r/EPA
United States
Environmental Protection
Agency
Office of Water
4301
EPA-820-B95-007
March 1995
Great Lakes Water
Quality Initiative
Technical Support
Document for
Human Health
Criteria and Values
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DISCLAIMER
This document has been reviewed by the Health and Ecological
Criteria Division, Office of Science and Technology, U.S.
Environmental Protection Agency, and approved for publication
as a support document for the Great Lakes Water Quality
Initiative. Mention of trade names and commercial products
does not constitute endorsement of their use.
AVAILABILITY NOTICE
This document is available for a fee upon written request or
telephone call to:
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(800) 553-6847
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1 NTIS Document Number: PB95187316
N
or
Education Resources Information Center/Clearinghouse for
Science, Mathematics, and Environmental Education (ERIC/CSMEE)
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(800) 276-0462
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ERIC Number: D051
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77 West Jackson Boulevard, 12tn Hoor
Chicago, IL 60604-3590
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GREAT LAKES WATER QUALITY INITIATIVE TECHNICAL
SUPPORT DOCUMENT FOR HUMAN HEALTH
I. INTRODUCTION 1
A. Goal 1
B. Level of Protection 1
C. Two Tiered Approach 4
D. Technical Background 4
II. MINIMUM DATA REQUIREMENTS 6
A. Carcinogens 6
1. Weight of Evidence 6
2. Appropriate Study Design and Data Development . 8
3. Borderline Conditions 12
B. Noncarcinogens 14
1. Appropriate Study Design 15
a. Acute Toxicity 15
b. 14 Day or 28 Day Repeated Dose Toxicity . 16
c. Subchronic and Chronic Toxicity 18
d. Reproductive and Developmental Toxicity . 19
C. Tier Designation 21
1. Carcinogens 21
2. Noncarcinogens 22
III. PRINCIPLES FOR CRITERIA DEVELOPMENT 24
A. General 24
B. Carcinogens 25
1. Mechanism/Mode of Action 25
2. Data Review 27
3. Model 28
a. Nonthreshold Approach 28
b. Threshold Approach 29
4. Lifespan Adjustment 30
5. Species Scaling 30
C. Noncarcinogens 31
1. Mechanism 31
2. Data Review 33
3. Uncertainty Factors 34
a. Intraspecies uncertainty factor 35
b. Interspecies uncertainty factor 35
c. Subchronic to chronic uncertainty factor . 35
d4. Less than subchronic duration uncertainty
factor 35
e. LOAEL to NOAEL uncertainty factor .... 35
f. Limited database uncertainty factor ... 36
D. Exposure Assumptions 36
1. Body Weight 36
2. Duration of Exposure 39
a. Population Mobility 39
b. Life Expectancy 39
3. Incidental Exposure . . • 40
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4. Drinking Water 41
5. Fish Consumption 42
6. Relative Source Contribution 44
IV. CRITERIA CALCULATIONS 46
A. Standard Exposure Assumptions 46
B. Carcinogens 46
C. Noncarcinogens 47
APPENDIX A Al
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I. INTRODUCTION
A. Goal
The goal of the human health criteria and values for the
Great Lakes and their tributaries is the protection - of humans
from unacceptable exposure to toxicants from consumption of
contaminated fish, drinking water and water related to
recreational activities. Emphasis is on protection of the
individual in evaluating toxicity information and its application
in the derivation of criteria and values. Exposure assumptions
follow trends and activities for the general population as a
region and also attempt to protect sensitive subpopulations.
Based on differences in behavior, there may be some individuals
who receive a greater level of protection or a lesser level of
protection via these procedures.
B. Level of Protection
Numeric criteria or interpretations of narrative criteria
developed for human health generally restrict chemical carcinogen
exposure in a target population to levels estimated to result in
a lifetime incremental risk of no greater than 1 in 100,000 of
developing cancer. The procedure generally used to estimate the
risk level leads to the development of a plausible upper limit of
the risk. This means that the actual risk is unknown, is
unlikely to exceed 10"5, and may even be as low as zero.
The selection of an "acceptable" target risk level does not
turn on scientific analysis, but on more subjective "risk
management" considerations. Differences in perception of risk,
opinions as to benefit versus risk reduction costs, as well as
distinctions between risks that are considered voluntary or
involuntary, all could play a meaningful role in determining risk
acceptability. For this initiative, a 10-5 cancer risk level has
been selected as acceptable. This is consistent with the
existing practice of the eight Great Lakes states, and therefore
is consistent with existing risk management policy in these
states. It is a.lso a risk level that EPA has found acceptable in
its review of state criteria in the past, and which EPA itself
has used as a basis for certain of its regulations. EPA notes
that States and Tribes are free to adopt a more stringent
approach than that contained in the final Guidance. It is also
instructive to note that this level of risk of developing cancer
appears to be roughly comparable to that which exists for death
due to natural phenomena. Table 1 represents data from a study
of everyday risks of death from several naturally occurring
incidents such as tornados, floods, lightning and animal bites or
stings. When extrapolated to lifetime risks, we see that these
risks range from 1.4 in 100,000 for animal bites or stings to 4
in 100,000 for floods and tornadoes. It is acknowledged that
risk of death (as described in Table 1) is not equatable with
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risk of cancer since many forms of cancer are now easily curable.
The comparison is made only for the purpose of illustrating the
potential background risk in the region.
For noncarcinogens, protection of human health is generally
centered on determining a level of daily exposure that is likely
to be without an appreciable risk of deleterious effects for a
lifetime. The concept of acceptable exposure incorporates the
potential for long term exposure of sensitive individuals in a
population to an environmental contaminant without any
anticipated adverse health effects.
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TABLE 1. EVERYDAY RISKS
4
• *•»*»•*•***••»»•*»•***»••
Motor vehicle accident
Falls
Drowning
Fires
Firearms
Electrocution
Tornadoes
Floods
Lightning
Animal bite or sting
^•^HHHHHHHHHfrtHHHHHHHHHHHHI
GENERAL
Manufacturing
Trade
Service and government
Transport and public utilities
Agriculture
Construction
Mining and quarrying
SPECIFIC
Coal mining (accidents)
Police duty
Railroad employment
Fire fighting
Time to Accumulate a 1 in
100 ,000 Risk of Death
*****
1 5 days
60 days
1 00 days
130 days
360 days
20 months
200 months
200 months
20 years
40 years
45 days
70 days
35 days
10 days
1 50 hours
140 hours
90 hours
.140 hours
15 days
15 days
110 hours
Average Annual
Risk per Capita
in the United State;
2x 10-4
6x 10'B
4 x 10-s
3 x 10'5
1 x 106
5 x 10-'
6x 10'7
6x10'7
5 x 10-7
2 x 10'7
8x10'5
5 x 10'6
1 x TO"4
4x ID"4
6x 10"4
6x 10"4
1 x 10'3
6x 10"4
2x 10"4
2x10-*
8x 10"4
Extrapolated to
Risk/Lifetime*
;
1.4x 10'2
4.2 x 10'3
2.8 x 10-3
2 x 10'3
7 x 10-4
3.5 x 10"4
4 x 10'6
4x 10'5
3.5 x 10'6
1.4x10-6
5.6 x 10'3
3.5 x 10'3
7 x 10'3
3 x 10'2
4x 10'2
4x 102
7 x 102
3 x 10'2
1.4x 10'2
1.4x 10'2
5.4 x 10 2
Some One in a Million Cancer Risks
Cosmic rays
Other radiation
Eating and drinking
Smoking
one transcontinental round trip by air
living 1.5 months in Colorado compared to New York
camping at 15,000 feet for 6 days compared to sea level
20 days of sea level natural background radiation
2.5 months in masonry rather than wood building
1H of a chest x-ray using modern equipment
40 diet sodas (saccharin)
6 pounds of peanut butter (aflataxia)
180 pints of milk (aflataxia)
200 gallons of drinking water from Miami or New Orleans
90 pounds of broiled steak (cancer risk only)
2 cigarettes
Adapted from Crouch and Wilson (1962)
Risk/Lifetime - 1 -(1-s)70
3
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C. Two Tiered Approach
A two tiered approach is used to derive ambient
concentration levels protective of human health. Both tiers rely
generally on the same standard procedure for data review and
criteria derivation. The difference between the two focuses
heavily on the certainty with which one can predict a level of
risk or a level of safety for humans from the data available.
The more adequate the database to estimate actual human risk or
to establish no adverse effect levels, the greater the certainty
in the appropriateness of the criterion or value. This level of
certainty depends heavily on the weight of experimental evidence
which includes factors such as: the quantity of studies or size
of the experimental database available for review; the quality of
study design, its conduct and range of effects evaluated; the
potency or range and type of adverse effects observed and, the
appropriateness of this data in predicting human effects, i.e.
evaluation of effects in humans or in animal species biologically
similar to human.
The greater the level of certainty in the database for
noncancer effects, generally the lower the need for adjustment of
the research findings to assure a level without appreciable risk.
The greater the weight of evidence for carcinogenicity, the
greater the strength in predicting cancer risk to humans.
Chemicals with databases providing a high level of certainty in
predicting a level of risk or safety for humans from adverse
health effects are suitable for Tier I numeric criteria
derivation. Tier I criteria are conceptionally those criteria
where the probability of change is low.
Chemicals with less extensive data or where the weight of
evidence toward predicting human health effects is less certain,
are subject to Tier II values. Under Tier II, the probability of
future change is greater than for Tier I as demonstrated by the
extent, level of quality and/or weight of evidence or
conclusiveness of effects demonstrated by the database. The
values derived via Tier II are more likely to change based on new
data and/or reinterpretation of effect or potency.
D. Technical Background
The process used to evaluate effects and in development of
criteria shall be based on currently acceptable scientific
methods and consider guidance offered by the various USEPA
methods. Particular attention should be paid to RfD and cancer
risk estimation development contained within the Integrated Risk
Information System (IRIS).
To promote consistency with other USEPA guidance for
chemical management, it is important to review and strongly
consider the IRIS values for chemicals undergoing criteria or
values development whenever available. Although consistency is
important, it is also important that the most current and
complete data should be used when generating criteria or values
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whether IRIS has considered this data or not. Further, since
IRIS values are developed under guidance and through the
judgement of workgroups, the final values may not always be
arrived at consistently, i.e., duration of studies, selected,
uncertainty factors applied, basis for derivation of potency
slopes, etc., may differ between decisions. If data used in
deriving Tier I criteria or Tier II values have not been
considered by the IRIS RfD or CRAVE workgroups, the appropriate
workgroup should be advised of the data. In cases where IRIS
RfDs or potency slopes have not been developed consistent with
these procedures, it is suggested that the rationale for RfD or
potency slope development be evaluated and determination made
whether 1) justification is sufficient to support deviating from
these procedures, or 2) justification exists to deviate from IRIS
guidance. When deviations from IRIS are contemplated, EPA
strongly urges States and Tribes to communicate these potential
changes to EPA, either through a Regional EPA Office or directly
to the EPA Reference Dose (RfD) and/or Cancer Risk Assessment
Verification Endeavor (CRAVE) workgroups, as soon as possible.
This will help foster consistency between EPA and the States and
Tribes. Additionally, when deviating from IRIS, States and
Tribes are encouraged to work with the Clearinghouse described in
Section II of the SID, to ensure that other States and Tribes are
aware of the deviations.
Specific references which should be reviewed and evaluated
for greater details on the basic parameters of the criteria and
values derivation methodology are as follows:
National Cancer Institute (NCI). 1976. Guidelines for
Carcinogen Bioassay in Small Rodents, Technical Report
Series No. 1, U.S. Department of Health, Education and
Welfare, NCI-CG-TR-1.
Office of Science and Technology Policy (OSTP). 1985. Chemical
Carcinogens; A Review of the Science and Its Associated
Principles, Federal Register, Vol. 50, No. 50. March 14,
1985, 10371-10442.
Organization for' Economic Cooperation and Development (OECD).
1987. Guidelines for Testing of Chemicals, Paris, France.
U.S. Environmental Protection Agency (EPA). 1989. Risk
Assessment Guidance for Superfund, Volume 1, Human Health
Evaluation Manual (Part A) - Interim Final, Office of
Emergency and Remedial Response, Washington, D.C.,
EPA/540/1-89/002.
U.S. Environmental Protection Agency (EPA). 1980. Water Quality
Criteria Availability, Appendix C Guidelines and Methodology
Used in the Preparation of Health Effects Assessment
Chapters of the Consent Decree Water Quality Criteria
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Documents, Federal Register, Vol. 45, November 28, 1980,
79347-79357.
U.S. Environmental Protection Agency (EPA). 1985. Toxic
Substances Control Act Test Guidelines; Final Rules, Federal
Register, Vol. 50, NO. 188. September 27, 1985, 39421- .
39425.
U.S. Environmental Protection Agency (EPA). 1986. Guidelines
for Carcinogen Risk Assessment. Federal Register, Vol. 51,
No. 185. September 24, 1986, 33992-34002.
U.S. Environmental Protection Agency (EPA). 1986. Guidelines
for the Health Assessment of Suspect Developmental
Toxicants, Federal Register, No. 51, No. 185. September 24,
1986 34028-34040.
This is by no means a complete list. Other sources of
information and guidance may also be considered as appropriate.
II. MINIMUM DATA REQUIREMENTS
A. Carcinogens
1. Weight of Evidence
Evidence of a chemical's possible carcinogenic effects in
humans shall be categorized according to the existing EPA weight
of evidence classification system, which is adapted from the
International Agency for Research on Cancer (IARC). The five
categories or groups are as follows:
Human Carcinogen (identified as Group A under existing
classification scheme)
"sufficient" evidence from epidemiologic studies to support
a causal association between exposure to the chemical and
cancer;
Probable Human Carcinogen (identified as Group B)
"limited" evidence from epidemiologic studies with or
without supporting animal data (Group Bl); or, "sufficient"
evidence of carcinogenicity based on animal studies, but for
which there may be "inadequate evidence" or "no data" from
epidemiologic studies (Group B2);
Possible Human Carcinogen (identified as Group C)
"limited" evidence of carcinogenicity in animals and the
absence of data for humans;
Not Classifiable as to Human Carcinogenicity (identified as
Group D)
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"inadequate" evidence of carcinogenicity in humans and
animals, or, for which no data are available; and
Evidence of Noncarcinogenicity for Humans (identified as
Group E)
"no evidence" for carcinogenicity in at least two adequate
animal tests in different species or in both adequate
epidemiologic and
animal studies.
The definitions of the EPA weight of evidence classifications are
as follows:
1. Humans
a. Sufficient evidence - a causal association can be
inferred between exposure to the chemical and human
cancer.
b. Limited evidence - a causal interpretation is credible,
but that alternative explanations, such as chance, bias
or confounding could not adequately be excluded.
c. Inadequate evidence - there were few pertinent data,
or, a causal interpretation is not credible from
available studies since they did not exclude change,
bias or confounding.
d. No evidence - no association was found between exposure
and an increased risk of cancer in well-designed and
well-conducted independent analytical epidemiologic
studies.
