EPA/600/FR-91/001

December 1991

Guidelines for 

Developmental Toxicity Risk Assessment

Published on December 5, 1991, Federal Register 56(234):63798-63826

Risk Assessment Forum

U.S. Environmental Protection Agency
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Washington, DC

DISCLAIMER

This document has been reviewed in accordance with U.S. Environmental Protection Agency

policy and approved for publication. Mention of trade names or commercial products does not

constitute endorsement or recommendation for use.

Note: This document represents the final guidelines. A number of editorial corrections have been made

during conversion and subsequent proofreading to ensure the accuracy of this publication.
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CONTENTS

Lists of Tables and Figures ........................................................ v

Federal Register Preamble ....................................................... vi

Part A: Guidelines for Developmental Toxicity Risk Assessment

1. Introduction ................................................................ 1

2. Definitions and Terminology .................................................... 3

3. Hazard Identification/Dose-Response Evaluation of Agents That Cause Developmental Toxicit 4 y 

3.1. Developmental Toxicity Studies: Endpoints and Their Interpretation .................. 5

3.1.1. Laboratory Animal Studies .......................................... 5

3.1.1.1 Endpoints of Maternal Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7 

3.1.1.2. Endpoints of Developmental Toxicity: Altered Survival, 

Growth, and Morphological Development ........................ 9

3.1.1.3. Endpoints of Developmental Toxicity: Functional Deficits ........... 13

3.1.1.4. Overall Evaluation of Maternal and Developmental Toxicity .......... 17

3.1.1.5. Short-Term Testing in Developmental Toxicity ................... 18

3.1.1.6. Statistical Considerations ................................... 20

3.1.2. Human Studies .................................................. 21

3.1.2.1. Epidemiologic Studies ..................................... 22

3.1.2.2. Examination of Clusters or Case Reports/Series .................. 31

3.1.3. Other Considerations ............................................. 31

3.1.3.1. Pharmacokinetics ......................................... 31

3.1.3.2. Comparisons of Molecular Structure ........................... 32

3.2. Dose-Response Evaluation ............................................... 32

3.3. Characterization of the Health-Related Database ............................... 37

3.4. Determination of the Reference Dose (RfDDT) or Reference Concentration (RfC DT) 

for Developmental Toxicity ............................................... 40

3.5. Summary ............................................................ 42
CONTENTS (continued)

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4. Exposure Assessment ........................................................ 43

5. Risk Characterization ........................................................ 45

5.1. Overview ............................................................ 45

5.2. Integration of the Hazard Identification/Dose-Response Evaluation and Exposure 

Assessment ........................................................... 45

5.3. Descriptors of Developmental Toxicity Risk ................................... 47

5.3.1. Estimation of the Number of Individuals Exposed to Levels of Concern ......... 47

5.3.2. Presenting Specific Scenarios ........................................ 47

5.3.3. Risk Characterization for Highly Exposed Individuals ...................... 47

5.3.4. Risk Characterization for Highly Sensitive or Susceptible Individuals ........... 48

5.3.5. Other Risk Descriptors ............................................. 48

5.4. Communicating Results .................................................. 49

6. Summary and Research Needs ................................................. 50

7. References ................................................................ 51

Part B: Response to Public and Science Advisory Board Comments

1. Introduction ............................................................... 63

2. Intent of the Guidelines ....................................................... 64

3. Basic Assumptions .......................................................... 64

4. Maternal/Developmental Toxicity ................................................ 65

5. Functional Developmental Toxicity ............................................... 65
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6. Weight-of-Evidence Scheme ................................................... 66

7. Applicability of the RfDDT Concept and the Benchmark Dose Approach .................. 67

LIST OF TABLES

Table 1. Endpoints of maternal toxicity .............................................. 8

Table 2. Endpoints of developmental toxicity ......................................... 10

Table 3. Categorization of the health-related database for hazard identification/dose-response

evaluation ............................................................ 38

LIST OF FIGURES

Figure 1. Graphical illustration of the benchmark dose approach ........................... 35
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GUIDELINES FOR DEVELOPMENTAL TOXICITY RISK ASSESSMENT

[FRL-4038-3]

AGENCY: U.S. Environmental Protection Agency (EPA).

ACTION: Final Guidelines for Developmental Toxicity Risk Assessment.

