[Federal Register Volume 61, Number 212 (Thursday, October 31, 1996)]
[Notices]
[Pages 56274-56322]
From the Federal Register Online via the Government Publishing Office [www.gpo.gov]
[FR Doc No: 96-27473]
[[Page 56273]]
_______________________________________________________________________
Part II
Environmental Protection Agency
_______________________________________________________________________
Reproductive Toxicity Risk Assessment Guidelines; Notice
Federal Register / Vol. 61, No. 212 / Thursday, October 31, 1996 /
Notices
[[Page 56274]]
ENVIRONMENTAL PROTECTION AGENCY
[FRL-5630-6]
Guidelines for Reproductive Toxicity Risk Assessment
AGENCY: U.S. Environmental Protection Agency.
ACTION: Notice of availability of final Guidelines for Reproductive
Toxicity Risk Assessment.
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SUMMARY: The U.S. Environmental Protection Agency (EPA) is today
publishing in final form a document entitled Guidelines for
Reproductive Toxicity Risk Assessment (hereafter ``Guidelines''). These
Guidelines were developed as part of an interoffice guidelines
development program by a Technical Panel of the Risk Assessment Forum.
They were proposed initially in 1988 as separate guidelines for the
female and male reproductive systems. Subsequently, based upon the
public comments and Science Advisory Board (SAB) recommendations,
changes made included combining those two guidelines, integrating the
hazard identification and dose-response sections, assuming as a default
that an agent for which sufficient data were available on only one sex
may also affect reproductive function in the other sex, expansion of
the section on interpretation of female endpoints, and consideration of
the benchmark dose approach for quantitative risk assessment. These
Guidelines were made available again for public comment and SAB review
in 1994. This notice describes the scientific basis for concern about
exposure to agents that cause reproductive toxicity, outlines the
general process for assessing potential risk to humans from exposure to
environmental agents, and addresses Science Advisory Board and public
comments on the 1994 Proposed Guidelines for Reproductive Toxicity Risk
Assessment. Subsequent reviews have included the Agency's Risk
Assessment Forum and interagency comment by members of subcommittees of
the Committee on the Environment and Natural Resources of the Office of
Science and Technology Policy. The EPA appreciates the efforts of all
participants in the process and has tried to address their
recommendations in these Guidelines.
EFFECTIVE DATE: The Guidelines will be effective October 31, 1996.
ADDRESSES: The Guidelines will be made available in the following ways:
(1) The electronic version will be accessible on EPA's Office of
Research and Development home page on the Internet at http://
www.epa.gov/ORD/WebPubs/repro/.
(2) 3\1/2\-inch high-density computer diskettes in WordPerfect 5.1
will be available from ORD Publications, Technology Transfer and
Support Division, National Risk Management Research Laboratory,
Cincinnati, OH; telephone: 513-569-7562; fax: 513-569-7566. Please
provide the EPA No. (EPA/630/R-96/009a) when ordering.
(3) This notice contains the full document. In addition, copies of
the Guidelines will be available for inspection at EPA headquarters in
the Air and Radiation Docket and Information Center and in EPA
headquarters and regional libraries. The Guidelines also will be made
available through the U.S. Government Depository Library program and
for purchase from the National Technical Information Service (NTIS),
Springfield, VA; telephone: 703-487-4650; fax: 703-321-8547. Please
provide the NTIS PB No. (PB97-100093) when ordering.
FOR FURTHER INFORMATION CONTACT: Dr. Eric D. Clegg, National Center for
Environmental Assessment--Washington Office (8623), U.S. Environmental
Protection Agency, 401 M Street, S.W., Washington, DC 20460; telephone:
202-260-8914; e-mail: clegg.eric@epamail.epa.gov.
SUPPLEMENTARY INFORMATION:
A. Application of the Guidelines
The EPA is authorized by numerous statutes, including the Toxic
Substances Control Act (TSCA), the Federal Insecticide, Fungicide, and
Rodenticide Act (FIFRA), the Clean Air Act, the Safe Drinking Water
Act, and the Clean Water Act, to regulate environmental agents that
have the potential to adversely affect human health, including the
reproductive system. These statutes are implemented through offices
within the Agency. The Office of Pesticide Programs and the Office of
Pollution Prevention and Toxics within the Agency have issued testing
guidelines (U.S. EPA, 1982, 1985b, 1996a) that provide protocols
designed to determine the potential of a test substance to produce
reproductive (including developmental) toxicity in laboratory animals.
Proposed revisions to these testing guidelines are in the final stages
of completion (U.S. EPA, 1996a). The Organization for Economic
Cooperation and Development (OECD) also has issued testing guidelines
(which are under revision) for reproduction studies (OECD, 1993b).
These Guidelines apply within the framework of policies provided by
applicable EPA statutes and do not alter such policies. They do not
imply that one kind of data or another is prerequisite for action
concerning any agent. The Guidelines are not intended, nor can they be
relied upon, to create any rights enforceable by any party in
litigation with the United States. This document is not a regulation
and is not intended to substitute for EPA regulations. These Guidelines
set forth current scientific thinking and approaches for conducting
reproductive toxicity risk assessments. EPA will revisit these
Guidelines as experience and scientific consensus evolve.
The procedures outlined here in the Guidelines provide guidance for
interpreting, analyzing, and using the data from studies that follow
the above testing guidelines (U.S. EPA 1982, 1985b, 1996a). In
addition, the Guidelines provide information for interpretation of
other studies and endpoints (e.g., evaluations of epidemiologic data,
measures of sperm production, reproductive endocrine system function,
sexual behavior, female reproductive cycle normality) that have not
been required routinely, but may be required in the future or may be
encountered in reviews of data on particular agents. The Guidelines
will promote consistency in the Agency's assessment of toxic effects on
the male and female reproductive systems, including outcomes of
pregnancy and lactation, and inform others of approaches that the
Agency will use in assessing those risks. More specific guidance on
developmental effects is provided by the Guidelines for Developmental
Toxicity Risk Assessment (U.S. EPA, 1991). Other health effects
guidance is provided by the Guidelines for Carcinogen Risk Assessment
(U.S. EPA, 1986a, 1996b), the Guidelines for Mutagenicity Risk
Assessment (U.S. EPA, 1986c), and the Proposed Guidelines for
Neurotoxicity Risk Assessment (U.S. EPA, 1995a). These Guidelines and
the four cited above are complementary.
The Agency has sponsored or participated in several conferences
that addressed issues related to evaluations of reproductive toxicity
data which provide some of the scientific bases for these risk
assessment guidelines. Numerous publications from these and other
efforts are available which provide background for these Guidelines
(U.S. EPA, 1982, 1985b, 1995b; Galbraith et al., 1983; OECD, 1983; U.S.
Congress, 1985, 1988; Kimmel, C.A. et al., 1986; Francis and Kimmel,
1988; Burger et al., 1989; Sheehan et al., 1989; Seed et al., 1996).
Also, numerous resources provide background information on the
[[Page 56275]]
physiology, biochemistry, and toxicology of the male and female
reproductive systems (Lamb and Foster, 1988; Working, 1989; Russell et
al., 1990; Atterwill and Flack, 1992; Scialli and Clegg, 1992; Chapin
and Heindel, 1993; Heindel and Chapin, 1993; Paul, 1993; Manson and
Kang, 1994; Zenick et al., 1994; Kimmel, G.L. et al., 1995; Witorsch,
1995). A comprehensive text on reproductive biology also has been
published (Knobil et al., 1994).
B. Environmental Agents and Reproductive Toxicity
Disorders of reproduction and hazards to reproductive health have
become prominent public health issues. A variety of factors are
associated with reproductive system disorders, including nutrition,
environment, socioeconomic status, lifestyle, and stress. Disorders of
reproduction in humans include but are not limited to reduced
fertility, impotence, menstrual disorders, spontaneous abortion, low
birth weight and other developmental (including heritable) defects,
premature reproductive senescence, and various genetic diseases
affecting the reproductive system and offspring.
The prevalence of infertility, which is defined clinically as the
failure to conceive after one year of unprotected intercourse, is
difficult to estimate. National surveys have been conducted to obtain
demographic information about infertility in the United States (Mosher
and Pratt, 1990). In their 1988 survey, an estimated 4.9 million women
ages 15-44 (8.4%) had impaired fertility. The proportion of married
couples that was infertile, from all causes, was 7.9%.
Carlsen et al. (1992) have reported from a meta analysis that human
sperm concentration has declined from 113 x 10\6\ per mL of semen prior
to 1960 to 66 x 10\6\ per mL subsequently. When combined with a
reported decline in semen volume from 3.4 mL to 2.75 mL, that suggests
a decline in total number of sperm of approximately 50%. Increased
incidence of human male hypospadias, cryptorchidism, and testicular
cancer have also been reported over the last 50 years (Giwercman et
al., 1993). Several other retrospective studies that examined semen
characteristics from semen donors have obtained conflicting results
(Auger et al., 1995; Bujan et al., 1996; Fisch et al., 1996; Ginsburg
et al., 1994; Irvine et al., 1996; Paulsen et al., 1996; Van Waeleghem
et al., 1996; Vierula et al., 1996). While concerns exist about the
validity of some of those conclusions, the data indicating an increase
in human testicular cancer, as well as possible occurrence of other
plausibly related effects such as reduced sperm production,
hypospadias, and cryptorchidism, suggest that an adverse effect may
have occurred. However, there is no definitive evidence that such
adverse human health effects have been caused by environmental
chemicals.
Endometriosis is a painful reproductive and immunologic disease in
women that is characterized by aberrant location of uterine endometrial
cells, often leading to infertility. It affects approximately five
million women in the United States between 15 and 45 years of age. Very
limited research has suggested a link between dioxin exposure and
development of endometriosis in rhesus monkeys (Rier et al., 1993).
Gerhard and Runnebaum (1992) reported an association in women between
occurrence of endometriosis and elevated blood PCB levels, while a
subsequent small clinical study found no significant correlations
between disease severity in women and serum levels of halogenated
aromatic hydrocarbons (Boyd et al., 1995).
Even though not all infertile couples seek treatment, and
infertility is not the only adverse reproductive effect, it is
estimated that in 1986, Americans spent about $1 billion on medical
care to treat infertility alone (U.S. Congress, 1988). With the
increased use of assisted reproduction techniques in the last 10 years,
that amount has increased substantially.
Disorders of the male or female reproductive system may also be
manifested as adverse outcomes of pregnancy. For example, it has been
estimated that approximately 50% of human conceptuses fail to reach
term (Hertig, 1967; Kline et al., 1989). Methods that detect pregnancy
as early as eight days after conception have shown that 32%-34% of
postimplantation pregnancies end in embryonic or fetal loss (Wilcox et
al., 1988; Zinaman et al., 1996). Approximately 3% of newborn children
have one or more significant congenital malformations at birth, and by
the end of the first post-natal year, about 3% more are recognized to
have serious developmental defects (Shepard, 1986). Of these, it is
estimated that 20% are of known genetic transmission, 10% are
attributable to known environmental factors, and the remaining 70%
result from unknown causes (Wilson, 1977). Also, approximately 7.4% of
children have low birth weight (i.e., below 2.5 kg) (Selevan, 1981).
A variety of developmental alterations may be detected after either
pre- or postnatal exposure. Several of these are discussed in the
Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA, 1991),
and developmental neurotoxicity is discussed in the Proposed Guidelines
for Neurotoxicity Risk Assessment (U.S. EPA, 1996a). Relative to
developmental reproductive alterations, chemical or physical agents can
affect the female and male reproductive systems at any time in the life
cycle, including susceptible periods in development. The reproductive
system begins to form early in gestation, but structural and functional
maturation is not completed until puberty. Exposure to toxicants early
in development can lead to alterations that may affect reproductive
function or performance well after the time of initial exposure.
Examples include the actions of estrogens, anti-androgens or dioxin in
interfering with male sexual differentiation (Gill et al., 1979; Gray
et al., 1994, 1995; Giusti et al., 1995; Gray and Ostby, 1995). Adverse
effects such as reduced fertility in offspring may appear as delayed
consequences of in utero exposure to toxicants. Effects of toxic agents
on other parameters such as sexual behavior, reproductive cycle
normality, or gonadal function can also alter fertility (Chapman, 1983;
Dixon and Hall, 1984; Schrag and Dixon, 1985b; U.S. Congress, 1985).
For example, developmental exposure to environmental compounds that
possess steroidogenic (Mattison, 1985) or antisteroidogenic (Schardein,
1993) activity affect the onset of puberty and reproductive function in
adulthood.
Numerous agents have been shown to cause reproductive toxicity in
adult male and female laboratory animals and in humans (Mattison, 1985;
Schrag and Dixon, 1985a, b; Waller et al., 1985; Lewis, 1991). In adult
males and females, exposure to agents of abuse, e.g., cocaine, disrupts
normal reproductive function in both test species and humans (Smith,
C.G. and Gilbeau, 1985). Numerous chemicals disrupt the ovarian cycle,
alter ovulation, and impair fertility in experimental animals and
humans. These include agents with steroidogenic activity, certain
pesticides, and some metals (Thomas, 1981; Mattison, 1985). In males,
estrogenic compounds can be testicular toxicants in rodents and humans
(Colborn et al., 1993; Toppari et al., 1995). Dibromochloropropane
(DBCP) impairs spermatogenesis in both experimental animals and humans
by another mechanism. These and other examples of toxicant-induced
effects on reproductive function have been reviewed (Katz and
Overstreet, 1981; Working, 1988).
[[Page 56276]]
Altered reproductive health is often manifested as an adverse
effect on the reproductive success or sexual behavior of the couple
even though only one of the pair may be affected directly. Often, it is
difficult to discern which partner has reduced reproductive capability.
For example, exposure of the male to an agent that reduces the number
of normal sperm may result in reduced fertility in the couple, but
without further diagnostic testing, the affected partner may not be
identified. Also, adverse effects on the reproductive systems of the
two sexes may not be detected until a couple attempts to conceive a
child.
For successful reproduction, it is critical that the biologic
integrity of the human reproductive system be maintained. For example,
the events in the estrous or menstrual cycle are closely interrelated;
changes in one event in the cycle can alter other events. Thus, a short
or inadequate luteal phase of the menstrual cycle is associated with
disorders in ovarian follicular steroidogenesis, gonadotropin
secretion, and endometrial integrity (McNatty, 1979; Scommegna et al.,
1980; Smith, S.K. et al., 1984; Sakai and Hodgen, 1987). Toxicants may
interfere with luteal function by altering hypothalamic or pituitary
function and by affecting ovarian response (La Bella et al., 1973a, b).
Fertility of the human male is particularly susceptible to agents
that reduce the number or quality of sperm produced. Compared with many
other species, human males produce fewer sperm relative to the number
of sperm required for fertility (Amann, 1981; Working, 1988). As a
result, many men are subfertile or infertile (Amann, 1981). The
incidence of infertility in men is considered to increase at sperm
concentrations below 20 x 10\6\ sperm per mL of ejaculate. As the
concentration of sperm drops below that level, the probability of a
pregnancy resulting from a single ejaculation declines. If the number
of normal sperm per ejaculate is sufficiently low, fertilization is
unlikely and an infertile condition exists. However, some men with low
sperm concentrations are able to achieve conception and many subfertile
men have concentrations greater than 20 x 10\6\ illustrating the
importance of sperm quality. Toxic agents may further decrease
production of sperm and increase risk of impaired fertility.
C. The Risk Assessment Process and Its Application To Reproductive
Toxicity
Risk assessment is the process by which scientific judgments are
made concerning the potential for toxicity to occur in humans. In 1983,
the National Research Council (NRC) defined risk assessment as
comprising some or all of the following components: hazard
identification, dose-response assessment, exposure assessment, and risk
characterization (NRC, 1983). In its 1994 report, Science and Judgment
in Risk Assessment, the NRC extended its view of the paradigm to
include characterization of each component (NRC, 1994). In addition, it
noted the importance of an interactive approach that deals with
recurring conceptual issues that cut across all stages of risk
assessment. These Guidelines adopt an interactive approach by
organizing the process around the components of hazard
characterization, the quantitative dose-response analysis, the exposure
assessment, and the risk characterization where hazard characterization
combines hazard identification with qualitative consideration of dose-
response relationships, route, timing, and duration of exposure. This
is done because, in practice, hazard identification for reproductive
toxicity and other noncancer health effects include an evaluation of
dose-response relationships, route, timing, and duration of exposure in
the studies used to identify the hazard. Determining a hazard often
depends on whether a dose-response relationship is present (Kimmel,
C.A. et al., 1990). This approach combines the information important in
comparing the toxicity of a chemical to potential human exposure
scenarios identified as part of the exposure assessment. Also, it
minimizes the potential for labeling chemicals inappropriately as
``reproductive toxicants'' on a purely qualitative basis.
In hazard characterization, all available experimental animal and
human data, including observed effects, associated doses, routes,
timing, and duration of exposure, are examined to determine if an agent
causes reproductive toxicity in that species and, if so, under what
conditions. From the hazard characterization and criteria provided in
these Guidelines, the health-related database can be characterized as
sufficient or insufficient for use in risk assessment (Section III.G.).
This approach does not preclude the evaluation and use of the data for
other purposes when adequate quantitative information for setting
reference doses (RfDs) and reference concentrations (RfCs) is not
available.
The next step, the quantitative dose-response analysis (Section
IV), includes determining the no-observed-adverse-effect-level (NOAEL)
and/or the lowest-observed-adverse-effect-level (LOAEL) for each study
and type of effect. Because of the limitations associated with the use
of the NOAEL, the Agency is beginning to use an additional approach,
the benchmark dose approach (Crump, 1984; U.S. EPA. 1995b), for a more
quantitative dose-response evaluation when allowed by the data. The
benchmark dose approach takes into account the variability in the data
and the slope of the dose-response curve, and thus, provides more
complete use of the data for calculation of the RfD or RfC. If the data
are considered sufficient for risk assessment, and if reproductive
toxicity occurs at the lowest toxic dose level (i.e., the critical
effect), an RfD or RfC, based on adverse reproductive effects, could be
derived. This RfD or RfC is derived using the NOAEL or benchmark dose
divided by uncertainty factors to account for interspecies differences
in response, intraspecies variability and deficiencies in the database.
Exposure assessment identifies and describes populations exposed or
potentially exposed to an agent, and presents the type, magnitude,
frequency, and duration of such exposures. Those procedures are
considered separately in the Guidelines for Exposure Assessment (U.S.
EPA, 1992). However, unique considerations for reproductive toxicity
exposure assessments are detailed in Section V.
A statement of the potential for human risk and the consequences of
exposure can come only from integrating the hazard characterization and
quantitative dose-response analysis with human exposure estimates in
the risk characterization. As part of risk characterization, the
strengths and weaknesses in each component of the risk assessment are
summarized along with major assumptions, scientific judgments, and to
the extent possible, qualitative descriptions and quantitative
estimates of the uncertainties.
In 1992, EPA issued a policy memorandum (Habicht, 1992) and
guidance package on risk characterization to encourage more
comprehensive risk characterizations, to promote greater consistency
and comparability among risk characterizations, and to clarify the role
of professional judgment in characterizing risk. In 1995, the Agency
issued a new risk characterization policy and guidance (Browner, 1995)
that refines and reaffirms the principles found in the 1992 policy and
outlines a process within the Agency for implementation. Although
specific program policies and procedures are still evolving, these
Guidelines discuss attributes of the Agency's risk
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characterization policy as it applies to reproductive toxicity.
Risk assessment is just one component of the regulatory process.
The other component, risk management, uses risk characterization along
with directives of the enabling regulatory legislation and other
factors to decide whether to control exposure to the suspected agent
and the level of control. Risk management decisions also consider
socioeconomic, technical, and political factors. Risk management is not
discussed directly in these guidelines because the basis for
decisionmaking goes beyond scientific considerations alone. However,
the use of scientific information in this process is discussed. For
example, the acceptability of the margin of exposure (MOE) is a risk
management decision, but the scientific bases for generating this value
are discussed here.
Dated: October 15, 1996.
Carol M. Browner,
Administrator.
Contents
List of Tables
Part A. Guidelines for Reproductive Toxicity Risk Assessment
I. Overview
II. Definitions and Terminology
III. Hazard Characterization for Reproductive Toxicants
III.A. Laboratory Testing Protocols
III.A.1. Introduction
III.A.2. Duration of Dosing
III.A.3. Length of Mating Period
III.A.4. Number of Females Mated to Each Male
III.A.5. Single- and Multigeneration Reproduction Tests
III.A.6. Alternative Reproductive Tests
III.A.7. Additional Test Protocols That May Provide Reproductive
Data
III.B. Endpoints for Evaluating Male and Female Reproductive
Toxicity In Test Species
III.B.1. Introduction
III.B.2. Couple-Mediated Endpoints
III.B.2.a. Fertility and Pregnancy Outcomes
III.B.2.b. Sexual Behavior
III.B.3. Male-Specific Endpoints
III.B.3.a. Introduction
III.B.3.b. Body Weight and Organ Weights
III.B.3.c. Histopathologic Evaluations
III.B.3.d. Sperm Evaluations
III.B.3.e. Paternally Mediated Effects on Offspring
III.B.4. Female-Specific Endpoints
III.B.4.a. Introduction
III.B.4.b. Body Weight, Organ Weight, Organ Morphology, and
Histology
III.B.4.b.1. Body weight
III.B.4.b.2. Ovary
III.B.4.b.3. Uterus
III.B.4.b.4. Oviducts
III.B.4.b.5. Vagina and external genitalia
III.B.4.b.6. Pituitary
III.B.4.c. Oocyte Production
III.B.4.c.1. Folliculogenesis
III.B.4.c.2. Ovulation
III.B.4.c.3. Corpus luteum
III.B.4.d. Alterations in the Female Reproductive Cycle
III.B.4.e. Mammary Gland and Lactation
III.B.4.f. Reproductive Senescence
III.B.5. Developmental and Pubertal Alterations
III.B.6. Endocrine Evaluations
III.B.7. In Vitro Tests of Reproductive Function
III.C. Human Studies
III.C.1. Epidemiologic Studies
III.C.1.a. Selection of Outcomes for Study
III.C.1.b. Reproductive History Studies
III.C.1.c. Community Studies and Surveillance Programs
III.C.1.d. Identification of Important Exposures for
Reproductive Effects
III.C.1.e. General Design Considerations
III.C.2. Examination of Clusters, Case Reports, or Series
III.D. Pharmacokinetic Considerations
III.E. Comparisons of Molecular Structure
III.F. Evaluation of Dose-Response Relationships
III.G. Characterization of the Health-Related Database
IV. QUANTITATIVE DOSE-RESPONSE ANALYSIS
V. EXPOSURE ASSESSMENT
VI. RISK CHARACTERIZATION
VI.A. Overview
VI.B. Integration of Hazard Characterization, Quantitative Dose-
Response, and Exposure Assessments
VI.C. Descriptors of Reproductive Risk
VI.C.1. Distribution of Individual Exposures
VI.C.2. Population Exposure
VI.C.3. Margin of Exposure
VI.C.4. Distribution of Exposure and Risk for Different
Subgroups
VI.C.5. Situation-Specific Information
VI.C.6. Evaluation of the Uncertainty in the Risk Descriptors
VI.D. Summary and Research Needs
VII. REFERENCES
PART B. RESPONSE TO SCIENCE ADVISORY BOARD AND PUBLIC COMMENTS
I. INTRODUCTION
II. RESPONSE TO SCIENCE ADVISORY BOARD COMMENTS
III. RESPONSE TO PUBLIC COMMENTS
List of Tables
1. Default Assumptions in Reproductive Toxicity Risk Assessment
2. Couple-Mediated Endpoints of Reproductive Toxicity
3. Selected Indices That May Be Calculated From Endpoints of
Reproductive Toxicity in Test Species
4. Male-Specific Endpoints of Reproductive Toxicity
5. Female-Specific Endpoints of Reproductive Toxicity
6. Categorization of the Health-Related Database
7. Guide for Developing Chemical-Specific Risk Characterizations for
Reproductive Effects
PART A. GUIDELINES FOR REPRODUCTIVE TOXICITY RISK ASSESSMENT
I. Overview
These Guidelines describe the procedures that the EPA follows in
using existing data to evaluate the potential toxicity of environmental
agents to the human male and female reproductive systems and to
developing offspring. These Guidelines focus on reproductive system
function as it relates to sexual behavior, fertility, pregnancy
outcomes, and lactating ability, and the processes that can affect
those functions directly. Included are effects on gametogenesis and
gamete maturation and function, the reproductive organs, and the
components of the endocrine system that directly support those
functions. These Guidelines concentrate on the integrity of the male
and female reproductive systems as required to ensure successful
procreation. They also emphasize the importance of maintaining the
integrity of the reproductive system for overall physical and
psychologic health. The Guidelines for Developmental Toxicity Risk
Assessment (U.S. EPA, 1991) focus specifically on effects of agents on
development and should be used as a companion to these Guidelines.
In evaluating reproductive effects, it is important to consider the
presence, and where possible, the contribution of other manifestations
of toxicity such as mutagenicity or carcinogenicity as well as other
forms of general systemic toxicity. The reproductive process is such
that these areas overlap, and all should be considered in reproductive
risk assessments. Although the endpoints discussed in these Guidelines
can detect impairment to components of the reproductive process, they
may not discriminate effectively between nonmutagenic (e.g., cytotoxic)
and mutagenic mechanisms. Examples of endpoints affected by either type
of mechanism are sperm head morphology and preimplantation loss. If the
effects seen may result from mutagenic events, then there is the
potential for transmissible genetic damage. In such cases, the
Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986c) should be
consulted in conjunction with these Guidelines. The Guidelines for
Carcinogen Risk Assessment (U.S. EPA, 1986a, 1996b) should be consulted
if reproductive system or developmentally induced cancer is detected.
For assessment of risk to the human reproductive systems, the most
appropriate data are those derived from human studies having adequate
study
[[Page 56278]]
design and power. In the absence of adequate human data, our
understanding of the mechanisms controlling reproduction supports the
use of data from experimental animal studies to estimate the risk of
reproductive effects in humans. However, some information needed for
extrapolation of data from experimental animal studies to humans is not
generally available. Therefore, to bridge these gaps in information, a
number of default assumptions are made. These default assumptions,
which are summarized in Table 1, should not preclude inquiry into the
relevance of the data to potential human risk and should be invoked
only after examination of the available information indicates that
necessity. These assumptions provide the inferential basis for the
approaches to risk assessment in these Guidelines. Each assumption
should be evaluated along with other relevant information in making a
final judgment as to human risk for each agent, and that information
summarized in the risk characterization.
Table 1.--Default Assumptions in Reproductive Toxicity Risk Assessment
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1. An agent that produces an adverse reproductive effect in experimental
animals is assumed to pose a potential threat to humans.
2. Effects of xenobiotics on male and female reproductive processes are
assumed generally to be similar unless demonstrated otherwise. For
developmental outcomes, the specific effects in humans are not
necessarily the same as those seen in the experimental species.
3. In the absence of information to determine the most appropriate
experimental species, data from the most sensitive species should be
used.
4. In the absence of information to the contrary, an agent that affects
reproductive function in one sex is assumed to adversely affect
reproductive function in the other sex.
5. A nonlinear dose-response curve is assumed for reproductive toxicity.
------------------------------------------------------------------------
An agent that produces an adverse reproductive effect in
experimental animal studies is assumed to pose a potential reproductive
threat to humans. This assumption is based on comparisons of data for
agents that are known to cause human reproductive toxicity (Thomas,
1981; Nisbet and Karch, 1983; Kimmel, C.A. et al., 1984, 1990; Hemminki
and Vineis, 1985; Meistrich, 1986; Working, 1988). In general, the
experimental animal data indicated adverse reproductive effects that
are also seen in humans.
Because similar mechanisms can be identified in the male and female
of many mammalian species, effects of xenobiotics on male and female
reproductive processes are assumed generally to be similar across
species unless demonstrated otherwise. However, for developmental
outcomes, it is assumed that the specific outcomes seen in experimental
animal studies are not necessarily the same as those produced in
humans. This latter assumption is made because of the possibility of
species-specific differences in timing of exposure relative to critical
periods of development, pharmacokinetics (including metabolism),
developmental patterns, placentation, or modes of action. However,
adverse developmental outcomes in laboratory mammalian studies are
presumed to predict a hazard for adverse developmental outcome in
humans.
When sufficient data are available (e.g., pharmacokinetic) to allow
a decision, the most appropriate species should be used to estimate
human risk. In the absence of such data, it is assumed that the most
sensitive species is most appropriate because, for the majority of
agents known to cause human reproductive toxicity, humans appear to be
as or more sensitive than the most sensitive animal species tested
(Nisbet and Karch, 1983; Kimmel, C.A. et al., 1984, 1990; Hemminki and
Vineis, 1985; Meistrich, 1986; Working, 1988), based on data from
studies that determined dose on a body weight or air concentration
basis.