2. Animals
a. Sufficient evidence - an increased incidence of
malignant or combined malignant and benign tumors: 1)
in multiple species or strains, 2) in multiple
experiments using different dosage levels and possible
different routes of exposure; or 3) in a single
experiment with a high incidence, unusual site or type
of tumor, or early onset.
b. Limited evidence - data suggest a carcinogenic effect
but are limited because: 1) the studies involve a
single species, strain or experiment which does not
demonstrate a high incidence, unusual site or type of
tumor, or early onset; 2) the experiments used
inadequate dosage levels, inadequate exposure duration,
inadequate follow-up periods, poor survival, too few
animals, or inadequate reporting; 3) an increase in
benign tumor incidence only and no response in a
variety of short-term tests for mutagenicity; or 4)
tumor responses of marginal statistical significance
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due to inadequate study design or reporting, or, in
tissue known to have a high or variable background
rate.
c. Inadequate evidence - because of major qualitative or
quantitative limitations, the studies cannot be
interpreted as showing either the presence or absence
of a carcinogenic effect.
d. No evidence - no increased tumor incidence in at least
two well-designed and well conducted animal studies in
different species.
Further detail regarding this classification system for
categorizing weight of evidence for carcinogenicity may be found
in the EPA Guidelines for Carcinogen Risk Assessment (EPA, 1986).
2. Appropriate Study Design and Data Development
The following discussion summarizes the process for
evaluating evidence of carcinogenicity and outlines an approach
study design by which one may measure the quality and adequacy of
data development. When available, human epidemiologic data with
quantifiable exposure levels are preferred for evaluating a
chemical's carcinogenic potential over use of animal data alone.
Epidemiological studies can provide direct evidence of a
chemical's carcinogenicity in humans (OSTP, 1985). The type of
epidemiologic study conducted indicates whether the study may be
useful in assessing carcinogenic risk to exposed humans
(analytical studies) or if it is merely hypothesis-generating and
inherently incapable of proving a causal association. Case
reports, descriptive studies and ecological (correlational)
studies generally cannot establish whether risks are associated
with particular exposures. Analytical studies can assess
carcinogenic risks to exposed humans, and can infer a casual
association (Mausner and Kramer, 1985; OSTP, 1985). The two
general types of analytical studies are case-control and cohort.
In case-control studies, a group of diseased "case" individuals
is initially identified and matched with nondiseased "controls".
Information on past exposure to reputed risk factors or causative
agents is then collected for both groups. If the proportion of
cases with a certain exposure is significantly different than
that of controls., an association between exposure and disease may
be indicated. A cohort study starts by identifying a group of
individuals with a particular exposure and a similar group of
unexposed persons and follows both groups over time to determine
subsequent health outcomes. The rates of disease in the exposed
and unexposed groups are then compared. Cohort studies may be
based on current exposure and future health outcomes (prospective
cohort study), or on past exposure information and disease
occurrence (historical cohort study). As with case-control
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studies, cohort studies that are well-designed, well-conducted,
and well-evaluated can test hypotheses and provide the basis for
causal inferences (OSTP, 1985; EPA, 1986). Factors such as proper
selection and characterization of exposed and control groups,
adequacy of duration and quality of follow-up, proper
identification and characterization of confounding factors and
bias,'appropriate consideration of latency effects, valid
ascertainment of. causes of morbidity and mortality, and the
ability to detect specific effects are all elements for
determining the adequacy of epidemiologic studies (EPA, 1986) .
In interpreting a reported causal association, reference may be
made to the following criteria, as described by IARC (1985), EPA
(1986), and the Tripartite Working Group (1985):
There is no identifiable positive bias which could
explain the association.
The possibility of positive confounding factors has
been considered and ruled out as explaining the
association.
The association is unlikely to be due to chance alone.
Although the weight of evidence increases with the number of
adequate studies, in some instances, a single epidemiologic study
may be strongly indicative of a cause-effect relationship (IARC,
1985; EPA, 1986). Confidence to infer a causal association is
increased by any of the following: when several independent
studies are concordant in showing the association; when the
association is strong; when there is a dose-response relationship
when a reduction in exposure is followed by a reduction in the
incidence of cancer; when the effect is biologically plausible;
or when the effect is specific for a particular chemical. When
epidemiological evidence based on analytical studies appears to
be significantly flawed, the evidence may then be downgraded to
being suggestive of an association based on scientific judgment.
This may still provide evidence that a causal interpretation is
credible, but that alternative explanations, such as chance, bias
or confounding factors, could not adequately be excluded.
Epidemiological studies are inherently capable of detecting
only comparatively large increases in the relative risk of cancer
(EPA, 1986). Other limitations of epidemiological studies
include the long latency of cancer, and the difficult task of
exposure assessment, including multiple exposures. Therefore,
negative results from such studies do not verify that a
particular agent is noncarcinogenic in humans (IARC, 1985; EPA,
1986; OSTP, 1985).
Although epidemiologic studies are preferable for assessing
carcinogenic potential for humans, the relative paucity of such
data necessitates the use of animal data as a surrogate for
humans in most situations. In the absence of adequate data on
humans, it is biologically plausible and prudent to regard agents
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for which there is sufficient evidence of carcinogenicity in
experimental animals as if they present a carcinogenic risk to
humans (IARC, 1991). The weight of evidence that an agent is
potentially carcinogenic in humans increases with: a) the
increase in tissue sites affected; b) the increase in number of
animal species, strains, sexes, doses and experiments showing a
carcinogenic response; c) the occurrence of clear-cut dose-
response relationships as well as a high level of statistical
significance of the increased tumor incidence in treated groups
as compared to controls; d) a dose related shortening of the
time-to-tumor occurrence or time to death with tumor; and e) a
dose-related increase in the proportion of tumors that are
malignant (EPA, 1986).
The guidelines detailed by EPA (1985), OSTP (1985) and NCI
(1976) for evaluating long-term carcinogenicity bioassays will be
utilized to determine the adequacy of design and the strength of
evidence provided by the study. Specific study design elements
of these guidelines are synopsized as follows:
Species used: The most widely used and accepted test
species is the rat. NCI/NTP bioassays routinely use the
Fischer inbred (F344) strain of rat and the B6C3F1 hybrid
mouse. Hamsters have also been frequently used. Other
animal species and strains may also be appropriate
surrogates to demonstrate a chemical's carcinogenic
potential.
Number of animals: At least 100 rodents (50 of each sex)
should be used at each dose level and concurrent control.
Age at start: Dosing of rodents should begin as soon as
possible after weaning to allow for the long latency of
cancer. For rats, dosing ideally begins before the age of 6
weeks and should not begin after 8 weeks of age.
Survival: All groups should have at least 50% survival at
the time of termination.
Concurrent control groups: These should be untreated, sham
treated, or, if a vehicle is used in administering the test
substance, vehicle control groups. The use of historical
control data is desirable for assessing the significance of
changes observed in exposed animals, but only if the strain
of animals and laboratory conditions have not changed. For
the evaluation of rare tumors, even small tumor responses
may be significant compared to historical data. The review
of tumor data at sites with high spontaneous background
requires special consideration (OSTP, 1985). For instance,
a response that is significant with respect to the
experimental control group may become questionable if the
historical control data indicate that the experimental
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control group had an unusually low background incidence
(NTP, 1984).
Dose levels': At least 3 dose levels are recommended in
addition to the concurrent control group, for the purpose of
risk assessment (OSTP, 1985; EPA, 1985). For the purpose of
hazard assessment, detection of a carcinogenic response is
possible with one dose level, although 2 dose levels are
preferred and are necessary to demonstrate a dose-response
relationship. The highest dose level should target the
maximum tolerated dose (MTD). The MTD is the dose which,
when given for the duration of the chronic study, elicits
signs of minimal toxicity (e.g., less than or equal to 10%
weight gain decrement) without substantially altering the
normal life span due to effects other than carcinogenicity.
The MTD is intended to provide an adequate statistical power
for the detection of carcinogenic activity. While not an
ideal solution to the problem of low bioassay sensitivity,
use of the MTD is appropriate if it is properly determined
(OSTP, 1985; EPA, 1986).
Dosing route: The test substance should be administered via
the oral,, dermal or inhalation route.
Dosing schedule: The animals should ideally be dosed on a 7
day per week basis. However, based primarily on practical
considerations, dosing on a 5 day per week basis is
acceptable. Treatment preferably should be continued for
the major portion of the animal's lifespan. This is at
least 18 months for mice and hamsters, and 24 months for
rats.
Data collection: During the study, animals should be
monitored for body weight and food intake, as well as for
the onset and progression of all toxic effects. Clinical
examinations, including hematology, biochemistry of blood,
urinalysis, and ophthalmological examination, should be
made. Gross necropsy and histopathology should be performed
on all animals. Specific requirements are too numerous to
list here, but may be reviewed via the EPA (1985) and NCI
(1976) guidelines.
All observed results should be evaluated by an appropriate
and generally accepted statistical method. Evidence for
carcinogenic action should be based on the observation of
statistically significant tumor responses in specific organs
or tissues. Appropriate statistical analysis should be
performed on data from long-term studies to help determine
whether the effects are treatment-related or possibly due to
chance. These should at least include a statistical test
for trend, including appropriate correction for differences
in survival. . The weight to be given to the level of
11
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statistical significance (the p-value) and. to other
available pieces of information is a matter of overall
scientific judgment. In a review of 25 NTP feeding studies
as discussed by OSTP (1985), a simple statistical rule was
derived by Haseman which appeared to mimic the scientific
judgment process used in those experiments. "Regard as
carcinogenic any chemical that produces a high dose increase
in a common tumor that is statistically significant at the
0.01 level or a high-dose increase in an uncommon tumor that
is statistically significant at the 0.05 level. The overall
false positive rate associated with this rule was estimated
to be no more than 7-8% for the NTP two-sex, two-species
protocol". A statistically significant excess of tumors of
all types in the aggregate, in the absence of a
statistically significant increase of any individual tumor
type, should be regarded as minimal evidence of carcinogenic
action unless there are persuasive reasons to the contrary
(OSTP, 1985).
These guidelines represent ideal parameters. Studies will
not be expected to meet all of these desirable conditions in
order to be further considered for use in the process. The
adequacy and appropriateness of all animal carcinogenicity
bioassays will be carefully considered. It is crucial that
judgment of adequate testing be based on sound scientific
principles. In general, it can be expected that most substances
tested for carcinogenicity have been reviewed by NCI/NTP, IARC,
and/or EPA. Historically the evaluations by these agencies have
been sufficient for decision-making. A thorough assessment of
the data should be performed regardless of the findings of those
independent agencies since these reviews might be dated in that
research data available subsequent to the date of review were not
considered by the reviewing group. The overall assessment of a
chemical's carcinogenic potential will depend on weight-of-
evidence based upon full consideration of all the evidence. Also
see Section III. Principles for Criteria Development for a
discussion of Mechanism/Mode of Action and the use of
mutagenicity studies in determining carcinogenicity.
3. Borderline Conditions
With regard to the overall database used in determining
carcinogenicity, a variety of studies may be encountered which
may be considered flawed or lacking in adequate design or
reporting. Such studies may only be able to be utilized
anecdotally and only considered suggestive evidence of
carcinogenicity. Examples of conditions meeting such a criteria
are:
1. Borderline conditions of:
12
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a. Statistical significance. A general example would be a
study in which the MTD was administered and the test
for positive dose-related trend (e.g., Cochran-Armitage
Test) determined that the slope of the dose-response
curve was different from zero; however, comparisons of
the tumor incidences in treated groups with that in the
control group (e.g., Fisher-Irwin exact test) were not
significant at p = 0.05.
b. Study design.
c. Study reporting. A general example would be a study
reporting a tumorigenic response, but lacking
statistical analyses to verify that an apparent
increase in incidence was statistically significant.
d. A tumor response in a tissue known to have a high and
variable background rate.
2. Tumor responses or lack of response which are more than
likely attributable to excessive doses that compromise major
organ systems. Positive studies at levels above the MTD
should be carefully reviewed to ensure that the responses
are not due to factors which do not operate at exposure
levels at or below the MTD. Evidence indicating that high
exposures alter tumor responses by indirect mechanisms that
may be unrelated to effects at lower exposures should be
dealt with pn an individual basis. As noted by the OSTP
(1985), "Normal metabolic activation of carcinogens may
possibly also be altered and carcinogenic potential reduced
as a consequence [of high-dose testing]." Negative long-
term animal studies at exposure levels above the MTD may not
be acceptable if animal survival is so impaired that the
sensitivity of the study is significantly reduced below that
of a conventional chronic animal study at the MTD.
"The carcinogenic effects of an agent may be influenced by
non-physiological responses (such as extensive organ damage,
radical disruption of hormonal function, saturation of
metabolic pathways, formation of stones in the urinary
tract, saturation of DNA repair with a functional loss of
the system) induced in the model systems. Testing regimes
inducing these responses should be evaluated for their
relevance to the human response to an agent and evidence
from such a study, whether positive or negative, must be
carefully reviewed." (OSTP, 1985).
3. Tumors at the site of oral, dermal or inhalation
administration attributable to irritation or frank tissue
damage.
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4. Tumor responses following administration by a route other
than oral, dermal or inhalation. Such tumors may be at the
site of administration or removed from it. Some general
examples are tumors induced following intraperitoneal,
intravenous or subcutaneous injection, or bladder
implantation.
Solid-state carcinogenesis is the occurrence of tumors
around an inserted inert object. It is a phenomenon that is
dependent primarily on the size and shape of the object,
rather than the chemical composition of the implanted
material (Williams and Weisburger, 1986). Therefore,
induction of solid-state tumors generally will not be
considered in the weight-of-evidence approach.
Data from all long-term animal studies should be considered
in evaluating carcinogenicity. However, carcinogenic
responses should be evaluated as to their relevance of
predicting cancer risks to humans. Therefore, data from
species that respond most like humans should be used
preferentially when such information exists. Data on tumors
in organs or as a result of effects on metabolic or
biochemical pathways that don't exist in humans should be
evaluated very carefully as to their inference of human
cancer risk-. Further, a positive carcinogenic response in
one species/strain or sex is not generally negated by
negative results in other species. Replicate negative
studies, however, that are essentially identical in all
other respects to a positive study may cast doubt on the
validity or reproducibility of a positive study. A variety
of other weight of evidence issues may make it difficult to
interpret the significance of tumor data and therefore
result in a lower classification of carcinogenicity.
Examples of such issues include: increased incidence of
tumors in the highest dose group only and/or only at the end
of the study; no substantial dose-related increase in the
proportion of tumors that are malignant; the occurrence of
tumors that are predominantly benign; no close-related
shortening of the time to the appearance of tumors; negative
or inconclusive results from a spectrum of short-term tests
for mutagenic activity; or, the occurrence of excess tumors
only in a single sex (EPA, 1985).
B. Noncarcinocrens
The full range of possible adverse health effects shall be
evaluated when establishing an acceptable exposure to
noncarcinogens. Acute/subacute, subchronic/chronic and
reproductive/developmental effects shall be considered. The
principles of data selection are similar to those for
carcinogenic effects, a well-conducted epidemiologic study which
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demonstrates a positive association between a quantifiable
exposure to a chemical and human disease is generally preferred
for evaluating adverse health effects. At present, however,
human data adequate to serve as a basis for quantitative risk
assessment are available for only a few chemicals. Frequently,
inference of adverse health effects to humans must be drawn from
toxicity information gained through animal experiments with human
data serving qualitatively as supporting evidence. Under this
condition, health effects data must be available from well
conducted studies in animals relevant to humans based on a
defensible biological rationale, i.e. similar metabolic pathways,
etc.
The following provides guidance on appropriate study design
for a variety of types of toxicity studies against which one may
evaluate the quality and adequacy of data development. This
evaluation of adequacy of data coupled with effects information
forms the basis for selection of uncertainty factors and
subsequent acceptable exposure levels.