SUMMARY: The U.S. Environmental Protection Agency (EPA) is today issuing final amended

guidelines for assessing the risks for developmental toxicity from exposure to environmental agents. As

background information for this guidance, this notice describes the scientific basis for concern about

exposure to agents that cause developmental toxicity, outlines the general process for assessing

potential risk to humans because of environmental contaminants, summarizes the history of these

guidelines, and addresses public and Science Advisory Board comments on the 1989 “Proposed

Amendments to the Guidelines for the Health Assessment of Suspect Developmental Toxicants” [54

FR 9386-9403]. These guidelines, which have been renamed “Guidelines for Developmental Toxicity

Risk Assessment” (hereafter “Guidelines”), outline principles and methods for evaluating data from

animal and human studies, exposure data, and other information to characterize risk to human

development, growth, survival, and function because of exposure prior to conception, prenatally, or to

infants and children. These Guidelines amend and replace EPA’s 1986 “Guidelines for the Health

Assessment of Suspect Developmental Toxicants” [51 FR 34028-34040] by adding new guidance on

the relationship between maternal and developmental toxicity, characterization of the health-related

database for developmental toxicity risk assessment, use of the reference dose or reference

concentration for developmental toxicity (RfDDT or RfCDT), and use of the benchmark dose approach. 

In addition, the Guidelines were reorganized to combine hazard identification and dose-response

evaluation since these are usually done together in assessing risk for human health effects other than

cancer. 

EFFECTIVE DATE: The Guidelines will be effective December 5, 1991.

FOR FURTHER INFORMATION, CONTACT: Dr. Carole A. Kimmel, Effects Identification

and Characterization Group, National Center for Environmental Assessment-Washington Division

(8623D), U.S. Environmental Protection Agency, 401 M Street, SW, Washington, DC 20460, TEL:

202-564-3307, FAX: 202-565-0078.
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SUPPLEMENTARY INFORMATION: The Clean Air Act (CAA), the Toxic Substances Control

Act (TSCA), the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), and other statutes

administered by the EPA authorize the Agency to protect public health against adverse effects from

environmental pollutants. One type of adverse effect of great concern is developmental toxicity, i.e.,

adverse effects produced prior to conception, or during pregnancy and childhood. Exposure to agents

affecting development can result in any one or more of the following manifestations of developmental

toxicity: death, structural abnormality, growth alteration, and/or functional deficit. These manifestations

encompass a wide array of adverse developmental endpoints, such as spontaneous abortions, stillbirths,

malformations, early postnatal mortality, reduced birth weight, mental retardation, sensory loss, and

other adverse functional or physical changes that are manifested postnatally.

The Role of Environmental Agents in Developmental Toxicity

Several environmental agents are established as causing developmental toxicity in humans (e.g.,

lead, polychlorinated biphenyls, methylmercury, ionizing radiation), while many others are suspected of

causing developmental toxicity in humans on the basis of data from experimental animal studies (e.g.,

some pesticides, other heavy metals, glycol ethers, alcohols, and phthalates). Data for several of the

agents identified as causing human developmental toxicity have been compared to the experimental

animal data (Nisbet and Karch, 1983; Kimmel et al., 1984; Hemminki and Vineis, 1985; Kimmel et al.,

1990a). In these comparisons, the agents causing human developmental toxicity in almost all cases

were found to produce effects in experimental animal studies and, in at least one species tested, types

of effects similar to those in humans were generally seen. This information provides a strong basis for

the use of animal data in conducting human health risk assessments. On the other hand, a number of

agents found to cause developmental toxicity in experimental animal studies have not shown clear

evidence of hazard in humans, but the available human data are often too limited to evaluate a causeand-effect relationship. The comparison of dose-response relationships is hampered by differences in

route, timing, and duration of exposure. When careful comparisons have been done taking these

factors into account, the minimally effective dose for the most sensitive animal species was generally

higher than that for humans, usually within 10-fold of the human effective dose, but sometimes was 100

times or more higher (e.g., polychlorinated biphenyls [Tilson et al., 1990]). Thus, the experimental

animal data were generally predictive of adverse developmental effects in humans, but in some cases,

the administered dose or exposure level required to achieve these adverse effects was much higher than

the effective dose in humans.
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In most cases, the toxic effects of an agent on human development have not been fully studied,

even though exposure of humans to that agent may have been established. At the same time, there are

many developmental effects in humans with unknown causes and no clear link with exposure to

environmental agents. The background incidence of human spontaneous abortion, for example, was

estimated by Hertig (1967) to be approximately 50% of all conceptions, and more recently, Wilcox et

al. (1985), using sensitive techniques for detecting pregnancy as early as 9 days postconception,

observed that 35% of postimplantation pregnancies ended in an embryonic or fetal loss. Of those

infants born alive, approximately 7.4% are reduced in weight at birth (i.e., below 2,500 g) (Selevan,