In the absence of specific information to the contrary, it is
assumed that a chemical that affects reproductive function in one sex
may also adversely affect reproductive function in the other sex. This
assumption for reproductive risk assessment is based on three
considerations: (1) For most agents, the nature of the testing and the
data available are limited, reducing confidence that the potential for
toxicity to both sexes and their offspring has been examined equally;
(2) Exposures of either males or females have resulted in developmental
toxicity; and (3) Many of the mechanisms controlling important aspects
of reproductive system function are similar in females and males, and
therefore could be susceptible to the same agents. Information that
would negate this assumption would demonstrate that either a
mechanistic difference existed between the sexes that would preclude
toxic action on the other sex or, on the basis of sufficient testing,
an agent did not produce an adverse reproductive effect when
administered to the other sex. Mechanistic differences could include
functions that do not exist in the other sex (e.g., lactation),
differences in endocrine control of affected organ development or
function, or pharmacokinetic and metabolic differences between sexes.
In a quantitative dose-response analysis, mode of action,
pharmacokinetic, and pharmacodynamic information should be used to
predict the shape of the dose-response curve when sufficient
information of that nature is available. When that information is
insufficient, it has generally been assumed that there is a nonlinear
dose-response for reproductive toxicity. This is based on known
homeostatic, compensatory, or adaptive mechanisms that must be overcome
before a toxic endpoint is manifested and on the rationale that cells
and organs of the reproductive system and the developing organism are
known to have some capacity for repair of damage. However, in a
population, background levels of toxic agents and preexisting
conditions may increase the sensitivity of some individuals in the
population. Thus, exposure to a toxic agent may result in an increased
risk of adverse effects for some, but not necessarily all, individuals
within the population. Although a threshold may exist for endpoints of
reproductive toxicity, it usually is not feasible to distinguish
empirically between a true threshold and a nonlinear low-dose
relationship. The shift to the term nonlinear does not change the RfD/
RfC methodology for reproductive system health effects, including the
use of uncertainty factors.
II. Definitions and Terminology
For the purposes of these Guidelines, the following definitions
will be used: Reproductive toxicity--The occurrence of biologically
adverse effects on the reproductive systems of females or males that
may result from exposure to environmental agents. The toxicity may be
expressed as alterations to the female or male reproductive organs, the
related endocrine system, or pregnancy outcomes. The manifestation of
such toxicity may include, but not be limited to, adverse effects on
onset of puberty, gamete production and transport, reproductive cycle
normality, sexual behavior, fertility, gestation, parturition,
lactation, developmental toxicity, premature reproductive senescence,
or modifications in other functions that are dependent on the integrity
of the reproductive systems.
Fertility--The capacity to conceive or induce conception.
[[Page 56279]]
Fecundity--The ability to produce offspring within a given period
of time. For litter-bearing species, the ability to produce large
litters is also a component of fecundity.
Fertile--A level of fertility that is within or exceeds the normal
range for that species.
Infertile--Lacking fertility for a specified period. The infertile
condition may be temporary; permanent infertility is termed sterility.
Subfertile--A level of fertility that is below the normal range for
that species but not infertile.
Developmental toxicity--The occurrence of adverse effects on the
developing organism that may result from exposure prior to conception
(either parent), during prenatal development, or postnatally to the
time of sexual maturation. Adverse developmental effects may be
detected at any point in the lifespan of the organism. The major
manifestations of developmental toxicity include (1) death of the
developing organism, (2) structural abnormality, (3) altered growth,
and (4) functional deficiency (U.S. EPA, 1991).
III. Hazard Characterization for Reproductive Toxicants
Identification and characterization of reproductive hazards can be
based on data from either human or experimental animal studies. Such
data can result from routine or accidental environmental or
occupational exposures or, for experimental animals, controlled
experimental exposures. A hazard characterization should evaluate all
of the information available and should:
Identify the strengths and limitations of the database,
including all available epidemiologic and experimental animal studies
as well as pharmacokinetic and mechanistic information.
Identify and describe key toxicological studies.
Describe the type(s) of effects.
Describe the nature of the effects (irreversible,
reversible, transient, progressive, delayed, residual, or latent
effects).
Describe how much is known about how (through what
biological mechanism) the agent produces adverse effects.
Discuss the other health endpoints of concern.
Discuss any nonpositive data in humans or experimental
animals.
Discuss the dose-response data (epidemiologic or
experimental animal) available for further dose-response analysis.
Discuss the route, level, timing, and duration of exposure
in studies as compared to expected human exposures.
Summarize the hazard characterization, including:
--Major assumptions used,
--Confidence in the conclusions,
--Alternative conclusions also supported by the data,
--Major uncertainties identified, and
--Significant data gaps.
Conduct of a hazard characterization requires knowledge of the
protocols in which data were produced and the endpoints that were
evaluated. Sections III.A. and III.B. present the traditional testing
protocols for rodents and endpoints used to evaluate male and female
reproductive toxicity along with evaluation of their strengths and
limitations. Because many endpoints are common to multiple protocols,
endpoints are considered separately from the discussion of the overall
protocol structures. These are followed by presentation of many of the
specific characteristics of human studies (Section III.C.) and limited
discussions of pharmacokinetic and structure-activity factors (Sections
III.D. and III.E.).
III.A. Laboratory Testing Protocols
III.A.1. Introduction
Testing protocols describe the procedures to be used to provide
data for risk assessments. The quality and usefulness of those data are
dependent on the design and conduct of the tests, including endpoint
selection and resolving power. A single protocol is unlikely to provide
all of the information that would be optimal for conducting a
comprehensive risk assessment. For example, the test design to study
reversibility of adverse effects or mechanism of toxic action may be
different from that needed to determine time of onset of an effect or
for calculation of a safe level for repeated exposure over a long term.
Ideally, results from several different types of tests should be
available when performing a risk assessment. Typically, only limited
data are available. Under those conditions, the limited data should be
used to the extent possible to assess risk.
Integral parts of the hazard characterization and quantitative
dose-response processes are the evaluation of the protocols from which
data are available and the quality of the resulting data. In this
section, design factors that are of particular importance in
reproductive toxicity testing are discussed. Then, standardized
protocols that may provide useful data for reproductive risk
assessments are described.
III.A.2. Duration of Dosing
To evaluate adequately the potential effects of an agent on the
reproductive systems, a prolonged treatment period is needed. For
example, damage to spermatogonial stem cells will not appear in samples
from the cauda epididymis or in ejaculates for 8 to 14 weeks, depending
on the test species. With some chemical agents that bioaccumulate, the
full impact on a given cell type could be further delayed, as could the
impact on functional endpoints such as fertility. In such situations,
adequacy of the dosing duration is a critical factor in the risk
assessment.
Conversely, adaptation may occur that allows tolerance to levels of
a chemical that initially caused an effect that could be considered
adverse. An example is interference with ovulation by chlordimeform
(Goldman et al., 1991); an effect for which a compensatory mechanism is
available. Thus, with continued dosing, the compensatory mechanism can
be activated so that the initial adverse effect is masked.
In these situations, knowledge of the relevant pharmacokinetic and
pharmacodynamic data can facilitate selection of dose levels and
treatment duration (see also section on Exposure Assessment). Equally
important is proper timing of examination of treated animals relative
to initiation and termination of exposure to the agent.
III.A.3. Length of Mating Period
Traditionally, pairs of rats or mice are allowed to cohabit for
periods ranging from several days to 3 weeks. Given a 4- or 5-day
estrous cycle, each female that is cycling normally should be in estrus
four or five times during a 21-day mating period. Therefore,
information on the interval or the number of cycles needed to achieve
pregnancy may provide evidence of reduced fertility that is not
available from fertility data. Additionally, during each period of
behavioral estrus, the male has the opportunity to copulate a number of
times, resulting in delivery of many more sperm than are required for
fertilization. When an unlimited number of matings is allowed in
fertility testing, a large impact to sperm production is necessary
before an adverse effect on fertility can be detected.
[[Page 56280]]
III.A.4. Number of Females Mated to Each Male
The EPA test guidelines prepared pursuant to FIFRA and TSCA specify
the use of 20 males and enough females to produce at least 20
pregnancies for each dose group in each generation in the
multigeneration reproduction test (U.S. EPA, 1982, 1985b, 1996a).
However, in some tests that were not designed to conform to EPA test
guidelines (OECD, 1983), 20 pregnancies may have been achieved by
mating two females with each male and using fewer than 20 males per
treatment group. In such cases, the statistical treatment of the data
should be examined carefully. With multiple females mated to each male,
the degree of independence of the observations for each female may not
be known. In that situation, when the cause of the adverse effect
cannot be assigned with confidence to only one sex, dependence should
be assumed and the male used as the experimental unit in statistical
analyses. Using fewer males as the experimental unit reduces ability to
detect an effect.
III.A.5. Single- and Multigeneration Reproduction Tests
Reproductive toxicity studies in laboratory animals generally
involve continuous exposure to a test substance for one or more
generations. The objective is to detect effects on the integrated
reproductive process as well as to study effects on the individual
reproductive organs. Test guidelines for the conduct of single- and
multigeneration reproduction protocols have been published by the
Agency pursuant to FIFRA and TSCA and by OECD (U.S. EPA, 1982, 1985b,
1996a; Galbraith et al., 1983; OECD, 1983).
The single-generation reproduction test evaluates effects of
subchronic exposure of peripubertal and adult animals. In the
multigeneration reproduction protocol, F1 and F2 offspring
are exposed continuously in utero from conception until birth and
during the preweaning period. This allows detection of effects that
occur from exposures throughout development, including the peripubertal
and young adult phases. Because the parental and subsequent filial
generations have different exposure histories, reproductive effects
seen in any particular generation are not necessarily comparable with
those of another generation. Also, successive litters from the same
parents cannot be considered as replicates because of factors such as
continuing exposure of the parents, increased parental age, sexual
experience, and parity of the females.
In a single- or multigeneration reproduction test, rats are used
most often. In a typical reproduction test, dosing is initiated at 5 to
8 weeks of age and continued for 8 to 10 weeks prior to mating to allow
effects on gametogenesis to be expressed and increase the likelihood of
detecting histologic lesions. Three dose levels plus one or more
control groups are usually included. Enough males and females are mated
to ensure 20 pregnancies per dose group for each generation. Animals
producing the first generation of offspring should be considered the
parental (P) generation, and all subsequent generations should be
designated filial generations (e.g., F1, F2). Only the P
generation is mated in a single-generation test, while both the P and
F1 generations are mated in a two-generation reproduction test.
In the P generation, both females and males are treated prior to
and during mating, with treatment usually beginning around puberty.
Cohabitation can be allowed for up to 3 weeks (U.S. EPA, 1982, 1985b),
during which the females are monitored for evidence of mating. Females
continue to be exposed during gestation and lactation.
In the two-generation reproduction test, randomly selected F1
male and female offspring continue to be exposed after weaning (day 21)
and through the mating period. Treatment of mated F1 females is
continued throughout gestation and lactation. More than one litter may
be produced from either P or F1 animals. Depending on the route of
exposure of lactating females, it is important to consider that
offspring may be exposed to a chemical by ingestion of maternal feed or
water (diet or drinking water studies), by licking of exposed fur
(inhalation study), by contact with treated skin (dermal study), or by
coprophagia, as well as via the milk.
In single- and multigeneration reproduction tests, reproductive
endpoints evaluated in P and F generations usually include visual
examination of the reproductive organs. Weights and histopathology of
the testes, epididymides, and accessory sex glands may be available
from males, and histopathology of the vagina, uterus, cervix, ovaries,
and mammary glands from females. Uterine and ovarian weights also are
often available. Male and female mating and fertility indices (Section
III.B.2.a.) are usually presented. In addition, litters (and often
individual pups) are weighed at birth and examined for number of live
and dead offspring, gender, gross abnormalities, and growth and
survival to weaning. Maturation and behavioral testing may also be
performed on the pups.
If effects on fertility or pregnancy outcome are the only adverse
effects observed in a study using one of these protocols, the
contributions of male- and female-specific effects often cannot be
distinguished. If testicular histopathology or sperm evaluations have
been included, it may be possible to characterize a male-specific
effect. Similarly, ovarian and reproductive tract histology or changes
in estrous cycle normality may be indicative of female-specific
effects. However, identification of effects in one sex does not exclude
the possibility that both sexes may have been affected adversely. Data
from matings of treated males with untreated females and vice versa
(crossover matings) are necessary to separate sex-specific effects.
An EPA workshop has considered the relative merits of one- versus
two-generation reproductive effects studies (Francis and Kimmel, 1988).
The participants concluded that a one-generation study is insufficient
to identify all potential reproductive toxicants, because it would
exclude detection of effects caused by prenatal and postnatal exposures
(including the prepubertal period) as well as effects on germ cells
that could be transmitted to and expressed in the next generation. For
example, adverse transgenerational effects on reproductive system
development by agents that disrupt endocrine control of sexual
differentiation would be missed. A one-generation test might also miss
adverse effects with delayed or latent onset because of the shorter
duration of exposure for the P generation. These limitations are shared
with the shorter-term ``screening'' protocols described below. Because
of these limitations, a comprehensive reproductive risk assessment
should include results from a two-generation test or its equivalent. A
further recommendation from the workshop was to include sperm analyses
and estrous cycle normality as endpoints in reproductive effects
studies. These endpoints have been included in the proposed revisions
to the EPA test guideline (U.S. EPA, 1996a).
In studies where parental and offspring generations are evaluated,
there are additional risk assessment issues regarding the relationships
of reproductive outcomes across generations. Increasing vulnerability
of subsequent generations is often, but not always, observed.
Qualitative predictions of increased risk of the filial generations
could be strengthened by
[[Page 56281]]
knowledge of the reproductive effects in the adult, the likelihood of
bioaccumulation of the agent, and the potential for increased
sensitivity resulting from exposure during critical periods of
development (Gray, 1991).
Occasionally, the severity of effects may be static or decreased
with succeeding generations. When a decrease occurs, one explanation
may be that the animals in the F1 and F2 generations
represent ``survivors'' who are (or become) more resistant to the agent
than the average of the P generation. If such selection exists, then
subsequent filial generations may show a reduced toxic response. Thus,
significant adverse effects in any generation may be cause for concern
regardless of results in other generations unless inconsistencies in
the data indicate otherwise.
III.A.6. Alternative Reproductive Tests
A number of alternative test designs have appeared in the
literature (Lamb, 1985; Lamb and Chapin, 1985; Gray et al., 1988, 1989,
1990; Morrissey et al., 1989). Although not necessarily viewed as
replacements for the standard two-generation reproduction tests, data
from these protocols may be used on a case-by-case basis depending on
what is known about the test agent in question. When mutually agreed on
by the testing organization and the Agency, such alternative protocols
may offer an expanded array of endpoints and increased flexibility
(Francis and Kimmel, 1988).
A continuous breeding protocol, Fertility (or Reproductive)
Assessment by Continuous Breeding (FACB or RACB), has been developed by
the National Toxicology Program (NTP) (Lamb and Chapin, 1985; Morrissey
et al., 1989; Gulati et al., 1991). As originally described, this
protocol (FACB) was a one-generation test. However, in the current
design (RACB), dosing is extended into the F1 generation to make
it compatible with the EPA workshop recommendations for a two-
generation design (Francis and Kimmel, 1988). The RACB protocol is
being used with both mice and rats. A distinctive feature of this
protocol is the continuous cohabitation of male-female pairs (in the P
generation) for 14 weeks. Up to five litters can be produced with the
pups removed soon after birth. This protocol provides information on
changes in the spacing, number, and size of litters over the 14-week
dosing interval. Treatment (three dose levels plus controls) is
initiated in postpubertal males and females (11 weeks of age) seven
days before cohabitation and continues throughout the test. Offspring
that are removed from the dam soon after birth are counted and examined
for viability, litter and/or pup weight, sex, and external
abnormalities and then discarded. The last litter may remain with the
dam until weaning to study the effects of in utero as well as perinatal
and postnatal exposures. If effects on fertility are observed in the P
or F generations, additional reproductive evaluations may be conducted,
including fertility studies and crossover matings to define the
affected gender and site of toxicity.
The sequential production of litters from the same adults allows
observation of the timing of onset of an adverse effect on fertility.
In addition, it improves the ability to detect subfertility due to the
potential to produce larger numbers of pregnancies and litters than in
a standard single- or multigeneration reproduction study. With
continuous treatment, a cumulative effect could increase the incidence
or extent of expression with subsequent litters. However, unless
offspring were allowed to grow and reproduce (as they are routinely in
the more recent version of the RACB protocol) (Gulati et al., 1991),
little or no information will be available on postnatal development or
reproductive capability of a second generation.
Sperm measures (including sperm number, morphology, and motility)
and vaginal smear cytology to detect changes in estrous cyclicity have
been added to the RACB protocol at the end of the test period and their
utility has been examined using model compounds in the mouse (Morrissey
et al., 1989).
Another test method combines the use of multiple endpoints in both
sexes of rats with initiation of treatment at weaning (Gray et al.,
1988). Thus, morphologic and physiologic changes associated with
puberty are included as endpoints. Both P sexes are treated (at least
three dose levels plus controls) continuously through breeding,
pregnancy, and lactation. The F1 generation is mated in a
continuous breeding protocol. Vaginal smears are recorded daily
throughout the test period to evaluate estrous cycle normality and
confirm breeding and pregnancy (or pseudopregnancy). Pregnancy outcome
is monitored in both the P and F1 generations at all doses, and
terminal studies on both generations include comprehensive assessment
of sperm measures (number, morphology, motility) as well as organ
weights, histopathology, and the serum and tissue levels of appropriate
reproductive hormones. As with the RACB, crossover mating studies may
be conducted to identify the affected sex as warranted. This protocol
combines the advantages of a continuous breeding design with
acquisition of sex-specific multiple endpoint data at all doses. In
addition, identification of pubertal effects makes this protocol
particularly useful for detecting compounds with hormone-mediated
actions such as environmental estrogens or antiandrogens.
III.A.7. Additional Test Protocols That May Provide Reproductive Data
Several shorter-term reproductive toxicity screening tests have
been developed. Among those are the Reproductive/Developmental Toxicity
Screening Test, which is part of the OECD's Screening Information Data
Set protocol (Scala et al., 1992; Tanaka et al., 1992; OECD, 1993a), a
tripartite protocol developed by the International Conference on
Harmonization (International Conference on Harmonization of Technical
Requirements of Pharmaceuticals for Human Use, 1994; Manson, 1994), and
the NTP's Short-Term Reproductive and Developmental Toxicity Screen
(Harris, M.W. et al., 1992). These protocols have been developed for
setting priorities for further testing and should not be considered
sufficient by themselves to establish regulatory exposure levels. Their
limited exposure periods do not allow assessment of certain aspects of
the reproductive process, such as developmentally induced effects on
the reproductive systems of offspring, that are covered by the
multigeneration reproduction protocols.
The male dominant lethal test was designed to detect mutagenic
effects in the male spermatogenic process that are lethal to the
offspring. A female dominant lethal protocol has also been used to
detect equivalent effects on oogenesis (Generoso and Piegorsch, 1993).
A review of the male dominant lethal test has been published as
part of the EPA's Gene-Tox Program (Green et al., 1985). Dominant
lethal protocols may use acute dosing (1 to 5 days) followed by serial
matings with one or two females per male per week for the duration of
the spermatogenic process. An alternative protocol may use subchronic
dosing for the duration of the spermatogenic process followed by
mating. Dose levels used with the acute protocol are usually higher
than those used with the subchronic protocol. Females are monitored for
evidence of mating, killed at approximately midgestation, and examined
for incidence of pre- and postimplantation loss (see Section III.B.2.
for discussions of these endpoints).
[[Page 56282]]
Pre- or postimplantation loss in the dominant lethal test is often
considered evidence that the agent has induced mutagenic damage to the
male germ cell (U.S. EPA, 1986c). A genotoxic basis for a substantial
portion of postimplantation loss is accepted widely. However, methods
used to assess preimplantation loss do not distinguish between
contributions of mutagenic events that cause embryo death and
nonmutagenic factors that result in failure of fertilization or early
embryo mortality (e.g., inadequate number of normal sperm, failure in
sperm transport or ovum penetration). Similar effects (fertilization
failure, early embryo death) could also be produced indirectly by
effects that delay the timing of fertilization relative to time of
ovulation. Such distinctions are important because cytotoxic effects on
gametogenic cells do not imply the potential for transmittable genetic
damage that is associated with mutagenic events. The interpretation of
an increase in preimplantation loss may require additional data on the
agent's mutagenic and gametotoxic potential if genotoxicity is to be
factored into the risk assessment. Regardless, significant effects may
be observed in a dominant lethal test that are considered reproductive
in nature.
An acute exposure protocol, combined with serial mating, may allow
identification of the spermatogenic cell types that are affected by
treatment. However, acute dosing may not produce adverse effects at
levels as low as with subchronic dosing because of factors such as
bioaccumulation. Conversely, if tolerance to an agent is developed with
longer exposure, an effect may be observed after acute dosing that is
not detected after longer-term dosing.
Subchronic toxicity tests may have been conducted before a detailed
reproduction study is initiated. In the subchronic toxicity test with
rats, exposure usually begins at 6-8 weeks of age and is continued for
90 days (U.S. EPA, 1982, 1985b). Initiation of exposure at 8 weeks of
age (compared with 6) and exposure for approximately 90 days allows the
animals to reach a more mature stage of sexual development and assures
an adequate length of dosing for observation of effects on the
reproductive organs with most agents. The route of administration is
often oral or by gavage but may be dermal or by inhalation. Animals are
monitored for clinical signs throughout the test and are necropsied at
the end of dosing.
The endpoints that are usually evaluated for the male reproductive
system include visual examination of the reproductive organs, plus
weights and histopathology for the testes, epididymides, and accessory
sex glands. For the females, endpoints may include visual examination
of the reproductive organs, uterine and ovarian weights, and
histopathology of the vagina, uterus, cervix, ovaries, and mammary
glands.
This test may be useful to identify an agent as a potential
reproductive hazard, but usually does not provide information about the
integrated function of the reproductive systems (sexual behavior,
fertility, and pregnancy outcomes), nor does it include effects of the
agent on immature animals.
Chronic toxicity tests provide an opportunity to evaluate toxic
effects of long-term exposures. Oral, inhalation, or dermal exposure is
initiated soon after weaning and is usually continued for 12 to 24
months. Because of the extended treatment period, data from interim
sacrifices may be available to provide useful information regarding the
onset and sequence of toxicity. In males, the reproductive organs are
examined visually, testes are weighed, and histopathologic examination
is done on the testes and accessory sex glands. In females, the
reproductive organs are examined visually, uterine and ovarian weights
may be obtained, and histopathologic evaluation of the reproductive
organs is done. The incidence of pathologic conditions is often
increased in the reproductive tracts of aged control animals.
Therefore, findings should be interpreted carefully.
III.B. Endpoints for Evaluating Male and Female Reproductive Toxicity
in Test Species
III.B.1. Introduction
The following discussion emphasizes endpoints that measure
characteristics that are necessary for successful sexual performance
and procreation. Other areas that are related less directly to
reproduction are beyond the scope of these Guidelines. For example,
secondary adverse health effects that may result from toxicity to the
reproductive organs (e.g., osteoporosis or altered immune function),
although important, are not included.
In these Guidelines, the endpoints of reproductive toxicity are
separated into three categories: couple-mediated, female-specific, and
male-specific. Couple-mediated endpoints are those in which both sexes
can have a contributing role if both partners are exposed. Thus,
exposure of either sex or both sexes may result in an effect on that
endpoint.
The discussions of endpoints and the factors influencing results
that are presented in this section are directed to evaluation and
interpretation of results with test species. Many of those endpoints
require invasive techniques that preclude routine use with humans.
However, in some instances, related endpoints that can be used with
humans are identified. Information that is specific for evaluation of
effects on humans is presented in Section III.C.
Although statistical analyses are important in determining the
effects of a particular agent, the biological significance of data is
most important. It is important to be aware that when many endpoints
are investigated, statistically significant differences may occur by
chance. On the other hand, apparent trends with dose may be
biologically relevant even though pair-wise comparisons do not indicate
a statistically significant effect. In each section, endpoints are
identified in which significant changes may be considered adverse.
However, concordance of results and known biology should be considered
in interpreting all results. Results should be evaluated on a case-by-
case basis with all of the evidence considered. Scientific judgment
should be used extensively. All effects that may be considered as
adverse are appropriate for use in establishing a NOAEL, LOAEL, or
benchmark dose.
III.B.2. Couple-Mediated Endpoints
Data on fertility potential and associated reproductive outcomes
provide the most comprehensive and direct insight into reproductive
capability. As noted previously, most protocols only specify
cohabitation of exposed males with exposed females. This complicates
the resolution of gender-specific influences. Conclusions may need to
be restricted to noting that the ``couple'' is at reproductive risk
when one or both parents are potentially exposed.
III.B.2.a. Fertility and Pregnancy Outcomes. Breeding studies with
test species are a major source of data on reproductive toxicants.
Evaluations of fertility and pregnancy outcomes provide measures of the
functional consequences of reproductive injury. Measures of fertility
and pregnancy outcome that are often obtained from multigeneration
reproduction studies are presented in Table 2. Many endpoints that are
pertinent for developmental toxicity are also listed and discussed in
the Agency's Guidelines for Developmental Toxicity Risk Assessment
(U.S. EPA, 1991). Also included in Table 2 are measures that
[[Page 56283]]
may be obtained from other types of studies (e.g., single-generation
reproduction studies, developmental toxicity studies, dominant lethal
studies) in which offspring are not retained to evaluate subsequent
reproductive performance.
Table 2.--Couple-Mediated End-points of Reproductive Toxicity
------------------------------------------------------------------------
-------------------------------------------------------------------------
Multigeneration studies:
Mating rate, time to mating (time to pregnancy*)
Pregnancy rate*
Delivery rate*
Gestation length*
Litter size (total and live)
Number of live and dead offspring (Fetal death rate*)
Offspring gender* (sex ratio)
Birth weight*
Postnatal weights*
Offspring survival*
External malformations and variations*
Offspring reproduction*
Other reproductive endpoints:
Ovulation rate
Fertilization rate
Preimplantation loss
Implantation number
Postimplantation loss*
Internal malformations and variations*
Postnatal structural and functional development*
------------------------------------------------------------------------
*Endpoints that can be obtained with humans.
Some of the endpoints identified above are used to calculate ratios
or indices (NRCl, 1977; Collins, 1978; Schwetz et al., 1980; U.S. EPA,
1982, 1985b; Dixon and Hall, 1984; Lamb et al., 1985; Thomas, 1991).
While the presentation of such indices is not discouraged, the
measurements used to calculate those indices should also be available
for evaluation. Definitions of some of these indices in published
literature vary substantially. Also, the calculation of an index may be
influenced by the test design. Therefore, it is important that the
methods used to calculate indices be specified. Some commonly reported
indices are in Table 3.
[[Page 56284]]
Table 3.--Selected Indices That May Be Calculated From Endpoints of
Reproductive Toxicity in Test Species
Mating Index
[GRAPHIC] [TIFF OMITTED] TN31OC96.000
Note: Mating is used to indicate that evidence of copulation
(observation or other evidence of ejaculation such as vaginal plug
or sperm in vaginal smear) was obtained.
Fertility Index
[GRAPHIC] [TIFF OMITTED] TN31OC96.001
Note: Because both sexes are often exposed to an agent,
distinction between sexes often is not possible. If responsibility
for an effect can be clearly assigned to one sex (as when treated
animals are mated with controls), then a female or male fertility
index could be useful.
Gestation (Pregnancy) Index
[GRAPHIC] [TIFF OMITTED] TN31OC96.002
Live Birth Index
[GRAPHIC] [TIFF OMITTED] TN31OC96.003
Sex Ratio
[GRAPHIC] [TIFF OMITTED] TN31OC96.004
4-Day Survival Index (Viability Index)
[GRAPHIC] [TIFF OMITTED] TN31OC96.005
Note: This definition assumes that no standardization of litter
size is done until after the day 4 determination is completed.
Lactation Index (Weaning Index)
[GRAPHIC] [TIFF OMITTED] TN31OC96.006
Note: If litters were standardized to equalize numbers of
offspring per litter, number of offspring after standardization
should be used instead of number born alive. When no standardization
is done, measure is called weaning index. When standardization is
done, measure is called lactation index.
