resident weighs 70 kg compared with 50 kg for the
typical Taiwanese, and the typical Taiwanese drinks just over 2 L of water per day compared
with 1 L per day in the United States. Therefore, it assumes that the Taiwanese exposure per
kilogram of body weight is 3 times that of the U.S. population.) It is possible to get higher and
lower estimates using other assumptions. Risk estimates are rounded to two significant figures.
All 95% confidence limits are less than +/- 12% ofthe maximum-likelihood estimate and are
not presented. It should be noted that those confidence limits are a function of sample size and
are not indicative of the true uncertainty associated with the risk estimates. The individual risk
estimates for bladder and lung cancer were added together to estimate combined risks.
b Risks are estimated using the U.S. background cancer rate (SEER 2001~.
c Risks are estimated using the Taiwanese background cancer rate. The Taiwanese background
cancer incidence data were estimated by multiplying the subcommittee's corresponding U.S.
lifetime incidence rate (Tables 5-7 and 5-8) by the ratio of the Taiwanese annualized rate (You
et al. 2001) to the U.S. annualized rate (Ferlay et al. 2001~.
relatively large sample size, end they are not indicative ofthe true uncertainty
associated with the risk assessment discussed in Chapters 4 and 5. The MLEs
of the lifetime excess risks for combined lung and bladder cancer incidence
for females range from 9 per 10,000 from exposure to drinking water with
arsenic at 3 ,ug/L to 60 per 10,000 from exposure to drinking water with ar-
senic at 20 ,ug/~. The corresponding risk estimates for males are ~ ~ to 72 per
10,000. Those values are estimates of the combined lifetime excess risk of
lung and bladder cancer (incidence) in a given population following lifetime
exposure to arsenic in drinking water at the given concentration.
OCR for page 219
HAZARD ASSESSMENT 2~9
As presented in Chapter 5, the subcommittee used data from a study per-
formed in northern Chile (Ferreccio et al. 2000) to estimate the theoretical
lifetime excess risk of incident lung cancer in U.S. males and females at ar-
senic concentrations in drinldug water of 3, 5, 10, and 20 ,ug/~. Using the
peak period of arsenic exposure in Chile from 1958 to 1970 as a dose metric,
the resulting estimates for excess lung cancer incidence in the United States
were 3 to 4 times higher than the risks derived from the Taiwanese data. in
contrast, when the dose metric used in the Chilean data was the average ar-
senic concentration in drinking water from 1930 to 1994, the corresponding
risk estimates were an order of magnitude higher.
The previous Subcommittee on Arsenic in Drinking Water presented
lifetime excess cancer risk estimates for bladder cancer mortality in males
based on its analyses of the southwestern Taiwanese data (Chen et al. 1985,
1992; Wu et al. 19894. Some ofthose risk estimates are presented in Table 5-
l. Those risks were estimated using an external comparison population and
a multiplicative linear model. At an arsenic concentration of 50 ~g/L of drink-
ing water, the excess risk of bladder cancer mortality for males was estimated
to be 10 tol5 per 10,000 (NRC 1999~. Assuming linearity and dividing by 5,
that corresponds to a mortality risk estimate of 2 to 3 per 10,000 at 10 ,ug/~.
If the U.S. mortality rate for bladder cancer is 20% (SEER 2001), that corre-
sponds to an estimated risk of bladder cancer incidence of 10-15 per 10,000.
Using the southwestern Taiwanese data, this subcommittee's estimate for
lifetime excess bladder cancer incidence in mates in the United States at an
arsenic concentration of 10 ,ug/L is 23 per 10,000 (see Table 6-~. Therefore,
although some analytical approaches were different, the estimates for bladder
cancer riskin males for arsenic at 10 ,ug/L, of drinking water determined by the
subcommittee in this report are generally consistent with those presented in
the previous NRC report.
As discussed in Chapter 5, EPA did not present theoretical lifetime excess
cancer risk estimates for arsenic in drinking water in its notices in the Federal
Register (2000, 2001~. The risk estimates it presents (EPA 2001) are adjusted
for the occurrence of arsenic in U.S. drinking water; consideration of such an
adjustment is beyond the charge to this subcommittee. It is not possible to
directly compare the theoretical lifetime cancer risks estimated by this sub-
committee with those presented by EPA. The different assumptions used by
EPA (2001) and this subcommittee are presented in Table 6-2.