1981), approximately 3% are found to have one or more congenital malformations at birth, and by the

end of the first postnatal year, about 3% more are found to have serious developmental defects

(Shepard, 1986). Of those children born with developmental defects, it has been estimated that 20%

are due to genetic transmission and 10% can be attributed to known exogenous factors (including

drugs, infections, ionizing radiation, and environmental agents), leaving the remaining 70% with

unknown causes (Wilson, 1977). In a recent hospital-based surveillance study (Nelson and Holmes,

1989), 50.7% of congenital malformations were estimated to be due to genetic or multifactorial causes,

while 3.2% were associated with exposure to exogenous agents and 2.9% to twinning or uterine

factors, leaving 43.2% to unknown causes. The proportion of the effects with unknown causes that

may be attributable to environmental agents or to a combination of factors, such as environmental

agents and genetic factors, nutritional deficiencies, alcohol consumption, direct or indirect exposure to

tobacco smoke, use of prescribed and illicit drugs, etc., is unknown.

The social and economic impact of developmental disabilities on the population is extremely

high. Close to one-half of the children in hospital wards are there because of prenatally acquired

malformations (Shepard, 1980). According to the Centers for Disease Control, congenital anomalies,

sudden infant death syndrome, and prematurity combined account for more than 50% of infant mortality

among all races in the United States (National Center for Health Statistics, 1988). In addition, among

the leading causes of estimated years of potential life lost (YPLL) due to death before the age of 65,

congenital anomalies, prematurity, and sudden infant death syndrome combined rank third (Centers for

Disease Control, 1988a,b). The YPLL estimates for developmental defects may actually underestimate

the public health impact because the estimates do not include prenatal deaths, they are based only on

those cases that die before age 65 and do not account for limited quality of life, and pregnancies may

be terminated early due to prenatal diagnosis of developmental defects.

These data provide the basis for a long-standing interest by Federal agencies that deal with

human health to protect against exposures to agents that cause developmental toxicity, and most of
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these regulatory agencies have provisions for considering data on developmental toxicity in protecting

human health. As a step in developing procedures for interpreting toxicity data in the regulatory

context, the National Academy of Sciences/National Research Council, in 1983, published a

framework for the risk assessment process, which EPA uses as the basis for its risk assessment

guidelines and for the assessment of risk due to environmental agents.

The Risk Assessment Process and Its Application to Developmental Toxicity

Risk assessment is the process by which scientific judgments are made concerning the potential

for toxicity to occur in humans. The National Research Council (1983) has defined risk assessment as

including some or all of the following components: hazard identification, dose-response assessment,

exposure assessment, and risk characterization. In general, the process of assessing the risk of human

developmental toxicity may be adapted to this format. In practice, however, hazard identification for

developmental toxicity and other noncancer health effects is usually done in conjunction with an

evaluation of dose-response relationships, since the determination of a hazard is often dependent on

whether a dose-response relationship is present (Kimmel et al., 1990b). One advantage of this

approach is that it reflects hazard within the context of dose, route, and duration and timing of

exposure, all of which are important in comparing the toxicity information available to potential human

exposure scenarios. Second, this approach avoids labeling of chemicals as developmental toxicants on

a purely qualitative basis. For these reasons, the Guidelines combine hazard identification and doseresponse evaluation under one section (Section 3), and characterize both hazard and dose information

as part of the health-related database for risk assessment. If data are considered sufficient for risk

assessment, an oral or dermal reference dose for developmental toxicity (RfDDT) or an inhalation

reference concentration for developmental toxicity (RfCDT) is then derived for comparison with human

exposure estimates. A statement of the potential for human risk and the consequences of exposure can

come only from integrating the hazard identification/dose-response evaluation with the human exposure

estimates in the final risk characterization. Combining hazard identification and dose-response

evaluation, as well as development of the RfDDT and RfCDT, are revisions of the 1986 Guidelines.

Hazard identification/dose-response evaluation involves examining all available experimental

animal and human data and the associated doses, routes, and timing and duration of exposures to

determine if an agent causes developmental toxicity and/or maternal or paternal toxicity in that species

and under what exposure conditions. The no-observed-adverse-effect-level (NOAEL) and/or the

lowest-observed-adverse-effect-level (LOAEL) are determined for each study and type of effect. 