Preweaning Index
[GRAPHIC] [TIFF OMITTED] TN31OC96.007
Note: If litters were standardized to equalize numbers of
offspring per litter, then number of offspring remaining after
standardization should be used instead of number born.
------------------------------------------------------------------------
Mating rate may be reported for the mated pairs, males only or
females only. Evidence of mating may be direct observation of
copulation, observation of copulatory plugs, or observation of sperm in
the vaginal fluid (vaginal lavage). The mating rate may be influenced
by the number of estrous cycles allowed or required for pregnancy to
occur. Therefore, mating rate and fertility data from the first estrous
cycle after initiation of cohabitation should be more discriminating
than measurements involving multiple cycles. Evidence of mating does
not necessarily mean successful impregnation.
A useful indicator of impaired reproductive function may be the
length of time required for each pair to mate after the start of
cohabitation (time to mating). An increased interval between initiation
of cohabitation and evidence of mating suggests abnormal estrous
cyclicity in the female or impaired sexual behavior in one or both
partners.
The time to mating for normal pairs (rat or mouse) could vary by 3
or 4 days depending on the stage of the estrous cycle at the start of
cohabitation. If the
[[Page 56285]]
stage of the estrous cycle at the time of cohabitation is known, the
component of the variance due to variation in stage at cohabitation can
be removed in the data analysis.
Data on fertilization rate, the proportion of available ova that
were fertilized, are seldom available because the measurement requires
necropsy very early in gestation. Pregnancy rate is the proportion of
mated pairs that have produced at least one pregnancy within a fixed
period where pregnancy is determined by the earliest available evidence
that fertilization has occurred. Generally, a more meaningful measure
of fertility results when the mating opportunity was limited to one
mating couple and to one estrous cycle (see Sections III.A.3. and
III.A.4.).
The timing and integrity of gamete and zygote transport are
important to fertilization and embryo survival and are quite
susceptible to chemical perturbation. Disruption of the processes that
contribute to a reduction in fertilization rate and increased early
embryo loss are usually identified simply as preimplantation loss.
Additional studies using direct assessments of fertilized ova and early
embryos would be necessary to identify the cause of increased
preimplantation loss (Cummings and Perreault, 1990). Preimplantation
loss (described below) occurs in untreated as well as treated rodents
and contributes to the normal variation in litter size.
After mating, uterine and oviductal contractions are critical in
the transport of spermatozoa from the vagina. In rodents, sufficient
stimulation during mating is necessary for initiation of those
contractions. Thus, impaired mating behavior may affect sperm transport
and fertilization rate. Exposure of the female to estrogenic compounds
can alter gamete transport. In women, low doses of exogenous estrogens
may accelerate ovum transport to a detrimental extent, whereas high
doses of estrogens or progestins delay transport and increase the
incidence of ectopic pregnancies.
Mammalian ova are surrounded by investments that the sperm must
penetrate before fusing with ova. Chemicals may block fertilization by
preventing this passage. Other agents may impair fusion of the sperm
with the oolemma, transformations of the sperm or ovum chromatin into
the male and female pronuclei, fusion of the pronuclei, or the
subsequent cleavage divisions. Carbendazim, an inhibitor of microtubule
synthesis, is an example of a chemical that can interfere with oocyte
maturation and normal zygote formation after sperm-egg fusion by
affecting meiosis (Perreault et al., 1992; Zuelke and Perreault, 1995).
The early zygote is also susceptible to detrimental effects of mutagens
such as ethylene oxide (Generoso et al., 1987).
Fertility assessments in test animals have limited sensitivity as
measures of reproductive injury. Therefore, results demonstrating no
treatment-related effect on fertility may be given less weight than
other endpoints that are more sensitive. Unlike humans, normal males of
most test species produce sperm in numbers that greatly exceed the
minimum requirements for fertility, particularly as evaluated in
protocols that allow multiple matings (Amann, 1981; Working, 1988). In
some strains of rats and mice, production of normal sperm can be
reduced by up to 90% or more without compromising fertility (Aafjes et
al., 1980; Meistrich, 1982; Robaire et al., 1984; Working, 1988).
However, less severe reductions can cause reduced fertility in human
males who appear to function closer to the threshold for the number of
normal sperm needed to ensure full reproductive competence (see
Supplementary Information). This difference between test species and
humans means that negative results with test species in a study that
was limited to endpoints that examined only fertility and pregnancy
outcomes would provide insufficient information to conclude that the
test agent poses no reproductive hazard in humans. It is unclear
whether a similar consideration is applicable for females for some
mechanisms of toxicity.
The limited sensitivity of fertility measures in rodents also
suggests that a NOAEL, LOAEL, or benchmark dose (see Section IV) based
on fertility may not reflect completely the extent of the toxic effect.
In such instances, data from additional reproductive endpoints might
indicate that an adverse effect could occur at a lower dose level. In
the absence of such data, the margin of exposure or uncertainty factor
applied to the NOAEL, LOAEL, or benchmark dose may need to be adjusted
to reflect the additional uncertainty (see Section IV).
Both the blastocyst and the uterus must be ready for implantation,
and their synchronous development is critical (Cummings and Perreault,
1990). The preparation of the uterine endometrium for implantation is
under the control of sequential estrogen and progesterone stimulation.
Treatments that alter the internal hormonal environment or inhibit
protein synthesis, mitosis, or cell differentiation can block
implantation and cause embryo death.
Gestation length can be determined in test animals from data on day
of mating (observation of vaginal plug or sperm-positive vaginal
lavage) and day of parturition. Significant shortening of gestation can
lead to adverse outcomes of pregnancy such as decreased birth weight
and offspring survival. Significantly longer gestation may be caused by
failure of the normal mechanism for parturition and may result in death
or impairment of offspring if dystocia (difficulty in parturition)
occurs. Dystocia constitutes a maternal health threat for humans as
well as test species. Lengthened gestation may result in higher birth
weight; an effect that could mask a slower growth rate in utero because
of exposure to a toxic agent. Comparison of offspring weights based on
conceptional age may allow insight, although this comparison is
complicated by generally faster growth rates postnatally than in utero.
Litter size is the number of offspring delivered and is measured at
or soon after birth. Unless this observation is made soon after
parturition, the number of offspring observed may be less than the
actual number delivered because of cannibalism by the dam. Litter size
is affected by the number of ova available for fertilization (ovulation
rate), fertilization rate, implantation rate, and the proportion of the
implanted embryos that survives to parturition. Litter size may include
dead as well as live offspring, therefore data on the numbers of live
and dead offspring should be available also.
When pregnant animals are examined by necropsy in mid- to late
gestation, pregnancy status, including pre- and postimplantation losses
can be determined. Postimplantation loss can be determined also by
examining uteri from postparturient females. Preimplantation loss is
the (number of corpora lutea minus number of implantation sites)/number
of corpora lutea. Postimplantation loss, determined following delivery
of a litter, is the (total number of implantation sites minus number of
full-term pups)/number of implantation sites.
Offspring gender in mammals is determined by the male through
fertilization of an ovum by a Y- or an X-chromosome-bearing sperm.
Therefore, selective impairment in the production, transport, or
fertilizing ability of either of these sperm types can produce an
alteration in the sex ratio. An agent may also induce selective loss of
male or female fetuses. Further, alteration of the external sexual
characteristics of offspring by agents that disrupt sexual development
may produce apparent
[[Page 56286]]
effects on sex ratios. Although not examined routinely, these factors
provide the most likely explanations for alterations in the sex ratio.
Birth weight should be measured on the day of parturition. Often
data from individual pups as well as the entire litter (litter weight)
are provided. Birth weights are influenced by intrauterine growth
rates, litter size, and gestation length. Growth rate in utero is
influenced by the normality of the fetus, the maternal environment, and
gender, with females tending to be smaller than males (Tyl, 1987).
Individual pups in large litters tend to be smaller than pups in
smaller litters. Thus, reduced birth weights that can be attributed to
large litter size should not be considered an adverse effect unless the
increased litter size is treatment related and the subsequent ability
of the offspring to survive or develop is compromised. Multivariate
analyses may be used to adjust pup weights for litter size (e.g.,
analysis of covariance, multiple regression). When litter weights only
are reported, the increased numbers of offspring and the lower weights
of the individuals tend to offset each other. When prenatal or
postnatal growth is impaired by an acute exposure, compensatory growth
after cessation of dosing could obscure the earlier effect.
Postnatal weights are dependent on birth weight, sex, and normality
of the individual, as well as the litter size, lactational ability of
the dam, and suckling ability of the offspring. With large litters,
small or weak offspring may not compete successfully for milk and show
impaired growth. Because it is not possible usually to determine
whether the effect was due solely to the increased litter size, growth
retardation or decreased survival rate should be considered adverse in
the absence of information to the contrary. Also, offspring weights may
appear normal in very small litters and should be considered carefully
in relation to controls.
Offspring survival is dependent on the same factors as postnatal
weight, although more severe effects are necessary usually to affect
survival. All weight and survival endpoints can be affected by toxicity
of an agent, either by direct effects on the offspring or indirectly
through effects on the ability of the dam to support the offspring.
Measures of malformations and variations, as well as postnatal
structural and functional development, are presented in the Guidelines
for Developmental Toxicity Risk Assessment and the Proposed Guidelines
for Neurotoxicity Risk Assessment (U.S. EPA, 1991, 1995a). These
documents should be consulted for additional information on those
parameters.
Adverse Effects
Table 2 lists couple-mediated endpoints that may be measured in
reproduction studies. Table 3 presents examples of indices that may be
calculated from couple-mediated reproductive toxicity data. Significant
detrimental effects on any of those endpoints or on indices derived
from those data should be considered adverse. Whether effects are on
the female reproductive system or directly on the embryo or fetus is
often not distinguishable, but the distinction may not be important
because all of these effects should be cause for concern.
III.B.2.b. Sexual Behavior. Sexual behavior reflects complex
neural, endocrine, and reproductive organ interactions and is therefore
susceptible to disruption by a variety of toxic agents and pathologic
conditions. Interference with sexual behavior in either sex by
environmental agents represents a potentially significant human
reproductive problem. Most human information comes from studies on
effects of drugs on sexual behavior or from clinical reports in which
the detection of exposure-effect associations is unlikely. Data on
sexual behavior are usually not available from studies of human
populations that were exposed occupationally or environmentally to
potentially toxic agents, nor are such data obtained routinely in
studies of environmental agents with test species.
In the absence of human data, the perturbation of sexual behavior
in test species suggests the potential for similar effects on humans.
Consistent with this position are data showing that central nervous
system effects can disrupt sexual behavior in both test species and
humans (Rubin and Henson, 1979; Waller et al., 1985). Although the
functional components of sexual performance can be quantified in most
test species, no direct evaluation of this behavior is done in most
breeding studies. Rather, copulatory plugs or sperm-positive vaginal
lavages are taken as evidence of sexual receptivity and successful
mating. However, these markers do not demonstrate whether male
performance resulted in adequate sexual stimulation of the female.
Failure of the male to provide adequate stimulation to the female may
impair sperm transport in the genital tract of female rats, thereby
reducing the probability of successful impregnation (Adler and Toner,
1986). Such a ``mating'' failure would be reflected in the calculated
fertility index as reduced fertility and could be attributed
erroneously to an effect on the spermatogenic process in the male or on
fertility of the female.
In the rat, a direct measure of female sexual receptivity is the
occurrence of lordosis. Sexual receptivity of the female rat is
normally cyclic, with receptivity commencing during the late evening of
vaginal proestrus. Agents that interfere with normal estrous cyclicity
also could cause absence of or abnormal sexual behavior that can be
reflected in reduced numbers of females with vaginal plugs or vaginal
sperm, alterations in lordosis behavior, and increased time to mating
after start of cohabitation. In the male, measures include latency
periods to first mount, mount with intromission, and first ejaculation,
number of mounts with intromission to ejaculation, and the
postejaculatory interval (Beach, 1979).
Direct evaluation of sexual behavior is not warranted for all
agents being tested for reproductive toxicity. Some likely candidates
may be agents reported to exert central or peripheral neurotoxicity.
Chemicals possessing or suspected to possess androgenic or estrogenic
properties (or antagonistic properties) also merit consideration as
potentially causing adverse effects on sexual behavior concomitant with
effects on the reproductive organs.
Adverse Effects
Effects on sexual behavior (within the limited definition of these
Guidelines) should be considered as adverse reproductive effects.
Included is evidence of impaired sexual receptivity and copulatory
behavior. Impairment that is secondary to more generalized physical
debilitation (e.g., impaired rear leg motor activity or general
lethargy) should not be considered an adverse reproductive effect,
although such conditions represent adverse systemic effects.
III.B.3. Male-Specific Endpoints
III.B.3.a. Introduction. The following sections (III.B.3. and
III.B.4.) describe various male-specific and female-specific endpoints
of reproductive toxicity that can be obtained. Included are endpoints
for which data are obtained routinely by the Agency and other endpoints
for which data may be encountered in the review of chemicals. Guidance
is presented for interpretation of results involving these endpoints
and their use in risk assessment. Effects are identified that should be
considered as adverse reproductive effects if significantly different
from controls.
The Agency may obtain data on the potential male reproductive
toxicity of
[[Page 56287]]
an agent from many sources including, but not limited to, studies done
according to Agency test guidelines. These may include acute,
subchronic, and chronic testing and reproduction and fertility studies.
Male-specific endpoints that may be encountered in such studies are
identified in Table 4.
Table 4.--Male-Specific Endpoints of Reproductive Toxicity
------------------------------------------------------------------------
------------------------------------------------------------------------
Organ weights................ Testes, epididymides, seminal vesicles,
prostate, pituitary.
Visual examination and Testes, epididymides, seminal vesicles,
histopathology. prostate, pituitary.
Sperm evaluation *........... Sperm number (count) and quality
(morphology, motility)
Sexual behavior *............ Mounts, intromissions, ejaculations.
Hormone levels *............. Luteinizing hormone, follicle stimulating
hormone, testosterone, estrogen,
prolactin.
Developmental effects........ Testis descent*, preputial separation,
sperm production*, ano-genital distance,
structure of external genitalia*.
------------------------------------------------------------------------
* Reproductive endpoints that can be obtained or estimated relatively
noninvasively with humans.
III.B.3.b. Body Weight and Organ Weights. Monitoring body weight
during treatment provides an index of the general health status of the
animals, and such information may be important for the interpretation
of reproductive effects (see also Section III.B.2.). Depression in body
weight or reduction in weight gain may reflect a variety of responses,
including rejection of chemical-containing food or water because of
reduced palatability, treatment-induced anorexia, or systemic toxicity.
Less than severe reductions in adult body weight induced by restricted
nutrition have shown little effect on the male reproductive organs or
on male reproductive function (Chapin et al., 1993a, b). When a
meaningful, biologic relationship between a body weight decline and a
significant effect on the male reproductive system is not apparent, it
is not appropriate to dismiss significant alteration of the male
reproductive system as secondary to the occurrence of nonreproductive
toxicity. Unless additional data provide the needed clarification,
alteration in a reproductive measure that would otherwise be considered
adverse should still be considered as an adverse male reproductive
effect in the presence of mild to moderate body weight changes. In the
presence of severe body weight depression or other severe systemic
debilitation, it should be noted that an adverse effect on a
reproductive endpoint occurred, but the effect may have resulted from a
more generalized toxic effect. Regardless, adverse effects would have
been observed in that situation and a risk assessment should be pursued
if sufficient data are available.
The male reproductive organs for which weights may be useful for
reproductive risk assessment include the testes, epididymides,
pituitary gland, seminal vesicles (with coagulating glands), and
prostate. Organ weight data may be presented as both absolute weights
and as relative weights (i.e., organ weight to body weight ratios).
Organ weight data may also be reported relative to brain weight since,
subsequent to development, the weight of the brain usually remains
quite stable (Stevens and Gallo, 1989). Evaluation of data on absolute
organ weights is important, because a decrease in a reproductive organ
weight may occur that was not necessarily related to a reduction in
body weight gain. The organ weight-to-body weight ratio may show no
significant difference if both body weight and organ weight change in
the same direction, masking a potential organ weight effect.
Normal testis weight varies only modestly within a given test
species (Schwetz et al., 1980; Blazak et al., 1985). This relatively
low interanimal variability suggests that absolute testis weight should
be a precise indicator of gonadal injury. However, damage to the testes
may be detected as a weight change only at doses higher than those
required to produce significant effects in other measures of gonadal
status (Berndtson, 1977; Foote et al., 1986; Ku et al., 1993). This
contradiction may arise from several factors, including a delay before
cell deaths are reflected in a weight decrease (due to preceding edema
and inflammation, cellular infiltration) or Leydig cell hyperplasia.
Blockage of the efferent ducts by cells sloughed from the germinal
epithelium or the efferent ducts themselves can lead to an increase in
testis weight due to fluid accumulation (Hess et al., 1991; Nakai et
al., 1993), an effect that could offset the effect of depletion of the
germinal epithelium on testis weight. Thus, while testis weight
measurements may not reflect certain adverse testicular effects and do
not indicate the nature of an effect, a significant increase or
decrease is indicative of an adverse effect.
Pituitary gland weight can provide valuable insight into the
reproductive status of the animal. However, the pituitary contains cell
types that are responsible for the regulation of a variety of
physiologic functions including some that are separate from
reproduction. Thus, changes in pituitary weight may not necessarily
reflect reproductive impairment. If weight changes are observed,
gonadotroph-specific histopathologic evaluations may be useful in
identifying the affected cell types. This information may then be used
to judge whether the observed effect on the pituitary is related to
reproductive system function and therefore an adverse reproductive
effect.
Prostate and seminal vesicle weights are androgen-dependent and may
reflect changes in the animal's endocrine status or testicular
function. Separation of the seminal vesicles and coagulating gland
(dorsal prostate) is difficult in rodents. However, the seminal vesicle
and prostate can be separated and results may be reported for these
glands separately or together, with or without their secretory fluids.
Differential loss of secretory fluids prior to weighing could produce
artifactual weights. Because the seminal vesicles and prostate may
respond differently to an agent (endocrine dependency and developmental
susceptibility differ), more information may be gained if the weights
were examined separately.
Adverse Effects
Significant changes in absolute or relative male reproductive organ
weights may constitute an adverse reproductive effect. Such changes
also may provide a basis for obtaining additional information on the
reproductive toxicity of that agent. However, significant changes in
other important endpoints that are related to reproductive function may
not be reflected in organ weight data. Therefore, lack of an organ
weight effect should not be used to negate significant changes in other
endpoints that may be more sensitive.
III.B.3.c. Histopathologic Evaluations. Histopathologic evaluations
of test animal tissues have a prominent role in male reproductive risk
assessment. Organs that are often evaluated include
[[Page 56288]]
the testes, epididymides, prostate, seminal vesicles (often including
coagulating glands), and pituitary. Tissues from lower dose exposures
are often not examined histologically if the high dose produced no
difference from controls. Histologic evaluations can be especially
useful by (1) providing a relatively sensitive indicator of damage; (2)
providing information on toxicity from a variety of protocols; and (3)
with short-term dosing, providing information on site (including target
cells) and extent of toxicity; and (4) indicating the potential for
recovery.
The quality of the information presented from histologic analyses
of spermatogenesis is improved by proper fixation and embedding of
testicular tissue. With adequately prepared tissue (Chapin, 1988;
Russell et al., 1990; Hess and Moore, 1993), a description of the
nature and background level of lesions in control tissue, whether
preparation-induced or otherwise, can facilitate interpreting the
nature and extent of the lesions observed in tissues obtained from
exposed animals. Many histopathologic evaluations of the testis only
detect lesions if the germinal epithelium is severely depleted or
degenerating, if multinucleated giant cells are obvious, or if sloughed
cells are present in the tubule lumen. More subtle lesions, such as
retained spermatids or missing germ cell types, that can significantly
affect the number of sperm being released normally into the tubule
lumen may not be detected when less adequate methods of tissue
preparation are used. Also, familiarity with the detailed morphology of
the testis and the kinetics of spermatogenesis of each test species can
assist in the identification of less obvious lesions that may accompany
lower dose exposures or lesions that result from short-term exposure
(Russell et al., 1990). Several approaches for qualitative or
quantitative assessment of testicular tissue are available that can
assist in the identification of less obvious lesions that may accompany
lower-dose exposures, including use of the technique of ``staging.'' A
book is available (Russell et al., 1990) which provides extensive
information on tissue preparation, examination, and interpretation of
observations for normal and high resolution histology of the germinal
epithelium of rats, mice, and dogs. Included is guidance for
identification and quantification of the various cell types and
associations for each stage of the spermatogenic cycle. Also, a
decision-tree scheme for staging with the rat has been published (Hess,
1990).
The basic morphology of other male reproductive organs (e.g.,
epididymides, accessory sex glands, and pituitary) has been described
as well as the histopathologic alterations that may accompany certain
disease states (Fawcett, 1986; Jones et al., 1987; Haschek and
Rousseaux, 1991). Compared with the testes, less is known about
structural changes in these tissues that are associated with exposure
to toxic agents. With the epididymides and accessory sex glands,
histologic evaluation is usually limited to the height and possibly the
integrity of the secretory epithelium. Evaluation should include
information on the caput, corpus, and cauda segments of the epididymis.
Presence of debris and sloughed cells in the epididymal lumen are
valuable indicators of damage to the germinal epithelium or the
excurrent ducts. The presence of lesions such as sperm granulomas,
leucocyte infiltration (inflammation) or absence of clear cells in the
cauda epididymal epithelium should be noted. Information from
examinations of the pituitary should include evaluation of the
morphology of the cell types that produce the gonadotropins and
prolactin.
The degree to which histopathologic effects are quantified is
usually limited to classifying animals, within dose groups, as either
affected or not affected by qualitative criteria. Little effort has
been made to quantify the extent of injury, and procedures for such
classifications are not applied uniformly (Linder et al., 1990).
Evaluation procedures would be facilitated by adoption of more uniform
approaches for quantifying the extent of histopathologic damage per
individual. In the absence of standardized tissue preparation
techniques and a standardized quantification system, the evaluation of
histopathologic data would be facilitated by the presentation of the
evaluation criteria and procedure by which the level of lesions in
exposed individuals was judged to be in excess of controls.
If properly obtained (i.e., proper preparation and analysis of
tissue), data from histopathologic evaluations may provide a relatively
sensitive tool that is useful for detection of low-dose effects. This
approach may also provide insight into sites and mechanisms of action
for the agent on that reproductive organ. When similar targets or
mechanisms exist in humans, the basis for interspecies extrapolation is
strengthened. Depending on the experimental design, information can
also be obtained that may allow prediction of the eventual extent of
injury and degree of recovery in that species and humans (Russell,
1983).
Adverse Effects
Significant and biologically meaningful histopathologic damage in
excess of the level seen in control tissue of any of the male
reproductive organs should be considered an adverse reproductive
effect. Significant histopathologic damage in the pituitary should be
considered as an adverse effect but should be shown to involve cells
that control gonadotropin or prolactin production to be called a
reproductive effect. Although thorough histopathologic evaluations that
fail to reveal any treatment-related effects may be quite convincing,
consideration should be given to the possible presence of other
testicular or epididymal effects that are not detected histologically
(e.g., genetic damage to the germ cell, decreased sperm motility), but
may affect reproductive function.
III.B.3.d. Sperm Evaluations. The parameters that are important for
sperm evaluations are sperm number, sperm morphology, and sperm
motility. Data on those parameters allow more adequate estimation of
the number of ``normal'' sperm; a parameter that is likely to be more
informative than sperm number alone. Although effects on sperm
production can be reflected in other measures such as testicular
spermatid count or cauda epididymal weight, no surrogate measures are
adequate to reflect effects on sperm morphology or motility. Similar
data can be obtained noninvasively from human ejaculates, enhancing the
ability to confirm effects seen in test species or to detect effects in
humans. Brief descriptions of these measures are provided below,
followed by a discussion of the use of various sperm measures in male
reproductive risk assessment.
Sperm Number
Measures of sperm concentration (count) have been the most
frequently reported semen variable in the literature on humans (Wyrobek
et al., 1983a). Sperm number or sperm concentration from test species
may be derived from ejaculated, epididymal, or testicular samples (Seed
et al., 1996). Of the common test species, ejaculates can only be
obtained readily from rabbits or dogs. Ejaculates can be recovered from
the reproductive tracts of mated females of other species (Zenick et
al., 1984). Measures of human sperm production are usually derived from
ejaculates, but could also be obtained from spermatid counts or
quantitative histology using testicular biopsy tissue samples. With
[[Page 56289]]
ejaculates, both sperm concentration (number of sperm/mL of ejaculate)
and total sperm per ejaculate (sperm concentration x volume) should be
evaluated.
Ejaculated sperm number from any species is influenced by several
variables, including the length of abstinence and the ability to obtain
the entire ejaculate. Intra- and interindividual variation are often
high, but are reduced somewhat if ejaculates were collected at regular
intervals from the same male (Williams et al., 1990). Such a
longitudinal study design has improved detection sensitivity and thus
requires a smaller number of subjects (Wyrobek et al., 1984). In
addition, if a pre-exposure baseline is obtained for each male (test
animal or human studies when allowed by protocol), then changes during
exposure or recovery can be better defined.
Epididymal sperm evaluations with test species usually use sperm
from only the cauda portion of the epididymis, but the samples for
sperm motility and morphology may be derived also from the vas
deferens. It has been customary to express the sperm count in relation
to the weight of the cauda epididymis. However, because sperm
contribute to epididymal weight, expression of the data as a ratio may
actually mask declines in sperm number. The inclusion of data on
absolute sperm counts can improve resolution. As is true for ejaculated
sperm counts, epididymal sperm counts are influenced directly by level
of sexual activity (Amann, 1981; Hurtt and Zenick, 1986).
Sperm production data may be derived from counts of the distinctive
elongated spermatid nuclei that remain after homogenization of testes
in a detergent-containing medium (Amann, 1981; Meistrich, 1982; Cassidy
et al., 1983; Blazak et al., 1993). The elongated spermatid counts are
a measure of sperm production from the stem cells and their ensuing
survival through spermatocytogenesis and spermiogenesis (Meistrich,
1982; Meistrich and van Beek, 1993). If evaluation was conducted when
the effect of a lesion would be reflected adequately in the spermatid
count, then spermatid count may serve as a substitute for quantitative
histologic analysis of sperm production (Russell et al., 1990).
However, spermatid counts may be misleading when duration of exposure
is shorter than the time required for a lesion to be fully expressed in
the spermatid count. Also, spermatid counts reported from some
laboratories have large coefficients of variation that may reduce the
statistical power and thus the usefulness of that measure.
The ability to detect a decrease in testicular sperm production may
be enhanced if spermatid counts are available. However, spermatid
enumerations only reflect the integrity of spermatogenic processes
within the testes. Posttesticular effects or toxicity expressed as
alterations in motility, morphology, viability, fragility, and other
properties of sperm can be determined only from epididymal, vas
deferens, or ejaculated samples.
Sperm Morphology
Sperm morphology refers to structural aspects of sperm and can be
evaluated in cauda epididymal, vas deferens, or ejaculated samples. A
thorough morphologic evaluation identifies abnormalities in the sperm
head and flagellum. Because of the suggested correlation between an
agent's mutagenicity and its ability to induce abnormal sperm, sperm
head morphology has been a frequently reported sperm variable in
toxicologic studies on test species (Wyrobek et al., 1983b). The
tendency has been to conclude that increased incidence of sperm head
malformations reflects germ-cell mutagenicity. However, not every
mutagen induces sperm head abnormalities, and other nonmutagenic
chemicals may alter sperm head morphology. For example, microtubule
poisons may cause increases in abnormal sperm head incidence,
presumably by interfering with spermiogenesis, a microtubule-dependent
process (Russell et al., 1981). Sperm morphology may be altered also
due to degeneration subsequent to cell death. Thus, the link between
sperm morphology and mutagenicity is not necessarily sensitive or
specific.
An increase in abnormal sperm morphology has been considered
evidence that the agent has gained access to the germ cells (U.S. EPA,
1986c). Exposure of males to toxic agents may lead to sperm
abnormalities in their progeny (Wyrobek and Bruce, 1978; Hugenholtz and
Bruce, 1983; Morrissey et al., 1988a, b). However, transmissible germ-
cell mutations might exist in the absence of any warning morphologic
indicator such as abnormal sperm. The relationships between these
morphologic alterations and other karyotypic changes remains uncertain
(de Boer et al., 1976).
The traditional approach to characterizing morphology in
toxicologic testing has relied on subjective categorization of sperm
head, midpiece, and tail defects in either stained preparations by
bright field microscopy (Filler, 1993) or fixed, unstained preparations
by phase contrast microscopy (Linder et al., 1992; Seed et al., 1996).