1. Appropriate Study Design
a. Acute Toxicity
Acute Toxicity Determination of an LD50 or LC50 is often an
initial step in experimental assessment and evaluation of a
chemical's toxic characteristics. Such studies are used in
establishing a dosage regimen in subchronic and other studies and
may provide initial information on the mode of toxic action of a
substance. Because LD50 or LC50 studies are of short duration,
inexpensive and easy to conduct, they are commonly used in hazard
classification systems. Acute lethality studies are of limited
use in this process. However, the data from such studies do
provide information on health hazards likely to arise from
individual short-term exposures. Although this process should
never allow exposures which approach such acute levels, such
studies provide high dose effects data from which to evaluate
potential effects from exposures which may temporarily exceed the
acceptable chronic exposure level. An evaluation of the data
should include the incidence and severity of all abnormalities,
the reversibility of abnormalities observed other than lethality,
gross lesions, body weight changes, effects on mortality, and any
other toxic effects.
In recent years guidelines have been established to,improve
quality and provide uniformity in test conditions.
Unfortunately, many published LD50 or LC50 tests were not
conducted in accordance with current EPA or OECD guidelines since
they were conducted prior to establishment of guidelines. For
this reason, it becomes necessary to examine each test or study
to determine if the study was conducted in an adequate manner.
The following is a list of ideal conditions compiled from
various testing guidelines which may.be used for determination of
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adequacy. Unfortunately, many published studies do not report
details of test conditions making such determinations difficult.
However, test conditions guidelines that might be considered
ideal may include:
animal- age and species identified;
minimum of 5 animals per sex per dose group (Both sexes
should be used.);
14 day or longer observation period following dosing;
minimum of 3 dose levels appropriately spaced. (Most
statistical methods require at least 3 dose levels.);
identification of purity or grade of test material used
(particularly important in older studies);
if a vehicle used, the selected vehicle is known to be
non-toxic;
gross necropsy results for test animals; or
acclimation period for test animals before initiating
study.
Specific conditions for oral LD50:
dosing by gavage or capsule;
total volume of vehicle plus test material remain
constant for all dose levels; and
animals were fasted before dosing.
Specific conditions for dermal LD50:
exposure on intact, clipped skin and involve
approximately 10% of body surface; and
animals prevented from oral access to test material by
restraining or covering test site.
Specific conditions for inhalation LC50:
duration of exposure at least 4 hours; and
if an aerosol (mist or particulate) the particle size
(median diameter and deviation) should be reported.
Although the above listed conditions would be included in an
ideally conducted study, not all of these conditions need to be
included in an adequately conducted study. Therefore, some
discretion is required on the part of the individual reviewing
these studies (EPA, 1985, OECD, 1987).
b. 14 Day or 28 Day Repeated Dose Toxicity
The following guidelines were derived using the OECD
Guideline for Testing of Chemicals (1987), for determining the
design and quality of a repeated dose short-term toxicity study.
The similarity between the conduct of a 14-day and 28-day study
is sufficient to consider them under the same guideline. The
main difference is the time period over which the dosing takes
place. These guidelines represent ideal conditions and studies
will not be expected to meet all standards in order to be
considered. For example, the National Toxicology Program's
cancer bioassay program has generated a substantial database of
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short-term repeated dose studies. The study periods for these
range from 14 days to 20 days with 12 to 15 doses administered
generally for 5 dose levels and a control. Since the quality of
this data is goo'd, it is desirable to consider these study
results even though they do not always identically follow the
protocol.
The purpose of short-term repeated dose studies is to
promote information on possible adverse health effects from
repeated exposures over a limited time period. Where subchronic
or chronic data are lacking, short-term repeated dose studies of
28 days or longer, with the application of appropriate
uncertainty factors, may be used by this initiative to estimate
acceptable long-term exposure levels.
According to OECD Guidelines, short-term repeated dose
studies should include the following:
minimum of 3 dose levels administered and an adequate
control group used;
minimum of 10 animals per sex, per dose group (both
sexes should be used);
the highest dose level should ideally elicit some signs
of toxjLcity without inducing excessive lethality and
the lowest dose should ideally produce no signs of
toxicity;
ideal dosing regimes include 7 days per week for a
period of 14 days or 28 days;
all animals should be dosed by the same method during
the entire experiment period;
animals should be observed daily for signs of toxicity
during the treatment period (i.e. 14 or 28 days).
Animals which die during the study are necropsied and
all survivors in the treatment groups are sacrificed
and necropsied at the end of the study period;
all observed results, quantitative and incidental,
should be evaluated by an appropriate statistical
method;
clinical examinations should include hematology and
clinical biochemistry, urinalysis may be required when
expected to provide an indication of toxicity.
Pathological examination should include gross necropsy
and histopathology.
The findings of short-term repeated dose toxicity studies
should be considered in terms of the observed toxic effects and
the necropsy and histopathological findings. The evaluation will
include the incidence and severity of abnormalities, gross
lesions, identified target organs, body weight changes, effects
on mortality and' other general or specific toxic effects (OECD,
1987).
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c. Subchronic and Chronic Toxicity
The following guidelines were derived using the EPA Health
Effects Testing Guidelines (1985), for determining the quality of
a subchronic or chronic (long term) study. Additional detailed
guidance may be found in that document. These guidelines
represent ideal conditions and studies will not be expected to
meet all standards in order to be considered. The subchronic and
chronic studies have been designed to permit determination of no-
observed-effect levels (NOEL) and toxic effects associated with
continuous or repeated exposure to a chemical. Subchronic
studies provide information on health hazards likely to arise
from repeated exposure over a limited period of time. They
provide information on target organs, the possibilities of
accumulation, and, with the appropriate uncertainty factors, may
be used in establishing safety criteria for human exposure.
Chronic studies provide information on potential effects
following prolonged and repeated exposure. Such effects might
require a long latency period or are cumulative in nature before
manifesting disease. The design and conduct of such tests should
allow for detection of general toxic effects including
neurological, physiological, biochemical and hematological
effects and exposure-related pathological effects.
According to the EPA Guidelines, high quality
subchronic/chronic studies include the following:
minimum of 3 dose levels administered and an adequate
control group used;
minimum of 10 animals for subchronic, 20 animals for
chronic studies per sex, per dose group (both sexes
should be used);
the highest dose level should ideally elicit some signs
of toxicity without inducing excessive lethality and
the lowest dose should ideally produce no signs of
toxicity;
ideal dosing regimes include dosing for 5-7 days per
week for 13 weeks or greater (90 days or greater) for
subchronic and at least 12 months or greater for
chronic studies in.rodents. For other species,
repeated dosing should ideally occur over 10% or
greater of animals lifespan for subchronic studies and
50% or greater of the animal's lifespan for chronic
studies;
all animals should be dosed by the same method during
the entire experimental period;
animals should be observed daily during the treatment
period (i.e., 90 days or greater);
animals which die during the study are necropsied and,
at the conclusion of the study, surviving animals are
sacrificed and necropsied and appropriate
histopathological examinations carried out;
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results should be evaluated by an appropriate
statistical method selected during experimental design;
and
such toxicity tests should evaluate the relationship
between the dose of the test substance and the
presence, incidence and severity of abnormalities
(including behavioral and clinical abnormalities),
gross lesions, identified target organs, body weight
changes, effects on mortality and any other toxic
effects noted (EPA, 1985).
d. Reproductive and Developmental Toxicity
Studies considered here can be evaluated for quality by
comparing the study protocol or methods section with accepted
testing guidelines prepared by EPA, OECD or Interagency
Regulatory Liaison Group (IRLG). The EPA Health Effects Testing
Guidelines (1985) include guidelines for both reproduction and
fertility studies and developmental studies. These EPA
guidelines can serve as the ideal experimental situation with
which to compare study quality. Studies being evaluated do not
need to match precisely but rather should be similar enough that
one can be assured that the chemical was adequately tested and
that the results closely reflect the true reproductive or
developmental toxicity of the chemical.
Developmental toxicity can be evaluated via a relatively
short-term study in which the compound is administered during the
period of organogenesis. Some of the specific guidelines for
developmental studies are cited below.
minimum of 20 young, adult, pregnant rats, mice or
hamsters or 12 young, adult, pregnant rabbits
recommended per dose group;
minimum of 3 dose levels with an adequate control group
used;
the highest dose should induce some slight maternal
toxicity but no more than 10% mortality. The lowest
dose should not produce grossly observable effects in
dams or fetuses. The middle dose level, in an ideal
situation, will produce minimal observable toxic
effects;
dose period should cover the major period of
organogenesis (days 6 to 15 gestation for rat and
mouse, 6 to 14 for hamster, and 6 to 18 for rabbit);
dams should be observed daily; weekly food consumption
and body weight measurements should be taken;
necropsy should include both gross and microscopic
examination of the dams; the uterus should be examined
so that the number of embryonic or fetal deaths and the
number of viable fetuses can be counted; fetuses should
be weighted; and
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one-third to one-half of each litter should be prepared
and examined for skeletal anomalies and the remaining
animals prepared and examined for soft tissue
anomalies.
The EPA.Health Effects Testing Guidelines (1985) recommend a
two-generation reproduction study to provide information on the
ability of a chemical to impact gonadal function, conception,
parturition and the growth and development of the offspring.
Additional information concerning the effects of a test compound
on neonatal morbidity, mortality and developmental toxicity may
also be provided. The recommendations for reproductive testing
are lengthy and quite detailed and may be reviewed further in the
Health Effects Testing Guidelines. In general, the test compound
is administered to the parental (P) animals (at least 20 males
and enough females to yield 20 pregnant females) at least 10
weeks before mating, through the resulting pregnancies and
through weaning of their Fl offspring. The compound is then
administered^ to the Fl generation similarly through the
production of their F2 offspring until weaning. Recommendations
for numbers of dose groups and dose levels are similar to those
reported for developmental studies. Details are also provided on
mating procedures, standardization of litter sizes (if possible,
4 males and 4 females from each litter are randomly selected),
observation, gross necropsy and histopathology. Full
histopathology is recommended on the following organs of all high
dose and control P and Fl animals used in mating: vagina,
uterus, testes, epididymides, seminal vesicles, prostate,
pituitary gland and target organs. Organs of animals from other
dose groups should be examined when pathology has been
demonstrated in high dose animals (EPA, 1985).
As with any other type of study, the appropriate statistical
analyses must be performed on the data for a study to qualify as
a good quality study. In addition, developmental studies are
unique in the sense that they yield two potential experimental
units for statistical analysis, the litter and the individual
fetus. The EPA testing guidelines do not provide any
recommendation on which unit to use, but the Guidelines for the
Health Assessment of Suspect Development Toxicants (EPA, 1986)
states that "since the litter is generally considered the
experimental unit in most developmental toxicity studies, the
statistical analyses should be designed to analyze the relevant
data based on incidence per litter or on the number of litters
with a particular end point". Others (Palmer, 1981 and Madson et
al., 1982) identify the litter as the preferred experimental unit
as well.
Information on maternal toxicity is very important when
evaluating developmental effects because it helps determine if
differential susceptibility exists for the offspring and mothers.
Since the conceptus relies on its mother for certain
physiological processes, interruption of maternal homeostasis
could result in abnormal prenatal development. Substances which
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affect prenatal development without compromising the dam are
considered to be a greater developmental hazard than chemicals
which cause developmental effects at maternally toxic doses.
Unfortunately, maternal toxicity information has not been
routinely presented in earlier studies and has become a routine
consideration in, studies only recently. In an attempt to use
whatever data are available, maternal toxicity information may
not be required if developmental effects are serious enough to
warrant consideration regardless of the presence of maternal
toxicity.
C. Tier Designation
1. Carcinogens .
Adequate weight-of-evidence of potential human carcinogenic
effects sufficient to calculate a Tier I Human Cancer Criterion
(HCC) generally consists of data sufficient to meet the
categorical definition of a Human Carcinogen and Probable Human
Carcinogen. Certain Possible Human Carcinogens may also be
suitable for Tier I criterion development. Designation of
Possible Carcinogens should be done on a case-by-case basis. For
example, where cancer bioassays have been well conducted, yet are
limited because they only involve a single animal species, strain
or experiment and do not demonstrate a high incidence, unusual
site or type of tumor, or early onset of tumorigenesis, such data
may be suitable for Tier I criterion development. In addition,
mode of action, the potential for the compound to interact
directly with DNA as discussed earlier, should be reviewed in
making a Tier designation.
As discussed earlier, data used for developing Tier I
criteria are expected to carry a high degree of certainty in
their ability to predict an effect. In this case, the quality of
data and the weight-of-evidence needs to be sufficient to
ascertain that the chemical holds at least a good potential of
producing carcinogenic effects in humans.
For chemicals where the weight-of-evidence and quality of
data is not sufficient for Tier I numeric criteria the database
may be adequate to develop Tier II values. In this case, the
data needs to be sufficient to ascertain that the chemical is at
least a possible4 human carcinogen, i.e. Group C. As discussed
previously under Weight-of-Evidence and Appropriate Study Design,
data on chemicals in this Group suggest only limited evidence of
carcinogenicity. Studies may be flawed or lacking adequate
design or reporting yet show strong enough evidence of
carcinogenicity or the potential for carcinogenic effects such
that the data should not be ignored. Examples of such data may
be studies where statistical analysis may be lacking or tumor
incidence may be only marginally significant; tumor responses or
lack of response* may be attributable to excessive dosing, or
there may be high mortality in the exposed groups also due to
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excessive dosing; increases exist for benign tumors only with no
evidence of mutagenicity, etc. Further discussion as to how
these data are treated in criteria derivation and what potential
differences may exist in such treatment will be discussed further
in the section on criteria development.
It is important to note that the Group C category may
contain chemicals with databases of highly variable quality.
Because of this," EPA has decided to allow States and Tribes to
address Group C chemicals on a case-by-case basis. As the final
GLWQI Guidance is written, States and Tribes have the discretion
to develop Tier I criteria or Tier II values for Group C
chemicals based on the overall toxicological database. The
final Guidance directs that this case-by-case determination be
made taking into account information on mode of action, including
mutagenicity, genotoxicity, structure activity and metabolism.
Those Group C chemicals (and all chemicals, in general) which act
via a genotoxic mechanism, that is through direct interaction
with DNA and in which a linear low-dose tumor incidence
relationship is expected, may be most appropriately dealt with
through use of a linearized multistage model (LMS) or other model
which appropriately reflect this type of mechanism of action.
The quality of data, as discussed above, would then determine the
Tier designation. If the chemical does not interact with DNA and
the dose response is considered nonlinear, it may be best dealt
with as a noncancer agent and an RfD should be developed. See
section on Mode of Action under Section III. Principles for
Criteria Development, B. Carcinogens.
2. Noncarcinogens
All available toxicity data should be evaluated considering
the full range of possible effects of a chemical. Unfortunately,
expansive data exists for a limited number of chemicals.
Although all data are evaluated, a line must be drawn below which
data are not sufficient for criteria development. Adequate data
necessary to develop a Human Noncancer Criterion (HNC) for
noncancer effects should ideally incorporate at least one well
conducted epidemiologic study which demonstrates a positive
association between a quantifiable exposure to a chemical and
human disease. Such data exist for only a few chemicals,
therefore, reliance on animal data in establishing noncancer
criteria and values is usually necessary. Although a more
extensive effects database is desirable, for this initiative, the
minimum database for a Tier I criterion must contain at least a
well conducted subchronic mammalian study. The duration of the
study must be at least 90 days in rodents or 10% of the lifespan
of other appropriate species with exposure preferably via the
oral route. Subchronic toxicity studies utilizing dosing periods
of approximately 10% of the test animal's lifespan (approx. 90
days in rodents) are sufficient to provide information on target
organ effects and can provide an estimate of a no effect level of
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exposure which can be used to establish human health criteria and
values (OECD, 1981).