Based upon the hazard identification/dose-response evaluation and criteria provided in these
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Guidelines, the health-related database can be characterized as sufficient or insufficient for use in risk

assessment (Section 3.3). Because of the limitations associated with the use of the NOAEL, the

Agency is evaluating the use of an additional approach, i.e., the benchmark dose approach (Crump,

1984), for more quantitative dose-response evaluation when sufficient data are available. The

benchmark dose provides an indication of the risk associated with exposures near the NOAEL, taking

into account the variability in the data and the slope of the dose-response curve.

For the determination of the RfDDT or the RfCDT, uncertainty factors are applied to the

NOAEL (or LOAEL, if a NOAEL has not been established) to account for extrapolation from

experimental animals to humans and for variability within the human population. The RfDDT or RfCDT is

generally based on a short duration of exposure as is typically used in developmental toxicity studies in

experimental animals. The use of the terms RfDDT and RfCDT distinguish them from the oral or dermal

reference dose (RfD) and the inhalation reference concentration (RfC), which refer primarily to chronic

exposure situations (U.S. EPA, 1991). Uncertainty factors may also be applied to a benchmark dose

for calculating the RfDDT or RfCDT, but the Agency has little experience with applying this approach

and is currently supporting research efforts to determine the appropriate methods. As more information

becomes available, guidance will be written and published as an addendum to these Guidelines. These

approaches are discussed further in Section 3.4. 

The exposure assessment identifies human populations exposed or potentially exposed to an

agent, describes their composition and size, and presents the types, magnitudes, frequencies, and

durations of exposure to the agent. The exposure assessment provides an estimate of human exposure

levels for particular populations from all potential sources. 

In risk characterization, the hazard identification/dose-response evaluation and the exposure

assessment for given populations are combined to estimate some measure of the risk for developmental

toxicity. As part of risk characterization, a summary of the strengths and weaknesses in each

component of the risk assessment is discussed along with major assumptions, scientific judgments, and,

to the extent possible, qualitative and quantitative estimates of the uncertainties. Confidence in the

health-related data is always presented in conjunction with information on dose-response and the

RfDDT or RfCDT. If human exposure estimates are available, the exposure basis used for the risk

assessment is clearly described, e.g., highly exposed individuals, or highly sensitive or susceptible

individuals. The NOAEL may be compared to the various estimates of human exposure to calculate

the margin(s) of exposure (MOE). The considerations for determining adequacy of the MOE are

similar to those used in determining the appropriate size of the uncertainty factor for calculating the

RfDDT or RfCDT.
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Risk assessment is just one component of the regulatory process and defines the potential

adverse health consequences of exposure to a toxic agent. The other component, risk management,

combines risk assessment with statutory directives regarding socioeconomic, technical, political, and

other considerations, to reach decisions about the appropriate regulation of the suspected toxic agents. 

Risk management is not dealt with directly in these Guidelines since the basis for decision making goes

beyond scientific consideration alone, but the use of scientific information in this process is discussed in

some cases. For example, the acceptability of the MOE is a risk management decision, but the

scientific bases for establishing this value are discussed here.

History of These Guidelines

In 1984, the Agency published “Proposed Guidelines for the Health Assessment of Suspect

Developmental Toxicants” [49 FR 46324-46331]. Following extensive scientific and public review,

final guidelines were issued on September 24, 1986 [51 FR 34028-34040]. The 1986 Guidelines set

forth principles and procedures to guide EPA scientists in the conduct of Agency risk assessments, to

help promote high scientific quality and Agencywide consistency, and to inform Agency decision

makers and the public about these scientific procedures. In publishing this guidance, EPA emphasized

that one purpose of its risk assessment guidelines was to “encourage research and analysis that will lead

to new risk assessment methods and data,” which in turn would be used to revise and improve the

guidelines, and better guide Agency risk assessors. Thus, the 1986 Guidelines were developed and

published with the understanding that risk assessment is an evolving science and that continued study

could lead to changes.

As expected, Agency experience with the 1986 Guidelines suggested that additional or

alternate approaches should be considered for certain aspects of the guidance. Proposals to amend the

guidelines were considered soon after their publication in September 1986, because of new reviews or

re-evaluations that focused on some of the issues identified for research in the guidelines. Included

were several workshops and symposia cited in the Introduction to these Guidelines. In addition, much

experience had been gained in using the 1986 Guidelines and in instructing others in their use.