Such an approach may be adequate for mice and rats with their
distinctly angular head shapes. However, the observable heterogeneity
of structure in human sperm and in nonrodent species makes it difficult
for the morphologist to define clearly the limits of normality. More
systematic, quantitative, and automated approaches have been offered
that can be used with humans and test species (Katz et al., 1982;
Wyrobek et al., 1984). Data that categorize the types of abnormalities
observed and quantify the frequencies of their occurrences are
preferred to estimation of overall proportion of abnormal sperm.
Objective, quantitative approaches that are done properly should result
in a higher level of confidence than more subjective measures.
Sperm morphology profiles are relatively stable and characteristic
in a normal individual (and a strain within a species) over time. Sperm
morphology is one of the least variable sperm measures in normal
individuals, which may enhance its use in the detection of
spermatotoxic events (Zenick et al., 1994). However, the reproductive
implications of the various types of abnormal sperm morphology need to
be delineated more fully. The majority of studies in test species and
humans have suggested that abnormally shaped sperm may not reach the
oviduct or participate in fertilization (Nestor and Handel, 1984; Redi
et al., 1984). The implication is that the greater the number of
abnormal sperm in the ejaculate, the greater the probability of reduced
fertility.
Sperm Motility
The biochemical environments in the testes and epididymides are
highly regulated to assure the proper development and maturation of the
sperm and the acquisition of critical functional characteristics, i.e.,
progressive motility and the potential to fertilize. With chemical
exposures, perturbation of this balance may occur, producing
alterations in sperm properties such as motility. Chemicals (e.g.,
epichlorohydrin) have been identified that selectively affect sperm
motility and also reduce fertility. Studies have examined rat sperm
motility as a reproductive endpoint (Morrissey et al., 1988a, b; Toth
et al., 1989b, 1991b), and sperm motility assessments are an integral
part of some reproductive toxicity tests (Gray et al., 1988; Morrissey
et al., 1989; U.S. EPA, 1996a).
[[Page 56290]]
Motility estimates may be obtained on ejaculated, vas deferens, or
cauda epididymal samples. Standardized methods are needed because
motility is influenced by a number of experimental variables, including
abstinence interval, method of sample collection and handling, elapsed
time between sampling and observation, the temperature at which the
sample is stored and analyzed, the extent of sperm dilution, the nature
of the dilution medium, and the microscopic chamber employed for the
observations (Slott et al., 1991; Toth et al., 1991a; Chapin et al.,
1992; Schrader et al., 1992; Weir and Rumberger, 1995; Seed et al.,
1996).
Sperm motility can be evaluated in fresh samples under phase
contrast microscopy, or sperm images can be recorded and stored in
video or digital format and analyzed later, either manually or by
computer-aided semen analysis (Linder et al., 1986; Boyers et al.,
1989; Toth et al., 1989a; Yeung et al., 1992; Slott and Perreault,
1993). For manual assessments, the percentage of motile and
progressively motile sperm can be estimated and a simple scale used to
describe the vigor of the sperm motion.
The recent application of video and/or digital technology to sperm
analysis allows a more detailed evaluation of sperm motion including
information about the individual sperm tracks. It also provides
permanent storage of the sperm tracks which can be re-analyzed as
necessary (manually or computer-assisted). With computer-assisted
technology, information about sperm velocity (straight-line and
curvilinear) as well as the amplitude and frequency of the track are
obtained rapidly and efficiently on large numbers of sperm. Using this
technology, chemically induced alterations in sperm motion have been
detected (Toth et al., 1989a, 1992; Slott et al., 1990; Klinefelter et
al., 1994a), and such changes have been related to the fertility of the
exposed animals (Toth et al., 1991a; Oberlander et al., 1994; Slott et
al., 1995). These preliminary studies indicate that significant
reductions in sperm velocity are associated with infertility, even when
the percentage of motile sperm is not affected. The ability to
distinguish between the proportion of sperm showing any type of motion
and those with progressive motility is important (Seed et al., 1996).
Changes in endpoints that measure effects on spermatogenesis and
sperm maturation have been related to fertility in several test
species, but the ability to predict infertility from these data (in the
absence of fertility data) is not reliable. This is in part due to the
observation, in both test species and humans, that fertility is
dependent not only on having adequate numbers of sperm, but also on the
degree to which those sperm are normal. If sperm quality is high, then
sperm number must be substantially reduced before fertility is
affected. For example, in a rat model that employs artificial
insemination of differing numbers of good quality sperm, sperm numbers
can be reduced substantially before fertility is affected (Klinefelter
et al., 1994b). In humans, the distribution of sperm counts for fertile
and infertile men overlap, with the mean for fertile men being higher
(Meistrich and Brown, 1983), but fertility is likely to be impaired
when counts drop below 20 million/mL (WHO, 1992). Similarly, if sperm
numbers are normal in rodents, a relatively large effect on sperm
motility is required before fertility is affected. For example, rodent
sperm velocity must be substantially reduced, in the presence of
adequate numbers of sperm, before fertility is affected (Toth et al.,
1991a; Slott et al., 1995). These models also show that relatively
modest changes in sperm numbers or quality may not cause infertility,
but can nevertheless be predictive of infertility. On the other hand,
fertility may be impaired by smaller decrements in both number and
motility (or other qualitative characteristics).
Thus, the process of reproductive risk assessment is facilitated by
having information on a variety of sperm measures and reproductive
organ histopathology in addition to fertility. Specific information
about reproductive organ and gamete function can then be used to
evaluate the occurrence and extent of injury, and the probable site of
toxicity in the reproductive system. The more information that is
available from supplementary endpoints, the more the risk assessment
can be based on science rather than uncertainty.
Adverse Effects
Human male fertility is generally lower than that of test species
and may be more susceptible to damage from toxic agents (see
Supplementary Information). Therefore, the conservative approach should
be taken that, within the limits indicated in the sections on those
parameters, statistically significant changes in measures of sperm
count, morphology, or motility as well as number of normal sperm should
be considered adverse effects.
III.B.3.e. Paternally Mediated Effects on Offspring. The concept is
well accepted that exposure of a female to toxic chemicals during
gestation or lactation may produce death, structural abnormalities,
growth alteration, or postnatal functional deficits in her offspring.
Sufficient data now exist with a variety of agents to conclude that
male-only exposure also can produce deleterious effects in offspring
(Davis et al., 1992; Colie, 1993; Savitz et al., 1994; Qiu et al.,
1995). Paternally mediated effects include pre- and postimplantation
loss, growth and behavioral deficits, and malformations. A large
proportion of the chemicals reported to cause paternally mediated
effects have genotoxic activity, and are considered to exert this
effect via transmissible genetic alterations. Low doses of
cyclophosphamide have resulted in induction of single strand DNA breaks
during rat spermatogenesis which, due in part to absence of subsequent
DNA repair capability, remain at fertilization (Qiu et al., 1995). The
results of such damage have been observed in the F2 generation
offspring (Hales et al., 1992). Other mechanisms of induction of
paternally mediated effects are also possible. Xenobiotics present in
seminal plasma or bound to the fertilizing sperm could be introduced
into the female genital tract, or even the oocyte directly, and might
also interfere with fertilization or early development. With humans,
the possibility exists that a parent could transport the toxic agent
from the work environment to the home (e.g., on work clothes), exposing
other adults or children. Further work is needed to clarify the extent
to which paternal exposures may be associated with adverse effects on
offspring. Regardless, if an agent is identified in test species or in
humans as causing a paternally mediated adverse effect on offspring,
the effect should be considered an adverse reproductive effect.
III.B.4. Female-Specific Endpoints
III.B.4.a. Introduction. The reproductive life cycle of the female
may be divided into phases that include fetal, prepubertal, cycling
adult, pregnant, lactating, and reproductively senescent. Detailed
descriptions of all phases are available (Knobil et al., 1994). It is
important to detect adverse effects occurring in any of these stages.
Traditionally, the endpoints that have been used have emphasized
ability to become pregnant, pregnancy outcome, and offspring survival
and development. Although reproductive organ weights may be obtained
and these organs examined histologically in test species, these
measures do not necessarily detect abnormalities in dynamic processes
such as estrous cyclicity or follicular atresia unless degradation is
severe. Similarly, toxic effects on onset of
[[Page 56291]]
puberty have not been examined, nor have the long-term consequences of
exposure on reproductive senescence. Thus, the amount of information
obtained routinely to detect toxic effects on the female reproductive
system has been limited.
The consequences of impairment in the nonpregnant female
reproductive system are equally important, and endpoints to detect
adverse effects on the nonpregnant reproductive system, when available,
can be useful in evaluating reproductive toxicity. Such measures may
also provide additional interrelated endpoints and information on
mechanism of action.
Adverse alterations in the nonpregnant female reproductive system
have been observed at dose levels below those that result in reduced
fertility or produce other overt effects on pregnancy or pregnancy
outcomes (Le Vier and Jankowiak, 1972; Barsotti et al., 1979; Sonawane
and Yaffe, 1983; Cummings and Gray, 1987). In contrast to the male
reproductive system, the status of the normal female system fluctuates
in adults. Thus, in nonpregnant animals (including humans), the ovarian
structures and other reproductive organs change throughout the estrous
or menstrual cycle. Although not cyclic, normal changes also accompany
the progression of pregnancy, lactation, and return to cyclicity during
or after lactation. These normal fluctuations may affect the endpoints
used for evaluation. Therefore, knowledge of the reproductive status of
the female at necropsy, including the stage of the estrous cycle, can
facilitate detection and interpretation of effects with endpoints such
as uterine weight and histopathology of the ovary and uterus. Necropsy
of all test animals at the same stage of the estrous cycle can reduce
the variance of test results with such measures.
A variety of measures to evaluate the integrity of the female
reproductive system has been used in toxicity studies. With appropriate
measures, a comprehensive evaluation of the reproductive process can be
achieved, including identification of target organs and possible
elucidation of the mechanisms involved in the agent's effect(s). Areas
that may be examined in evaluations of the female reproductive system
are listed in Table 5.
Table 5.--Female-Specific Endpoints of Reproductive Toxicity
----------------------------------------------------------------------------------------------------------------
----------------------------------------------------------------------------------------------------------------
Organ weights............................................................... Ovary, uterus, vagina, pituitary.
Visual examination and histopathology....................................... Ovary, uterus, vagina, pituitary,
oviduct, mammary gland.
Estrous (menstrual *) cycle normality....................................... Vaginal smear cytology.
Sexual behavior............................................................. Lordosis, time to mating, vaginal
plugs, or sperm.
Hormone levels *............................................................ LH, FSH, estrogen, progesterone,
prolactin.
Lactation *................................................................. Offspring growth, milk quantity
and quality.
Development................................................................. Normality of external genitalia *,
vaginal opening, vaginal smear
cytology, onset of estrous
behavior (menstruation *).
Senescence.................................................................. Vaginal smear cytology, ovarian
histology (menopause *).
----------------------------------------------------------------------------------------------------------------
* Endpoints that can be obtained relatively noninvasively with humans.
Reproductive function in the female is controlled through complex
interactions involving the central nervous system (particularly the
hypothalamus), pituitary, ovaries, the reproductive tract, and the
secondary sexual organs. Other nongonadotrophic components of the
endocrine system may also modulate reproductive system function.
Because it is difficult to measure certain important aspects of female
reproductive function (e.g., increased rate of follicular atresia,
ovulation failure), assessment of the endocrine status may provide
needed insight that is not otherwise available.
To understand the significance of effects on the reproductive
endpoints, it is critical that the relationships between the various
reproductive hormones and the female reproductive organs be understood.
Although certain effects may be identified routinely as adverse, all of
the results should be considered in the context of the known biology.
The format used below for presentation of the female reproductive
endpoints is altered from that used for the male to allow examination
of events that are linked and that fluctuate with the changing
endocrine status. Particularly, the organ weight, gross morphology, and
histology are combined for each organ. Endpoints and endocrine factors
for the individual female reproductive organs are discussed, with
emphasis on the nonpregnant animal. This is followed by examination of
measures of cyclicity and their interpretation. Then, considerations
relevant to prepubertal, pregnant, lactating, and aging females are
presented.
III.B.4.b. Body Weight, Organ Weight, Organ Morphology, and Histology
III.B.4.b.1. Body weight. Toxicologists are often concerned about
how a change in body weight may affect reproductive function. In
females, an important consideration is that body weight fluctuates
normally with the physiologic state of the animal because estrogen and
progesterone are known to influence food intake and energy expenditure
to an important extent (Wang, 1923; Wade, 1972). Water retention and
fat deposition rates are also affected (Galletti and Klopper, 1964;
Hervey and Hervey, 1967). Food consumption is elevated during
pregnancy, in part because of the elevated serum progesterone level.
One of the most sensitive noninvasive indicators of a compound with
estrogenic action in the female rat is a reduction in food intake and
body weight. Also, growth retardation induced by effects on
extragonadal hormones (e.g., thyroid or growth hormone) can cause a
delay in pubertal development, and induce acyclicity and infertility.
Because of these endocrine-related fluctuations, the weights of the
reproductive organs are poorly correlated with body weight, except in
extreme cases. Thus, actual organ weight data, rather than organ to
body weight ratios, should be reported and evaluated for the female
reproductive system.
Chapin et al. (1993a, b) have studied the influence of food
restriction on female Sprague-Dawley rats and Swiss CD-1 mice when body
weights were 90%, 80%, or 70% of controls. Female rats were resistant
to effects on reproductive function at 80% of control weight whereas
mice showed adverse effects at 80% and a marginal effect at 90%. These
results indicate that differences exist between species (and probably
between strains) in the response of the female rodent reproductive
system to reduced food intake or body weight reduction.
III.B.4.b.2. Ovary. The ovary serves a number of functions that are
critical to reproductive activity, including production and ovulation
of oocytes.
[[Page 56292]]
Estrogen is produced by developing follicles and progesterone is
produced by corpora lutea that are formed after ovulation.
Ovarian Weight
Significant increases or decreases in ovarian weight compared with
controls should be considered an indication of female reproductive
toxicity. Although ovarian function shifts throughout the estrous
cycle, ovarian weight in the normal rat does not show significant
fluctuations. Still, oocyte and follicle depletion, persistent
polycystic ovaries, inhibition of corpus luteum formation, luteal cyst
development, reproductive aging, and altered hypothalamic-pituitary
function may all be associated with changes in ovarian weight.
Therefore, it is important that ovarian gross morphology and histology
also be examined to allow correlation of alterations in those
parameters with changes in ovarian weight. However, not all adverse
histologic alterations in the ovary are concurrent with changes in
ovarian weight. Therefore, a lack of effect on organ weights does not
preclude the need for histologic evaluation.
Histopathology
Histologic evaluation of the three major compartments of the ovary
(i.e., follicular, luteal, and interstitial) plus the epithelial
capsule and ovarian stroma may indicate ovarian toxicity. A number of
pathologic conditions can be detected by ovarian histology (Kurman and
Norris, 1978; Langley and Fox, 1987). Methods are available to quantify
the number of follicles and their stages of maturation (Plowchalk et
al., 1993). These techniques may be useful when a compound depletes the
pool of primordial follicles or alters their subsequent development and
recruitment during the events leading to ovulation.
Adverse Effects
Significant changes in the ovaries in any of the following effects
should be considered adverse:
Increase or decrease in ovarian weight.
Increased incidence of follicular atresia.
Decreased number of primary follicles.
Decreased number or lifespan of corpora lutea.
Evidence of abnormal folliculogenesis or luteinization,
including cystic follicles, luteinized follicles, and failure of
ovulation.
Evidence of altered puberty or premature reproductive
senescence.
III.B.4.b.3. Uterus.
Uterine Weight
An alteration in the weight of the uterus may be considered an
indication of female reproductive organ toxicity. Compounds that
inhibit steroidogenesis and cyclicity can dramatically reduce the
weight of the uterus so that it appears atrophic and small. However,
uterine weight fluctuates three- to four-fold throughout the estrous
cycle, peaking at proestrus when, in response to increased estrogen
secretion, the uterus is fluid filled and distended. This increase in
uterine weight has been used as a basis for comparing relative potency
of estrogenic compounds in bioassays (Kupfer, 1987). As a result of the
wide fluctuations in weight, uterine weights taken from cycling animals
have a high variance, and large compound-related effects are required
to demonstrate a significant effect unless interpreted relative to that
animal's estrous cycle stage. A number of environmental compounds
(e.g., pesticides such as methoxychlor and chlordecone, mycotoxins,
polychlorinated biphenyls, alkylphenols, and phytoestrogens) possess
varying degrees of estrogenic activity and have the potential to
stimulate the female reproductive tract (Barlow and Sullivan, 1982;
Bulger and Kupfer, 1985; Hughes, 1988).
When pregnant or postpartum animals are examined, the numbers of
implantation sites or implantation scars should be counted. This
information, along with corpus luteum counts, can be used to calculate
pre- and postimplantation losses.
Histopathology
The histologic appearance of the normal uterus fluctuates with
stage of the estrous cycle and pregnancy. The uterine endometrium is
sensitive to influences of estrogens and progestogens (Warren et al.,
1967), and extended treatment with these compounds leads to hypertrophy
and hyperplasia. Conversely, inhibition of ovarian activity and reduced
steroid secretion results in endometrial hypoplasia and atrophy, as
well as altered vaginal smear cytology. Effects induced during
development may delay or prevent puberty, resulting in persistence of
infantile genitalia.
Adverse Effects
Effects on the uterus that may be considered adverse include
significant dose-related alteration of weight, as well as gross
anatomic or histologic abnormalities. In particular, any of the
following effects should be considered as adverse.
Infantile or malformed uterus or cervix.
Decreased or increased uterine weight.
Endometrial hyperplasia, hypoplasia, or aplasia.
Decreased number of implantation sites.
III.B.4.b.4. Oviducts.
Typically, the oviducts are not weighed or examined histologically
in tests for reproductive toxicity. However, information from visual
and histologic examinations is of value in detecting morphologic
anomalies. Descriptions of pathologic effects within the oviducts of
animals other than humans are not common. Hypoplasia of otherwise well-
formed oviducts and loss of cilia result most commonly from a lack of
estrogen stimulation, and for this reason, this condition may not be
recognized until after puberty. Hyperplasia of the oviductal epithelium
results from prolonged estrogenic stimulation. Anomalies induced during
development have also been described, including agenesis, segmental
aplasia, and hypoplasia.
Anatomic anomalies in the oviduct occurring in excess of control
incidence should be considered as adverse effects. Hypoplasia or
hyperplasia of the oviductal epithelium may be considered as an adverse
effect, particularly if that result is consistent with observations in
the uterine histology.
III.B.4.b.5. Vagina and external genitalia.
Vaginal Weight
Vaginal weight changes should parallel those seen in the uterus
during the estrous cycle, although the magnitude of the changes is
smaller.
Histopathology
In rodents, cytologic changes in the vaginal epithelium (vaginal
smear) may be used to identify the different stages of the estrous
cycle (see Section III.B.4.d.). The vaginal smear pattern may be useful
to identify conditions that would delay or preclude fertility, or
affect sexual behavior. Other histologic alterations that may be
observed include aplasia, hypoplasia, and hyperplasia of the vaginal
epithelial cell lining.
Developmental Effects
Developmental abnormalities, either genetic or related to prenatal
exposure to compounds that disrupt the endocrine balance, include
agenesis, hypoplasia, and dysgenesis. Hypoplasia of the vagina may be
concomitant with hyperplasia of the external genitalia and can be
induced by gonadal or adrenal steroid exposure. In rodents,
[[Page 56293]]
malpositioning of the vaginal and urethral ducts is common in steroid-
treated females. Such developmentally induced lesions are irreversible.
The sex ratio observed at birth may be affected by exposure of
genotypic females in utero to agents that disrupt reproductive tract
development. In cases of incomplete sex reversal because of such
exposures, female rodents may appear more male-like and have an
increased ano-genital distance (Gray and Ostby, 1995).
At puberty, the opening of the vaginal orifice normally provides a
simple and useful developmental marker. However, estrogenic or
antiestrogenic chemicals can act directly on the vaginal epithelium and
alter the age at which vaginal patency occurs without truly affecting
puberty.
Adverse Effects
Significant effects on the vagina that may be considered adverse
include the following:
Increases or decreases in weight
Infantile or malformed vagina or vulva, including
masculinized vulva or increased ano-genital distance
Vaginal hypoplasia or aplasia
Altered timing of vaginal opening
Abnormal vaginal smear cytology pattern
III.B.4.b.6. Pituitary.
Pituitary Weight
Alterations in weight of the pituitary gland should be considered
an adverse effect. The discussion on pituitary weight and histology for
males (see Section III.B.3.b.) is pertinent also for females. Pituitary
weight increases normally with age, as well as during pregnancy and
lactation. Changes in pituitary weight can occur also as a consequence
of chemical stimulation. Increased pituitary weight often precedes
tumor formation, particularly in response to treatment with estrogenic
compounds. Increased pituitary size associated with estrogen treatment
may be accompanied by hyperprolactinemia and constant vaginal estrus.
Decreased pituitary weight is less common but may result from decreased
estrogenic stimulation (Cooper et al., 1989).
Histopathology
In histologic evaluations with rats and mice, the relative size of
cell types in the anterior pituitary (acidophils and basophils) has
been reported to vary with the stages of the reproductive cycle and in
pregnancy (Holmes and Ball, 1974). Therefore, the relationship of
morphologic pattern to estrous or menstrual cycle stage or pregnancy
status should be considered in interpreting histologic observations on
the female pituitary.
Adverse Effects
A significant increase or decrease in pituitary weight should be
considered an adverse effect. Significant histopathologic damage in the
pituitary should be considered an adverse effect, but should be shown
to involve cells that control gonadotropin or prolactin production to
be called a reproductive effect.
III.B.4.c. Oocyte Production
III.B.4.c.1. Folliculogenesis. In normal females, all of the
follicles (and the resident oocytes) are present at or soon after
birth. The large majority of these follicles undergo atresia and are
not ovulated. If the population of follicles is depleted, it cannot be
replaced and the female will be rendered infertile. In humans,
depletion of oocytes leads to premature menopause. Ovarian follicle
biology and toxicology have been reviewed by Crisp (1992).
In rodents, lead, mercury, cadmium, and polyaromatic hydrocarbons
have all been implicated in the arrest of follicular growth at various
stages of the life cycle (Mattison and Thomford, 1989). Susceptibility
to oocyte toxicity varies considerably between species (Mattison and
Thorgeirsson, 1978).
Environmental agents that affect gonadotropin-mediated ovarian
steroidogenesis or follicular maturation can prolong the follicular
phase of the estrous or menstrual cycle and cause atresia of follicles
that would otherwise ovulate. Estrogenic as well as antiestrogenic
agents can produce this effect. Also, normal follicular maturation is
essential for normal formation and function of the corpus luteum formed
after ovulation (McNatty, 1979).
III.B.4.c.2. Ovulation. Chemicals can delay or block ovulation by
disrupting the ovulatory surge of luteinizing hormone (LH) or by
interfering with the ability of the maturing follicle to respond to
that gonadotropic signal. Examples for rats include compounds that
interfere with normal central nervous system (CNS) norepinephrine
receptor stimulation such as the pesticides chlordimeform and amitraz
(Goldman et al., 1990, 1991) and compounds that interfere with
norepinephrine synthesis such as the fungicide thiram (Stoker et al.,
1993). Compounds that increase central opioid receptor stimulation also
decrease serum LH and inhibit ovulation in monkeys and rats (Pang et
al., 1977; Smith, C.G., 1983). Delayed ovulation can alter oocyte
viability and cause trisomy and polyploidy in the conceptus (Fugo and
Butcher, 1966; Butcher and Fugo, 1967; Butcher et al., 1969, 1975; Na
et al., 1985). Delayed ovulation induced by exposure to the pesticide
chlordimeform has also been shown to alter fetal development and
pregnancy outcome in rats (Cooper et al., 1994).
III.B.4.c.3. Corpus luteum. The corpus luteum arises from the
ruptured follicle and secretes progesterone, which has an important
role in the estrous or menstrual cycle. Luteal progesterone is also
required for the maintenance of early pregnancy in most mammalian
species, including humans (Csapo and Pulkkinen, 1978). Therefore,
establishment and maintenance of normal corpora lutea are essential to
normal reproductive function. However, with the exception of
histopathologic evaluations that may establish only their presence or
absence, these structures are not evaluated in routine testing.
Additional research is needed to determine the importance of
incorporating endpoints that examine direct effects on luteal function
in routine toxicologic testing.
Adverse Effects
Increased rates of follicular atresia and oocyte toxicity leads to
premature menopause in humans. Altered follicular development,
ovulation failure, or altered corpus luteum formation and function can
result in disruption of cyclicity and reduced fertility, and, in
nonprimates, interference with normal sexual behavior. Therefore,
significant increases in the rate of follicular atresia, evidence of
oocyte toxicity, interference with ovulation, or altered corpus luteum
formation or function should be considered adverse effects.
III.B.4.d. Alterations in the Female Reproductive Cycle. The
pattern of events in the estrous cycle may provide a useful indicator
of the normality of reproductive neuroendocrine and ovarian function in
the nonpregnant female. It also provides a means to interpret hormonal,
histologic, and morphologic measurements relative to stage of the
cycle, and can be useful to monitor the status of mated females.
Estrous cycle normality can be monitored in the rat and mouse by
observing the changes in the vaginal smear cytology (Long and Evans,
1922; Cooper et al., 1993). To be most useful with cycling females,
vaginal smear cytology should be examined daily for at least three
normal estrous cycles prior to treatment, after onset of treatment, and
before necropsy (Kimmel, G.A. et al., 1995). However, practical
[[Page 56294]]
limitations in testing may limit the examination to the period before
mating or necropsy.
Daily vaginal smear data from rodents can provide useful
information on (1) cycle length, (2) occurrence or persistence of
estrus, (3) duration or persistence of diestrus, (4) incidence of
spontaneous pseudopregnancy, (5) distinguishing pregnancy from
pseudopregnancy (based on the number of days the smear remains
leukocytic), and (6) indications of fetal death and resorption by the
presence of blood in the smear after day 12 of gestation. The technique
also can detect onset of reproductive senescence in rodents (LeFevre
and McClintock, 1988). It is useful further to detect the presence of
sperm in the vagina as an indication of mating.
In nonpregnant females, repetitive occurrence of the four stages of
the estrous cycle at regular, normal intervals suggests that
neuroendocrine control of the cycle and ovarian responses to that
control are normal. Even normal, control animals can show irregular
cycles. However, a significant alteration compared with controls in the
interval between occurrence of estrus for a treatment group is cause
for concern. Generally, the cycle will be lengthened or the animals
will become acyclic. Lengthening of the cycle may be a result of
increased duration of either estrus or diestrus. Knowing the affected
phase can provide direction for further investigation.
The persistence of regular vaginal cycles after treatment does not
necessarily indicate that ovulation occurred, because luteal tissue may
form in follicles that have not ruptured. This effect has been observed
after treatment with anti-inflammatory agents (Walker et al., 1988).
However, that effect should be reflected in reduced fertility.
Conversely, subtle alterations of cyclicity can occur at doses below
those that alter fertility (Gray et al., 1989).
Irregular cycles may reflect impaired ovulation. Extended vaginal
estrus usually indicates that the female cannot spontaneously achieve
the ovulatory surge of LH (Huang and Meites, 1975). A number of
compounds have been shown to alter the characteristics of the LH surge
including anesthetics (Nembutal), neurotransmitter receptor binding
agents (Drouva et al., 1982), and the pesticides chlordimeform and
lindane (Cooper et al., 1989; Morris et al., 1990). Persistent or
constant vaginal cornification (or vaginal estrus) may result from one
or several effects. Typically, in the adult, if the vaginal epithelium
becomes cornified and remains so in response to toxicant exposure, it
is the result of the agent's estrogenic properties (i.e., DES or
methoxychlor), or the ability of the agent to block ovulation. In the
latter case, the follicle persists and endogenous estrogen levels bring
about the persistent vaginal cornification. Histologically, the ovaries
in persistent estrus will be atrophied following exposure to estrogenic
substances. In contrast, the ovaries of females in which ovulation has
been blocked because of altered gonadotropin secretion will contain
several large follicles and no corpora lutea. Females in constant
estrus may be sexually receptive regardless of the mechanism
responsible for this altered ovarian condition. However, if ovulation
has been blocked by the treatment, an LH surge may be induced by mating
(Brown-Grant et al., 1973; Smith, E.R. and Davidson, 1974) and a
pregnancy or pseudopregnancy may ensue. The fertility of such matings
is reduced (Cooper et al., 1994). Significant delays in ovulation can
result in increased embryonic abnormalities and pregnancy loss (Fugo
and Butcher, 1966; Cooper et al., 1994).