It has been observed, with up to a 95% degree of certainty,
that as little as a 6 fold difference may exist between chemical
effect levels observed at 90 days exposure and at lifetime (7
years) in rodents. (Weil,et al.,1969.) Such a study (90-day or
otherwise used to develop a HNV) should ideally establish a
frank-effect-level (PEL), a lowest-observed-adverse-effect-level
(LOAEL) and a no-observed-adverse-effect-level (NOAEL). The
study must be conducted in an animal species relevant to humans
(for example, birds, reptiles, and fish are not considered
biologically relevant to humans due to incompatible
pharmacokinetics, organ structure, toxicokinetics, etc.) based on
a defensible biological rationale and generally follow the study
protocol previously discussed. To further reduce uncertainty,
data from longer studies approaching the lifetime of the test
animal are preferable. In some cases, chronic studies of one
year or longer in rodents or 50% of the lifespan or greater in
other appropriate test species may b$ sufficient. Dose response
must be demonstrated in these longer term studies, however a
LOAEL involving relatively mild and reversible effects may be
considered an acceptable data point for decision making. For
example, there are many studies for which only one dose has been
tested with resulting minimal, reversible effects such as minimal
enzyme changes or slight body weight decreases. These minimal
changes or effects, on their own, may not be thought of as
adverse but may be indicators or precursors to more severe
effects which result from extended exposure and or higher doses.
In those cases, while it can be argued that such an effect may be
a LOAEL, it may also be very close to the NOAEL and is therefore
suitable for criteria derivation.
Reproductive/developmental effects data as well as evidence
of effects seen in test animals consistent with human
epidemiologic data are also highly desirable in order to evaluate
the full range of potential adverse effects to humans. When data
are not sufficient to meet the minimum requirements for deriving
Tier I numeric criteria, such data may be considered for
development of Tier II values. As with Tier I, all available
data should be considered, however, a minimum database suitable
for Tier II must, contain at least a well conducted subacute
mammalian study with an exposure period of at least 28 days,
preferably via the oral route of exposure. The 28-day study was
chosen as a minimally acceptable test that can yield sufficient
information upon which to derive a Tier II value. Please refer
to Appendix A for further discussion of the use of less than
chronic data to predict chronic endpoints. The study should,
ideally, establish a dose-response relationship including a
frank-effect-level (PEL), a lowest-observed-adverse-effect-level
(LOAEL), and a no-observed-adverse-effect-level (NOAEL).
Acceptable protocol for conducting such 28 day studies may be
found in the OECD Testing Guidelines (OECD, 1987) as discussed
previously. Although the effects observed from short duration
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studies are usually fewer than normally evaluated in longer
duration studies., such effects should at least include mortality,
clinical observations, body weight changes and necropsy of major
organs with whatever histopathology that may be available. The
minimum data point for decision making on such short term
exposure data must be a NOAEL. A NOAEL was chosen over a LOAEL
since it is believed the use of a LOAEL may result in
underprotective Tier II values. A LOAEL from a 28-day study may
not capture the most critical toxic endpoint or be predicting of
chronic endpoints. Structure-activity relationship (SAR) review
should also accompany the minimum data evaluation. SAR compares
a chemical with substances that have structural similarities in
order to predict whether the chemical might cause similar toxic
effects. Such information may then be used in deciding what
uncertainty factors may be appropriate to apply to such limited
data in order to protect against potential similar effects.
Studies of longer duration than 28 days and with greater
evaluation of effects are more desirable for use in Tier II and
may allow the use of a LOAEL for decision making, depending on
the quality and duration of the study. As with Tier I,
reproductive/developmental effects data as well as any supportive
epidemiologic evidence is highly desirable in order to evaluate
the full range of potential adverse effects of the chemical. As
with carcinogens, further discussion as to how these data will be
applied in the derivation of acceptable exposure levels and what
adjustments must be made to account for uncertainty will be
discussed in further detail in the section on criteria
development.
III. PRINCIPLES FOR CRITERIA DEVELOPMENT
A. General
The process to derive Tier I criteria or Tier II values is
generally the same. The weight of evidence and level of
certainty in the^ data available for calculating acceptable
exposure levels establishes the major difference between the two.
For risk assessment of noncarcinogenic effects, the minimum data
requirements differ between tiers. Therefore, differences in
adjustments to the data (i.e., uncertainty factors) may also
occur between tiers. These differences reflect differing levels
of certainty in the data base and an attempt to estimate a level
without appreciable risk of deleterious effects over a lifetime.
In the case of carcinogens, the same quantitative risk assessment
approach generally followed for Tier I is used as well for Tier
II when the data allow. When the bioassay data for Tier II
carcinogens are not suitable for quantitative risk assessment and
the chemical does not appear to interact with DNA, yet the
weight-of-evidence supports concern for possible threshold
carcinogenic effects, an additional uncertainty factor may be
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applied to the LOAEL or NOAEL for the chemical in order to
account for carcinogenicity.
All available appropriate human epidemiologic data and
animal toxicologic data shall be considered. Data specific to an
environmentally appropriate route of exposure shall be used for
criteria and values development, i.e. oral, dermal or inhalation
versus injection, implantation, etc. Findings from studies using
less than appropriate routes of exposure may be considered
supportive of data obtained through more appropriate routes.
Although local effects are important, for the purposes of this
initiative oral exposure should be considered preferential to
dermal and inhalation data since ingestion is the primary route
of exposure, i.e. water and fish consumption. Caution must be
exercised in the use of dermal and inhalation data. Strong
consideration must be given for pharmacokinetic information on
absorption, distribution and metabolism in establishing
equivalent doses with oral exposure. Effects produced through
exposure via a non-oral route generally should be as a result of
systemic distribution of a toxicant rather than as local effects
to the skin or the respiratory tract.
In general, study results shall be converted, as necessary,
to the standard unit of milligrams of toxicant per kilogram
of body weight per day (mg/kg/day).
If a study does not specify water or food consumption rates,
or body weight of the test animals, standard values may be
used for the test species, such as may be obtained from the
National Institute of Occupational Safety and Health,
Registry of Toxic Effects of Chemical Substances (RTECS) or
similar appropriate references.
Study results from multiple exposures shall be adjusted, as
necessary, to a daily dose exposure as if received daily for
the duration of the exposure period. The exposure period
shall be defined as the interval beginning with
administration of the first dose through the last dose,
inclusively.
B. Carcinogens
1. Mechanism/Mode of Action
The mechanism by which chemicals cause cancer is not
completely known4, and may involve a variety of mechanisms
occurring at various stages in the carcinogenic process. A
chemical may act at a single stage or more than one stage.
Currently, the dominant theory regarding the process by which a
chemical causes cancer is based on two stages: initiation and
promotion (Borzsonyi, 1984; OSTP, 1985; Trosko, 1983; Williams,
1986). The concept of two-stage carcinogenesis has been
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supported by investigations involving skin and liver systems
(Argyris, 1985; Pitot and Sirica, 1980). This operational theory
allows the classification of carcinogens according to their
apparent biological activity. Some chemicals are capable, by a
variety of genotoxic mechanisms, of triggering the carcinogenic
process (initiation). Other chemicals may only alter the
expression of the initiated genome and enhance tumor development
by a variety of non-genetic mechanisms (promotion). Complete
carcinogens operate by both processes. Initiators are capable of
directly altering in an irreversible manner the native structure
of the DNA. Promotion may be reversible in the early stages,
appears to be highly dose-dependent, and apparently requires
prolonged or repeated exposure (Pitot, 1981; Slaga, 1984; Thomas,
1986) .
Calling an agent a promoter does not eliminate the
carcinogenic hazard potential of a chemical. Indeed, data
indicate that promoting phenomena are largely responsible for the
expression of many human cancer types (Williams, 1986). However,
it is very difficult both in principle and in practice to confirm
the assertion that a given chemical acts by promotion alone
(OSTP, 1985) .
Currently, for most if not all chemicals, data are not
available to determine the exact mechanism by which they cause
cancer. As a result, significant controversy exists regarding
the existence of thresholds for carcinogens. Therefore, for the
purpose of routine cancer risk assessment, agents that are
positive in long-term animal experiments should be considered as
complete carcinogens unless there is evidence to the contrary
because, at present, it is difficult to determine whether an
agent is acting only as a promoting or cocarcinogenic agent (EPA,
1986) . However, in making all judgements with regard to
mechanism, all data related to mode of action should be
considered
EPA, in revising its Guidelines for Carcinogenic Risk
Assessment, is suggesting that mode of action information,
reflecting the manner in which an agent causes cancer, be used
more extensively in carcinogen assessments than has been done in
the past. As a result, the final GLWQI Guidance now includes a
requirement to review all possible evidence including available
information on mode of action including
mutagenicity/genotoxicity, structure activity, and metabolism.
Mode of action should be used in the assessment and
characterization of the potential human carcinogenicity of a
substance, and in the selection of a model for quantifying its
risks, especially at low doses. This change in emphasis, while
still draft and in the formative stages, is being recommended so
that all relevant scientific data can be used to carry out cancer
risk assessments. Of particular importance, in determining mode
of action, is distinguishing between carcinogens that are
mutagenic (i.e., interact directly with DNA) and those that are
non-mutagenic.
26
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To distinguish between carcinogenic agents that are
mutagenic and non-mutagenic, many test systems can be used.
These include assays for changes in DNA base pairs of a gene such
as gene mutation- tests in bacteria or mammalian cells (see 40 CFR
798:5265, USEPA 1991) and chromosomal aberrations, such as in
vivo cytogenetics tests. Initial consideration is usually given
to mammalian-bone marrow using either micronucleus assays to
detect damage of chromosomes or mitotic apparatus by agents (See
40 CFR 798:5398, USEPA 1991; Dearfield et al. 1991) or metaphase
chromosomal analysis for detection of structural aberrations
(also see 40 CFR 798:5385, USEPA 1991). Mutagenicity assessment
guidelines are provided in USEPA (1991). Other assays that do
not measure gene- mutations or chromosomal aberrations per se
(e.g., tests for DNA adducts, unscheduled DNA synthesis, sister
chromatid exchange, strand breaks, repair and recombination) are
not sufficient in and of themselves to make a determination of
mutagenicity; they only provide supportive evidence of
mutagenicity.
For the purpose of this initiative, unless adequate
mechanistic data demonstrate otherwise, a nonthreshold mechanism
will be assumed for those chemicals classified as Group A, B and
C carcinogens.
2. Data Review
If acceptable human epidemiologic data are available, a risk
associated dose shall be set equal to the lifetime exposure which
would produce an incremental increased cancer risk of 1 in
100,000. If more than one study is judged acceptable, the study
resulting in the' most protective risk associated dose is
generally used to calculate the human cancer criterion.
In the absence of appropriate human studies, data from a
species that responds most like humans should be used, if
information to this effect exists. Where several studies are
available, which may involve different animal species, strains,
and sexes at several doses and by different routes of exposure,
the following approach to selecting the data sets is used:
The tumor incidence data are separated according to organ
site and tumor type.
All biologically and statistically acceptable data sets are
presented.
The range of the risk estimates is presented with due regard
to biological relevance (particularly in the case of animal
studies) and appropriateness of route of exposure.
Because is it possible that human sensitivity is as high as
the most sensitive responding animal species, in the absence
of evidence to define the most relevant species to humans,
27
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the biologically acceptable data set from long-term animal
studies showing the greatest sensitivity should generally be
given the greatest emphasis, again with due regard to
biological and statistical consideration (EPA, 1986).
Exceptions to the above may exist as follows:
If two or more studies exist which are identical with
respect to species, strain, sex and tumor type and are of
equal quality, the geometric mean of the potency from these
studies may be used (EPA, 1980). In certain instances where
there are several studies in various strains and even
several species and where there is no indication of a single
study or species judged most appropriate, the geometric mean
estimates from all studies may be used to determine the
potency. This ensures that all relevant data are included
in the derivation (EPA, 1989d).
Where two or more significantly elevated tumor sites or
types are observed in the same study, extrapolations may be
conducted on selected sites or types. These selections will
be made on biological grounds. To obtain a total estimate
of carcinogenic risk, animals with one or more tumor sites
or types showing significantly elevated tumor incidence
should be pooled and used for extrapolation so long as
double-counting of tumor-bearing animals is prevented. The
pooled estimates will generally be used in preference to
risk estimates based on single sites or types. Quantitative
risk extrapolations will generally not be done on the basis
of totals that include tumor sites without statistically
significant elevations (EPA, 1986).
Benign tumors should generally be combined with malignant
tumors for risk estimates unless the benign tumors are not
considered to have the potential to progress to the
associated malignancies of the same histogenic origin. The
contribution of the benign tumors, however, to the total
risk should, be indicated (EPA, 1986).
3. Model
a. Nonthreshold Approach
When acceptable human epidemiologic data are not available
and a nonthreshold mechanism is assumed, carcinogenesis bioassay
data, as appropriate, are fitted to a linearized multistage
computer model. (Note: Other models, such as time-to-tumor,
modifications or variations of the multistage model may be used
which consider the data more appropriately under case-by-case
circumstances.)
28
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Since risks at low exposure levels cannot be measured
directly either by animal experiments or by epidemiologic
studies, a number of mathematical models have been developed to
extrapolate from high to low dose. Different extrapolation
models, however, may fit the observed data reasonably well but
may lead to large differences in the projected risk at low doses
(EPA, 1986).
"No single mathematical procedure is recognized as the most
appropriate for low-dose extrapolation in carcinogenesis. When
relevant biological evidence on mechanism of action exists (e.g.,
pharmacokinetics, target organ dose), the models or procedures
employed should be consistent with the evidence. When data and
information are limited, however, and when much uncertainty
exists regarding the mechanism of carcinogenic action, models or
procedures which incorporate low-dose linearity are preferred
when compatible with the limited information." (OSTP, 1985)
In an attempt to characterize the underlying dose-response
relationship, models which use the nonthreshold assumption of
carcinogenicity are commonly used. The linearized multistage
(LMS) model calculates an upper bound based on the theory that a
developing tumor goes through several different stages which can
be affected by a chemical carcinogen. The LMS model is forced to
be linear in the low-dose region, regardless of the shape of the
dose response curve, and therefore LMS-based risk estimates may
be regarded as relatively conservative when used for public
health protection.
In calculating upper bounds on potency from the LMS model,
the bioassay data are fitted to the LMS model, e.g. Global 86
developed by Howe et al. (1986). The 95 percent upper bound
estimate on the linear term, q.^*! is used to calculate the upper
confidence bound on risk for a given dose, or the lower
confidence bound on dose for a given risk. The slope factor
(q-L*) is taken as an upper bound of the potency of the chemical
in inducing cancer at low doses. When pharmacokinetic or
metabolism data are available, or when other substantial evidence
on the mechanistic aspects of the carcinogenesis process exists,
a low-dose extrapolation model other than the linearized
multistage procedure might be considered more appropriate on
biological grounds. When a different model is chosen, the risk
assessment should clearly discuss the nature and weight-of-
evidence that led to the choice. Considerable uncertainty will
remain concerning response at low doses; therefore, in most cases
an upper-limit risk estimate using the linearized multistage
procedure should also be presented for comparison (EPA, 1986).
b. Threshold Approach
Whenever appropriate human epidemiological data are not
available, and the preponderance of data suggest that a chemical
causes cancer via a threshold mechanism, the risk associated dose
29
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may be calculated via a method other than a linearized multistage
model on a case-by-case basis.