Based on this experience, amendments to the 1986 Guidelines were proposed for public

comment in March 1989 [54 FR 9386-9403]. Following receipt and review of the public comments,

they were collated, summarized, and reviewed by scientists within the Agency. On October 27, 1989,

EPA’s Science Advisory Board (SAB) met to review the Proposed Amendments and the summarized

public comments, and to be briefed by Agency scientists concerning proposed responses. 
xii

During this same period, several issues with implications for health effects other than cancer

were under discussion in the Agency and elsewhere. These issues included use of the benchmark dose

(see Section 3.2), exposure descriptors (see Section 5.3), and risk characterization (see Section 5). 

Thus, generic discussions on risk assessment issues, along with comments from the public and the SAB,

have influenced the structure and content of these Guidelines. 

These revised Guidelines were then reviewed by a number of Agency scientists and official

panels, including the Risk Assessment Forum and the Risk Assessment Council. The revised

Guidelines also were presented to the SAB on March 27, 1991, for final comment. In addition, a

review was conducted by the interagency Working Party on Reproductive Toxicology, Subcommittee

on Risk Assessment of the Federal Coordinating Committee on Science, Engineering and Technology. 

Comments of these groups have been considered in the revision of these Guidelines. The full text of the

final “Guidelines for Developmental Toxicity Risk Assessment” is published here.

These Guidelines were developed as part of an interoffice guidelines development program

under the auspices of the Risk Assessment Forum and the Office of Health and Environmental

Assessment (OHEA) in the Agency’s Office of Research and Development. The Agency is continuing

to study risk assessment issues raised in these Guidelines, and will revise them in line with new

information as appropriate.

Following this Preamble are two parts: Part A is the Guidelines and Part B is the Response to

the Public and Science Advisory Board Comments. Part B includes a summary of the issues raised by

the public and the SAB, and the Agency’s responses to those comments.

References, supporting documents, and comments received on the Proposed Amendments, as

well as a copy of the final Guidelines, are available for inspection and copying at the Public Information

Reference Unit Docket (202-260-5926), EPA Headquarters Library, 401 M Street, S.W.,

Washington, DC, between the hours of 8:00 a.m. and 4:30 p.m.

______________________ _______________________________________

Dated: November 26, 1991 Signed by EPA Administrator

William K. Reilly
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PART A: GUIDELINES FOR DEVELOPMENTAL TOXICITY RISK ASSESSMENT

1. INTRODUCTION

These Guidelines describe the procedures that EPA follows in evaluating potential

developmental toxicity associated with human exposure to environmental agents. The Agency has

sponsored or participated in several conferences that addressed issues related to such evaluations and

that provide some of the scientific basis for these Guidelines (U.S. EPA, 1982a; Kimmel et al., 1982b,

1987; Hardin, 1987; Perlin and McCormack, 1988; Kimmel et al., 1989; Kimmel and Francis, 1990;

Kimmel et al., 1990a). The Agency’s authority to regulate substances that have the potential to

interfere with human development is derived from a number of statutes that are implemented through

multiple offices within the EPA. The procedures described herein are intended to promote consistency

in the assessment of developmental toxic effects across program offices within the Agency.

These Guidelines provide a general format for analyzing and organizing the available data for

conducting risk assessments. The Agency previously has issued testing guidelines (U.S. EPA, 1982b,

1985a, 1989a, 1991a) that provide protocols designed to determine the potential of a test substance to

induce structural and/or other adverse effects during development. These risk assessment Guidelines

do not change any prescribed statutory or regulatory standards for the type of data necessary for

regulatory action, but rather provide guidance for the interpretation of studies that follow the testing

guidelines and, in addition, provide limited information for interpretation of other studies (e.g.,

epidemiologic data, functional developmental toxicity studies, and short-term tests) that are not routinely

required, but may be encountered when reviewing data on particular agents.

Since the purpose of risk assessment is to make inferences about potential risks to human

health, the most appropriate data to be used are those deriving from studies of humans. If adequate

human data are not available, then it is necessary to use data obtained from other species. There are a

number of unknowns in the extrapolation of data from animal studies to humans. Therefore, a number

of assumptions must be made on the relevance of effects to potential human risk that are generally

applied in the absence of data. These assumptions provide the inferential basis for the approaches

taken to risk assessment in these Guidelines.