Persistent diestrus indicates temporary or permanent cessation of
follicular development and ovulation, and thus at least temporary
infertility. Prolonged vaginal diestrus, or anestrus, may be indicative
of agents (e.g., polyaromatic hydrocarbons) that interfere with
follicular development or deplete the pool of primordial follicles
(Mattison and Nightingale, 1980) or agents such as atrazine that
interrupt gonadotropin support of the ovary (Cooper et al., 1996).
Pseudopregnancy is another altered endocrine state reflected by
persistent diestrus. A pseudopregnant condition also has been shown to
result in rats following single or multiple doses of atrazine (Cooper
et al., 1996). The ovaries of anestrous females are atrophic, with few
primary follicles and an unstimulated uterus (Huang and Meites, 1975).
Serum estradiol and progesterone are abnormally low.
Adverse Effects
Significant evidence that the estrous cycle (or menstrual cycle in
primates) has been disrupted should be considered an adverse effect.
Included should be evidence of abnormal cycle length or pattern,
ovulation failure, or abnormal menstruation.
III.B.4.e. Mammary Gland and Lactation. The mammary glands of
normal adults change dramatically during the period around parturition
because of the sequential effects of a number of gonadal and
extragonadal hormones. Milk letdown is dependent on the suckling
stimulus and the release of oxytocin from the posterior pituitary.
Thus, mammary tissue is highly endocrine dependent for development and
function (Wolff, 1993; Imagawa et al., 1994; Tucker, 1994).
Mammary gland size, milk production and release, and histology can
be affected adversely by toxic agents, and many exogenous chemicals and
drugs are transferred into milk (American Academy of Pediatrics
Committee on Drugs, 1994; Oskarsson et al., 1995; Sonawane, 1995).
Reduced growth of young could be caused by reduced milk availability,
palatability or quality, by ingestion of a toxic agent secreted into
the milk, or by other factors unrelated to lactational ability (e.g.,
deficient suckling ability or deficient maternal behavior). Perinatal
exposure to steroid hormones and other chemicals can alter mammary
gland morphology and tumor potential in adulthood. Because of the
tendency for mobilization of lipids from adipose tissue and secretion
of those lipids into milk by lactating females, milk may contain
lipophilic agents at concentrations equal to or higher than those
present in the blood or organs of the dam. Thus, suckling offspring may
be exposed to elevated levels of such agents.
Techniques for measuring mammary tissue development, nucleic acid
content, milk production and milk composition in rodents are discussed
by Tucker (1994). During lactation, the mammary glands can be dissected
and weighed only with difficulty. RNA content of the mammary glands may
be measured as an index of lactational potential. More direct estimates
of milk production may be obtained by measuring litter weights of milk-
deprived pups taken before and after nursing. Milk from the stomachs of
pups treated similarly can also be weighed at necropsy. Cleared and
stained whole mounts of the mammary gland can be prepared at necropsy
for histologic examination. The DNA, RNA, and lipid content of the
mammary gland and the composition of the milk have been measured
following toxicant administration as indicators of toxicity to this
target organ.
Significant reductions in milk production or negative effects on
milk quality, whether measured directly or reflected in impaired
development of young, should be considered adverse reproductive
effects.
III.B.4.f. Reproductive Senescence. With advancing age, there is a
loss of the regular ovarian cycles and associated normal cyclical
changes in the uterine and vaginal epithelium that
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are typical of the young-adult female rat (Cooper and Walker, 1979).
Although the mechanisms responsible for this loss of cycling are not
thoroughly understood, age-dependent changes occur within the
hypothalamic-pituitary control of ovulation (Cooper et al., 1980; Finch
et al., 1984). Cumulative exposure to estrogen secreted by the ovary
may play a role, as treatment with estrogens during adulthood can
accelerate the age-related loss of ovarian function (Brawer and Finch,
1983). In contrast, the principal cause of the loss of ovarian cycling
in humans appears to be the depletion of oocytes (Mattison, 1985).
Prenatal or postnatal treatment of females with estrogens or
estrogenic pesticides can also cause impaired ovulation and sterility
(Gorski, 1979). These observations imply that alterations in ovarian
function may not be noticeable immediately after treatment but may
become evident at puberty or influence the age at which reproductive
senescence occurs.
Adverse Effects
Significant effects on measures showing a decrease in the age of
onset of reproductive senescence in females should be considered
adverse. Cessation of normal cycling, which is measured by vaginal
smear cytology, ovarian histopathology, or an endocrine profile that is
consistent with this interpretation, should be included as an adverse
effect.
III.B.5. Developmental and Pubertal Alterations
Developmental Effects
Alterations of reproductive differentiation and development,
including those produced by endocrine system disruption, can result in
infertility, functional and morphologic alterations of the reproductive
system, and cancer (Steinberger and Lloyd, 1985; Gray, 1991). Prenatal
and postnatal exposure to toxicants can produce changes that may not be
predicted from effects seen in adults, and those effects are often
irreversible. Adverse developmental outcomes in either sex can result
from exposure to toxic agents in utero, through contact with exposed
dams, or in milk. Dosing of dams during lactation also can result in
developmental effects through impaired nursing capability of the dams.
Effects observed in rodents following developmental exposure to
agents can include alterations in the genitalia (including ano-genital
distance), inhibited (female) or retained (male) nipple development,
impaired sexual behavior, delay or acceleration of the onset of
puberty, and reduced fertility (Gray et al., 1985, 1994, 1995; Gray and
Ostby, 1995; Kelce et al., 1995). Effects may include altered sexual
behavior or ability to produce gametes normally that are not observed
until after puberty. Hepatic enzyme systems for steroid metabolism that
are imprinted during development may be altered in males. Testis
descent from the abdominal cavity into the scrotum may be delayed or
may not occur. Generally, the type of effect seen may differ depending
on the stage of development at which the exposure occurred.
Many of these effects have been detected in human females and males
exposed prenatally to diethylstilbestrol (DES), other estrogens,
progestins, androgens, and anti-androgens (Giusti et al., 1995;
Harrison et al., 1995). Accelerated reproductive aging and tumors of
the reproductive tract have been observed in laboratory animal and
human females after pre- or perinatal exposure to hormonally active
agents. However, capability to alter sexual differentiation is not
limited to agents with known direct hormonal activity. Other agents,
for which the mode of action is not known (e.g., busulfan, nitrofen),
or which affect the endocrine system indirectly (e.g., PCBs, dioxin),
may act via different mechanisms during critical periods of development
to alter sexual differentiation and reproductive system development.
Effects on Puberty
In female rats and mice, the age at vaginal opening is the most
commonly measured marker of puberty. This event results from an
increase in the blood level of estradiol. The ages and weights of
females at the first cornified (estrous) vaginal smear, the first
diestrous smear, and the onset of vaginal cycles have also been used as
endpoints for onset of puberty. In males, preputial separation or
appearance of sperm in expressed urine or ejaculates can serve as
markers of puberty. Body weight at puberty may provide a means to
separate specific delays in puberty from those that are related to
general delays in development. Agents may differentially affect the
endpoints related to puberty onset, so it is useful to have information
on more than one marker.
Puberty can be accelerated or delayed by exogenous agents, and both
types of effects may be adverse (Gray et al., 1989, 1995; Gray and
Ostby, 1995; Kelce et al., 1995). For example, an acceleration of
vaginal opening may be associated with a delay in the onset of
cyclicity, infertility, and with accelerated reproductive aging
(Gorski, 1979). Delays in pubertal development in rodents are usually
related to delayed maturation or inhibition of function of the
hypothalamic-pituitary axis. Adverse reproductive outcomes have been
reported in rodents when puberty is altered by a week or more, but the
biologic relevance of a change in these measures of a day or two is
unknown (Gray, 1991).
Adverse Effects
Effects induced or observed during the pre- or perinatal period
should be judged using guidance from the Guidelines for Developmental
Toxicity Risk Assessment (U.S. EPA, 1991) as well as from these
Guidelines. Significant effects on ano-genital distance or age at
puberty, either early or delayed, should be considered adverse as
should malformations of the internal or external genitalia. Included as
adverse effects for females should be effects on nipple development,
age at vaginal opening, onset of cyclic vaginal smears, onset of estrus
or menstruation, or onset of an endocrine or behavioral pattern
consistent with estrous or menstrual cyclicity. Included as adverse
effects for males should be delay or failure of testis descent, as well
as delays in age at preputial separation or appearance of sperm in
expressed urine or ejaculates.
III.B.6. Endocrine Evaluations
Toxic agents can alter endocrine system function by affecting any
part of the hypothalamic-pituitary-gonadal-reproductive tract axis.
Effects may be induced in either sex by altering hormone synthesis,
storage, release, transport, or clearance, as well as by altering
hormone receptor recognition or postreceptor responses. The involvement
of the endocrine system in female reproductive physiology and
toxicology has been presented to a substantial degree as a necessary
component in Section III.B.4. (Female-specific Endpoints). The
information in that section should be considered together with the
following material.
The male reproductive system can be affected adversely by
disruption of the normal endocrine balance. In adults, effects that
result in interference with normal concentrations or action of LH and/
or follicle stimulating hormone (FSH) can decrease or abolish
spermatogenesis, affect secondary sex organ (e.g., epididymis) and
accessory sex gland (e.g., prostate, seminal vesicle) function, and
impair sexual behavior (Sharpe, 1994). In mammals, a female
reproductive tract develops unless androgen is produced and utilized
normally by the fetus (Byskov and Hoyer, 1994; George and Wilson,
1994).
[[Page 56296]]
Therefore, the consequences of disruption of the normal endocrine
pattern during development of the male reproductive system pre- and
postnatally are of particular concern. Differentiation and development
of the male reproductive system are especially sensitive to substances
that interfere with the production or action of androgens (testosterone
and dihydrotestosterone). Sexual differentiation of the CNS can be
affected also. Therefore, interference with normal production or
response to androgens can result in a range of abnormal effects in
genotypic males ranging from a pseudohermaphrodite condition to
reduction in sperm production or altered sexual behavior. Chemicals
with estrogenic or anti-androgenic activity have been identified that
are capable, with sufficient exposure levels, of causing effects of
these types in males (Gray et al., 1994; Harrison et al., 1995; Kelce
et al., 1995). While sensitivity may differ, it is likely that
mechanisms of action for these endocrine disrupting agents will be
consistent across mammalian species. Chemicals with the ability to
interact with the Ah receptor (e.g., dioxin or PCBs) may also disrupt
reproductive system development or function (Brouwer et al., 1995;
Safe, 1995). Several of the effects seen with exposure of male and
female rats and hamsters differ from those caused by estrogens,
indicating a different mechanism of action.
The developing nervous system can be a target of chemicals. In
rats, sexual differentiation of the CNS can be modified by hormonal
treatments or exposure to environmental agents that mimic or interfere
with the action of certain hormones. Prior to gender differentiation,
the brain is inherently female or at least bipotential (Gorski, 1986).
Thus, the functional and structural sex differences in the CNS are not
due directly to sex differences in neuronal genomic expression, but
rather are imprinted by the gonadal steroid environment during
development.
Chemicals with endocrine activity have been shown to masculinize
the CNS of female rats. Examples include chlordecone (Gellert, 1978),
DDT (Bulger and Kupfer, 1985), and methoxychlor (Gray et al., 1989).
Exposure of newborn female rats to these agents during the critical
period of sexual differentiation can alter the timing of puberty and
perturb subsequent reproductive function, presumably by altering the
development of the neural mechanisms that regulate gonadotropin
secretion.
In females, the situation is more complex than in males due to the
female cycle, the fertilization process, gestation and lactation. All
of the functions of the female reproductive system are under endocrine
control, and therefore can be susceptible to disruption by effects on
the reproductive endocrine system.
As with males, disturbance of the normal endocrine patterns during
development can result in abnormal development of the female
reproductive tract at exposure levels that tend to be lower than those
affecting adult females (Gellert, 1978; Brouwer et al., 1995).
Consistent with the differentiation mechanism described above, exposure
of genotypic females to androgens causes formation of
pseudohermaphrodite reproductive tracts with varying degrees of
severity as well as alteration of brain imprinting. However, exposure
to estrogenic substances during development also results in adverse
effects on anatomy and function including, in rats, malformations of
the genitalia. Exposure of human females to diethylstilbestrol in utero
has been shown to cause an increased incidence of vaginal clear cell
adenoma (Giusti et al., 1995). Dioxin, presumably acting through the Ah
receptor, also disrupts development of the female reproductive system
(Gray and Ostby, 1995).
Endpoints can be included in standardized toxicity testing that are
capable of detecting, but are not specific for, effects of reproductive
endocrine system disruption. For effects of exposure on adults,
endpoints can be incorporated into the subchronic toxicity protocol or
into reproductive toxicity protocols. For effects that are induced
during development, protocols that include exposure throughout the
development process and allow evaluation of the offspring
postpubertally are needed. Data from specialized testing, including in
vitro screening tests, may be useful to evaluate further the site,
timing, and mechanism of action.
Endpoints that can detect endocrine-related effects with adult-only
exposure in standardized testing include evaluation of fertility,
reproductive organ appearance, weights, and histopathology, oocyte
number, cycle normality and mating behavior. Endpoints that can detect
effects induced by endocrine system disruption during development
include, in addition to those identified for adult-exposed animals, the
reproductive developmental endpoints identified in Section III.B.5.
Significant effects on any of these measures may be considered to be
adverse if the results are consistent and biologically plausible.
Levels of the reproductive hormones are not available routinely
from toxicity testing. However, measurements of the reproductive
hormones in males offer useful supplemental information in assessing
potential reproductive toxicity for test species (Sever and Hessol,
1984; Heywood and James, 1985; NRC, 1989). Such measurements have
increased importance with humans where invasiveness of approaches must
be limited. The reproductive hormones measured often are circulating
levels of LH, FSH, and testosterone. Other useful measures that may be
available include prolactin, inhibin, and androgen binding protein
levels. In addition, challenge tests with exogenous agents (e.g.,
gonadotropin releasing hormone, LH, or human chorionic gonadotropin)
may provide insight into the functional responsiveness of the pituitary
or Leydig cells.
Interpretation of endocrine effects is facilitated if information
is available on a battery of hormones. However, in evaluating such
data, it is important to consider that serum hormones such as FSH, LH,
prolactin, and androgens exhibit cyclic variations within a 24-hour
period (Fink, 1988). Thus, the time of sampling should be controlled
rigorously to avoid excessive variability (Nett, 1989). Sequential
sampling can allow detection of treatment-related changes in circadian
and pulsatile rhythms.
The pattern seen in levels of reproductive system hormones can
provide useful information about the possible site and type of effect
on reproductive system function. For example, if a compound acts at the
level of the hypothalamus or pituitary, then serum LH and FSH may be
decreased, leading to decreased testosterone levels. On the other hand,
severe interference with Sertoli cell function or spermatogenesis would
be expected to elevate serum FSH levels. An agent having antiandrogenic
activity in adults might elevate serum LH and testosterone. Testis
weight might be unaffected, while the weight and size of the accessory
sex glands may be reduced. The endocrine profile presented by exposure
to specific antiandrogens can differ markedly because of differences in
tissue specificity and receptor kinetics, as well as age at which
exposure occurred.
Adverse Effects
In the absence of endocrine data, significant effects on
reproductive system anatomy, sexual behavior, pituitary, uterine or
accessory sex gland
[[Page 56297]]
weights or histopathology, female cycle normality, or Leydig cell
histopathology may suggest disruption of the endocrine system. In those
instances, additional testing for endocrine effects may be indicated.
Significant alterations in circulating levels of estrogen,
progesterone, testosterone, prolactin, LH, or FSH may be indicative of
existing pituitary or gonadal injury. When significant alterations from
control levels are observed in those hormones, the changes should be
considered cause for concern because they are likely to affect, occur
in concert with, or result from alterations in gametogenesis, gamete
maturation, mating ability, or fertility. Such effects, if compatible
with other available information, may be considered adverse and may be
used to establish a NOAEL, LOAEL, or benchmark dose. Furthermore,
endocrine data may facilitate identification of sites or mechanisms of
toxicant action, especially when obtained after short-term exposures.
III.B.7. In Vitro Tests of Reproductive Function
Numerous in vitro tests are available and under development to
measure or detect chemically induced changes in various aspects of both
male and female reproductive systems (Kimmel, G.L. et al., 1995). These
include in vitro fertilization using isolated gametes, whole organ
(e.g., testis, ovary) perfusion, culture of isolated cells from the
reproductive organs (e.g., Leydig cells, Sertoli cells, granulosa
cells, oviductal or epididymal epithelium), co-culture of several
populations of isolated cells, ovaries, quarter testes, seminiferous
tubule segments, various receptor binding assays on reproductive cells
and transfected cell lines, and others.
Tests of sperm properties and function that have been applied to
reproductive toxicology include penetration of sperm through viscous
medium (Yeung et al., 1992), in vitro capacitation and fertilization
assays (Holloway et al., 1990a, b; Perreault and Jeffay, 1993; Slott et
al., 1995), and evaluation of sperm nuclear integrity (Darney, 1991).
In addition, evaluation of human sperm function may include sperm
penetration of cervical mucus, ability of sperm to undergo an acrosome
reaction, and ability to penetrate zona pellucida-free hamster oocytes
or bind to human hemi-zona pellucidae (Franken et al., 1990; Liu and
Baker, 1992).
The diagnostic information obtained from such tests may help to
identify potential effects on the reproductive systems. However, each
test bypasses essential components of the intact animal system and
therefore, by itself, is not capable of predicting exposure levels that
would result in toxicity in intact animals. While it is desirable to
replace whole animal testing to the extent possible with in vitro
tests, the use of such tests currently is to screen for toxicity
potential and to study mechanisms of action and metabolism (Perreault,
1989; Holloway et al., 1990a, b).
III.C. Human Studies
In principle, human data are scientifically preferable for risk
assessment since test animal to human extrapolation is not required. At
this time, reproductive data for humans are available for only a
limited number of toxicants. Many of these are from occupational
settings in which exposures tend to be higher than in environmental
settings. As more data become available, expanding the number of agents
and endpoints studied and improving exposure assessment, more risk
assessments will include these data. The following describes the
methods of generation and evaluation of human data and the relative
weight the various types of human data should be given in risk
assessments.
``Human studies'' include both epidemiologic studies and other
reports of individual cases or clusters of events. Typical
epidemiologic studies include (1) cohort studies in which groups are
defined by exposure and health outcomes are examined; (2) case-referent
studies in which groups are defined by health status and prior
exposures are examined; (3) cross-sectional studies in which exposure
and outcome are determined at the same time; and (4) ecologic studies
in which exposure is presumed based typically on residence. Greatest
weight should be given to carefully designed epidemiologic studies with
more precise measures of exposure, because they can best evaluate
exposure-response relationships. This assumes that human exposures
occur in broad enough ranges for observable differences in response to
occur. Epidemiologic studies in which exposure is presumed, based on
occupational title or residence (e.g., some case-referent and all
ecologic studies), may contribute data for hazard characterization, but
are of limited use for quantitative risk determination because of the
generally broad categorical groupings of exposure. Reports of
individual cases or clusters of events may generate hypotheses of
exposure-outcome associations, but require further confirmation with
well-designed epidemiologic or laboratory studies. These reports of
cases or clusters may support associations suggested by other human or
test animal data, but cannot stand by themselves in risk assessments.
III.C.1. Epidemiologic Studies
Good epidemiologic studies provide valuable data for assessment of
human risk. As there are many different designs for epidemiologic
studies, simple rules for their evaluation do not exist. Risk assessors
should seek the assistance of professionals trained in epidemiology
when conducting a detailed analysis. The following is an overview of
key issues to consider in evaluation for risk assessment of
reproductive effects.
III.C.1.a. Selection of Outcomes for Study. As already discussed, a
number of endpoints can be considered in the evaluation of adverse
reproductive effects. However, some of the outcomes are not easily
observed in humans, such as early embryonic loss, reproductive capacity
of the offspring, and invasive evaluations of reproductive function
(e.g., testicular biopsies). Currently, the most feasible endpoints for
epidemiologic studies are (1) indirect measures of fertility/
infertility; (2) reproductive history studies of some pregnancy
outcomes (e.g., embryonic/fetal loss, birth weight, sex ratio,
congenital malformations, postnatal function, and neonatal growth and
survival); (3) semen evaluations; (4) menstrual history; and (5) blood
or urinary hormone measures. Factors requiring control in the design or
analysis (such as effect modifiers and confounders, described below)
may vary depending on the specific outcomes selected for study.
The reproductive outcomes available for epidemiologic examination
are limited by a number of factors, including the relative magnitude of
the exposure, the size and demographic characteristics of the
population, and the ability to observe the outcome in humans. Use of
improved methods for identifying some outcomes, such as embryonic loss
detected by more sensitive urinary hCG (human chorionic gonadotropin)
assays, change the spectrum of outcomes available for study (Wilcox et
al., 1985; Sweeney et al., 1988; Zinaman et al., 1996). Other, less
accessible, endpoints may require invasive techniques to obtain samples
(e.g., histopathology) or may have high intra- or interindividual
variability (e.g., serum hormone levels, sperm count).
Demographic characteristics of the population, such as marital
status, age, education, socioeconomic status (SES), and prior
reproductive history are associated with the probability of
[[Page 56298]]
whether couples will attempt to have children. Differences in birth
control practices would also affect the number of outcomes available
for study.
In addition to the above-mentioned factors, reproductive endpoints
may be envisioned as effects recognized at various points in a
continuum starting before conception and continuing through death of
the progeny. Many studies, however, are limited to evaluating endpoints
at a particular time in this continuum. For example, in a study of
defects observed at live birth, a malformed stillbirth would not be
included, even though the etiology could be identical (Bloom, 1981).
Also, a different spectrum of outcomes could result from differences in
timing or in level of exposure (Selevan and Lemasters, 1987).
Human Reproductive Endpoints
The following section discusses various human male and female
reproductive endpoints. These outcomes may be an indicator of sub- or
infertility. These are followed by a discussion of reproductive history
studies.
Male Endpoints--Semen Evaluations
The use of semen analysis was discussed in Section III.B.3.d. Most
epidemiologic studies of potential effects of agents on semen
characteristics have been conducted in occupational groups and patients
receiving drug therapy. Obtaining a high level of participation in the
workforce has been difficult, because social and cultural attitudes
concerning sex and reproduction may affect cooperation of the study
groups. Increased participation may occur in men who are planning to
have children or who are concerned about existing reproductive problems
or possible ill effects of their exposures. Unless controlled, such
biased participation may yield unrepresentative estimates of risk
associated with exposure, resulting in data that are less useful for
risk assessment. While some studies have response rates greater than
70% (Ratcliffe et al., 1987; Welch et al., 1988), response rates are
often less than 70% in such studies and may be even lower in the
comparison group (Egnatz et al., 1980; Lipshultz et al., 1980; Milby
and Whorton, 1980; Lantz et al., 1981; Meyer, 1981; Milby et al., 1981;
Rosenberg et al., 1985; Ratcliffe et al., 1989). Some of the low
response rates may be caused by inclusion of vasectomized men in the
total population, although this could vary widely by population (Milby
and Whorton, 1980). Participation in the comparison group may be biased
toward those with preexisting reproductive problems. The response rate
may be improved substantially with proper education and payment of
subjects (Ratcliffe et al., 1986, 1987).
Several factors may influence the semen evaluation, including the
period of abstinence preceding collection of the sample, health status,
and social habits (e.g., alcohol, recreational drugs, smoking). Data on
these factors may be collected by interview, subject to the limitations
described for pregnancy outcome studies.
Reports of studies with semen analyses have rarely included an
evaluation of endocrine status (hormone levels in blood or urine) of
exposed males (Lantz et al., 1981; Ratcliffe et al., 1989). Conversely,
studies that have examined endocrine status typically do not have data
on semen quality (Mason, 1990; McGregor and Mason, 1991; Egeland et
al., 1994).
Female Endpoints
Reproductive effects may result from a variety of exposures. For
example, environmental exposures may be toxic to the oocyte, producing
a loss of primary oocytes that irreversibly affects the woman's
fecundity. The exposures of importance may occur during the prenatal
period, and beyond. Oocyte depletion is difficult to examine directly
in women because of the invasiveness of the tests required; however, it
can be studied indirectly through evaluation of the age at reproductive
senescence (menopause) (Everson et al., 1986).
Numerous diagnostic methods have been developed to evaluate female
reproductive dysfunction. Although these methods have been used rarely
for occupational or environmental toxicologic evaluations, they may be
helpful in defining biologic parameters and the mechanisms related to
female reproductive toxicity. If clinical observations are able to link
exposures to the reproductive effect of concern, these data will aid
the assessment of adverse female reproductive toxicity. The following
clinical observations include endpoints that may be reported in case
reports or epidemiologic research studies.
Reproductive dysfunction also can be studied by the evaluation of
irregularities of menstrual cycles. However, menstrual cyclicity is
affected by many parameters such as age, nutritional status, stress,
exercise level, certain drugs, and the use of contraceptive measures
that alter endocrine feedback. Vaginal bleeding at menstruation is a
reflection of withdrawal of steroidogenic support, particularly
progesterone. Vaginal bleeding can occur at midcycle, in early
miscarriage, after withdrawal of contraceptive steroids, or after an
inadequate luteal phase. The length of the menstrual cycle,
particularly the follicular phase (before ovulation), can vary between
individuals and may make it difficult to determine significant effects
on length in populations of women (Burch et al., 1967; Treloar et al.,
1967). Human vaginal cytology may provide information on the functional
state of reproductive cycles. Cytologic evaluations, along with the
evaluation of changes in cervical mucus viscosity, can be used to
estimate the occurrence of ovulation and determine different stages of
the reproductive cycle (Kesner et al., 1992). Menstrual dysfunction
data have been used to examine adverse reproductive effects in women
exposed to potentially toxic agents occupationally (Lemasters, 1992),
Reports of prospective clinical evaluations of menstrual function
(Kesner et al., 1992; Wright et al., 1992), have shown urinary
endocrine measures to be practical and useful. The endocrine status of
a woman can be evaluated by the measurement of hormones in blood and
urine. Progesterone can also be measured in saliva. Because the female
reproductive endocrine milieu changes in a cyclic pattern, single
sample analysis does not provide adequate information for evaluating
alterations in reproductive function. Still, a single sample for
progesterone determination some 7 to 9 days after the estimated
midcycle surge of gonadotropins in a regularly cycling woman may
provide suggestive evidence for the presence of a functioning corpus
luteum and prior follicular maturation and ovulation. Clinically
abnormal levels of gonadotropins, steroids, or other biochemical
parameters may be detected from a single sample. However, a much
stronger design involves collection of multiple samples and their
observation in conjunction with events in the menstrual cycle.
The day of ovulation can be estimated by the biphasic shift in
basal body temperature. Ovulation can also be detected by serial
measurement of hormones in the blood or urine and analyses of estradiol
and gonadotropin status at midcycle. After ovulation, luteal phase
function can be assessed by analysis of progesterone secretion and by
evaluation of endometrial histology. Tubal patency, which could be
affected by abnormal development, endometriosis or infection, is an
endpoint that can be observed in clinical evaluations of reproductive
[[Page 56299]]
function (Forsberg, 1981). These latter evaluations of endometrial
histology and tubal patency are less likely to be present in
epidemiologic studies or surveillance programs because of the
invasiveness of the procedures.
III.C.1.b. Reproductive History Studies
Measures of Fertility
Subfertility may be thought of as nonevents: a couple is unable to
have children within a specific time frame. Therefore, the
epidemiologic measurement of reduced fertility or fecundity is
typically indirect and is accomplished by comparing birth rates or time
intervals between births or pregnancies. These outcomes have been
examined using several methods: the Standardized Birth Ratio (SBR; also
referred to as the Standardized Fertility Ratio) and the length of time
to pregnancy or birth. In these evaluations, the couple's joint ability
to procreate is estimated. The SBR compares the number of births
observed to those expected based on the person-years of observation
preferably stratified by factors such as time period, age, race,
marital status, parity, and (if possible) contraceptive use (Wong et
al., 1979; Levine et al., 1980, 1981, 1983; Levine, 1983; Starr et al.,
1986). The SBR is analogous to the Standardized Mortality Ratio (SMR),
a measure frequently used in studies of occupational cohorts and has
similar limitations in interpretation (Gaffey, 1976; McMichael, 1976;
Tsai and Wen, 1986). The SBR was found to be less sensitive in
identifying an effect when compared to semen analyses (Welch et al.,
1991). These data can also be analyzed using Poisson regression.