As a default, a safety factor approach may be pursued after
thorough evaluation of. the toxicologic and pharmacologic data on
the compound including, but not limited to: mechanism of
carcinogenesis, number and type of tumors induced, the
spontaneous incidence of tumors, the number of animal species
tested and affected, metabolic considerations, epidemiologic
data, extent of data supporting a non-genotoxic mechanism of
tumor induction,' i.e. mutagenicity assay data, initiation/
promotion assay data, etc.
4. Lifespan Adjustment
If the duration of the study (Le) is significantly less than
the natural lifespan for the species (L), the slope factor, q *
can be adjusted to account for unobserved tumors due to the snort
study duration. * The assumption is that if the duration of the
study was increased, tumor incidence would continue to increase
as a constant function of the background rate. EPA believes this
adjustment should be made on a case-by-case basis taking into
consideration factors such as mechanism of action, the type of
tumor and the organ affected. One option for correcting for less
than lifetime duration is to use the method described by EPA
(1980). Based on the EPA (1980) method it is assumed that the
cumulative tumor rate would increase at least by the 3rd power of
age since age specific rates for humans increase at least by the
2nd power of age and often considerably higher.
For mice and rats, the natural lifespan (L) is defined as 90
weeks and 104 weeks, respectively. The slope factor adjustment
may be conducted for mice and rat data if the study duration (Le)
is significantly less than natural lifetime such as less than 78
weeks for mice or 90 weeks for rats, by multiplying the slope
factor by the factor (L/Le) . For other species, this adjustment
factor may also be used whenever appropriate, using species-
specific values for L and the Le trigger level. The latter may
be determined using the trigger levels for mice and rats as a
guideline.
5. Species Scaling
Low-dose risk estimates derived from laboratory animal data
extrapolated to humans are complicated by a variety of factors
that differ among species and potentially affect the response to
carcinogens. Included among these factors are differences
between humans and experimental test animals with respect to life
span, body size, genetic variability, population, homogeneity,
existence of concurrent disease, pharmacokinetic effects such as
metabolism and excretion patterns, and the exposure regimen.
30
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The usual approach for making interspecies comparisons has
been to use standardized scaling factors. Commonly employed
standardized dosage scales include3mg per kg body weight per dayj
ppm in the diet or water, mg per m body surface area per day,
and mg per kg body weight per lifetime. In the absence of
comparative toxicological, physiological, metabolic and
pharmacokinetic data for a given chemical, extrapolation on the
basis of surface area is considered to be most appropriate
because certain pharmacological effects commonly scale according
to surface area '(Dedrick, R.L., J. Pharmacokin. Biopharm. 1:435-
461; Freireich et al., Cancer Chemother. Rep. 50:219-244 (1966);
Pinkel, D., Cancer Res., 18:853-856 (1958)).
The species scaling factor is calculated by dividing the
average weight of a human (Wh) by the weight of the test species
(Wa) and taking the cube root of the resultant value. This is
based on the premise that a close approximation of the surface
area is 2/3 the power of weight, and that the exposure in mg-2/3
the power of body weight/day is similarly considered to be an
equivalent exposure (EPA, 1980). The animal slope factor is
multiplied by this factor to obtain the human slope factor.
yi^ (human) = q.* (animal) X 3 J "" kg
\ Wa kg
The weight (Wa) of the test species should be the average adult
weight from the particular bioassay if possible, or derived from
available data tables or standard assumed weights.
EPA also believes other scaling factors may be used as long
as there is justification on the basis of species-specific
pharmacokinetic data.
C. Noncarc inogens
1. Mechanism
Noncarcinogens generally are assumed to have a threshold
dose or level below which no adverse effects should be observed
(NOAEL). For many noncarcinogenic effects, protective mechanisms
are believed to exist that must be overcome before an adverse
effect is manifested. For example, where a large number of cells
perform the same or similar function, the cell population may
have to be significantly depleted before the effect is seen. As
a result, a range of exposures exists from zero to some finite
value that can be tolerated by the organism with essentially no
expression of adverse effects. In the development of an estimate
without appreciable risk of deleterious effect from exposure to a
chemical, the effort exists to find the upper bound of this
tolerance range (i.e., the maximum subthreshold level). Because
variability exists in the human population, attempts are made to
31
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assure a subthreshold level which would not result in appreciable
risk to sensitive individuals in the population. For most
chemicals, this level can only be estimated and incorporates the
use of uncertainty factors indicating the degree of extrapolation
used to derive the estimated value (EPA, 1989d).
Exceptions to this principle exist. Noncarcinogenic
chemicals may exist with no identifiable threshold. One example
of this phenomenon appears to be nickel, for which there is no
apparent threshold for subsequent dermal effects of the chemical.
Another example is the effects of lead exposure, where no
discernable threshold has been identified. Other examples of
this exception may include genotoxic teratogens and germline
mutagens. These agents have been specifically identified to
differentiate between chemicals thought to produce reproductive
and/or developmental effects via a genetically linked effect from
those chemicals more routinely considered to act via a nongenetic
mechanism. There are few chemicals, if any, which currently have
sufficient mechanistic information about their mode of action to
link teratogenic or developmental effects to mutational events
during organogenesis, histogenesis or other stages of
development. These chemicals may also interact with germ cells
to produce mutations which may be transmitted to the zygote and
also be expressed during one or more of these stages of
development.
EPA has recognized this potential and discussed this issue
in their 1989 Proposed Amendments to Agency Guidelines for Health
Assessments of Suspect Developmental Toxicants (EPA, 1989c) and
in their 1986 Guidelines for Mutagenicity Risk Assessment (EPA,
1986) . Various statements within these guidelines should raise
concern for the potential for future generations inheriting
chemically induced germline mutations or suffering from
mutational events occurring in utero:
"It is estimated that at least 10% of all human disease is
related to specific genetic abnormalities...11
"Life in our technological society results in exposure to
many natural and synthetic chemicals. Some have been shown
to have mutagenic activity in mammalian and sub-mammalian
test systems, and these may have the potential to increase
genetic damage in the human population... The extent to
which exposure to natural and synthetic environmental agents
may have increased the frequency of genetic disorders in the
present human population and contributed to the mutational
"load" that will be transmitted to future generations is
unknown at this time. However, for the reasons cited above,
it seems prudent to limit exposure to potential mutagens."
"Approximately 3% of newborn children are found to have one
or more significant congenital malformation at birth, and by
the end of the first postnatal year, about 3% more are found
to have serious developmental defects. Of these, it is
32
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estimated that 20% are of known genetic transmission, 10%
are attributable to known exogenous factors (including
drugs, infections, radiation and environmental agents) ..."
An awareness of the potential for such teratogenic/mutagenic
effects should be established in order to deal with such data
should it occur in the future. However, without adequate data to
support a genetic or mutational basis for developmental or
reproductive effects, the default becomes an uncertainty factor
approach. This approach follows the procedure identified for
noncarcinogens assumed to have a threshold. Genotoxic teratogens
and germline mutagens should be considered an exception while the
traditional uncertainty factor approach is the general rule for
calculating criteria or values for chemicals demonstrating
developmental/ reproductive effects.
A nonthreshpld mechanism shall be assumed for genotoxic
teratogens and germline mutagens. Since there is no well
established mechanism for calculating criteria protective of
human health from the effects of these agents, criteria will be
established on a case-by-case basis. For more information on
this phenomenon, it is recommended that the reader refer to the
EPA Drinking Water Criteria Documents for Nickel and Lead.
2. Data Review .
All toxicity data on a chemical should be evaluated for
criterion or level of concern development. Those studies
representing the best quality and most appropriate data as
discussed previously under appropriate study design should be
selected for defining adverse effects and their level of
occurrence. As previously discussed, adequate human
epidemiologic data should be used in evaluating the adverse
health effects of a chemical whenever available. When adequate
human data are not available, animal data from species most
relevant to humans should be used. In the absence of data on the
"most relevant" species or the inability to identify the most
relevant species, data from the most sensitive animal species
tested, i.e., the species demonstrating an adverse health effect
at the lowest administered dose via a relevant route of exposure,
shall generally be used.
For guidance, adverse health effects are those deleterious
effects which are or may become debilitating, harmful or toxic to
the normal functions of an organism including reproductive and
developmental effects. These do not include such effects as
tissue discoloration without other noted effects, or the
induction of enzymes involved in the metabolism of the substance.
Guidelines for defining the severity of adverse effects have been
suggested by Hartung and Durkin (1985) which proposes a ranking
from slight to severe effects. Distinguishing slight effects
such as reversible enzyme induction and reversible subcellular
change from more' severe effects is critical in distinguishing
33
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between a no observed adverse effect level (NOAEL) and a low-
observed-adverse-effect (LOAEL).
The experimental exposure level representing the highest
dosage level tested at which no-adverse-effects were demonstrated
(NOAEL) shall be. used in the formula for criteria development.
In the absence of such data, the dosage level at which the
lowest-observed adverse-effect-level was demonstrated may be used
in some circumstances for criteria development. .
Preference should be given to studies involving exposure
over a significant portion of the animal's lifespan since this is
anticipated to reflect the most relevant environmental exposure.
An exception to this is where reproductive and/or developmental
effects may be demonstrated to have a lower NOAEL over a shorter
exposure period.. When two or more studies of equal quality and
relevance exist, the geometric means of the NOAEL or LOAEL may be
used.
3. Uncertainty Factors
The choice of appropriate uncertainty and modifying factors
reflects a case-by-case judgement by experts and should account
for each of the applicable areas of uncertainty and any nuances
in the available data that might change the magnitude of any
factor. Several reports describe the underlying basis of
uncertainty factors (Zielhuis et al., 1979; Dourson and Stara,
1983) and research into this area (Calabrese, 1985; Hattis et
al., 1987; Hartley and Ohanian, 1988; Lewis et al., 1990; Dourson
et al., 1992).
The following are examples of where uncertainty exists as a
result of weakness either in the data base or the process which
needs accommodation:
using dose-response information from effects observed at
high doses to predict the adverse health effects that may
occur following exposure to the low levels expected from
human contact with the agent in the environment;
using dose-response information from short-term exposure
studies to predict the effects of long-term exposures, and
vice-versa;'
using dose-response information from animal studies to
predict effects in humans; and
using dose-response information from homogeneous animal
populations or healthy human populations to predict the
effects likely to be observed in the general population
consisting of individuals with a wide range of
sensitivities. (EPA, 1989d)
34
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For this initiative, accommodation for these uncertainties will
be handled in the following process. For further detail in the
selection of these uncertainty factors, please see Appendix A..
a. Intraspecies. uncertainty factor
An uncertainty factor of 10 shall generally be used when
extrapolating from valid experimental results from studies on
prolonged exposure to average healthy humans. This 10-fold
factor is used to protect sensitive members of the human
population.
b. Interspecies. uncertainty factor
An uncertainty factor of 100 shall generally be used when
extrapolating from valid results of long-term studies on
experimental animals when results of studies of human exposure
are not available or are inadequate. In comparison to a, above,
this represents an additional 10-fold uncertainty factor in
extrapolating data from the average animal to the average human.
c. Subchronic to chronic uncertainty factor
An uncertainty factor of up to 1000 shall generally be used
when extrapolating from animal studies for which the exposure
duration is less than chronic (but greater than subchronic, e.g.,
90 days or more in length) or when other significant deficiencies
in study quality are present, and when useful long-term human
data are not available. In comparison to b, above, this
represents an additional uncertainty factor of up to 10-fold for
less than chronic (but greater than subchronic) studies.
d. Less than subchronic duration uncertainty factor
An uncertainty factor of up to 3000 shall generally be used
when extrapolating from animal studies for which the exposure
duration is less than subchronic (<90 days, e.g., 28 days). In
comparison to b,' above, this represents an additional uncertainty
factor of up to 30-fold for less than subchronic studies (<90
days, e.g., 28-day). The level of additional uncertainty applied
for less than chronic exposures depends on the duration of the
study used relative to the lifetime of the experimental animal.
e. LOAEL to NOAEL uncertainty factor
An additional uncertainty factor of between one and ten may
be used when deriving a criterion from a lowest observable
35
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adverse effect level (LOAEL). This uncertainty factor accounts
for the lack of an identifiable no observable adverse effect
level (NOAEL). The level of additional uncertainty applied may
depend upon the severity and the incidence of the observed
adverse effect.
f. Limited database uncertainty factor
An additional uncertainty factor of between one and ten may
be applied when there are limited effects data or incomplete
subacute or chronic toxicity data (e.g.,
reproductive/developmental data). The level of quality and
quantity of the experimental data available as well as structure-
activity relationships may be used to determine the factor
selected.
When deriving an uncertainty factor in developing a Tier I
criterion or Tier II value, the total uncertainty, as calculated
following the guidance of a-f, cited above, shall not exceed
10,000 for Tier I criteria and 30,000 for Tier II values.
D. Exposure Assumptions
When dealing with site specific and individual specific
exposure, it is more accurate to use actual available exposure
information to estimate an individual's specific risk.
Individual behaviors can be assessed and specific activity
information compiled to address quantity, frequency and duration
of exposure. When dealing with such diverse populations of
individuals covering as large an area as the Great Lakes Basin,
extreme ranges of behaviors and activities are likely.
Therefore, deriving default assumptions that can estimate
reasonable exposures which address the vast majority of the Basin
population becomes necessary.
1. Body Weight
National body weight data has been compiled by the National
Center for Health Statistics from a survey conducted from 1976
through 1980 entitled the second National Health and Nutrition
Examination Survey (NHANES II). Approximately 28,000 people aged
6 months to 74 years were surveyed with other 20,000 individuals
actually interviewed and examined. Weighted mean body weights
have been determined from this data. Since body weights change
so rapidly during childhood, it is reasonable to use mean adult
body weight to reflect population body weights when assuming a
long exposure duration. From national survey data, the mean
adult body weight appears to be approximately 72 kg (EPA, 1989).
If NHANES data are separated out by Great Lakes regional data, it
appears that the mean may even be higher for the Great Lakes
36
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Basin population. However, as a matter of convention, 70 kg has
been used for many years in chemical regulatory programs and
still appears appropriate for this initiative.
EPA believes 70 kg is an appropriate body weight because it
represents a reasonable measurement for the entire population.
If a State believes that use of a lower body weight is
appropriate (which yields a more stringent criterion), the State
or Tribe may adopt such an assumption in calculating their
criteria and values under their authority to establish more
stringent requirements pursuant to section 510 of the Act.
As to whether lower body weights should be used to protect
women of childbearing age, children and fetuses, EPA believes
that categorically adopting more conservative body weight
assumptions may not be appropriate. Each chemical must be
addressed separately since some chemicals may be generically
toxic to both adult sexes, while others may be specifically toxic
to one sex more than the other, or children, specifically. It
therefore would not be appropriate to require generally that all
criteria be based on conservative body weight assumptions. In
the case of mercury, however, a fetotoxic chemical, to be
protective of women of child bearing age, EPA has assumed a body
weight of 65 kg (as opposed to 70 kg) which results in a Tier I
mercury criterion of 1.8 ng/L, which is slightly less than the
proposed criterion of 2 ng/L. EPA has set a final Tier I
criterion for mercury at 1.8 ng/L.