First, it is assumed that an agent that produces an adverse developmental effect in experimental

animal studies will potentially pose a hazard to humans following sufficient exposure during

development. This assumption is based on the comparisons of data for agents known to cause human

developmental toxicity (Nisbet and Karch, 1983; Kimmel et al., 1984; Hemminki and Vineis, 1985;
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Kimmel et al., 1990a), which indicate that, in almost all cases, experimental animal data are predictive

of a developmental effect in humans.

It is assumed that all of the four manifestations of developmental toxicity (death, structural

abnormalities, growth alterations, and functional deficits) are of concern. In the past, there has been a

tendency to consider only malformations or malformations and death as endpoints of concern. From

the data on agents that are known to cause human developmental toxicity (Nisbet and Karch, 1983;

Kimmel et al., 1984; Hemminki and Vineis, 1985; Kimmel et al., 1990a), there is usually at least one

experimental species that mimics the types of effects seen in humans, but in other species tested, the

type of developmental perturbation may be different. Thus, a biologically significant increase in any of

the four manifestations is considered indicative of an agent’s potential for disrupting development and

producing a developmental hazard.

It is assumed that the types of developmental effects seen in animal studies are not necessarily

the same as those that may be produced in humans. This assumption is made because it is impossible

to determine which will be the most appropriate species in terms of predicting the specific types of

effects seen in humans. The fact that every species may not react in the same way could be due to

species-specific differences in critical periods, differences in timing of exposure, metabolism,

developmental patterns, placentation, or mechanisms of action.

The most appropriate species is used to estimate human risk when data are available (e.g.,

pharmacokinetics). In the absence of such data, it is assumed that the most sensitive species is

appropriate for use, based on observations that humans are as sensitive or more so than the most

sensitive animal species tested for the majority of agents known to cause human developmental toxicity

(Nisbet and Karch, 1983; Kimmel et al., 1984; Hemminki and Vineis, 1985; Kimmel et al., 1990a).

In general, a threshold is assumed for the dose-response curve for agents that produce

developmental toxicity. This is based on the known capacity of the developing organism to compensate

for or to repair a certain amount of damage at the cellular, tissue, or organ level. In addition, because

of the multipotency of cells at certain stages of development, multiple insults at the molecular or cellular

level may be required to produce an effect on the whole organism. 
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2. DEFINITIONS AND TERMINOLOGY

The Agency recognizes that there are differences in the use of terms in the field of

developmental toxicology. For the purposes of these Guidelines the following definitions will be used. 

Developmental toxicology - The study 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.

Altered growth - An alteration in offspring organ or body weight or size. Changes in one endpoint

may or may not be accompanied by other signs of altered growth (e.g., changes in body weight may or

may not be accompanied by changes in crown-rump length and/or skeletal ossification). Altered

growth can be induced at any stage of development, may be reversible, or may result in a permanent

change.

Functional developmental toxicology - The study of alterations or delays in the physiological and/or

biochemical competence of an organism or organ system following exposure to an agent during critical

periods of development pre- and/or postnatally.

Structural abnormalities - Structural alterations in development that include both malformations and

variations.

Malformations and variations - A malformation is usually defined as a permanent structural change

that may adversely affect survival, development, or function. The term teratogenicity is used in these

Guidelines to refer only to malformations. The term variation is used to indicate a divergence beyond

the usual range of structural constitution that may not adversely affect survival or health. Distinguishing

between variations and malformations is difficult since there exists a continuum of responses from the

normal to the extremely deviant. There is no generally accepted classification of malformations and

variations. Other terms that are often used, but no better defined, include anomalies, deformations, and

aberrations.
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3. HAZARD IDENTIFICATION/DOSE-RESPONSE EVALUATION OF AGENTS THAT

CAUSE DEVELOPMENTAL TOXICITY

This section discusses the evaluation and interpretation of hazards for a variety of endpoints of

developmental toxicity seen in both human and animal studies, and describes the criteria for

characterizing the sufficiency of the health-related database for conducting a developmental toxicity risk

assessment. It also details the use of dose-response data for determining potential hazards, and

describes the calculation of the RfDDT or RfCDT, a dose or concentration that is assumed to be without

appreciable risk of deleterious developmental effects for a given agent.

Developmental toxicity is expressed as one or more of a number of possible endpoints that may

be used for evaluating the potential of an agent to cause abnormal development. Developmental

toxicity generally occurs in a dose-related manner, may result from short-term exposure (including

single exposure situations) or from longer term low-level exposure, may be produced by various routes

of exposure, and the types of effects may vary depending on the timing of exposure because of a

number of critical periods of development for various organs and functional systems.