Analysis of the time between recognized pregnancies or live births
is a more recent approach to indirect measurement of fertility (Dobbins
et al., 1978; Baird and Wilcox, 1985; Baird et al., 1986; Weinberg and
Gladen, 1986; Rowland et al., 1992). Because the time between births
increases with increasing parity (Leridon, 1977), comparisons within
birth order (parity) are more appropriate. A statistical method (Cox
regression) can stratify by birth or pregnancy order to help control
for nonindependence of these events in the same woman or couple.
Fertility may also be affected by alterations in sexual behavior.
However, data linking toxic exposures to these alterations in humans
are limited and are not obtained easily in epidemiology studies (see
Section III.C.1.d.).
Developmental Outcomes
Developmental outcomes examined in human studies of parental
exposures may include embryo or fetal loss, congenital malformations,
birth weight effects, sex ratio at birth, and possibly postnatal
effects (e.g., physical growth and development, organ or system
function, and behavioral effects of exposure). Developmental effects
are discussed in more detail in the Guidelines for Developmental
Toxicity Risk Assessment (U.S. EPA, 1991). As mentioned above,
epidemiologic studies that focus on only one type of developmental
outcome or exposures to only one parent may miss a true effect of
exposure.
Evidence of a dose-response relationship is usually an important
criterion in the assessment of exposure to a potentially toxic agent.
However, traditional dose-response relationships may not always be
observed for some endpoints (Wilson, 1973; Selevan and Lemasters,
1987). For example, with increasing dose, a pregnancy might end in
embryo or fetal loss, rather than a live birth with malformations. A
shift in the patterns of outcomes could result from differences either
in level of exposure or in timing (Wilson, 1973; Selevan and Lemasters,
1987) (for a more detailed description, see Section III.C.1.d.).
Therefore, a risk assessment should, when possible, attempt to look at
the relationship of different reproductive endpoints and patterns of
exposure.
In addition to the above effects, exposure may produce genetic
damage to germ cells. Outcomes resulting from germ-cell mutations could
include reduced probability of fertilization and increased probability
of embryo or fetal loss and postnatal developmental effects. Based on
studies with test species, germ cells or early zygotes are critical
targets of potentially toxic agents. Germ-cell mutagenicity could be
expressed also as genetic diseases in future generations.
Unfortunately, these studies are difficult to conduct in human
populations because of the long time between exposure and outcome and
the large study groups needed. For more information and guidance on the
evaluation of these data, refer to the Guidelines for Mutagenicity Risk
Assessment (U.S. EPA, 1986c).
III.C.1.c. Community Studies and Surveillance Programs.
Epidemiologic studies may be based on broad populations such as a
community, a nationwide probability sample, or surveillance programs
(such as birth defects registries). Some studies have examined the
effects of environmental exposures such as potential toxic agents in
outdoor air, food, water, and soil. These studies may assume certain
exposures through these routes due to residence (ecologic studies). The
link between environmental measurements and critical periods of
exposure for a given reproductive effect may be difficult to make.
Other studies may go into more detail, evaluating the above routes and
also indoor air, house dust, and occupational exposures on an
individual basis (Selevan, 1991). Such environmental studies, relating
individual exposures to health outcomes should have less
misclassification of exposure.
Exposure definition in community studies has some limitations in
the assessment of exposure-effect relationships. For example, in many
community-based studies, it may not be possible to distinguish
maternally mediated effects from paternally mediated effects since both
parents spend time in the same home environment. In addition, the
presumably lower exposure levels (compared with industrial settings)
may require very large groups for the study. A number of case-referent
studies have examined the relationship between broad classes of
parental occupation in certain communities or countries and embryo/
fetal loss (Silverman et al., 1985; McDonald et al., 1989; Lindbohm et
al., 1991), birth defects (Hemminki et al., 1980; Kwa and Fine, 1980;
Papier, 1985), and childhood cancer (Fabia and Thuy, 1974; Hemminki et
al., 1981; Peters et al., 1981; Gardner et al., 1990a, b). In these
reports, jobs are classified typically into broad categories based on
the probability of exposure to certain classes or levels of exposure.
Such studies are most helpful in the identification of topics for
additional study. However, because of the broad groupings of types or
levels of exposure, these studies are not typically useful for risk
assessment of any one particular agent.
Surveillance programs may also exist in occupational settings. In
this case, reproductive histories (including menstrual cycles) or semen
evaluations could be followed to monitor reproductive effects of
exposures. With adequate exposure information, these could yield very
useful data for risk assessment. Reproductive histories tend to be
easier and less costly to collect, whereas, a semen evaluation program
would be rather costly. Success with such programs in the workplace
will be determined by the confidence the worker has that reproductive
data are kept confidential and will not affect employment status
(Samuels, 1988; Lemasters and Selevan, 1993).
III.C.1.d. Identification of Important Exposures for Reproductive
Effects. For all examinations of the relationship between reproductive
effects and
[[Page 56300]]
potentially toxic exposures, defining the exposure that produces the
effect is crucial. Preconceptional exposures of either parent and in
utero exposures have been associated with the more commonly examined
outcomes (e.g., fetal loss, malformations, low birth weight, and
measures of in- or subfertility). These exposures, plus postnatal
exposure via breast milk, food, and the environment, may also be
associated with postnatal developmental effects (e.g., changes in
growth or in behavioral and cognitive function).
A number of factors affect the intensity and duration of exposure.
General environmental exposures are typically lower than those found in
industrial or agricultural settings. However, this relationship may
change as exposures are reduced in workplaces and as more is learned
about environmental exposures (e.g., indoor air exposures, home
pesticide usage). Larger populations are necessary to achieve
sufficient power in settings with lower exposures which are likely to
have lower measures of risk (Lemasters and Selevan, 1984). In addition,
exposure to individuals may change as they move in and out of areas
with differing levels and types of exposures, thus affecting the number
of exposed and comparison events for study.
Data on exposure from human studies are frequently qualitative,
such as employment or residence histories. More quantitative data may
be difficult to obtain because of the nature of certain study designs
(e.g., retrospective studies) and limitations in estimates of historic
exposures. Many reproductive effects result from exposures during
certain critical times. The appropriate exposure classification depends
on the outcomes studied, the biologic mechanism affected by exposure,
and the biologic half-life of the agent. The half-life, in combination
with the patterns of exposure (e.g., continuous or intermittent)
affects the individual's body burden and consequently the actual dose
during the critical period. The probability of misclassification of
exposure status may affect the ability to recognize a true effect in a
study (Selevan, 1981; Hogue, 1984; Lemasters and Selevan, 1984; Sever
and Hessol, 1984; Kimmel, C.A. et al., 1986). As more prospective
studies are done, better estimates of exposure should be developed.
III.C.1.e. General Design Considerations. The factors that enhance
a study and thus increase its usefulness for risk assessment have been
noted in a number of publications (Selevan, 1980; Bloom, 1981; Hatch
and Kline, 1981; Wilcox, 1983; Sever and Hessol, 1984; Axelson, 1985;
Tilley et al., 1985; Kimmel, C.A. et al., 1986; Savitz and Harlow,
1991). Some of the more prominent factors are discussed below.
The Power of the Study
The power, or ability of a study to detect a true effect, is
dependent on the size of the study group, the frequency of the outcome
in the general population, and the level of excess risk to be
identified. In a cohort study, common outcomes, such as recognized
fetal loss, require hundreds of pregnancies to have a high probability
of detecting a modest increase in risk (e.g., 133 pregnancies in both
exposed and unexposed groups to detect a twofold increase;
=0.05, power=80%), while less common outcomes, such as the
total of all malformations recognized at birth, require thousands of
pregnancies to have the same probability (e.g., more than 1,200
pregnancies in both exposed and unexposed groups) (Bloom, 1981;
Selevan, 1981, 1985; Sever and Hessol, 1984; Stein, Z. et al., 1985;
Kimmel, C.A. et al., 1986). Semen evaluation may require fewer subjects
depending on the sperm parameters evaluated, especially when each man
is used as his own control (Wyrobek, 1982, 1984). In case-referent
studies, study sizes are dependent upon the frequency of exposure
within the source population. The confidence one has in the results of
a study showing no effect is related directly to the power of the study
to detect meaningful differences in the endpoints.
Power may be enhanced by combining populations from several studies
using a meta-analysis (Greenland, 1987). The combined analysis could
increase confidence in the absence of risk for agents showing no
effect. However, caution must be exercised in the combination of
potentially dissimilar study groups.
Results of a negative study should be carefully evaluated,
examining the power of the study and the degree of concordance or
discordance between that study and other studies (including careful
examination of comparability in the details such as similarity of
adverse endpoints and study design). The consistency among results of
different studies could be evaluated by comparing statistical
confidence intervals for the effects found in different studies.
Studies with lower power will tend to yield wider confidence intervals.
If the confidence intervals from a negative study and a positive study
overlap, then there may be no conflict between the results of the two
studies.
Potential Bias in Data Collection
Bias may result from the way the study group is selected or
information is collected (Rothman, 1986). Selection bias may occur when
an individual's willingness to participate varies with certain
characteristics relating to exposure or health status. In addition,
selection bias may operate in the identification of subjects for study.
For example, in studies of very early pregnancy loss, use of hospital
records to identify the study group will under-ascertain events,
because women are not always hospitalized for these outcomes. More
weight would be given in a risk assessment to a study in which a more
complete list of pregnancies is obtained by, for example, collecting
biologic data (e.g., human chorionic gonadotropin [hCG] measurements)
of pregnancy status from study members. The representativeness of these
data may be affected by selection factors related to the willingness of
different groups of women to continue participation over the total
length of the study. Interview data result in more complete
ascertainment than hospital records; however this strategy carries with
it the potential for recall bias, discussed in further detail below.
Other examples of different levels of ascertainment of events include:
(1) use of hospital records to study congenital malformations since
hospital records contain more complete data on malformations than do
birth certificates (Mackeprang et al., 1972; Snell et al., 1992) and
(2) use of sperm bank or fertility clinic data for semen studies. Semen
data from either source are selected data because semen donors are
typically of proven fertility, and men in fertility clinics are part of
a subfertile couple who are actively trying to conceive. Thus, studies
using the different record sources to identify reproductive outcomes
need to be evaluated for ascertainment patterns prior to use in risk
assessment.
Studies of women who work outside the home present the potential
for additional bias because some factors that influence employment
status may also affect reproductive endpoints. For example, because of
child-care responsibilities, women may terminate employment, as might
women with a history of reproductive problems who wish to have children
and are concerned about workplace exposures (Joffe, 1985; Lemasters and
Pinney, 1989). Thus, retrospective studies of female exposure that do
not include terminated women workers may be of
[[Page 56301]]
limited use in risk assessment because the level of risk for these
outcomes is likely to be overestimated (Lemasters and Pinney, 1989).
Information bias may result from misclassification of
characteristics of individuals or events identified for study. Recall
bias, one type of information bias, may occur when respondents with
specific exposures or outcomes recall information differently than
those without the exposures or outcomes. Interview bias may result when
the interviewer knows a priori the category of exposure (for cohort
studies) or outcome (for case-referent studies) in which the respondent
belongs. Use of highly structured questionnaires and/or ``blinding'' of
the interviewer reduces the likelihood of such bias. Studies with lower
likelihood of such bias should carry more weight in a risk assessment.
When data are collected by interview or questionnaire, the
appropriate respondent depends on the type of data or study. For
example, a comparison of husband-wife interviews on reproduction found
the wives' responses to questions on pregnancy-related events to be
more complete and valid than those of the husbands, and the
individual's self-report of his/her occupational exposures and health
characteristics more reliable than his/her mate's report (Selevan,
1980; Selevan et al., 1982). Studies based on interview data from the
appropriate respondents would carry more weight than those from proxy
respondents.
Data from any source may be prone to errors or bias. All types of
bias are difficult to assess; however, validation with an independent
data source (e.g., vital or hospital records), or use of biomarkers of
exposure or outcome, where possible, may suggest the degree of bias
present and increase confidence in the results of the study. Those
studies with a low probability of biased data should carry more weight
(Axelson, 1985; Stein, A. and Hatch, 1987; Weinberg et al., 1994).
Differential misclassification (i.e., when certain subgroups are
more likely to have misclassified data than others) may either raise or
lower the risk estimate. Nondifferential misclassification will bias
the results toward a finding of ``no effect'' (Rothman, 1986).
Collection of Data on Other Risk Factors, Effect Modifiers, and
Confounders
Risk factors for reproductive toxicity include such characteristics
as age, smoking, alcohol or caffeine consumption, drug use, and past
reproductive history. Groups of individuals may represent susceptible
subpopulations based on genetic, acquired (e.g., behavioral), or
developmental characteristics (e.g., greater effect of childhood
exposures). Known and potential risk factors should be examined to
identify those that may be confounders or effect modifiers. An effect
modifier is a factor that produces different exposure-response
relationships at different levels of that factor. For example, age
would be an effect modifier if the risk associated with a given
exposure changed with age (e.g., if older men had semen changes with
exposure while younger ones did not). A confounder is a variable that
is a risk factor for the outcome under study and is associated with the
exposure under study, but is not a consequence of the exposure. A
confounder may distort both the magnitude and direction of the measure
of association between the exposure of interest and the outcome. For
example, smoking might be a confounder in a study of the association of
socioeconomic status and fertility because smoking may be associated
with both.
Both effect modifiers and confounders need to be controlled in the
study design and/or analysis to improve the estimate of the effects of
exposure (Kleinbaum et al., 1982). A more in-depth discussion may be
found elsewhere (Epidemiology Workgroup for the Interagency Regulatory
Liaison Group, 1981; Kleinbaum et al., 1982; Rothman, 1986). The
statistical techniques used to control for these factors require
careful consideration in their application and interpretation
(Kleinbaum et al., 1982; Rothman, 1986). Studies that fail to account
for these important factors should be given less weight in a risk
assessment.
Statistical Factors
As in studies of test animals, pregnancies experienced by the same
woman are not fully independent events. For example, women who have had
fetal loss are reported to be more likely to have subsequent losses
(Leridon, 1977). In test animal studies, the litter can be used as the
unit of measure to deal with nonindependence of response within the
litter. In studies of humans, pregnancies are sequential, requiring
analyses which consider nonindependence of events (Epidemiology
Workgroup for the Interagency Regulatory Liaison Group, 1981; Kissling,
1981; Selevan, 1981; Zeger and Liang, 1986). If more than one pregnancy
per woman is included, as is often necessary with small study groups,
the use of nonindependent observations overestimates the true size of
the groups being compared, thus artificially increasing the probability
of reaching statistical significance (Stiratelli et al., 1984).
Analysis problems may occur when (1) prior adverse outcomes are due to
the same exposures or (2) when prior adverse outcomes could result in
changes in behaviors that could reduce exposures. Some approaches to
deal with these issues have been suggested (Kissling, 1981; Stiratelli
et al., 1984; Selevan, 1985; Zeger and Liang, 1986). These approaches
include selecting one pregnancy per family (Selevan, 1985) or using
generalized estimating equations (Zeger and Liang, 1986).
III.C.2. Examination of Clusters, Case Reports, or Series
The identification of cases or clusters of adverse reproductive
effects is generally limited to those identified by the individuals
involved or clinically by their physicians. The likelihood of
identification varies with the gender of the exposed person.
Identification of subfecundity in either gender is difficult. This
might be thought of as identification of a nonevent (e.g., lack of
pregnancies or children), and thus is much harder to recognize than are
some developmental effects, including malformations, resulting from in
utero exposure.
The identification of cases or clusters of adverse male
reproductive outcomes may be limited because of cultural norms that may
inhibit the reporting of impaired fecundity in men. Identification is
also limited by the decreased likelihood of recognizing adverse
developmental effects in their offspring as resulting from paternal
exposure rather than maternal exposure. Thus far, only one agent
causing human male reproductive toxicity, dibromochloropropane (DBCP),
has been identified after observation of a cluster of infertility that
resulted from male subfecundity. This cluster was identified because of
an atypically high level of communication among the workers' wives
(Whorton et al., 1977, 1979; Biava et al., 1978; Whorton and Milby,
1980).
Adverse effects identified in females through clusters and case
reports have, thus far, been limited to adverse pregnancy outcomes such
as fetal loss and congenital malformations. Identification of other
effects, such as subfertility/subfecundity or menstrual cycle
disorders, may be more difficult, as noted above.
Case reports may have importance in the recognition of agents that
cause reproductive toxicity. However, they are
[[Page 56302]]
probably of greatest use in suggesting topics for further
investigation. Reports of clusters and case reports/series are best
used in risk assessment in conjunction with strong laboratory data to
suggest that effects observed in test animals also occur in humans.
III.D. Pharmacokinetic Considerations
Extrapolation of toxicity data between species can be aided
considerably by the availability of data on the pharmacokinetics of a
particular agent in the species tested and, when available, in humans.
Information on absorption, half-life, steady-state or peak plasma
concentrations, placental metabolism and transfer, comparative
metabolism, and concentrations of the parent compound and metabolites
in target organs may be useful in predicting risk for reproductive
toxicity. Information on the variability between humans and test
species also may be useful in evaluating factors such as age-related
differences in the balance between activation and deactivation of a
toxic agent. These types of data may be helpful in defining the
sequence of events leading to an adverse effect and the dose-response
curve, developing a more accurate comparison of species sensitivity,
including that of humans (Wilson et al., 1975, 1977), determining
dosimetry at target sites, and comparing pharmacokinetic profiles for
various dosing regimens or routes of exposure. EPA's Office of
Prevention, Pesticides, and Toxic Substances has published protocols
for metabolism studies that may be adapted to provide information
useful in reproductive toxicity risk assessment for a suspect agent.
Pharmacokinetic studies in reproductive toxicology are most useful if
the data are obtained with animals that are at the same reproductive
status and stage of life (e.g., pregnant, nonpregnant, embryo or fetus,
neonate, prepubertal, adult) at which reproductive insults are expected
to occur in humans.
Specific guidance regarding both the development and application of
pharmacokinetic data was agreed on by the participants of the Workshop
on Dermal Developmental Toxicity Studies (Kimmel, C.A. and Francis,
1990). This guidance is also applicable to nondermal reproductive
toxicity studies. Participants of the Workshop concluded that
absorption data are needed both when a dermal study does or does not
show effects. The results of a dermal study showing no effects and
without blood level data are potentially misleading and are inadequate
for risk assessment, especially if interpreted as a ``negative'' study.
In studies where adverse effects are detected, regardless of the route
of exposure, pharmacokinetic data can be used to establish the internal
dose in maternal and paternal animals for risk extrapolation purposes.
The existence of a Sertoli cell barrier (formerly called the blood-
testis barrier) in the seminiferous tubules may influence the
pharmacokinetics of an agent with potential to cause testicular
toxicity by restricting access of compounds to the adluminal
compartment of seminiferous tubules. The Sertoli cell barrier is formed
by tight junctions between Sertoli cells and divides the seminiferous
epithelium into basal and adluminal compartments (Russell et al.,
1990). The basal compartment contains the spermatogonia and primary
spermatocytes to the preleptotene stage, whereas more advanced germ
cells are located on the adluminal side. This selectively permeable
barrier is most effective in limiting the access of large, hydrophilic
molecules in the intertubular lymph to cells on the adluminal side. An
analogous barrier in the ovary has not been found, although the zona
pellucida and granulosa cells may modulate access of chemicals to
oocytes (Crisp, 1992).
The reproductive organs appear to have a wide range of metabolic
capabilities directed at both steroid and xenobiotic metabolism.
However, there are substantial differences between compartments within
the organs in types and levels of enzyme activities (Mukhtar et al.,
1978). Recognition of these differences can be important in
understanding the potential of agents to have specific toxic effects.
Most pharmacokinetic studies have incompletely characterized the
distribution of toxic agents and their subsequent metabolic fate within
the reproductive organs. Generalizations based on hepatic metabolism
are not necessarily adequate to predict the fate of the agent in the
testis, ovary, placenta, or conceptus. For example, the metabolic
profile for a given agent may differ in the male between the liver and
the testis and in the female between the maternal liver, ovary, and
placenta. Detailed interspecies comparisons of the metabolic
capabilities of the testis, ovary, placenta, and conceptus also have
not been conducted. For some xenobiotics, significant differences in
metabolism have been identified between males and females (Harris, R.Z.
et al., 1995). This is, in part, attributable to organizational effects
of the gonadal steroids in the developing liver (Gustafsson et al.,
1980; Skett, 1988). Also, in adults, the sex steroids have been shown
to affect the activity of a number of enzymes involved in the
metabolism of administered compounds. Thus, the blood levels of a toxic
agent, as well as the final concentration in the target tissue, may
differ significantly between sexes. If data are to be used effectively
in interspecies comparisons and extrapolations for these target
systems, more attention should be directed to the pharmacokinetic
properties of chemicals in the reproductive organs and in other organs
that are affected by reproductive hormones.
III.E. Comparisons of Molecular Structure
Comparisons of the chemical or physical properties of an agent with
those of agents known to cause reproductive toxicity may provide some
indication of a potential for reproductive toxicity. Such information
may be helpful in setting priorities for testing of agents or for
evaluation of potential toxicity when only minimal data are available.
Structure-activity relationships (SAR) have not been well studied in
reproductive toxicology, and have had limited success in predicting
reproductive toxicity. The early literature has been reviewed and a set
of classifications offered relating structure to reported male
reproductive system activity (Bernstein, 1984). Data are available that
suggest structure-activity relationships with limited utility in risk
assessment for certain classes of chemicals (e.g., glycol ethers, some
estrogens, androgens, other steroids, substituted phenols, retinoids,
phthalate esters, short-chain halogenated hydrocarbon pesticides,
alkyl-substituted polychlorinated dibenzofurans, PCBs, vinylcyclohexene
and related olefins, halogenated propanes, metals, and azo dyes).
McKinney and Waller (1994) have studied the qualitative SAR properties
of PCBs with respect to their recognition by thyroxine, Ah and estrogen
receptors. Although generally limited in scope and in need of
validation, such relationships provide hypotheses that can be tested.
In spite of the limited information available on SAR in
reproductive toxicology, under certain circumstances (e.g., in the case
of new chemicals), this procedure can be used to evaluate the potential
for toxicity when little or no other data are available.
III.F. Evaluation of Dose-Response Relationships
The description and evaluation of dose-response relationships is a
critical component of the hazard characterization. Evidence for a dose-
response relationship is an important
[[Page 56303]]
criterion in establishing a toxic reproductive effect. It includes the
evaluation of data from both human and laboratory animal studies. When
possible, pharmacokinetic data should be used to determine the
effective dose at the target organ(s). When adequate dose-response data
are available in humans and with a sufficient range of exposure, dose-
response relationships in humans may be examined. Because quantitative
data on human dose-response relationships are available infrequently,
the dose-response evaluation is usually based on the assessment of data
from tests performed in laboratory animals.
The dose-response relationships for individual endpoints, as well
as the combination of endpoints, must be examined in data
interpretation. Dose-response evaluations should consider the effects
that competing risks between different endpoints may have on outcomes
observed at different exposure levels. For example, an agent may
interfere with cell function in such a manner that, at a low dose
level, an increase in abnormal sperm morphology is observed. At higher
doses, cell death may occur, leading to a decrease in sperm counts and
a possible decrease in proportion of abnormal sperm.
When data on several species are available, the selection of the
data for the dose-response evaluation is based ideally on the response
of the species most relevant to humans (e.g., comparable physiologic,
pharmacologic, pharmacokinetic, and pharmacodynamic processes), the
adequacy of dosing, the appropriateness of the route of administration,
and the endpoints selected. However, availability of information on
many of those components is usually very limited. For dose-response
assessment, no single laboratory animal species can be considered the
best in all situations for predicting risk of reproductive toxicity to
humans. However, in some cases, such as in the assessment of
physiologic parameters related to menstrual disorders, higher nonhuman
primates are considered generally similar to the human. In the absence
of a clearly most relevant species, data from the most sensitive
species (i.e., the species showing a toxic effect at the lowest
administered dose) are used, because humans are assumed to be at least
as sensitive generally as the most sensitive animal species tested
(Nisbet and Karch, 1983; Kimmel, C.A. et al., 1984, 1990; Hemminki and
Vineis, 1985; Meistrich, 1986; Working, 1988).
The evaluation of dose-response relationships includes the
identification of effective dose levels as well as doses that are
associated with low or no increased incidence of adverse effects
compared with controls. Much of the focus is on the identification of
the critical effect(s) (i.e., the adverse effect occurring at the
lowest dose level) and the LOAEL and NOAEL or benchmark dose associated
with the effect(s) (see Section IV).
Generally, in studies that do not evaluate reproductive toxicity,
only adult male and nonpregnant females are examined. Therefore, the
possibility that pregnant females may be more sensitive to the agent is
not tested. In studies in which reproductive toxicity has been
evaluated, the effective dose range should be identified for both
reproductive and other forms of systemic toxicity, and should be
compared with the corresponding values from other adult toxicity data
to determine if the pregnant or lactating female may be more sensitive
to an agent.
In addition to identification of the range of doses that is
effective in producing reproductive and other forms of systemic
toxicity for a given agent, the route of exposure, timing and duration
of exposure, species specificity of effects, and any pharmacokinetic or
other considerations that might influence the comparison with human
exposure scenarios should be identified and evaluated. This information
should always accompany the characterization of the health-related
database (discussed in the next section).
Because the developing organism is changing rapidly and is
vulnerable at a number of stages, an assumption is made with
developmental effects that a single exposure at a critical time in
development may produce an adverse effect (U.S. EPA, 1991). Therefore,
with inhalation exposures, the daily dose is usually not adjusted to a
24-hour equivalent duration with developmental toxicity unless
appropriate pharmacokinetic data are available. However, for other
reproductive effects, daily doses by the inhalation route may be
adjusted for duration of exposure. The Agency is planning to review
these stances to determine the most appropriate approach for the
future.
III.G. Characterization of the Health-Related Database
This section describes evaluation of the health-related database on
a particular chemical and provides criteria for judging the potential
for that chemical to produce reproductive toxicity under the exposure
conditions inherent in the database. This determination provides the
basis for judging whether the available data are sufficient to
characterize a hazard and to conduct quantitative dose-response
analyses. It also should provide a summary and evaluation of the
existing data and identify data gaps for an agent that is judged to
have insufficient information to proceed with a quantitative dose-
response analysis. Characterizing the available evidence in this way
clarifies the strengths and uncertainties in a particular database. It
does not address the level of concern, nor does it completely address
determining relevance of available data for estimating human risk.
Issues concerning relevance of mechanisms of action and types of
effects observed should be included in the hazard characterization.
Both level of concern and relevance are discussed further as part of
the final characterization of risk, taking into account the information
concerning potential human exposure. Data from all potentially relevant
studies, whether indicative of potential hazard or not, should be
included in the hazard characterization.
A complex interrelationship exists among study design, statistical
analysis, and biologic significance of the data. Thus, substantial
scientific judgment, based on experience with reproductive toxicity
data and with the principles of study design and statistical analysis,
may be required to evaluate the database adequately. In some cases, a
database may contain conflicting data. In these instances, the risk
assessor must consider each study's strengths and weaknesses within the
context of the overall database to characterize the evidence for
assessing the potential hazard for reproductive toxicity. Scientific
judgment is always necessary and, in many cases, interaction with
scientists in specific disciplines (e.g., reproductive toxicology,
epidemiology, genetic toxicology, statistics) is recommended.
A scheme for judging the available evidence on the reproductive
toxicity of a particular agent is presented below (Table 6). The scheme
contains two broad categories, ``Sufficient'' and ``Insufficient,''
which are defined in Table 6. Data from all available studies, whether
or not indicative of potential concern, are evaluated and used in the
hazard characterization for reproductive toxicity. The primary
considerations are the human data, if available, and the experimental
animal data. The judgment of whether data are sufficient or
insufficient should consider a variety of parameters that contribute to
the overall quality of the data, such as the power of the studies
(e.g., sample size and variation in the data), the number and types of
endpoints examined,
[[Page 56304]]
replication of effects, relevance of route and timing of exposure for
both human and experimental animal studies, and the appropriateness of
the test species and dose selection in experimental animal studies. In
addition, pharmacokinetic data and structure-activity considerations,
data from other toxicity studies, as well as other factors that may
affect the overall decision about the evidence, should be taken into
account.
Table 6.--Categorization of the Health-Related Database
------------------------------------------------------------------------
-------------------------------------------------------------------------
Sufficient Evidence
The Sufficient Evidence category includes data that collectively
provide enough information to judge whether or not a reproductive
hazard exists within the context of effect as well as dose, duration,
timing, and route of exposure. This category may include both human and
experimental animal evidence.