The subcommittee did, however, use a linear extrapolation from the EDo~s
estimated in the analysis on which EPA based its risk estimates (Morales et
al. 2000) to estimate the theoretical lifetime excess bladder and lung cancer
risks at 3, 5, 10, and 20 ,ug/L, presented in Table 5-2. Thus, the subcommittee
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220 ARSENIC IN DRINKING WA TER: 2001 UPDA TE
TABLE 6-2 Summary of Assumptions Used by EPA and the Subcommittee for
Dose-Response Analyses and Their Impact on the EPA's Risk Estimates Relative to
the Subcomm~ttee's Risk Estimatesa
Study
Parameter EPA (2001) Subcommittee Impact
Choice of End Lung and bladder cancer Lung and bladder cancer No
Point difference
Choice of Study Southwestern Taiwanesecan- Southwestern Taiwanese No
cer mortality data from Chen et cancer mortality data from difference
al. (1985, 1988, 1992) Chen et al. (1985, 1988,
1 992)
Model Choice Multiplicative Poisson regres-
sion model with linear extrapo-
lation
Additive Poisson regression Decrease
with linear extrapolation;
BEIR IV (NRC 1988)
Selection of No external comparison group External comparison group Decrease
Comparison used used
Group
Adjustments for U.S. population: Monte Carlo
Water Intake analysis of CSFII (EPA 2000)
water intakes
Taiwan population: water con-
sumption is 3.5 L for males
and 2.0 L for females
U.S. population: Mean
daily average from CSFII of
1 today for males and fe-
males
Taiwan population: expo-
sures equal to 3 times U.S.
default value, i.e., 3 IJday
for males and females
Decrease
Adjustmentsfor Taiwan: Adjustedlower Taiwan: added a constant Decrease
Dietary Intake bound estimates to account for concentration of arsenic (30
of Arsenic arsenic from cooking water by payday) to exposure rates
adding 1 L of water; therefore, for all individuals in study
total water intake for males villages
was 4.5 LJday and for females
was 3.0 IJday. Also to
account for intake from food
directly, multiplied lower
bound estimate by fraction of
arsenic consumed per kilogram
contributed by drinking water
Adjustments for Used Taiwanese mortality data Used U.S. background inci- Decrease
Mortality ver- for bladder and lung cancers; dence data for bladder and
sus Incidence adjusted upper bound by 1.25 lung cancers from SEER
for bladder cancer to reflect (2001 ) database
mortality, assumed all lung
cancer is fatal in Taiwan
a More detailed information about these assumptions can be found in Chapter 5.
Abbreviations: CSFII, Continuing Survey of Food Intakes by Individuals; SEER, Surveillance,
Epidemiology, and End Results.
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HAZARD ASSESSMENT 221
compared its risk estimates with those estimates calculated from the published
analyses (Morales et al. 2000) on which EPA based its risk estimates (Table
5-2~. The subcommittee notes, however, that the estimates in Table 5-2 are
not adjusted for water consumption or arsenic in food in the same manner as
used by EPA, nor by this subcommittee in its analysis in Chapter 5. (The
adjustments used by EPA for food and water consumption would decrease the
risk estimates.) However, even without those adjustments, the risk estimates
on which EPA based its analyses are Tower than this subcommittee's esti-
mates, regardless of whether the U.S. or Taiwanese background cancer rates
are used to estimate the risks. Several factors contribute to that difference.
Unlike the subcommittee ' s estimates, EPA' s analyses were based on estimates
that were calculated without using an external comparison population. The
subcommittee also used a different statistical method than EPA to estimate
lifetime cancer risks. The subcommittee has presented lifetime excess cancer
risk estimates calculated using either the U.S. or the Taiwanese background
rates; Morales et al. (2000) estimated the EDo~s using Taiwanese background
rates. The magnitude of the difference between the estimates can be seen in
Table 6-1 . In addition, the method the subcommittee used to adjust for arsenic
in food and its assumptions regarding water intake in the U.S. and Taiwanese
populations were different from those used by EPA in its analyses.
It should be noted that the subcommittee was split on whether using the
U. S. background rates was preferable to using the Taiwanese background rates
for estimating arsenic risks in the United States. Some members of the sub-
committee felt strongly that using U.S. background rates was the preferred
approach, while others felt that there was not sufficient justification to select
one set of background rates over the other, and that both should be presented.
Thus, the results from both approaches are presented in Table 6-l . The sub-
committee agreed, however, that if there was a multiplicative interaction
between a complex array of risk factors, including smoking, that establish the
background rates, then using the U.S. background cancer incidence rates
would be preferred over the Taiwanese background rates for estimating ar-
senic cancer risks in the U.S. population.
PLAUSIBILITY OF CANCER RISK ESTIMATES
Upon completion of an assessment of the potential health effects of an
environmental contaminant, it is wise to compare the results ofthe assessment
with a real-world situation that is, the adverse health effects observed among
the people most exposed to the contaminant. The key factors triggering public-
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222 ARSENIC IN DRINKING WA TER. 2001 UPDA TE
health concern regarding arsenic in drinking water have been the high inci-
dences of different types of cancer in populations exposed to increased con-
centrations of arsenic in drinking water (greater than 100 ,ug/~) in Taiwan,
Chile, and Argentina. The cancer with the highest increases in relative risk in
these countries is cancer of the bladder.