The four major manifestations of developmental toxicity are death, structural abnormality,

altered growth, and functional deficit. The relationship among these manifestations may vary with

increasing dose and, especially at higher doses, death of the conceptus may preclude expression of

other manifestations. Of these, all four manifestations have been evaluated in human studies, but only

the first three are traditionally measured in laboratory animals using the conventional developmental

toxicity (also called teratogenicity or Segment II) testing protocol as well as in other study protocols,

such as the multigeneration study or the continuous breeding study. Although functional deficits seldom

have been evaluated in routine testing studies in experimental animals, functional evaluations are

beginning to be required in certain regulatory situations (U.S. EPA, 1986a, 1988a, 1989b, 1991a).

Developmental toxicity can be considered a component of reproductive toxicity, and often it is

difficult to distinguish between effects mediated through the parents versus direct interaction with

developmental processes. For example, developmental toxicity may be influenced by the effects of

toxic agents on the maternal system when exposure occurs during pregnancy or lactation. In addition,

following parental exposure prior to conception, developmental toxicity may result in their offspring and,

potentially, in subsequent generations. Therefore, it is useful to consult the “Proposed Guidelines for

Assessing Male Reproductive Risk” (U.S. EPA, 1988b) and the “Proposed Guidelines for Assessing

Female Reproductive Risk” (U.S. EPA, 1988c) in conjunction with these Guidelines. Mutational

events that occur as a result of exposure to agents that cause developmental toxicity may be difficult to
5

discriminate from other possible mechanisms in standard studies of developmental toxicity. When

mutational events are suspected, the “Guidelines for Mutagenicity Risk Assessment” (U.S. EPA,

1986c), which specifically address the risks of heritable mutation, should be consulted.

Carcinogenic effects have occurred in humans following developmental exposures to

diethylstilbestrol (Herbst et al., 1971). Several additional agents (e.g., direct-acting alkylating agents)

have been shown to cause cancer following developmental exposures in experimental animals, and it

appears from the data collected thus far that agents capable of causing cancer in adults may also cause

transplacental or neonatal carcinogenesis (Anderson et al., 1985). Currently, there is no way to predict

whether the developing offspring or adult will be more sensitive to the carcinogenic effects of an agent. 

At present, testing for carcinogenesis following developmental exposure is not routinely required. 

However, if this type of effect is reported for an agent, it is considered appropriate to use the

“Guidelines for Carcinogen Risk Assessment” (U.S. EPA, 1986b) for assessing human risk.

3.1. DEVELOPMENTAL TOXICITY STUDIES: ENDPOINTS AND THEIR

INTERPRETATION

3.1.1. Laboratory Animal Studies

This section discusses the endpoints examined in routinely used protocols as well as the use of

other types of studies, including functional studies and short-term tests. 

The most commonly used protocol for assessing developmental toxicity in laboratory animals

involves the administration of a test substance to pregnant animals (usually mice, rats, or rabbits) during

the period of major organogenesis, evaluation of maternal responses throughout pregnancy, and

examination of the dam and the uterine contents just prior to term (U.S. EPA, 1982b, 1985a; Food and

Drug Administration [FDA], 1966, 1970; Organization for Economic Cooperation and Development

[OECD], 1981). Some studies may use exposures of one to a few days to investigate periods of

particular sensitivity for induction of abnormalities in specific organs or organ systems. In addition,

developmental toxicity may be evaluated in studies involving exposure to one or both parents prior to

conception, to the conceptus during pregnancy and over several generations, or to offspring during the

prenatal and preweaning periods (U.S. EPA, 1982b, 1985a, 1986a, 1988a, 1991a; FDA, 1966,

1970; OECD, 1981; Lamb, 1985). These Guidelines are intended to provide information for

interpreting developmental effects related to any of these types of exposure. 

Appropriate study designs include a number of important factors. For example, test animal

selection is generally based on considerations of species, strain, age, weight, and health status. 

Assignment of animals to dose groups by stratified randomization (on the basis of body weight) reduces
6

bias and provides a basis for performing valid statistical tests. At a minimum, a high dose, a low dose,

and one intermediate dose are included. The high dose is selected to produce some minimal maternal

or adult toxicity (i.e., a level that at the least produces marginal but significantly reduced body weight,

reduced weight gain, or specific organ toxicity, and at the most produces no more than 10% mortality). 