37
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Body Weights of Adults (kilograms)
Age
18
25
35
45
55
65
18
25
35
45
55
65
75
75
Mean
73.7
78.7
80.8
81.0
78.8
74.8
78.1
Men
Std. Error
Women
Men and Women
Std. Error
of Mean Mean
' 0.0035
0.0034
0.0040
0.0041
0.0041
0.0051
0.0016
.
Body Weights
Age
3
6
9
12
15
3
6
9
12
15
18
Mean
11.9
17.6
25.3
35.7
50.5
64.9
Boys
Std. Error
60.6
64.2
67.1
67.9
67.9
66.6
65.4
of Mean
0.0032
0.0037
0.0043
0.0044
0.0045
0.0048
0.0017
Mean
67.2
71.5
74.0
74.5
73.4
70.7
71.8
Std. Error
of Mean
...
—
—
—
—
—
_ _ _
(USEPA, 1989a)
of Children (kilograms)
Girls
Std . Error
of Mean Mean
. 0.0016
0.0014
0.0023
0.0038
0.0051
0.0047
11.2
17.1
24.6
36.1
50.7
57.4
of Mean
0.0011
0.0015
0.0024
0.0043
0.0049
0.0042
Mean
11.6
17.4
25.0
36.0
50.6
61.2
Boys and Girls
Std. Error
of Mean
...
—
—
—
—
_ _ _
(USEPA, 1989a)
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2. Duration of Exposure
a. Population Mobility
The default assumptions for mobility is to consider that an
individual remains in the same residence for a "lifetime".
Movement of individuals from individual residences, communities,
or even regions of the country may influence exposure duration to
contaminants from sources such as drinking water and sport caught
fish dramatically. If movement occurs within the same community,
the influence by drinking water may not change. If movement is
still within the region, the influence of contaminated sport fish
may not change. Be that as it may, mobility may lower or
increase exposure duration and intensity.
Based on a survey conducted by the Oxford Development
Corporation, a property management company, the average residence
time for an apartment dweller is estimated to range from 18 to 24
months. A survey conducted by the Bureau of the Census in 1983,
determined that 93% of householders moved into their present home
between 1950 and 1983. Using this information, the following
time of residence ranges have been determined:
Years in Current Home Total % of Householders
0 - 1 7.5
1 - 3 16.9
'3 - 13 40.2
13-18 11.0
18 - 23 7.9
23 - 33 9.5
33 7.0
Based on these statistics, the 50th percentile of
householders living in their current residence is 9.4 years and
the 90th percentile is 29.8 years. This data does not, of
course, indicate how far people move or whether they will
increase or decrease their exposure by moving. Accordingly, it
is only of limited relevance in determining exposure patterns.
b. Life Expectancy
Life Expectancy Statistical data on life expectancy is
gathered annually by the U.S. Department of Commerce. Data
presented by the Bureau of Census for 1985 show that life
expectancy for the total U.S. population is 74.7 years. The
breakdown of this average is as follows:
Male Female Total
white 71.8 78.7 75.3
black and other 67.2 75.2 71.2
black 65.3 73.7 69.5
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total average 71.2 78.2 74.7
(USEPA, 1989a)
Although the average life expectancy now is approximately 75
years, it is probable that over the course of a lifetime there
should be periods of no exposure that add up to at least five
years. Accordingly, the traditional default value for "lifetime"
exposure of 70 years appears adequate for considering chronic
"lifetime" exposure.
3. Incidental Exposure
The suggested 0.01 liter/day adjustment for recreational
exposure is based on an assumption of 123 hours of recreational
exposure equivalent to swimming, and consumption of an average
mouthful (30 mL) of water per hour of such recreational exposure.
Exposure potential, when averaged over a year, equals 0.01
liters/day. Such exposures could result from an average of one
hour swimming per day during the four month warm weather period
starting in mid May and ending in mid September (i.e., 123 days).
EPA has recently estimated a national average frequency of
swimming to be 7 days/year with a 2.6 hour duration (EPA, 1989d).
An earlier EPA publication estimated an average annual frequency
of 9 days/year with a 2 hour duration of exposure. Other total
body contact recreation such as water skiing was also identified
as having approximately 20 million participants with a total
exposure of 260 million hours per year or an average of 14 hours
exposure per participant. Partial body contact was identified as
20% body exposure for fishing and 40% body exposure for boating.
This earlier reference listed 68 million people involved
nationally in boating with an average duration of 1600 million
person hours per. year and 54 million people involved nationally
in fishing with 6600 million person hours duration per year. The
resulting individual participant average exposure duration equals
approximately 24 hours and 122 hours of participation,
respectively (EPA, 1979). If each hour of total body contact
equivalent for bathing, water skiing, boating and fishing were
calculated (18, 14, 10 and 24 hours, respectively), the total
equals 66 hours of average body contact exposure.
Various recreational surveys have been conducted in Michigan
and may serve as- a typical example of Great Lakes Basin activity.
Estimations similar to EPA's for activities per participant and
hours per participation may be calculated from this information.
If we were to assume an individual were to participate in all
activities for the number of days listed from the 1981 Michigan
Travel and Recreation Survey and for the duration of hours per
participation as identified in the 1976 Recreation survey, and
the percentage adjustment made for total body contact exposure
from the older EPA reference, the following calculations may be
made:
40
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Activity Days
per
Participant
Hours per
Participati
on
Body
Contact
Adjustment
Hours of
Exposure
Swimming
Fishing
Power
Boating
Water
Skiing
Sailing
Canoeing
13.3
14.3
24.5 (total)
2.1 (ave.)
3.7 (ave.)
3.2
1.0
0.2
0.4
27.9
10.6
31.4
9.6 1.5
10.4 (total) 3.2
4.8 3.9
1.0
0.4
0.4
TOTAL
14.4
13.3
7.5
105.1
(Wells, 1990)
Given these comparisons of water recreation activities, the
suggested incidental exposure level appears appropriate, given
the variability in individual behavior.
4. Drinking Water
Two liters of water has been the nationwide conventional
estimate of aduljb human's daily water consumption. The 2 liters
of water per day is a historical figure set by the U.S. Army in
determining the amount of water needed for each person in the
field. The National Academy of Sciences (NAS) estimates that
daily water consumption may vary with physical exercise and
fluctuations in temperature and humidity. It is reasonable to
assume those living in a more arid, hot climate will consume
higher levels of water. NAS has calculated the average per
capita water consumption to be 1.64 liters per day. The National
Cancer Institute. (NCI) in a study, also known as the Cantor
(1987) study, also has looked at this issue with an overall tap
water consumption rate of 1.39 liters of water per day as their
study average. The NCI study is of particular interest since
data were compiled from Detroit, Iowa, New Jersey and Connecticut
giving a database of over 3500 respondents with similar weather
conditions to the Great Lakes Basin. The consumption rate of
less than or equal to 1.96 liters per day is equated to the 100%
cumulative frequency level as seen in the following table:
41
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Frequency Distribution of Tap Water Consumption Rates*
Consumption Rate (L/day) Cumulative Frequency
0.80 19.2
0.81-.- 1.12 39.6
1.13 - 1.33 59.7
1.45 - 1.95 79.9
1.96 100.0
*Represents consumption in a "typical" week.(Cantor et al.,
1987)
Other researchers have discovered average levels both higher and
lower than NCI. The Food and Drug Administration's (FDA) Total
Diet Study estimated rates for water and water-based foods for two
groups of adults to be 1.07 and 1.3 liters per day with an average
of 1.2 liters per day. The U.S. Department of Agriculture (USDA)
in the 1977-78 Nationwide Food Consumption Survey identified daily
beverage intakes of from 1.24 to 1.73 liters per day. In a more
recent study specifically characterizing tap water intake by Ershow
and Cantor (1989), 2 liters/day represents approximately the 85th
percentile value of drinking water consumption. After review of
all these studies, EPA has judged the average adult drinking water
consumption rate to be 1.4 liters per day with a reasonably
conservative assumption of 2 liters per day as being the 90th
percentile value (USEPA, 1989a). This compensates in part for
parts of the population, such as manual or migrant laborers, who
drink much more than 2 liters a day.
5. Fish Consumption
Much debate has occurred over the years as to the appropriate
regionally caught fish consumption rate for the Great Lakes Basin.
This is one area where extreme differences exist in the region's
consumption behavior. A large segment of the population consumes
little or no fish caught from the region, while a small segment of
the population c6nsumes a significant quantity of regionally caught
fish.
Several studies of fish consumption and sport angler behavior
have been evaluated to estimate an appropriate fish consumption
value for the region. Three regional surveys; Michigan (West,
1989), Wisconsin (Fiore, 1989) and New York (Connelly, 1990) ; have
been selected for consideration. In summary, the results of the
Michigan survey suggest that approximately 65% of the licensed
anglers consume less than one meal per week of all fish. This is
consistent with' Wisconsin data which estimates the mean annual
total number of all fish meals consumed by anglers to be 41. This
is also consistent with New York anglers who consume 45.2 meals
statewide and approximately 41.6 total meals in the regions with
42
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the greatest number of sport anglers and greatest sport fishing
effort. Based on the Michigan and Wisconsin surveys, approximately
43% of the fish meals consumed are sport caught, or approximately
18-19 meals per year. .Estimates of meal sizes range up to 8 ounces
(0.5 pounds) or an approximate total of 9-9.5 pounds per year. This
equates to a daily fish consumption rate of 11-12 g/day. The
Michigan survey data indicate a mean annual total fish consumption
rate of 17 gm/day or (at 43%) approximately 7 gm of sport caught
fish. There is poor data on the proportion of the nonsport caught
(commercial) fish consumed within the region which is actually
caught within the region. Using the Michigan survey data, at least
22% of the fish consumed are species from outside the region.
Thereby, the maximum proportion of regionally caught and consumed
fish in Michigan may be estimated to be only 78% or 13 grams per
day. All those contacted familiar with commercial fishing within
the region estimated the major amount of regionally caught
commercial fish are sold outside of the region and therefore,
generally not available to regional anglers. If one assumes a
conservative mean total of regionally caught meals to equal 24
meals per year at 8 ounces per meal or up to 48 meals per year at
4 ounces per meal, the mean daily consumption rate is 15 gm/day.
A second study conducted by West et al. (1993) for the State of
Michigan provided results which were very supportive of the use of
15 grams/day. This study is a full year (February 1991 to February
1992) fish consumption survey of 7000 licensed Michigan anglers.
The survey found that the average sport fish consumption rate,
adjusted for non-response bias, was 14.5 grams/day. The average
total fish (all fish, not just Great Lakes sport fish) consumption
rate, adjusted for non-response bias, was 24.4 grams/day. This
study indicated that fish consumption rates may differ according to
race and income level. The lowest income group (< $14,999/year)
averaged 21 grams/day sport fish consumption as compared to 14.7
grams/day for those making $40,000 or more/year. The average sport
fish consumption'rate for minorities was 23.2 grams/day as compared
with 16.3 grams/day for non-minority individuals. Lower income
($24,999 or less) minorities averaged the highest consumption rate
of all groups in the survey: 43.1 grams/day sport caught fish and
57.9 grams/day total fish; Non-minority individuals of lower
income averaged 18.6 grams/day sport fish and 25.8 grams/day total
fish. The study also indicated that minorities eat less fish from
the Great Lakes and more fish from the inland tributaries than non-
minority individuals. For greater detail on the West et al. (1993)
study and a statistical analysis of the west study findings refer
to U.S. EPA (1995).
For this initiative, the assumption of 15 g/day of regionally
caught fish should adequately estimate the consumption rate of the
mean angler population and their families for all sport caught
fish. A much larger segment of the sport angler population is
included if this consumption is attributed totally to species of
fish more susceptible to persistent and bioaccumulative
contaminants, i ..e., the salmonids. Based on the Regional Survey
data, including number of licenses bought and used, members per
43
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family, and fish consumption rates for sport anglers, 15 g/day
approximates at least the 90% consumption level of regionally
caught fish for the regional population as a whole, i.e.,
fisherpersons as. well as nonfisherpersons.
6. Relative Source Contribution
In the final GLWQI Guidance, the Agency assumes an 80 percent
relative source contribution (RSC) from surface water pathways
(water and fish) for all chemicals, bioaccumulative chemicals of
concern (BCCs) and non-BCCs, in deriving noncancer criteria/values.
A 100 percent RSC is assumed for all chemicals in deriving cancer
criteria/values. EPA also recommends that actual data be used in
developing an RSC when available. As stated in the 1980 National
Guidelines, to account for exposures from other sources, actual
exposure data can be subtracted from the RfD (ADI, as it was called
in 1980) to account for contributions of the pollutant from diet
and air (ADI - (DT + IN) where DT is the estimated non-fish dietary
intake and IN is the estimated daily intake by inhalation (U.S.
EPA, 1980). Therefore, where data are available, if States or
Tribes want to use actual data in developing their RSC, they may do
so, following the procedure outlined in the 1980 National
Guidelines. It is important to note, however, that EPA's policy on
how to use exposure data in developing an RSC is now under review.
Once EPA has finalized its policy review on the RSC, EPA will
address the application of the RSC during the triennial review of
Water Quality Standards under section 303 of the Clean Water Act.
Until such time, the Agency has decided to apply an RSC of 80% to
all noncarcinogenic chemicals (both BCCs and non-BCCs).
With regard to using different RCSs for BCCs and non-BCCs which
was presented in the Proposed Human Health TSD, EPA does not
believe there is a clear difference in RSC development for BCCs as
opposed to non-BCCs. While it may be true that surface water may
be the major route of exposure for bioaccumulatives (through fish
consumption), even though a pollutant is not bioaccumulative, it
does not preclude the possibility that there may be other
significant sources of exposure.
With regard to the use of a 80 percent default value, EPA
believes that the assumption helps to provide some measure of
protection against the possibility that exposures from other
sources may contribute to the overall exposure of the public to a
particular contaminant. Available data indicate that non-water
sources contribute varying amounts to overall exposure to a
particular chemical (U.S. EPA 1982, U.S. EPA 1983). Such exposures
can occur through air and the diet. Since available data indicate
that such exposures can and do occur, but these data are often
limited in their ability to predict with precision the relative
source contribution, EPA believes it is prudent not to assume that
all exposure to a pollutant occurs from one medium. The 80
percent default was chosen because it reflects the approximate
contribution from surface water pathways (fish consumption) to the
44
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overall exposure to BCCs such as PCBs in the Basin. For PCBs, the
FDA Total Diet 'Study estimates that consumption of pollutant-
bearing fish represents the most significant exposure. The average
adult's daily intake of PCBs via diet is estimated to be 560 ng,
versus estimated inhalation levels of 100 ng per day. Based on
these estimates, diet contributes approximately 85% of exposure
(ATSDR, 1987). .It appears likely that, for other highly
bioaccumulative chemicals, a similar estimate may be made as well.
For nonbioaccumulatives, 80 percent was also chosen as a default
value to account,for the other possible non-water sources which may
contribute to the overall exposure of the chemical. However,
actual exposure data may also be used in the final Guidance by
States and Tribes to calculate a relative source contribution. EPA
recognizes that the choice of a default value of 80% in these cases
is fundamentally a policy judgment that criteria development should
reflect the fact that exposures to a pollutant occur through other
media, rather than an empirically-based calculation of the precise
proportion of exposure via water versus non-water sources, since
such values vary, on a case-by-case basis. EPA also acknowledges
that use of a 80% default for non-BCCs is a conservative measure,
however, if other significant exposures are not accounted for, the
criteria could underestimate overall exposure to the chemical and
thus could underestimate the risk of adverse health effects. In
addition, in the absence of data, it is prudent and consistent with
the health protection goals of the CWA to include a margin of
safety in the event that there are exposures from other sources.