Sufficient Human Evidence
This category includes agents for which there is convincing evidence
from epidemiologic studies (e.g., case control and cohort) to judge
whether exposure is causally related to reproductive toxicity. A case
series in conjunction with other supporting evidence also may be judged
as Sufficient Evidence. An evaluation of epidemiologic and clinical
case studies should discuss whether the observed effects can be
considered biologically plausible in relation to chemical exposure.
Sufficient Experimental Animal Evidence/Limited Human Data
This category includes agents for which there is sufficient evidence
from experimental animal studies and/or limited human data to judge if
a potential reproductive hazard exists. Generally, agents that have
been tested according to EPA's two-generation reproductive effects test
guidelines (but not limited to such designs) would be included in this
category. The minimum evidence necessary to determine if a potential
hazard exists would be data demonstrating an adverse reproductive
effect in a single appropriate, well-executed study in a single test
species. The minimum evidence needed to determine that a potential
hazard does not exist would include data on an adequate array of
endpoints from more than one study with two species that showed no
adverse reproductive effects at doses that were minimally toxic in
terms of inducing an adverse effect. Information on pharmacokinetics,
mechanisms, or known properties of the chemical class may also
strengthen the evidence.
Insufficient Evidence
This category includes agents for which there is less than the minimum
sufficient evidence necessary for assessing the potential for
reproductive toxicity. Included are situations such as when no data are
available on reproductive toxicity; as well as for data bases from
studies on test animals or humans that have a limited study design or
conduct (e.g., small numbers of test animals or human subjects,
inappropriate dose selection or exposure information, other
uncontrolled factors); data from studies that examined only a limited
number of endpoints and reported no adverse reproductive effects; or
data bases that were limited to information on structure-activity
relationships, short-term or in vitro tests, pharmacokinetic data, or
metabolic precursors.
------------------------------------------------------------------------
In general, the characterization is based on criteria defined by
these Guidelines as the minimum evidence necessary to characterize a
hazard and conduct dose-response analyses. Establishing the minimum
human evidence to proceed with quantitative analyses based on the human
data is often difficult because there may be considerable variations in
study designs and study group selection. The body of human data should
contain convincing evidence as described in the ``Sufficient Human
Evidence'' category. Because the human data necessary to judge whether
or not a causal relationship exists are generally limited, few agents
can be classified in this category. Agents that have been tested in
laboratory animals according to EPA's two-generation reproductive
effects test guidelines (U.S. EPA, 1982, 1985b, 1996a), but not limited
to such designs (e.g., a continuous breeding study with two
generations), generally would be included in the ``Sufficient
Experimental Animal Evidence/Limited Human Data'' category. There are
occasions in which more limited data regarding the potential
reproductive toxicity of an agent (e.g., a one-generation reproductive
effects study, a standard subchronic or chronic toxicity study in which
the reproductive organs were well examined) are available. If
reproductive toxicity is observed in these limited studies, the data
may be used to the extent possible to reach a decision regarding hazard
to the reproductive system, including determination of an RfD or RfC.
In cases in which such limited data are available, it would be
appropriate to adjust the uncertainty factor to reflect the attendant
increased uncertainty regarding the use of these data until more
definitive data are developed. Identification of the increased
uncertainty and justification for the adjustment of the uncertainty
factor should be stated clearly.
Because it is more difficult, both biologically and statistically,
to support a finding of no apparent hazard, more data are generally
required to support this conclusion than a finding for a potential
hazard. For example, to judge whether a hazard for reproductive
toxicity could exist for a given agent, the minimum evidence could be
data from a single appropriate, well-executed study in a single test
species that demonstrates an adverse reproductive effect, or suggestive
evidence from adequately conducted clinical or epidemiologic studies.
As in all situations, it is important that the results be biologically
plausible and consistent. On the other hand, to judge whether an agent
is unlikely to pose a hazard for reproductive toxicity, the minimum
evidence would include data on an array of endpoints and from studies
with more than one species that showed no reproductive effects at doses
that were otherwise minimally toxic to the adult animal. In addition,
there may be human data from appropriate studies that are supportive of
no apparent hazard. In the event that a substantial database exists for
a given chemical, but no single study meets current test guidelines,
the risk assessor should use scientific judgment to determine whether
the composite database may be viewed as meeting the ``Sufficient''
criteria.
Some important considerations in determining the confidence in the
health database are as follows:
Data of equivalent quality from human exposures are given
more weight than data from exposures of test species.
Although a single study of high quality could be
sufficient to achieve a relatively high level of confidence,
replication increases the confidence that may be placed in such
results.
Data are available from one or more in vivo studies of
acceptable quality with humans or other mammalian species that are
believed to be predictive of human responses.
Data exhibit a dose-response relationship.
Results are statistically significant and biologically
plausible.
When multiple studies are available, the results are
consistent.
Sufficient information is available to reconcile
discordant data.
Route, level, duration, and frequency of exposure are
appropriate.
An adequate array of endpoints has been examined.
The power and statistical treatment of the studies are
appropriate.
Any statistically significant deviation from baseline levels for an
in vivo effect warrants closer examination. To determine whether such a
deviation constitutes an adverse effect requires an understanding of
its role within a complex system and the determination of whether a
``true effect'' has beenobserved. Application of the above criteria,
combined with guidance presented in Section III.B. can facilitate such
determinations.
The greatest confidence for identification of a reproductive hazard
should be placed on significant adverse effects on sexual behavior,
fertility or development, or other endpoints that are directly related
to reproductive function such as menstrual (estrous) cycle normality,
sperm evaluations, reproductive histopathology, reproductive organ
weights, and reproductive endocrinology. Agents producing adverse
effects on these endpoints can be assigned to the ``Sufficient
Evidence'' category if study quality is adequate.
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Less confidence should be placed in results when only measures such
as in vitro tests, data from nonmammals, or structure-activity
relationships are available, but positive results may trigger follow-up
studies that extend the preliminary data and determine the extent to
which function might be affected. Results from these types of studies
alone, whether or not they demonstrate an effect, may be suggestive,
but should be assigned to the ``Insufficient Evidence'' category.
The absence of effects in test species on the endpoints that are
evaluated routinely (i.e., fertility, histopathology, prenatal
development, and organ weights) may constitute sufficient evidence to
place a low priority on the potential reproductive toxicity of a
chemical. However, in such cases, careful consideration should be given
to the sensitivity of these endpoints and to the quality of the data on
these endpoints. Consideration also should be given to the possibility
of adverse effects that may not be reflected in these routine measures
(e.g., germ-cell mutation, alterations in estrous cyclicity or sperm
measures such as motility or morphology, functional effects from
developmental exposures).
Judging that the health database indicates a potential reproductive
hazard does not mean that the agent will be a hazard at every exposure
level (because of the assumption of a nonlinear dose-response) or in
every situation (e.g., the type and degree of hazard may vary
significantly depending on route and timing of exposure). In the final
risk characterization, the summary of the hazard characterization
should always be presented with information on the quantitative dose-
response analysis and, if available, with the human exposure estimates.
IV. Quantitative Dose-Response Analysis
In quantitative dose-response assessment, a nonlinear dose-response
is assumed for noncancer health effects unless mode of action or
pharmacodynamic information indicate otherwise. If sufficient data are
available, a biologically based approach should be used on a chemical-
specific basis to predict the shape of the dose-response curve below
the observable range. It is plausible that certain biologic processes
(e.g., Sertoli cell barrier selectivity, metabolic and repair
capabilities of the germ cells) may impede the attainment or
maintenance of concentrations of the agent at the target site following
exposure to low-dose levels that would be associated with adverse
effects. The assumption of a nonlinear dose-response suggests that the
application of adequate uncertainty factors to a NOAEL, LOAEL, or
benchmark dose will result in an exposure level for all humans that is
not attended with significant risk above background. With a linear
dose-response, it is assumed that some risk exists at any level of
exposure, with risk decreasing as exposure decreases.
The NOAEL is the highest dose at which there is no significant
increase in the frequency of an adverse effect in any manifestation of
reproductive toxicity compared with the appropriate control group in a
database having sufficient evidence for use in a risk assessment. The
LOAEL is the lowest dose at which there is a significant increase in
the frequency of adverse reproductive effects compared with the
appropriate control group in a database having sufficient evidence. A
significant increase may be based on statistical significance or on a
biologically significant trend. Evidence for biological significance
may be strengthened by mode of action or other biochemical evidence at
lower exposure levels that supports the causation of such an effect.
The existence of a NOAEL in an experimental animal study does not show
the shape of the dose-response below the observable range; it only
defines the highest level of exposure under the conditions of the study
that is not associated with a significant increase in an adverse
effect. Alternatively, mathematical modeling of the dose-response
relationship may be performed in the experimental range. This approach
can be used to determine a benchmark dose, which may be used in place
of the NOAEL as a point of departure for calculating an RfD, RfC, MOE,
or other exposure estimates.
Several limitations in the use of the NOAEL have been described
(Kimmel, C.A. and Gaylor, 1988; U.S. EPA, 1995b): (1) Use of the NOAEL
focuses only on the dose that is the NOAEL and does not incorporate
information on the slope of the dose-response curve or the variability
in the data; (2) Because data variability is not taken into account
(i.e., confidence limits are not used), the NOAEL will likely be higher
with decreasing sample size or poor study conduct, either of which are
usually associated with increasing variability in the data; (3) The
NOAEL is limited to one of the experimental doses; (4) The number and
spacing of doses in a study can influence the dose that is chosen for
the NOAEL; and (5) Because the NOAEL is defined as a dose that does not
produce an observable change in adverse responses from control levels
and is dependent on the power of the study, theoretically the risk
associated with it may fall anywhere between zero and an incidence just
below that detectable from control levels (usually in the range of 7%
to 10% for quantal data). The 95% upper confidence limit on
developmental toxicity risk at the NOAEL has been estimated for several
data sets to be 2% to 6% (Crump, 1984; Gaylor, 1989); similar
evaluations have not been conducted on data for other reproductive
effects. Because of the limitations associated with the use of the
NOAEL, the Agency is beginning to use the benchmark dose approach for
quantitative dose-response evaluation when sufficient data are
available.
Calculation and use of the benchmark dose are described in the EPA
document The Use of the Benchmark Dose Approach in Health Risk
Assessment (U.S. EPA, 1995b). The Agency is currently developing
guidance for application of the benchmark dose, including a decision
process to use for the various steps in the analysis (U.S. EPA, 1996c).
The benchmark dose is based on a model-derived estimate of a particular
incidence level, such as a 5% or 10% incidence. The BMD/C for a given
endpoint serves as a consistent point of departure for low-dose
extrapolation. In some cases, mode of action data may be sufficient to
estimate a BMD/C at levels below the observable range for the health
effect of concern. A benchmark response (BMR) of 5% is usually the
lowest level of risk that can be estimated adequately for binomial
endpoints from standard developmental toxicity studies (Allen et al.,
1994a, b). For fetal weight, a continuous endpoint, a 5% change from
the control mean was near the limit of detection for standard prenatal
toxicity studies (Kavlock et al., 1995). The modeling approaches that
have been proposed for developmental toxicity (U.S. EPA, 1995b) are,
for the most part, curve-fitting models that have biological
plausibility, but do not incorporate mode of action. Similar approaches
can be applied to other reproductive toxicity data to derive dose-
response curves for data in the observed dose range, but may or may not
accurately predict risk at low levels of exposure. Further guidance on
the use of the BMD/C is being developed by the Agency (U.S. EPA,
1996c).
The RfD or RfC for reproductive toxicity is an estimate of a daily
exposure to the human population that is assumed to be without
appreciable risk of deleterious reproductive effects over a lifetime of
exposure. The RfD or RfC is derived by applying uncertainty factors to
the NOAEL, or the LOAEL if a NOAEL is not available, or to the
[[Page 56306]]
benchmark dose. Because of the short duration of most studies of
developmental toxicity, a unique value (RfDDT or RfCDT) is
determined for adverse developmental effects. For adverse reproductive
effects on endpoints other than those of developmental toxicity, no
special designator is attached. Data on reproductive toxicity
(including developmental toxicity) are considered along with other data
on a particular chemical in deriving an RfD or RfC.
The effect used for determining the NOAEL, LOAEL, or benchmark dose
in deriving the RfD or RfC is the most sensitive adverse reproductive
endpoint (i.e., the critical effect) from the most appropriate or, in
the absence of such information, the most sensitive mammalian species
(see Sections II and III.B.1.). Uncertainty factors for reproductive
and other forms of systemic toxicity applied to the NOAEL or benchmark
dose generally include factors of 3 or 10 each for interspecies
variation and for intraspecies variation. Additional factors may be
applied to account for other uncertainties that may exist in the
database. In circumstances where only a LOAEL is available, the use of
an additional uncertainty factor of up to 10 may be required, depending
on the sensitivity of the endpoints evaluated, adequacy of dose levels
tested, or general confidence in the LOAEL.
Other areas of uncertainty may be identified and modifying factors
used depending on the characterization of the database (e.g., if the
only data available are from a one-generation reproductive effects
study; see Section III.G.), data on pharmacokinetics, or other
considerations that may alter the level of confidence in the data (U.S.
EPA, 1987). The total size of the uncertainty factor will vary from
agent to agent and requires scientific judgment, taking into account
interspecies differences, variability within species, the slope of the
dose-response curve, the types of reproductive effects observed, the
background incidence of the effects, the route of administration, and
pharmacokinetic data.
The NOAEL, LOAEL, or the benchmark dose is divided by the total
uncertainty factor selected for the critical effect in the most
appropriate or most sensitive mammalian species to determine the RfD or
RfC. If the NOAEL, LOAEL, or benchmark dose for other forms of systemic
toxicity is lower than that for reproductive toxicity, this should be
noted in the risk characterization, and this value should be compared
with data from other studies in which adult animals are exposed. Thus,
reproductive toxicity data should be discussed in the context of other
toxicity data.
It has generally been assumed that there is a nonlinear dose-
response for reproductive toxicity. This is based on known homeostatic,
compensatory, or adaptive mechanisms that must be overcome before a
toxic endpoint is manifested and on the rationale that cells and organs
of the reproductive system and the developing organism are known to
have some capacity for repair of damage. However, in a population,
background levels of toxic agents and preexisting conditions may
increase the sensitivity of some individuals in the population. Thus,
exposure to a toxic agent may result in an increased risk of adverse
effects for some, but not necessarily all, individuals within the
population.
Efforts are underway to develop models that are more biologically
based. These models should provide a more accurate estimation of low-
dose risk to humans. The development of biologically based dose-
response models in reproductive toxicology has been impeded by a number
of factors, including limited understanding of the biologic mechanisms
underlying reproductive toxicity, intra- and interspecies differences
in the types of reproductive events, lack of appropriate
pharmacokinetic data, and inadequate information on the influence of
other types of systemic toxicity on the dose-response curve. Current
research on modes of action in reproductive toxicology is promising and
may provide data that are useful for appropriate modeling in the near
future.
Utilization of Information in Risk Characterization
The hazard characterization and quantitative dose-response
evaluations are incorporated into the final characterization of risk
along with information on estimates of human exposure. The analysis
depends on and should describe scientific judgments as to the accuracy
and sufficiency of the health-related data in experimental animals and
humans (if available), the biologic relevance of significant effects,
and other considerations important in the interpretation and
application of data to humans. Scientific judgment is always necessary,
and in many cases, interaction with scientists in specific disciplines
(e.g., reproductive toxicology, epidemiology, statistics) is
recommended.
V. Exposure Assessment
To obtain a quantitative estimate of risk for the human population,
an estimate of human exposure is required. The Guidelines for Exposure
Assessment (U.S. EPA, 1992) have been published separately and will not
be discussed in detail here. Rather, issues important to reproductive
toxicity risk assessment are addressed. In general, the exposure
assessment describes the magnitude, duration, schedule, and route of
human exposure. Ideally, existing body burden as well as internal
circulating and target organ exposure information for the agent of
concern and other synergistic or antagonistic agents should be
described. It should include information on the purpose, scope, level
of detail and approach used, including estimates of exposure and dose
by pathway and route for populations, subpopulations, and individuals
in a manner that is appropriate for the intended risk characterization.
It also should provide an evaluation of the overall level of confidence
in the estimate(s) of exposure and dose and the conclusions drawn. This
information is usually developed from monitoring data, from estimates
based on modeling of environmental exposures, and from application of
paradigms to exposure data bases. Often quantitative estimates of
exposures may not be available (e.g., workplace or environmental
measurements). In such instances, employment or residential histories
also may be used in characterizing exposure in a qualitative sense. The
potential use of biomarkers as indicators of exposure is an area of
active interest.
Studies of occupational populations may provide valuable
information on the potential environmental health risks for certain
agents. Exposures among environmentally exposed human populations tend
to be lower (but of longer duration) than those in studies of
occupationally exposed populations and therefore may require more
observations to assure sufficient statistical power. Also,
reconstruction of exposures is more difficult in an environmental study
than in those done in workplace settings where industrial hygiene
monitoring may provide more detailed exposure data.
The nature of the exposure may be defined at a particular point in
time or may reflect cumulative exposure. Each approach makes an
assumption about the underlying relationship between exposure and
outcome. For example, a cumulative exposure measure assumes that total
exposure is important, with a greater probability of effect with
greater total exposure or body burden. A dichotomous exposure measure
(ever exposed versus never exposed) assumes an irreversible effect of
exposure. Models that define exposure only at a
[[Page 56307]]
specific time may assume that only the present exposure is important
(Selevan and Lemasters, 1987). The appropriate exposure model depends
on the biologic processes affected and the nature of the chemical under
study. Thus, a cumulative or dichotomous exposure model may be
appropriate if injury occurs in cells that cannot be replaced or
repaired (e.g., oocytes); on the other hand, a concurrent exposure
model may be appropriate for cells that are being generated continually
(e.g., spermatids).
There are a number of unique considerations regarding the exposure
assessment for reproductive toxicity. Exposure at different stages of
male and female development can result in different outcomes. Such age-
dependent variation has been well documented in both experimental
animal and human studies. Prenatal and neonatal treatment can
irreversibly alter reproductive function and other aspects of
development in a manner or to an extent that may not be predicted from
adult-only exposure. Moreover, chemicals that alter sexual
differentiation in rodents during these periods may have similar
effects in humans, because the mechanisms underlying these
developmental processes appear to be similar in all mammalian species
(Gray, 1991).
The susceptibility of elderly males and females to chemical insult
has not been well studied. Although procreative competence may not be a
major health concern with elderly individuals, other biologic functions
maintained by the gonads (e.g., hormone production) are of significance
(Walker, 1986). An exposure assessment should characterize the
likelihood of exposure of these different subgroups (embryo or fetus,
neonate, juvenile, young adult, older adult) and the risk assessment
should factor in the susceptibility of different age groups to the
extent possible.
The relationship between time or duration of exposure and
observation of male reproductive effects has particular significance
for short-term exposures. Spermatogenesis is a temporally synchronized
process. In humans, germ cells that were spermatozoa, spermatids,
spermatocytes, or spermatogonia at the time of an acute exposure
require 1 to 2, 3 to 5, 5 to 8, or 8 to 12 weeks, respectively, to
appear in an ejaculate. That timing may vary somewhat depending on
degree of sexual activity. It is possible that an endpoint may be
examined too early or too late to detect an effect if only a particular
cell type was affected during a relatively brief exposure to an agent.
The absence of an effect when observations were made too late suggests
either a reversible effect or no effect. However, an effect that is
reversible at lower exposures might become irreversible with higher or
longer exposures or exposure of a more susceptible individual. Thus,
the failure to detect transient effects because of improper timing of
observations may be important. If information is available on the type
of effect expected from a class of agents, it may be possible to
evaluate whether the timing of endpoint measurement relative to the
timing of the short-term exposure is appropriate. Some information on
the appropriateness of the protocol can be obtained if test animal data
are available to identify the most sensitive cell type or the putative
mechanism of action for a given agent.
Compared with acute exposures, the link between exposure and
outcome may be more apparent with relatively constant subchronic or
longer exposures that are of sufficient duration to cover all phases of
spermatogenesis (Russell et al., 1990). Assessments may be made at any
time after this point as long as exposure remains constant. Time
required for the agent or metabolite to attain steady-state levels
should also be considered. Again, application of models of exposure
(e.g., dichotomous, concurrent, or cumulative) depends on the suspected
target and chemical mechanism of action.
The reversibility of an adverse effect on the reproductive system
can be affected by the degree and duration of exposure (Clegg, 1995).
The degree of stem cell loss is inversely related to the degree of
restoration of sperm production, because repopulation of the germinal
epithelium is dependent on the stem cells (Meistrich, 1982; Foote and
Berndtson, 1992). For agents that bioaccumulate, increasing duration of
exposure may also increase the extent of damage to the stem cell
population. Damage to other spermatogenic cell types reduces the number
of sperm produced, but recovery should occur when the toxic agent is
removed. Less is known about the effects of toxicity on the Sertoli
cells. Temporary impairment of Sertoli cell function may produce long-
lasting effects on spermatogenesis. Destruction of Sertoli cells or
interference with their proliferation before puberty are irreversible
effects because replication ceases after puberty. Sertoli cells are
essential for support of the spermatogenic process and loss of those
cells results in a permanent reduction of spermatogenic capability
(Foster, 1992).
When recovery is possible, the duration of the recovery period is
determined by the time for regeneration (for stem cells) and
repopulation of the affected spermatogenic cell types and appearance of
those cells as sperm in the ejaculate. The time required for these
events to occur varies with the species, the pharmacokinetic properties
of the agent, the extent to which the stem cell population has been
destroyed, and the degree of sublethal toxicity inflicted on the stem
cells or Sertoli cells. When the stem cell population has been
partially destroyed, humans require more time than mice to reach the
same degree of recovery (Meistrich and Samuels, 1985).
Unique considerations in the assessment of female reproductive
toxicity include the duration and period of exposure as related to the
development or stage of reproductive life (e.g., prenatal,
prepubescent, reproductive, or postmenopausal) or considerations of
different physiologic states (e.g., nonpregnant, pregnant, lactating).
For infertility, a cumulative exposure measure assumes destruction of
increasing numbers of primary oocytes with greater lifetime exposure or
increasing body burden. However, humans may be exposed to varying
levels of an agent within the study period. Exposures during certain
critical points in the reproductive process may affect the outcomes
observed in humans (Lemasters and Selevan, 1984). In test species,
perinatal exposure to androgens or estrogens such as zearalenone,
methoxychlor, and DDT (Bulger and Kupfer, 1985; Gray et al., 1985) have
been shown to advance puberty and masculinize females. Similar effects
have been reported in humans (both sexes) exposed neonatally to
synthetic estrogens or progestins (Steinberger and Lloyd, 1985;
Schardein, 1993). Studies using test species also have shown that
exposure to some environmental agents such as ionizing radiation
(Dobson and Felton, 1983) and glycol ethers (Heindel et al., 1989) can
deplete the pool of primordial follicles and thus significantly shorten
the female's reproductive lifespan. Furthermore, exposure to compounds
at different stages of the ovarian cycle can disrupt or delay
follicular recruitment and development (Armstrong, 1986), ovulation
(Everett and Sawyer, 1950; Terranova, 1980), and ovum transport
(Cummings and Perreault, 1990). Compounds that delay ovulation can lead
to significant alterations in egg viability (Peluso et al., 1979),
fertilizability of the egg (Fugo and Butcher, 1966; Butcher and Fugo,
1967; Butcher et al., 1975), and a reduction in litter size (Fugo and
Butcher, 1966). After ovulation, single exposures to
[[Page 56308]]
microtubule poisons such as carbendazim may impair the completion of
meiosis in the fertilized oocyte with adverse developmental
consequences (Perreault et al., 1992; Zuelke and Perreault, 1995).
Thus, knowledge of when acute exposures occur relative to the female's
lifespan and reproductive cycle can provide insight into how an agent
disrupts reproductive function.
DES is a classic example of an agent causing different effects on
the reproductive system in the developing organism compared with those
in adults (McLachlan, 1980). DES, as well as other agents with
estrogenic or anti-androgenic activity, interferes with the development
of the Mullerian and Wolffian duct systems and thereby causes
irreversible structural and functional damage to the developing
reproductive system. In adults, the reproductive effects that are
caused by the estrogenic activity of DES do not necessarily result in
permanent damage.
Unique considerations for developmental effects are duration and
period of exposure as related to stage of development (i.e., critical
periods) and the possibility that even a single exposure may be
sufficient to produce adverse developmental effects. Repeated exposure
is not a necessary prerequisite for developmental toxicity to be
manifested, although it should be considered in cases where there is
evidence of cumulative exposure or where the half-life of the agent is
long enough to produce an increasing body burden over time. For these
reasons, it is assumed that, in most cases, a single exposure at the
critical time in development is sufficient to produce an adverse
developmental effect. Therefore, the human exposure estimates used to
calculate the MOE for an adverse developmental effect or to compare to
the RfD or RfC are usually based on a single daily dose that is not
adjusted for duration or pattern (e.g., continuous or intermittent) of
exposure. For example, it would be inappropriate to use time-weighted
averages or adjustment of exposure over a different time frame than
that actually encountered (such as the adjustment of a 6-hour
inhalation exposure to account for a 24-hour exposure scenario) unless
pharmacokinetic data were available to indicate an accumulation with
continuous exposure. In the case of intermittent exposures, examination
of the peak exposures as well as the average exposure over the time of
exposure would be important.
It should be recognized that, based on the definitions used in
these Guidelines, almost any segment of the human population may be at
risk for a reproductive effect. Although the reproductive effects of
exposures may be manifested while the exposure is occurring (e.g.,
menstrual disorder, decreased sperm count, spontaneous abortion) some
effects may not be detectable until later in life (e.g., endocrine
disruption of reproductive tract development, premature reproductive
senescence due to oocyte depletion), long after exposure has ceased.
VI. Risk Characterization
VI.A. Overview
A risk characterization is an essential part of any Agency report
on risk whether the report is a preliminary one prepared to support
allocation of resources toward further study, a site-specific
assessment, or a comprehensive one prepared to support regulatory
decisions. A risk characterization should be prepared in a manner that
is clear, reasonable, and consistent with other risk characterizations
of similar scope prepared across programs in the Agency. It should
identify and discuss all the major issues associated with determining
the nature and extent of the risk and provide commentary on any
constraints limiting more complete exposition. The key aspects of risk
characterization are: (1) bridging risk assessment and risk management,
(2) discussing confidence and uncertainties, and (3) presenting several
types of risk information. In this final step of a risk assessment, the
risk characterization involves integration of toxicity information from
the hazard characterization and quantitative dose-response analysis
with the human exposure estimates and provides an evaluation of the
overall quality of the assessment, describes risk in terms of the
nature and extent of harm, and communicates results of the risk
assessment to a risk manager. A risk manager can then use the risk
assessment, along with other risk management elements, to make public
health decisions. The information should also assist others outside the
Agency in understanding the scientific basis for regulatory decisions.
Risk characterization is intended to summarize key aspects of the
following components of the risk assessment:
The nature, reliability, and consistency of the data used.
The reasons for selection of the key study(ies) and the
critical effect(s) and their relevance to human outcomes.
The qualitative and quantitative descriptors of the
results of the risk assessment.
The limitations of the available data, the assumptions
used to bridge knowledge gaps in working with those data, and
implications of using alternative assumptions.
The strengths and weaknesses of the risk assessment and
the level of scientific confidence in the assessment.
The areas of uncertainty, additional data/research needs
to improve confidence in the risk assessment, and the potential impacts
of the new research.
The risk characterization should be limited to the most significant
and relevant data, conclusions, and uncertainties. When special
circumstances exist that preclude full assessment, those circumstances
should be explained and the related limitations identified.
The following sections describe these aspects of the risk
characterization in more detail, but do not attempt to provide a full
discussion of risk characterization. Rather, these Guidelines point out
issues that are important to risk characterization for reproductive
toxicity. Comprehensive general guidance for risk characterization is
provided by Habicht (1992) and Browner (1995).
VI.B. Integration of Hazard Characterization, Quantitative Dose-
Response, and Exposure Assessments
In developing each component of the risk assessment, risk assessors
must make judgments concerning human relevance of the toxicity data,
including the appropriateness of the various test animal models for
which data are available, and the route, timing, and duration of
exposure relative to the expected human exposure. These judgments
should be summarized at each stage of the risk assessment process. When
data are not available to make such judgments, as is often the case,
the background information and assumptions discussed in the Overview
(Section I) provide default positions. The default positions used and
the rationale behind the use of each default position should be clearly
stated. In integrating the parts of the assessment, risk assessors must
determine if some of these judgments have implications for other
portions of the assessment, and whether the various components of the
assessment are compatible.