It has been suggested that, if the risks of bladder cancer from arsenic in
drinking water were indeed as high as estimated in this report (see Table 6-~),
high cancer rates would have been anticipated in areas of the United States
with increased concentrations of arsenic in groundwater, and these high rates
would have readily attracted public-health attention. Some simple calcula-
tions demonstrate how risk estimates for low-level arsenic exposure in this
report might be difficult to detect by observing geographical differences in
cancer incidence or mortality. To illustrate that point, the subcommittee used
its risk estimate of 45 per 10,000 for bladder cancer incidence in U.S. males
(based on the Taiwanese data, U.S. cancer incidence data, and a ratio of 3 for
water ingestion on a per-body-weight basis for the Taiwanese population
compared with the U.S. population) exposed to arsenic at a concentration of
20 ,ug/L (Table 6-~. The lifetime risk of being diagnosed with bladder cancer
in U.S. mates is 3.42% for the period of 1 996-1998 (or 342 per 1 0,000) (SEER
2001~. An increased risk of 45 per 10,000 over a background risk of 342 cases
in 10,000 mates would be difficult to detect. In terms of bladder cancer mor-
taTity, if it is assumed that only about one in five bladder cancer cases in the
United States results in death (the ratio of mortality to incidence is approxi-
mately 20% for U.S. males, SEER 2001), a lifetime excess risk for mortality
from bladder cancer in U.S. mates is about 9 in 10,000 following lifetime
exposure to arsenic in drinking water at 20 ,ug/~. The subcommittee further
explored how that risk contributes to overall U.S. mortality for bladder cancer.
Lifetime mortality for bladder cancer in the United States for mates is 0.72%
(72 per 10,000) for the period of 1996-1998 (SEER 2001~. That increase in
mortality risk of 9 per 10,000 would be difficult to detect against that back-
ground rate of 72 per 10,000. Indeed, it would represent only about 13% of
the total risk ofbladder cancermortality. Furthermore, the denominator ofthe
risk estimate for arsenic assumes that all 10,000 individuals are at risk (e.g.,
all consume arsenic at 20 ,ug/L of their drinking water for a lifetime). Detec-
tion is further complicated by the variability in the actual exposure to arsenic
in drinking water (not considered by this subcommittee), the unknown distri-
bution of other risk factors (especially smoking), and the mobility of the U.S.
population. However, if the risks for arsenic-related bladder cancer were
OCR for page 223
HAZARD ASSESSMENT 223
higher than the estimate used in this example, then bladder cancer incidence
and mortality at exposures of 20 ~g/L would be proportionately higher and
thus might be easier to detect in a population. Because background lung
cancer mortality is almost 10-fold greater than bladder cancer, it would be
even more difficult to demonstrate an association between Tow concentrations
of arsenic in drinking water and lung cancer risk. Therefore, although the
subcommittee's risk estimates are of public-health concern, they are not high
enough to be easily detected in U.S. populations by comparing geographical
differences in the rates of specific cancers with geographical differences in the
concentrations of arsenic in drinking water.
In accordance with its charge, the subcommittee has not conducted an
exposure assessment and subsequent risk characterization and risk assessment.
The theoretical lifetime excess cancer risks estimated by the subcommittee
and presented in this report, however, should be interpreted in a public-health
context using an appropriate risk-management framework, such as that pro-
posed by the Presidential/Congressional Commission on Risk Assessment and
Risk Management (1997a,b).
SUMMARY AND CONCLUSIONS
· The subcommittee's evaluation and analyses of the data from south-
western Taiwan indicate that the lifetime excess cancer risks in the United
States for bladder and lung cancers combined at arsenic concentrations in
Winking wafer between 3 and 20 ~g/L (ppb) are estimated to be between 9 and
72 per 10,000 people based on U.S. background cancer incidence data. (The
corresponding range based on Taiwanese background cancer incidence data
is 4 to 24 per 10,000.) These estimates can be interpreted in light of EPA's
stated goals for public-health protection (EPA ~ 992~.
· Depending on the dose metric used in the study, excess risk estimates
for cancer in the United States derived from a recent investigation in Chile are
either similar to or higher than risk estimates derived from the Taiwanese data.
· Although these risk estimates are high, they would not be detected in
U.S. populations by comparing geographical differences in the rates of spe-
cif~c cancers with geographical differences in the concentration of arsenic in
drinking water.
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224 ARSENIC IN DRINKING WA TER: 2001 UPDA TE
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HAZARD ASSESSMENT 225
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OCR for page 226
Representative terms from entire chapter:
risk estimates