The important fact, EPA believes, is to take some accounting of
other possible exposure pathways.
With regard to the concern that point sources should not be
expected to compensate for the failure to address other pollutant
sources, EPA does not believe that the relative source contribution
factor in the final methodology unduly burdens point source
dischargers. It is common practice in EPA programs (e.g., in
establishing maximum contaminant level goals under the SDWA) to
take into account other routes of exposure to a chemical when
establishing health-based standards for a particular route of
exposure. If this step is not taken, and EPA were always to assume
that no exposures occurred through other media (in spite of
evidence to the contrary), then the totality of exposures could
obviously result in adverse health effects, contrary to EPA's goal
of establishing standards that insure that such effects do not
occur. EPA agrees, however, that it is important to take steps to
address all routes of exposure to pollutants in order to achieve
the greatest overall public health protection at the least cost.
45
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IV. CRITERIA CALCULATIONS
A. Standard Exposure Assumptions
BW = weight of an average human (BW = 70 kg).
WC = per capita water consumption for surface waters
classified as public water supplies (WC, = 2 liters/day)
-or-
average per capita incidental daily water exposure for
surface waters not classified as drinking water supplies
(WCr =0.01 liters/day)
FC = per capita daily consumption of regionally caught
fish = 0.015 kg/day
BAF = bioaccumulation factor.
B. Carcinogens *
When a linear, nonthreshold dose-response relationship is
assumed, the human cancer value shall be calculated using the
following equation:
HCV = RAD x BW
WC + [(Peru x BAF^) + (PC™ x BAF^) ]
Where:
4
HCV = Human Cancer Value in milligrams per liter (mg/L).
RAD = RAD in milligrams toxicant per kilogram body weight per
day (mg/kg/day) that is associated with a lifetime
incremental cancer risk equal to 1 in 100,000.
BW = Body weight of an average human (BW = 70kg).
WC = average per capita water consumption (both drinking and
incidental exposure) for surface waters classified as
public water supplies (WCd = 2 L/day) and average per
capita incidental daily water exposure for surface waters
not used as public water supplies (WCr =0.01 liters/day)
= mean consumption of trophic level 3 fish by regional
sport fishers = 0.0036 kg/day
= mean consumption of trophic level 4 fish by regional
sport fishers = 0.0114 kg/day
BAFTL3 « BAF for trophic level 3 fish
= BAF for trophic level 4 fish
46
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C. Noncar c inogens
The human noncancer value shall be calculated as follows:
HNV = APE x BW x RSC
WC + [(FCru x BAFTLa) + (FC^ x BAF^) ]
Where :
HNV = HNV in milligrams per liter (mg/L) .
ADE = ADE in milligrams toxicant per kilogram body weight per
day (mg/kg/day) .
RSC = RCS factor of 0.8 for all chemicals of concern. This
is used to allow for potential exposure via sources
other than consumption of contaminated water and fish
recreational exposure. States may develop an RSC using
actual exposure data following the procedures specified
in the 1980 National Guidelines.
An ADE may be derived directly from the following example methods
depending on the type and quality of the toxicity database:
1. a scientifically valid reference dose (RfD) as identified
through best available information sources, such as IRIS; and
2. a scientifically valid acceptable daily intake (ADI) as
identified from the U.S. Food and Drug Administration. Both
sources should be updated with the most recent data available.
3 . a chronic or subchronic NOAEL for humans exposed to the
toxicant via contaminated drinking water as follows:
ADE = NOAEL (ma/1) x WC ,
U x Wh ~
Where : •
U = Uncertainty factor of 10-100 depending on the quality
of the data.
4. a chronic or subacute NOAEL from a mammalian test species
exposed to toxicant contaminated drinking water as follows:
ADE = NOAEL (mg/1) x
wa
U
Where:
Vw = Volume of water consumed per day by test animals
(L/day) .
47
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Wa = Weight of test animal (kg) .
U = Uncertainty factor of 100-1000 depending on quality
of data. An additional uncertainty factor of up to 10
may be used to account for studies of very short term,
e.g., 28 days .
5. a chronic or subacute NOAEL from a mammalian test species
exposed to toxicant -contaminated food as follows:
ADE - NOAEL (mg/kg food) x
U
Where :
f c = Daily food consumption by test animal (kg) .
Wa = Weight of test animal (kg) .
U = Uncertainty factor of 100-1000 depending on quality
of data. An additional uncertainty factor of up to 10
may be used to account for studies of very short term,
i.e., 28 days .
6. a chronic or subacute NOAEL from a mammalian test species
exposed to a toxicant by gavage as follows:
ADE = NOAEL (mq/kq) x Fw
U
Where:
Fw = Fraction of week dose.
U = Uncertainty factor of 100-1000 depending on quality
of data. An additional uncertainty factor of up to 10
may be used to account for studies of very short term,
i.e., 28 days .
7. A chronic or subacute NOAEL from a mammalian test species
exposed to a toxicant by inhalation:
ADE
NOAEL (me
T^)
U
x
X
I X
Wa
fw
x
fd
x
r
Where : 3
I = Inhalation rate for test species (m /day) .
fw = Fraction of week exposed.
fd = Fraction of day exposed.
r = Absorption coefficient.
Wa = Weight of test animal (kg) .
U = Uncertainty factor of 100-1000 depending on quality
of data. An additional uncertainty factor of up to 10 may
be used to account for studies of very short term,
i.e., 28 days .
48
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8. Similar approaches shall be followed when data is limited to
a LOAEL with an appropriate increase in uncertainty factor. For
example, a subacute LOAEL from a mammalian test species exposed
to toxicant contaminated drinking water would be calculated as
follows :
ADE = LOAEL (mg/1) x
wa
U
Where :
Vw = Volume of water consumed per day by test animal
(L/day) .
Wa = Weight of the test animal (kg) .
U = Uncertainty factor of 1000-30,000 depending on
quality of data and severity of effect.
49
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APPENDIX A
UNCERTAINTY FACTORS
A. INTRODUCTION
Uncertainty factors (also called safety factors) are intended
for use in extrapolating toxic responses thought to have a
threshold (i.e., noncarcinogenic effects) . "Uncertainty factor" is
defined as a number that reflects, the degree or amount of
uncertainty that must be considered when experimental data in
animals are extrapolated to man (EPA, 1980). In addition,
uncertainty factors are used when extrapolating from small
populations of humans to the entire heterogeneous human population
and when extrapolating from a single animal species to wildlife
communities. The use of uncertainty factors in extrapolating
animal toxicity data to acceptable exposure levels for humans has
been the cornerstone of regulatory toxicology (National Academy of
Sciences, 1980) .. This appendix will provide the risk assessor with
additional guidelines, rationale and information concerning the
selection of uncertainty factors.
Because of the high degree of judgment involved in the selection
of uncertainty factors, the risk assessment justification should
include a detailed discussion of the selection of the uncertainty
factors along with the data to which they are applied.
This report is organized with the recommended uncertainty
factors listed in Part B for quick reference, and a discussion of
those factors and their support in Part C. Also included in Part
C is a discussion of the exposure duration terms "subacute",
"subchronic", and "chronic".
Al
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B. RECOMMENDED UNCERTAINTY FACTORS
1. An uncertainty factor of 10 shall generally be used when
extrapolating from valid experimental results from studies on
prolonged exposure to average healthy humans. This 10-fold factor
is used to protect sensitive members of the human population.
2. An uncertainty factor of 100 shall generally be used when
extrapolating from valid results of long-term studies on
experimental animals when results of studies of human exposure are
not available or are inadequate. In comparison to 1, above, this
represents an additional 10-fold uncertainty factor in
extrapolating data from the average animal to the average human.
3. An uncertainty factor of up to 1000 shall generally be used
when extrapolating from animal studies for which the exposure
duration is less than chronic (but greater than subchronic, e.g.,
90 days or more in length) or when other significant deficiencies
in study quality are present, and when useful long-term human data
are not available. In comparison to 2, above, this represents an
additional uncertainty factor of up to 10-fold for less than
chronic (but greater than subchronic) studies.
4. An additional uncertainty factor of between one and ten may
be used when deriving a criterion from a lowest observable adverse
effect level (LOAEL). This uncertainty factor accounts for the
lack of an identifiable no observable adverse effect level (NOAEL).
The level of additional uncertainty applied may depend upon the
severity and the incidence of the observed adverse effect.
5. An uncertainty factor of up to 3000 shall generally be used
when extrapolating from animal studies for which the exposure
duration is less than subchronic (<90 days, e.g., 28 days). In
comparison to 2, above, this represents an additional uncertainty
factor of up to 3*0-fold for less than subchronic studies (<90 days,
e.g., 28-day). The level of additional uncertainty applied for
less than chronic exposures depends on the duration of the study
used relative to the lifetime of the experimental animal.
6. An additional uncertainty factor of between one and ten may
be applied when there are limited effects data or incomplete
subacute or chronic toxicity data (e.g., reproductive/developmental
data). The level of quality and quantity of the experimental data
available as well as structure-activity relationships may be used
to determine the" factor selected.
When deriving an uncertainty factor in developing a Tier I
criterion or Tier II value, the total uncertainty, as calculated
following the guidance of 1-6, cited above, shall not exceed 10,000
for Tier I criteria and 30,000 for Tier II values.
The following discussion is generalized for categories of
commonly applied uncertainty factors which are used in the
developing the specific uncertainty factors used in GLWQI,
described above..
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C. DISCUSSION
Dourson and Stara (1983) reviewed available literature on
uncertainty factors which are used to estimate acceptable daily
intakes (ADIs) for toxicants. They found that the use and choice
of these factors is supported by reasonable qualitative biological
premises and specific biological data. Therefore, in the absence
of adequate chemical-specific data, uncertainty factors for
criteria derivation may be selected according to reasonable
assumptions and .approximations rather than total arbitrariness.
They presented a set of guidelines for the use of uncertainty
factors based on those utilized by the FDA, WHO, NAS, and EPA,
indicating consistency and widespread acceptance among the
scientific community. Those guidelines have been adapted herein
for use in risk assessment under the Great Lakes Initiative. Their
rationale and experimental support are discussed below. The
guidelines should not be misconstrued as being unalterable and
inflexible. They are intended to help ensure appropriateness and
consistency of -risk assessments. They should be regarded as
general recommendations, with the realization that the data for a
particular chemical may be such that a different uncertainty factor
would be more appropriate.
A 10-fold factor is recommended when extrapolating from valid
experimental results from human studies of prolonged exposure.
People of all ages, states of health, and genetic predispositions
may be exposed to environmental contaminants. The 10-fold factor
is intended to offer protection for the sensitive subpopulations
(the very young, the aged, medically indigent, genetically
predisposed, etc.), since the observed no-effect level is generally
based on average healthy individuals. Experimental support for
this 10-fold factor is provided by log-probit analysis and the
study of composite human sensitivity (Dourson and Stara, 1983).
However, Calabrese (1985) has presented data on human
variability in several physiological parameters and in
susceptibility to several diseases, and concluded that human
variation may range up to two or three orders of magnitude. While
human variation in the metabolism of various xenobiotics may have
a 1000-fold range, Calabrese (1985) noted that the vast majority of
the responses addressed fell clearly within a factor of 10.
Another study on key human pharmacokinetic parameters indicates
that the 10-fold factor to encompass human variability may only
capture the variability among normal healthy adult humans. That
report recommends further study to determine the degree of
additional susceptibility among sensitive subpopulations (EPA,
1986).
Given the heterogeneous and highly outbred state of the human
population, and the multifactorial nature of disease
susceptibility, reliance on the adequacy of the 10-fold factor for
extrapolation to "safe" levels appears somewhat precarious. But
because of its history of use and current widespread acceptance,
this factor may continue to be used until the availability of new
data indicating quantitatively a more acceptable factor.
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A 100-fold factor is recommended when extrapolating from valid
results of long-term studies on experimental animals with
results of studies of human exposure not available or scanty
(e.g.. acute exposure only). This represents the 10-fold factor for
intraspecies extrapolation (see C.I) and an additional 10-fold
uncertainty factor for extrapolating data from the average animal
to the average man.
The 100-fold uncertainty factor has been justified for use with
the risk extrapolation for food additives. That justification has
been based on differences in body size, differences in food
requirements varying with age, sex, muscular expenditure, and
environmental conditions within a species, differences in water
balance of exchange between the body and its environment among
species, and differences among species in susceptibility to the
toxic effect of a given contaminant (Bigwood, 1973). The use of
the 100-fold uncertainty factor has also been substantiated by
citing differences in susceptibility between animals and humans to
toxicants, variations in sensitivities in the human population, the
fact that the number of animals tested is small compared with the
size of the human population that may be exposed, the difficulty in
estimating human intake, and the possibility of synergistic action
among chemicals (Vettorazzi, 1976).
On a dose per unit of body weight basis, large animals (e.g.,
man) are generally more sensitive to toxic effects than small
animals (e.g., rats, mice). This principle is attributed to the
relationship between animal size and pharmacokinetics, whereby the
tissues of a large animal are exposed to a substance (mg/kg dose)
for a much longer time than the tissues of a small animal. This
principle has been demonstrated experimentally. The
pharmacokinetic processes underlying this phenomenon include: in
general, large animals metabolize compounds more slowly than do
small animals; large animals have many more susceptible cells; in
large animals, substances are distributed more slowly and tend to
persist longer; the blood volume circulates much more rapidly in
small animals. Thus, for the same mg/kg dose, human tissues are
exposed to a substance for a much longer time than rodent tissues
(National Academy of Sciences, 1977).
Experimental support for the additional 10-fold uncertainty
factor when extrapolating from animal data to humans is provided by
studies on body-surface area dose equivalence and toxicity
comparisons between humans and different animal species (Dourson
and Stara, 1983). On a dose per unit of body-surface area basis,
the effects seen in man are generally in the same range as those
seen in experimental animals. An interspecies adjustment factor
accounts for differences in mg per kg body weight doses due to
different body-surface areas between experimental animals and man.
The factor may be calculated by dividing the average weight of a
human (70 kg) by the weight of the test species (in kg) and taking
the cube root of this value. Thus on a body weight basis, man is
assumed to be more sensitive than the experimental animals by
factors of approximately 5 and 13 for rats and mice, respectively.
For most experimental animal species (i..e., all species larger than
A4
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mice) , the 10-'fold decrease in dose therefore appears to
incorporate a margin of safety. For mice, the interspecies
adjustment factor suggests that the additional 10-fold uncertainty
factor for interspecies extrapolation to humans is not large enough
(Dourson and Stara, 1983). Nevertheless, the additional 10-fold
factor is considered adequate to adjust from mice to humans when
chemical-specific data are not available.