The description of the relevant data should convey the major
strengths and weaknesses of the assessment that arise from availability
and quality of data and the current limits of understanding of the
mechanisms of toxicity. Confidence in the results of a risk assessment
is a function of confidence in the results of
[[Page 56309]]
these analyses. Each section (hazard characterization, quantitative
dose-response analysis, and exposure assessment) should have its own
summary, and these summaries should be integrated into the overall risk
characterization. Interpretation of data should be explained, and risk
managers should be given a clear picture of consensus or lack of
consensus that exists about significant aspects of the assessment. When
more than one interpretation is supported by the data, the alternative
plausible approaches should be presented along with the strengths,
weaknesses, and impacts of those options. If one interpretation or
option has been selected over another, the rationale should be given;
if not, then both should be presented as plausible alternatives.
The risk characterization should not only examine the judgments,
but also should explain the constraints of available data and the state
of knowledge about the phenomena studied in making them, including:
The qualitative conclusions about the likelihood that the
chemical may pose a specific hazard to human health, the nature of the
observed effects, under what conditions (route, dose levels, time, and
duration) of exposure these effects occur, and whether the health-
related data are sufficient and relevant to use in a risk assessment.
A discussion of the dose-response patterns for the
critical effect(s) and their relationships to the occurrence of other
toxicity data, such as the shapes and slopes of the dose-response
curves for the various other endpoints; the rationale behind the
determination of the NOAEL, LOAEL, and/or benchmark dose; and the
assumptions underlying the estimation of the RfD, RfC, or other
exposure estimate.
Descriptions of the estimates of the range of human
exposure (e.g., central tendency, high end), the route, duration, and
pattern of the exposure, relevant pharmacokinetics, and the size and
characteristics of the various populations that might be exposed.
The risk characterization of an agent being assessed for
reproductive toxicity should be based on data from the most appropriate
species or, if such information is not available, on the most sensitive
species tested. It also should be based on the most sensitive indicator
of an adverse reproductive effect, whether in the male, the female
(nonpregnant or pregnant), or the developing organism, and should be
considered in relation to other forms of toxicity. The relevance of
this indicator to human reproductive outcomes should be described. The
rationale for those decisions should be presented.
If data to be used in a risk characterization are from a route of
exposure other than the expected human exposure, then pharmacokinetic
data should be used, if available, to extrapolate across routes of
exposure. If such data are not available, the Agency makes certain
assumptions concerning the amount of absorption likely or the
applicability of the data from one route to another (U.S. EPA, 1985a,
1986b). Discussion of some of these issues may be found in the
Proceedings of the Workshop on Acceptability and Interpretation of
Dermal Developmental Toxicity Studies (Kimmel, C.A. and Francis, 1990)
and Principles of Route-to-Route Extrapolation for Risk Assessment
(Gerrity et al., 1990). The risk characterization should identify the
methods used to extrapolate across exposure routes and discuss the
strengths and limitations of the approach.
The level of confidence in the hazard characterization and
quantitative dose-response evaluation should be stated to the extent
possible, including placement of the agent into the appropriate
category regarding the sufficiency of the health-related data (see
Section III.G.). A comprehensive risk assessment ideally includes
information on a variety of endpoints that provide insight into the
full spectrum of potential reproductive responses. A profile that
integrates both human and test species data and incorporates both
sensitive endpoints (e.g., properly performed and fully evaluated
histopathology) and functional correlates (e.g., fertility) allows more
confidence in a risk assessment for a given agent.
Descriptions of the nature of potential human exposures are
important for prediction of specific outcomes and the likelihood of
persistence or reversibility of the effect in different exposure
situations with different subpopulations (U.S. EPA, 1992; Clegg, 1995).
In the risk assessment process, risk is estimated as a function of
exposure, with the risk of adverse effects increasing as exposure
increases. Information on the levels of exposure experienced by
different members of the population is key to understanding the range
of risks that may occur. Where possible, several descriptors of
exposure such as the nature and range of populations and their various
exposure conditions, central tendencies, and high-end exposure
estimates should be presented. Differences among individuals in
absorption rates, metabolism, or other factors mean that individuals or
subpopulations with the same level and pattern of exposure may have
differing susceptibility. For example, the consequences of exposure can
differ markedly between developing individuals, young adults and aged
adults, including whether the effects are permanent or transient. Other
considerations relative to human exposures might include pregnancy or
lactation, potential for exposures to other agents, concurrent disease,
nutritional status, lifestyle, ethnic background and genetic
polymorphism, and the possible consequences. Knowledge of the molecular
events leading to induction of adverse effects may be of use in
determining the range of susceptibility in sensitive populations.
An outline to serve as a guide and formatting aid for developing
reproductive risk characterizations for chemical-specific risk
assessments can be found in Table 7. A common format will assist risk
managers in evaluating and using reproductive risk characterization.
The outline has two parts. The first part tracks the reproductive risk
assessment to bring forward its major conclusions. The second part
pulls the information together to characterize the reproductive risk.
Table 7.--Guide for Developing Chemical-Specific Risk Characterizations
for Reproductive Effects
------------------------------------------------------------------------
-------------------------------------------------------------------------
Part One
Summarizing Major Conclusions in Risk Characterization
I. Hazard Characterization
A. What is (are) the key toxicological study (or studies) that provides
the basis for health concerns for reproductive effects?
How good is the key study?
Are the data from laboratory or field studies? In a single or
multiple species?
What adverse reproductive endpoints were observed, and what
is the basis for the critical effect?
Describe other studies that support this finding.
[[Page 56310]]
Discuss any valid studies which conflict with this finding.
B. Besides the reproductive effect observed in the key study, are there
other health endpoints of concern? What are the significant data gaps?
C. Discuss available epidemiological or clinical data. For
epidemiological studies:
What types of data were used (e.g., human ecologic, case-
control or cohort studies, or case reports or series)?
Describe the degree to which exposures were described.
Describe the degree to which confounding factors were
considered.
Describe the degree to which other causal factors were
excluded.
D. How much is known about how (through what biological mechanism) the
chemical produces adverse reproductive effects?
Discuss relevant studies of mechanisms of action or
metabolism.
Does this information aid in the interpretation of the
toxicity data?
What are the implications for potential adverse reproductive
effects?
E. Comment on any nonpositive data in animals or people, and whether
these data were considered in the hazard characterization.
F. If adverse health effects have been observed in wildlife species,
characterize such effects by discussing the relevant issues as in A
through E above.
G. Summarize the hazard characterization and discuss the significance of
each of the following:
Confidence in conclusions
Alternative conclusions that are also supported by the data
Significant data gaps
Highlights of major assumptions
II. Characterization of Dose-Response
A. What data were used to develop the dose-response curve? Would the
result have been significantly different if based on a different data
set?
If laboratory animal data were used:
Which species were used?
Most sensitive, average of all species, or other?
Were any studies excluded? Why?
If epidemiological data were used:
Which studies were used?
Only positive studies, all studies, or some other combination?
Were any studies excluded? Why?
Was a meta-analysis performed to combine the epidemiological
studies?
What approach was used?
Were studies excluded? Why?
B. Was a model used to develop the dose-response curve and, if so, which
one? What rationale supports this choice? Is chemical-specific
information available to support this approach?
How was the RfD/RfC (or the acceptable range) calculated?
What assumptions and uncertainty factors were used?
What is the confidence in the estimates?
C. Discuss the route, level, and duration of exposure observed, as
compared to expected human exposures.
Are the available data from the same route of exposure as the
expected human exposures? If not, are pharmacokinetic data available
to extrapolate across route of exposure?
How far does one need to extrapolate from the observed data
to environmental exposures? One to two orders of magnitude? Multiple
orders of magnitude? What is the impact of such an extrapolation?
D. If adverse health effects have been observed in wildlife species,
characterize dose-response information using the process outlined in A
through C above.
III. Characterization of Exposure
A. What are the most significant sources of environmental exposure?
Are there data on sources of exposure from different media?
What is the relative contribution of different sources of exposure?
What are the most significant environmental pathways for exposure?
B. Describe the populations that were assessed, including the general
population, highly exposed groups, and highly susceptible groups.
C. Describe the basis for the exposure assessment, including any
monitoring, modeling, or other analyses of exposure distributions such
as Monte Carlo or krieging.
D. What are the key descriptors of exposure?
Describe the (range of) exposures to: ``average'' individuals, ``high-
end'' individuals, general population, high exposure group(s),
children, susceptible populations, males, females (nonpregnant,
pregnant, lactating).
How was the central tendency estimate developed?
What factors and/or methods were used in developing this estimate?
How was the high-end estimate developed?
Is there information on highly exposed subgroups?
Who are they?
What are their levels of exposure?
How are they accounted for in the assessment?
E. Is there reason to be concerned about cumulative or multiple
exposures because of biological, ethnic, racial, or socioeconomic
reasons?
F. If adverse reproductive effects have been observed in wildlife
species, characterize wildlife exposure by discussing the relevant
issues as in A through E above.
G. Summarize exposure conclusions and discuss the following:
Results of different approaches, i.e., modeling, monitoring,
probability distributions;
Limitations of each, and the range of most reasonable values;
Confidence in the results obtained, and the limitations to
the results
[[Page 56311]]
Part Two
Risk Conclusions and Comparisons
IV. Risk Conclusions
A. What is the overall picture of risk, based on the hazard,
quantitative dose-response, and exposure characterizations?
B. What are the major conclusions and strengths of the assessment in
each of the three main analyses (i.e., hazard characterization,
quantitative dose-response, and exposure assessment)?
C. What are the major limitations and uncertainties in the three main
analyses?
D. What are the science policy options in each of the three major
analyses?
What are the alternative approaches evaluated?
What are the reasons for the choices made?
V. Risk Context
A. What are the qualitative characteristics of the reproductive hazard
(e.g., voluntary vs. involuntary, technological vs. natural, etc.)?
Comment on findings, if any, from studies of risk perception that
relate to this hazard or similar hazards.
B. What are the alternatives to this reproductive hazard? How do the
risks compare?
C. How does this reproductive risk compare to other risks?
How does this risk compare to other risks in this regulatory program,
or other similar risks that the EPA has made decisions about?
Where appropriate, can this risk be compared with past Agency
decisions, decisions by other federal or state agencies, or common
risks with which people may be familiar?
Describe the limitations of making these comparisons.
D. Comment on significant community concerns which influence public
perception of risk.
VI. Existing Risk Information
Comment on other reproductive risk assessments that have been done on
this chemical by EPA, other federal agencies, or other organizations.
Are there significantly different conclusions that merit discussion?
VII. Other Information
Is there other information that would be useful to the risk manager or
the public in this situation that has not been described above?
------------------------------------------------------------------------
VI.C. Descriptors of Reproductive Risk
Descriptors of reproductive risk convey information and answer
questions about risk, with each descriptor providing different
information and insights. There are a number of ways to describe risk.
Details on how to use these descriptors can be obtained from the
guidance on risk characterization (Browner, 1995) from which some of
the information below has been extracted.
In most cases, the state of the science is not yet adequate to
define distributions of factors such as population susceptibility. The
guidance principles below discuss a variety of risk descriptors that
primarily reflect differences in estimated exposure. If a full
description of the range of susceptibility in the population cannot be
presented, an effort should be made to identify subgroups that, for
various reasons, may be particularly susceptible.
VI.C.1. Distribution of Individual Exposures
Risk managers are interested generally in answers to questions such
as: (1) Who are the people at the highest risk and why? (2) What is the
average risk or distribution of risks for individuals in the population
of interest? and (3) What are they doing, where do they live, etc.,
that might be putting them at this higher risk?
Exposure and reproductive risk descriptors for individuals are
intended to provide answers to these questions. To describe the range
of risks, both high-end and central tendency descriptors are used to
convey the distribution in risk levels experienced by different
individuals in the population. For the Agency's purposes, high-end risk
descriptors are plausible estimates of the individual risk for those
persons at the upper end of the risk distribution. Given limitations in
current understanding of variability in individuals' sensitivity to
agents that cause reproductive toxicity, high-end descriptors will
usually address high-end exposure or dose. Conceptually, high-end
exposure means exposure above approximately the 90th percentile of the
population distribution, but not higher than the individual in the
population who has the highest exposure. Central tendency descriptors
generally reflect central estimates of exposure or dose. The descriptor
addressing central tendency may be based on either the arithmetic mean
exposure (average estimate) or the median exposure (median estimate),
either of which should be clearly labeled. The selection of which
descriptor(s) to present in the risk characterization will depend on
the available data and the goals of the assessment.
VI.C.2. Population Exposure
Population risk refers to assessment of the extent of harm for the
population as a whole. In theory, it can be calculated by summing the
individual risks for all individuals within the subject population.
That task requires more information than is usually available.
Questions addressed by descriptors of population risk for reproductive
effects would include: What portion of the population is within a
specified range of some reference level, e.g., exceeds the RfD (a
dose), the RfC (a concentration), or other health concern level?
For reproductive effects, risk assessment techniques have not been
developed generally to the point of knowing how to add risk
probabilities, although Hattis and Silver (1994) have proposed
approaches for certain case-specific situations. Therefore, the
following descriptor is usually appropriate: An estimate of the
percentage of the population, or the number of persons, above a
specified level of risk or within a specified range of some reference
level (e.g., exceeds the RfD, RfC, LOAEL, or other specific level of
interest). The RfD or RfC is assumed to be a level below which no
significant risk occurs. Therefore, information from the exposure
assessment on the populations below the RfD or RfC (``not likely to be
at risk'') and above the RfD or RfC (``may be at risk'') may be useful
information for risk managers. Estimating the number of persons
potentially removed from the ``may be at risk'' category after a
contemplated action is taken may be particularly
[[Page 56312]]
useful to a risk manager considering possible actions to ameliorate
risk for a population. This descriptor must be obtained through
measuring or simulating the population distribution.
VI.C.3. Margin of Exposure
In the risk characterization, dose-response information and the
human exposure estimates may be combined either by comparing the RfD or
RfC and the human exposure estimate or by calculating the margin of
exposure (MOE). The MOE is the ratio of the NOAEL or benchmark dose
from the most appropriate or sensitive species to the estimated human
exposure level from all potential sources (U.S. EPA, 1985a). If a NOAEL
is not available, a LOAEL may be used in the calculation of the MOE,
but consideration for the acceptability would be different than when a
NOAEL is used. Considerations for the acceptability of the MOE are
similar to those for the selection of uncertainty factors applied to
the NOAEL, LOAEL, or the benchmark dose for the derivation of an RfD.
The MOE is presented along with the characterization of the database,
including the strengths and weaknesses of the toxicity and exposure
data, the number of species affected, and the information on dose-
response, route, timing, and duration. The RfD or RfC comparison with
the human exposure estimate and the calculation of the MOE are
conceptually similar, but may be used in different regulatory
situations.
The choice of approach is dependent on several factors, including
the statute involved, the situation being addressed, the database used,
and the needs of the decisionmaker. The RfD, RfC, or MOE are considered
along with other risk assessment and risk management issues in making
risk management decisions, but the scientific issues that should be
taken into account in establishing them have been addressed here.
VI.C.4. Distribution of Exposure and Risk for Different Subgroups
A risk manager might also ask questions about the distribution of
the risk burden among various segments of the subject population such
as the following: How do exposure and reproductive risk impact various
subgroups? and What is the population risk of a particular subgroup?
Questions about the distribution of exposure and reproductive risk
among such population segments require additional risk descriptors.
Highly Exposed
The purpose of this measure is to describe the upper end of the
exposure distribution, allowing risk managers to evaluate whether
certain individuals are at disproportionately high or unacceptably high
risk. The objective is to look at the upper end of the exposure
distribution to derive a realistic estimate of relatively highly
exposed individual(s). The ``high end'' of the risk distribution has
been defined (Habicht, 1992; Browner, 1995) as above the 90th
percentile of the actual (either measured or estimated) distribution.
Whenever possible, it is important to express the number or proportion
of individuals who comprise the selected highly exposed group and, if
data are available, discuss the potential for exposure at still higher
levels.
Highly exposed subgroups can be identified and, where possible,
characterized, and the magnitude of risk quantified. This descriptor is
useful when there is (or is expected to be) a subgroup experiencing
significantly different exposures or doses from those of the larger
population. These subpopulations may be identified by age, sex,
lifestyle, economic factors, or other demographic variables. For
example, toddlers who play in contaminated soil and consumers of large
amounts of fish represent subpopulations that may have greater
exposures to certain agents.
If population data are absent, it will often be possible to
describe a scenario representing high-end exposures using upper
percentile or judgment-based values for exposure variables. In these
instances, caution should be taken not to overestimate the high-end
values if a ``reasonable'' exposure estimate is to be achieved.
Highly Susceptible
Highly susceptible subgroups also can be identified and, if
possible, characterized, and the magnitude of risk quantified. This
descriptor is useful when the sensitivity or susceptibility to the
effect for specific subgroups is (or is expected to be) significantly
different from that of the larger population. Therefore, the purpose of
this measure is to quantify exposure of identified sensitive or
susceptible populations to the agent of concern. Sensitive or
susceptible individuals are those within the exposed population at
increased risk of expressing the adverse effect. Examples might be
pregnant or lactating women, women with reduced oocyte numbers, men
with ``borderline'' sperm counts, or infants. To calculate risk for
these subgroups, it will be necessary sometimes to use a different
dose-response relationship; e.g., upon exposure to a chemical, pregnant
or lactating women, elderly people, children of varying ages, and
people with certain illnesses may each be more sensitive than the
population as a whole.
In general, not enough is understood about the mechanisms of
toxicity to identify sensitive subgroups for most agents, although
factors such as age, nutrition, personal habits (e.g., smoking,
consumption of alcohol, and abuse of drugs), existing disease (e.g.,
diabetes or sexually transmitted diseases), or genetic polymorphisms
may predispose some individuals to be more sensitive to the
reproductive effects of various agents.
It is important to consider, however, that the Agency's current
methods for developing reference doses and reference concentrations
(RfDs and RfCs) are designed to protect sensitive populations. If data
on sensitive human populations are available (and there is confidence
in the quality of the data), then the RfD is based on the dose level at
which no adverse effects are observed in the sensitive population. If
no such data are available (for example, if the RfD is developed using
data from humans of average or unknown sensitivity), then an additional
3- to 10-fold factor may be used to account for variability between the
average human response and the response of more sensitive individuals
(see Section IV).
Generally, selection of the population segments to consider for
high susceptibility is a matter of either a prior interest in the
subgroup (e.g., environmental justice considerations), in which case
the risk assessor and risk manager can jointly agree on which subgroups
to highlight, or a matter of discovery of a sensitive or highly exposed
subgroup during the assessment process. In either case, once
identified, the subgroup can be treated as a population in itself and
characterized in the same way as the larger population using the
descriptors for population and individual risk.
VI.C.5. Situation-Specific Information
Presenting situation-specific scenarios for important exposure
situations and subpopulations in the form of ``what if?'' questions may
be particularly useful to give perspective to risk managers on possible
future events. The question being asked in these cases is, for any
given exposure level, what would be the resulting number or proportion
of individuals who may be exposed to levels above that value?
``What if * * *?'' questions, such as those that follow, can be
used to examine candidate risk management options:
[[Page 56313]]
What are the reproductive risks if a pesticide applicator
applies this pesticide without using protective equipment?
What are the reproductive risks if this site becomes
residential in the future?
What are the reproductive risks if we set the standard at
100 ppb?
Answering such ``what if?'' questions involves a calculation of
risk based on specific combinations of factors postulated within the
assessment. The answers to these ``what if?'' questions do not, by
themselves, give information about how likely the combination of values
might be in the actual population or about how many (if any) persons
might be subjected to the potential future reproductive risk. However,
information on the likelihood of the postulated scenario would be
desirable to include in the assessment.
When addressing projected changes for a population (either expected
future developments or consideration of different regulatory options),
it usually is appropriate to calculate and consider all the
reproductive risk descriptors discussed above. When central tendency or
high-end estimates are developed for a scenario, these descriptors
should reflect reasonable expectations about future activities. For
example, in site-specific risk assessments, future scenarios should be
evaluated when they are supported by realistic forecasts of future land
use, and the reproductive risk descriptors should be developed within
that context.
VI.C.6. Evaluation of the Uncertainty in the Risk Descriptors
Reproductive risk descriptors are intended to address variability
of risk within the population and the overall adverse impact on the
population. In particular, differences between high-end and central
tendency estimates reflect variability in the population but not the
scientific uncertainty inherent in the risk estimates. As discussed
above there will be uncertainty in all estimates of reproductive risk.
These uncertainties can include measurement uncertainties, modeling
uncertainties, and assumptions to fill data gaps. Risk assessors should
address the impact of each of these factors on the confidence in the
estimated reproductive risk values.
Both qualitative and quantitative evaluations of uncertainty
provide useful information to users of the assessment. The techniques
of quantitative uncertainty analysis are evolving rapidly and both the
SAB (Loehr and Matanoski, 1993) and the NRC (1994) have urged the
Agency to incorporate these techniques into its risk analyses. However,
it should be noted that a probabilistic assessment that uses only the
assessor's best estimates for distributions of population variables
addresses variability, but not uncertainty. Uncertainties in the
estimated risk distribution need to be evaluated separately. An
approach has been proposed for estimating distribution of uncertainty
in noncancer risk assessments (Baird et al., 1996).
VI.D. Summary and Research Needs
These Guidelines summarize the procedures that the EPA will follow
in evaluating the potential for agents to cause reproductive toxicity.
They discuss the assumptions that must be made in risk assessment for
reproductive toxicity because of gaps in our knowledge about underlying
biologic processes and how these compare across species. Research to
improve the interpretation of data and interspecies extrapolation is
needed. This research includes studies that: (1) more completely
characterize and define female and male reproductive endpoints, (2)
more completely characterize the types of developmental toxicity
possible, (3) evaluate the interrelationships among endpoints, (4)
examine quantitative extrapolation between endpoints (e.g., sperm
count) and function (e.g., fertility), (5) provide a better
understanding of the relationships between reproductive toxicity and
other forms of toxicity, (6) explore pharmacokinetic disposition of the
target, and (7) examine mechanistic phenomena related to
pharmacokinetic disposition. These types of studies, along with further
evaluation of a nonlinear dose-response for susceptible populations,
should provide methods to more precisely assess risk.
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Part B. Response to Science Advisory Board and Public Comments
I. Introduction
A notice of availability for public comment of these Guidelines was
published in the Federal Register (FR) in February 1994. Seven
responses were received. These Guidelines were presented to the
Environmental Health Committee of the Science Advisory Board (SAB) on
July 19, 1994. The report of the SAB was provided to the Agency in May
1995, with further communication from the SAB Executive Committee
provided in December 1995.
The SAB and public comments were diverse and represented varying
perspectives. Many of the comments were favorable and expressed
agreement with positions taken in the proposed guidelines. A number of
the comments addressed items that were more pertinent to testing
guidance than risk assessment guidance or were otherwise beyond the
scope of these Guidelines. Some of those were generic issues that are
not system specific. Others were topics that have not been developed
sufficiently and should be viewed as research issues. There were
conflicting views about the need to provide additional detailed
guidance about decision-making in the evaluation process as opposed to
promoting extensive use of scientific judgment. Also, comments provided
specific suggestions for clarification of details.
II. Response to Science Advisory Board Comments
In general, the SAB found ``the overall scientific foundations of
the draft guidelines' positions to be generally sound.'' However,
recommendations were made to improve specific areas.
The SAB recommended that EPA retain separate sections for
identification and dose-response assessment in the draft guidelines. In
subsequent meetings involving the SAB Executive Committee, members of
the Clean Air Scientific Advisory Committee, and the Environmental
Health Committee, this issue was explored further. After discussion,
the SAB agreed with expanding the hazard identification to include
certain components of the dose-response assessment. The resulting
hazard characterization provides an evaluation of hazard within the
context of the dose, route, timing, and duration of exposure. The next
step, the dose-response analysis, quantitatively evaluates the
relationship between dose or exposure and severity or probability of
effect in humans. EPA has revised these Guidelines to reflect that
position which is consistent also with the 1994 NRC report, Science and
Judgment in Risk Assessment. The SAB suggested an alternative scheme
for characterizing health effects data in Table 5. The Agency's intent
for Table 5 is not to characterize the available data, but rather to
judge whether the database is sufficient to proceed further in the risk
assessment process. The text has been modified to clarify the intended
use of this table and to ensure that it is consistent with the
reorganization of the Guidelines into separate hazard characterization
and quantitative dose-response analysis sections.
The SAB supported the concept of using a gender neutral default
assumption, but indicated that more discussion to support this
assumption was needed. In particular, the Committee indicated that a
fuller discussion is needed on ``information to the contrary'' (to
obviate the need for making this default assumption), as well as
additional guidance for using this and other default assumptions in
risk characterization. The Agency agrees with this recommendation and
provides further guidance on the use of the gender neutral default
assumption. In keeping with recent Agency guidance on risk
characterization, discussion on the use of default assumptions has been
expanded in the risk characterization section of these Guidelines.
The SAB in its reviews of the reproductive toxicity and
neurotoxicity risk assessment guidelines discussed assumptions about
the behavior of the dose-response curve. The SAB's advice has been that
the Agency examine available data first, and only use
[[Page 56322]]
nonlinear behavior as a default if available data do not define the
dose-response curve. The SAB also recommended that the benchmark dose
method be considered as a possible alternative to the NOAEL/LOAEL
approach. The Agency agrees.
The SAB recommended that more discussion be devoted to the issue of
disruption of endocrine systems by environmental agents. The section on
Endocrine Evaluations has been expanded to include endocrine disruption
of the reproductive system during development in addition to effects on
adults.
The SAB supported the principle in the Guidelines that more than
one negative study is necessary to judge that a chemical is unlikely to
pose a reproductive hazard. That principle has been retained and, as
recommended by the SAB, an explicit statement included that data from a
second species are necessary to determine that sufficient information
is available to indicate that an agent is unlikely to pose a hazard.
The SAB recommended that the topic of susceptible populations be
expanded and that the Guidelines should indicate that relevant
information be incorporated into risk assessments when possible. To
address this issue, the Agency has emphasized potential differences in
risks in children at different stages of development, females
(including pregnant and lactating females), and males, and indicated
that relevant information on differential risks for susceptible
populations should be included in the risk characterization section
when available. When specific information on differential risks is not
available, the Agency will continue to apply a default uncertainty
factor to account for potential differences in susceptibility.
The SAB recommended that the Agency provide more specific guidance
for exposure assessment issues that arise when characterizing exposure
for reproductive toxicants. The Agency agrees and has indicated that an
exposure assessment: include a statement of purpose, scope, level of
detail, and approach used; present the estimate of exposure and dose by
pathway and route for individuals, population segments, and populations
in a manner appropriate for the intended risk characterization; and
provide an evaluation of the overall level of confidence (including
consideration of uncertainty factors) in the estimate of exposure and
dose and the conclusions drawn. The SAB recommended that the MOE
discussion be modified to address specific circumstances where the
administered dose and the ``effective dose'' are known to be different.
The discussion has been modified to emphasize that pharmacokinetic
data, when available, be utilized to address such instances.
The SAB recommended that the Agency expand substantially the
discussion of overall strategy to evaluate exposure from mixtures,
exposures to multiple single agents, and exposures to the same agent
via different routes. It is anticipated that this type of information
will be addressed in the Agency's upcoming revisions to the chemical
mixture guidelines.
III. Response to Public Comments
In addition to numerous supportive statements, several issues were
indicated although each issue was raised by a very limited number of
submissions. Use of the benchmark dose was supported along with the
suggestion that the amount of text could be reduced on that subject.
The text has been reduced and reference made to the report, The Use of
the Benchmark Dose Approach in Health Risk Assessment (U.S. EPA,
1995b). A request was made for increased emphasis on paternally
mediated effects on offspring. The text in that section has been
expanded to provide additional discussion and references. Concern was
expressed about the existence of constraints on the use of professional
judgment in the risk assessment process, particularly in determining
the relevance and sufficiency of the database, in evaluating biological
plausibility of statistically different effects, and in the
determination of uncertainty factors. Requests also have been made to
provide additional criteria for when and under what conditions the risk
assessment process will be used. These Guidelines emphasize the
importance of using scientific judgment throughout the risk assessment
process. They provide flexibility to permit EPA's offices and regions
to develop specific guidance suited to their particular needs. The
comment was made that the exposure assessment and risk characterization
sections were not developed as well as the rest of the document. In
1992, EPA published Guidelines for Exposure Assessment (U.S. EPA, 1992)
that were intended to apply generically to noncancer risk assessments.
These Guidelines only address aspects of exposure that are specific to
reproduction and have been developed sufficiently. The risk
characterization section has been expanded substantially to reflect the
recent guidance provided within EPA for application in all risk
assessments.
[FR Doc. 96-27473 Filed 10-30-96; 8:45 am]
BILLING CODE 6560-50-P