A factor of up to 1000 is recommended when extrapolating from
animal studies for which the exposure duration is less than chronic
(i.e..less than'50% of the lifespan) or when other significant
deficiencies in study quality are present, with no useful long-term
or acute human data. This represents the 10-fold factors for
intraspecies and interspecies extrapolation (see C.2), and an
additional uncertainty factor of up to 10-fold for extrapolating
from less than chronic to chronic animal exposures (or when the
data are significantly flawed in some other way). Injury from
chronic exposure may occur in at least three ways: by accumulation
of the chemical.to a critical concentration at sites of action
sufficient to induce detectable injury; by accumulation of injury
until physiological reserves can no longer compensate (i.e., repair
is never complete); or after a long, latent period beginning with
an exposure that has an unrecognized biological effect and
precipitates the eventual appearance of injury (National Academy of
Sciences, 1977). Obviously, sufficient duration of exposure is
necessary in order for the effects seen in chronic toxicity to
become manifest. Subchronic toxicology studies may not offer
reliable means .for assessment of long-term toxic effects in
animals, let along extrapolation to chronic effects in man
(National Academy of Sciences, 1977). However, it is often the
case that a good quality, chronic exposure study for a particular
chemical is unavailable. The intention of this additional
uncertainty factor is to enable the use of subchronic or flawed
studies to protect against the risk of adverse effects which might
only appear with chronic dosing.
Experimental support for the additional uncertainty factor is
given by literature reviews which compare subchronic NOAELs and
chronic NOAELs for many compounds (McNamara, 1971; Weil and
McCollister, 1963). The studies reviewed by those investigators
employed a variety of rodent and non-rodent species. The duration
of the subchronic exposures was usually 90 days, but ranged from 30
to 210 days. Wide variations in endpoints and criteria for adverse
effects were encountered in these literature reviews. However,
their findings do give a rough indication of the general subchronic
and chronic NOAELs for other than carcinogenic or reproductive
effects. For over 50% of the compounds tested, the chronic NOAEL
was less than the 90-day NOAEL by a factor of 2 or less. There was
some indication that chronic dosing may result in the development
of tolerance toward certain chemicals, as the chronic NOAEL was
larger than the 90-day NOAEL in a few cases. However, it was also
found that the chronic NOAEL may be less than the 90-day NOAEL by
a factor of 10 or more. The latter situation appeared to be
uncommon. Therefore, these reviews report that the additional ID-
AS
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fold uncertainty factor appears to be adequate or incorporate a
margin of safety in the majority of cases.
As the literature reviews by McNamara (1971) and Weil .and
McCollister (1963) are limited and the studies reviewed utilized a
variety of toxicologic endpoints with questionable sensitivities,
one must be cautious in interpreting their conclusions. But for
lack of data to -the contrary, it appears that application of the
additional 10-fold uncertainty factor is appropriate and justified
when extrapolating a NOAEL from a 90-day study to a chronic NOAEL
estimate. This practice may underestimate the true chronic NOAEL
far more often than overestimating it, thus adding a margin of
safety to the risk calculations.
One remaining question regarding exposure duration is: At what
point is the duration considered adequate, such that the additional
uncertainty factor of up to 10 is unnecessary? In other words, how
is "chronic" defined for the sake of this guideline?
At this point, further discussion of the terms "chronic"
"subchronic", and "subacute", is necessary. The term "subacute"
has been used to describe a duration less than subchronic, while it
has also been used as a term analogous to subchronic. EPA (1980)
describes "subacute" exposures (in this case, analogously to
"subchronic") as often exceeding 10% of the lifespan, e.g., 90 days
for the rat with an average lifespan of 30 months. However, as
pointed out by the Organization for Economic Cooperation and
Development (OECD, 1981), the term "subacute" is semantically
incorrect. The OECD prefers to use the phrase "short-term repeated
dose studies", referring to 14, 21 and 28 day studies, to
distinguish from "subchronic" studies of greater duration.
"Subchronic" is generally defined as part of the lifespan of the
test species, although opinions differ on the precise definition.
Klaassen (1986) defines "subacute" as repeated exposure to a
chemical for one month or less, and "subchronic" as repeated
exposure for 1-3 months. Chan et al. (1982) describe "subchronic"
exposure durations as generally ranging from 1 to 3 months in
rodents and one year in longer-lived animals (dogs, monkeys), or
for part (not exceeding 10%) of the lifespan. Stevens and Gallo
(1982) define "long-term toxicity tests" (encompassing subchronic
and chronic toxicity studies) as studies of longer than 3 months
duration, i.e., greater than 10% of the lifespan in the laboratory
rat. EPA (1985) describes "subchronic" toxicity testing as
involving continuous or repeated exposure for a period of 90 days,
or approximately. 10% of the lifespan for rats.
The various definitions offered for "chronic" are generally
inconsistent. Klaassen (1986) defines "chronic" as repeated
exposure for more than 3 months. According to the National Academy
of Sciences (1977), chronic exposure in animals is generally
considered to be at least half the life span. In estimating
chronic SNARLs, the National Academy of Sciences (1980) in most
cases utilized data from studies lasting a "major portion of the
lifetime of the experimental animal". According to the EPA's
Health Effects Testing Guidelines (EPA, 1985), chronic toxicity
tests should involve dosing over a period of at least 12 months.
A6
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The application of their guidelines, they add, should generate data
on which to identify the majority of chronic effects and shall
serve to define long-term dose-response relationships. The OECD
(1981) states that the division between subchronic and chronic
dosing regimes is sometimes taken as 10% of the test animal's life
span. They also state that the duration of the exposure period for
chronic toxicity studies should be at least 12 months. They
describe "chronic" as prolonged and repeated exposure capable of
identifying the majority of chronic effects and to determine dose-
response relationships.
Others have investigated the delayed appearance of toxic effects
which might be missed under shorter dosing regimes. Frederick
(1986) conducted a pilot survey of new drug evaluators for
incidences of delayed (greater than 12 month) drug-induced
pathology. It was concluded that new toxic effects "not
infrequently" arise after one year of dosing in rodents. It was
further stated that those findings formed the basis for the
conclusion of the Bureau of Human Prescription Drugs: the duration
of the long-term-toxicity tests of drugs that are likely to be used
in man for more than a few days should be at least 18 months.
Glocklin (1986) reviewed the issues regarding testing requirements
for new drugs, and concluded that 12 month chronic toxicity studies
seemed to be an appropriate requirement for characterization of the
dose-response.
It is evident that there are discrepancies in the qualitative
and quantitative characterization of "chronic" animals studies. An
appropriate and reasonable working definition for "chronic" would
appear to be at 'least half the life span (therefore, at least 52
weeks for rats and at least 45 weeks for mice) . Qualitatively,
"chronic" means that the exposure duration was sufficient to
represent a full lifetime exposure, in terms of dose-response
relationships. For example, a study providing an experimental
NOAEL which approximates a lifetime NOAEL is considered a chronic
study. It is recognized that the above quantitative definition (at
least half the life span) does not demonstrate the flexibility
inherent in the above qualitative description. That flexibility
reflects the vast differences in the toxicology of various
chemicals: demonstration of a lifetime NOAEL for some chemicals
may require dosing for half the life span, while the toxicology of
most chemicals may allow demonstration of a lifetime NOAEL under a
much shorter dosing regime. It may be argued that the lifetime
NOAEL for noncarcinogenic effects of many chemicals can be
demonstrated in rodent studies of much less than one year. While
the previously-discussed works of McNamara (1971) and Weil and
McCollister (1963) support that view, they also demonstrate that
the chronic NOAEL may be less than the 90-day NOAEL by a factor of
10 or more, for some chemicals.
This discussion is necessary in order to properly interpret the
uncertainty factor guideline, which recommends that the additional
uncertainty factor of up to 10 be applied when the exposure
duration is less than "chronic". The intent of the uncertainty
factor is to adjust the experimental NOAEL to a lifetime NOAEL in
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those cases where the lifetime NOAEL was presumably not adequately
demonstrated. The key issues are summarized in the following
points and recommendations:
a. An acceptable quantitative definition of "chronic" is
elusive. Due to differing toxicological properties, .the
'necessary minimum exposure duration to demonstrate a
lifetime NOAEL differs widely among .chemicals. A
qualitative, philosophical definition of chronic is:
"Chronic" is when the exposure duration is sufficient for
the identification of the majority of long-term effects and
their dose-response relationships. Therefore, a "chronic"
study reporting a NOAEL is one which can be reasonably
presumed to predict the lifetime NOAEL.
b. The use of scientific judgment is predominant in the
decision of when chronic exposure conditions exist, and
hence, when the additional uncertainty factor is no longer
appropriate.
c. That scientific judgment should be guided by a review of all
available pertinent data, e.g., metabolism,
pharmacokinetics, bioaccumulation, mechanism of action,
target organ characteristics, potential for latent effects,
etc.
d. Available'reviews of rodent studies indicate that, for many
chemicals, studies of much less than one year duration can
provide reasonable estimates of lifetime NOAELs. However,
it is also recognized that the toxicological characteristics
of some • chemicals will prevent the qualitative and
quantitative demonstration of latent adverse effects and a
lifetime NOAEL if the duration is less than one year. If
the lack of additional data prevents scientific judgment in
these cases, 50% of the lifespan (52 weeks for rats; 45
weeks for mice) may be considered the minimum necessary
duration for a "chronic" exposure. Application of the
additional uncertainty factor for these apparently
"subchronic" studies may later provide to be excessively
conservative in some cases. But, if the toxicologic
database is inadequate, the additional uncertainty factor
should be included, both as a matter of prudent public
policy and as an incentive to others to generate the
appropriate data.
e. Ordinarily, the additional 10-fold factor may be applied for
all rodent studies of 90 days duration, unless there is
chemical-specific data indicating that would be unnecessary
and overly conservative.
f. For rodent studies of between 90 days and 12 months
duration, the use of the additional 10-fold uncertainty
A3
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factor is best determined by professional judgment. As
described above, if data are not available to sufficiently
guide professional judgment, then such studies may. be
subject to part.or all of the additional 10-fold factor. A
"sliding scale" or between 1 and 10 is a reasonable means of
selecting a lesser factor when 10 appears excessive. Under
this concept, the additional uncertainty factor applied may
vary on a scale of one to ten according to how closely the
dosing duration approached 50% of the lifespan. Of course,
consideration must be given of the study quality and the
other pertinent data mentioned in 3.c above. A 90-day
rodent study would be subject to a 10-fold additional
factor, if study quality is otherwise nominal and other
chemical-specific data are lacking. A nominal-quality
study, with exposure over 50% of the lifespan, would be
subject to a "1", i.e., no additional adjustment.
Situations where the exposure duration is between 90 days
and 50% of the lifespan, and/or study quality is flawed,
must be handled on a case-by-case basis. This "sliding
scale" concept may offer guidance to the scientific judgment
that will be necessary.
Dosing duration is but one parameter upon which to assess the
adequacy of a study. Other deficiencies in the study design may
cause increased concern about the validity of the reported NOAEL or
LOAEL. Therefore, risk assessors may utilize part or all of this
additional 10-fold uncertainty factor to compensate for data which
appears less-than-adequate. Factors which may affect the degree of
confidence in the data include the number of animals per dose
group, the sensitivity and appropriateness of the endpoints, the
quality of the control group, the exposure route, the dosing
schedule, the age and sex of the exposed animals, and the
appropriateness of the surrogate species tested, among others.
EPA's Health Effects Testing Guidelines (EPA, 1985) provide
specific information on the desirable qualities of subchronic and
chronic toxicity' tests.
An additional uncertainty factor of between 1 and 10 is recommended
depending on the severity and sensitivity of the adverse effect
when extrapolating from a LOAEL rather than a NOAEL. This
uncertainty factor reduces the LOAEL into the range of a NOAEL,
according to comparisons of LOAELs and NOAELs for specific
chemicals. There is evidence available which indicates, for a
select set of chemicals, 96% have LOAEL/NOAEL ratios of 5 or less,
and that all are'10 or less (Dourson and Stara, 1983) . In practice
the value for this variable uncertainty factor has been chosen by
the U.S. EPA from values among 1 through 10 based on the severity
and sensitivity of the adverse effect of the LOAEL. For example,
if the LOAEL represents liver cell necrosis, a higher value is
suggested for this uncertainty factor (perhaps 10). If the LOAEL
is fatty infiltration of the liver (less severe than liver cell
necrosis), then a lower value is suggested (perhaps 3; see the
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following discussion). The hypothesized NOAEL should be closer to
the LOAEL showing less severe effects (Dourson and Stara, 1983) .
In some cases the data do not completely fulfill the conditions
for one category of .the above guidelines, and appear to be
intermediate between two categories. Although one order of
magnitude is generally the smallest unit of accuracy that is
reasonable for uncertainty factors, an intermediate value may be
used if felt necessary (Dourson, 1987). According to EPA (1980),
such an intermediate uncertainty factor may be developed based on
a logarithmic scale rather than a linear scale. Calculating the
mean logarithmically may be the more appropriate option, because
the precision of all uncertainty factor estimates is poor, and a
logarithmic scale is the best way to estimate the mean of two
imprecise estimates (Dourson, 1987). Halfway between l and 10 is
approximately 3.16 on a logarithmic scale. However, so as not to
imply excessive accuracy in the estimate, that mean value should be
rounded-off to 3 (Dourson, 1987).
An additional uncertainty factor of up to 10 may be applied when
there are limited or incomplete subacute or chronic toxicity data.
such as with short-term repeated dose animal studies where the
exposure regime involves a limited period that is markedly short-
term relative to the lifespan of the test species (e.g.. 28-
day rodent NOAEL). As previously noted (see C.3) the OECD (1981)
distinguishes between 14, 21 or 28 day studies and "subchronic"
studies of greater duration, by referring to the former as "short-
term repeated dose studies". The short-term studies are commonly
conducted by the NTP to enable appropriate dose selection in
subchronic studies (NCI, 1976). When a limited database exists,
short-term animal studies of 28 days or longer may be of sufficient
quality to support risk assessment of potential chronic exposure.
Because the duration of exposure is substantially less than the 90-
day period discussed under C.3, the risk assessment may require an
additional uncertainty factor in conjunction with the 1000-fold
factor recommended under C.3. As guidance, an additional factor of
up to 10 is recommended when extrapolating from a short-term NOAEL
(e.g., 28 days) to subchronic duration (e.g., 90 days).
Although the. extrapolation from oral LDcns to chronic oral
NOAELs has been reported by several investigators (Venman and
Flaga, 1985; Layton et al., 1987; McNamara, 1971), there has been
relatively little investigation of the extrapolation from short-
term NOAELs (much less than 90 days in rodents) to chronic NOAELs.
EPA (1989) states that when experimental data are available only
for shorter durations than desired for subchronic RfD derivation an
additional uncertainty factor is applied. However, further details
on the selection of an adequate and appropriate uncertainty factor
for those "shorter durations" are not provided. Weil et al. (1969)
evaluated the relationship between 7-day, 90-day and 2-year minimum
effect levels (MiE) for 20 materials via feed exposure. They found
that the median value for a 90-day MiE was obtained by dividing the
7-day MiE by a factor of 3. The 95th percentile for the 90-day MiE
was obtained by dividing the 7-day MiE by 6.2. Also noteworthy is
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the finding that the 95th percentile for the 2-year MiE was
obtained by dividing the 7-day MiE by a factor of 35.3.
These data, Albeit limited, support the general principle that
as exposure duration, decreases, the ability of the data to
demonstrate chronic dose-response relationships also decreases.
While an additional 10-fold uncertainty factor may reasonably and
appropriately convert a 90-day NOAEL to a surrogate chronic NOAEL,
an additional uncertainty factor may be , necessary when
extrapolating from short-term exposures. Applying an additional
uncertainty factor of up to 10 will help ensure that the risk
assessment for potential chronic exposures is adequately
conservative, i.e., the true chronic NOAEL will generally not be
overestimated.
All
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