At doses that cause excessive maternal toxicity (that is, significantly greater than the minimal toxic level),

information on developmental effects may be difficult to interpret and of limited value. The low dose is

generally a NOAEL for adult and offspring effects, although if the low dose produces a biologically or

statistically significant increase in response, it is considered a LOAEL (see Section 3.1.1.6 for a

discussion of biological versus statistical significance). A concurrent control group treated with the

vehicle used for agent administration is a critical component of a well-designed study.

The route of exposure in these studies is usually oral, unless the chemical or physical

characteristics of the test substance or pattern of human exposure suggest a more appropriate route of

administration. In the case of dermal exposure, developmental toxicity studies showing no indication of

maternal or developmental toxicity are considered insufficient for risk assessment unless accompanied

by absorption data (Kimmel and Francis, 1990). Dermal developmental toxicity studies in which skin

irritation is too marked (moderate erythema and/or moderate edema, i.e., raised approximately 1 mm)

also are considered insufficient, since excessive maternal toxicity may be produced from the irritation

rather than from systemic exposure to the agent. Assessment of maternal toxicity is based on signs of

systemic toxicity rather than on local effects such as skin irritation. Absorption data and limited

pharmacokinetic data collected in dermal developmental toxicity studies provide very useful information

in the evaluation of study design and data interpretation (Kimmel and Francis, 1990). Many of these

points also are pertinent to studies by other routes of exposure.

The evaluation of specific endpoints of maternal and developmental toxicity is discussed in the

next several sections. Appropriate historical control data sometimes can be very useful in the

interpretation of these endpoints. Comparison of data from treated animals with concurrent study

controls should always take precedence over comparison with historical control data. The most

appropriate historical control data are those from the same laboratory in which studies were conducted. 

Even data from the same laboratory, however, should be used cautiously and examined for subtle

changes over time that may result from genetic alterations in the strain or stock of the species used,

changes in environmental conditions both in the breeding colony of the supplier and in the laboratory,

and changes in personnel conducting studies and collecting data (Kimmel and Price, 1990). Study data

should be compared with recent as well as cumulative historical data. Any change in laboratory
7

procedure that might affect control data should be noted and the data accumulated separately from

previous data.

The next three sections (3.1.1.1, 3.1.1.2, and 3.1.1.3) discuss individual endpoints of maternal

and developmental toxicity as measured in the conventional developmental toxicity study, the

multigeneration study, and, when available, in postnatal studies. Other endpoints specifically related to

reproductive toxicity are covered in the relevant risk assessment guidelines (U.S. EPA, 1988b, 1988c). 

The fourth section (3.1.1.4) deals with the integrated evaluation of all data, including the relative effects

of exposure on maternal animals and their offspring, which is important in assessing the level of concern

about a particular agent.

3.1.1.1. Endpoints of Maternal Toxicity

A number of endpoints that may be observed as possible indicators of maternal toxicity are

listed in Table 1. Maternal mortality is an obvious endpoint of toxicity; however, a number of other

endpoints can be observed that may give an indication of the more subtle adverse effects of an agent. 

For example, in well-conducted studies, the mating and fertility indices provide information on the

general fertility rate of the animal stock used and are important indicators of toxic effects to adults if

treatment begins prior to mating or implantation. Changes in gestation length may indicate effects on the

process of parturition.

Body weight and change in body weight are viewed collectively as indicators of maternal

toxicity for most species, although these endpoints may not be as useful in rabbits, because body weight

changes are usually more variable (Kimmel and Price, 1990) and, in some strains of rabbits, body

weight is not a good indicator of pregnancy status. Body weight changes may provide more

information than a daily body weight measured during treatment or during gestation. Changes in weight

gain during treatment could occur that would not be reflected in the total weight change throughout

gestation, because of compensatory weight gain that may occur following treatment but before sacrifice. 

For this reason, changes in weight gain during treatment can be examined as another indicator of

maternal toxicity.

Changes in maternal body weight corrected for gravid uterine weight at sacrifice may indicate

whether the effect is primarily maternal or intrauterine. For example, a significant reduction in weight

gain throughout gestation and in gravid uterine weight without any change in corrected maternal weight

gain generally would indicate an intrauterine effect. Conversely, a change in corrected weight gain and

no change in gravid uterine weight generally would suggest maternal toxicity and little or no intrauterine

effect. An alternate estimate of maternal weight change during gestation can be obtained by subtracting
8

the sum of the weights of the fetuses. However, this weight does not include the uterine or placental

tissue, or the amniotic fluid.
9

Table 1. Endpoints of maternal toxicity

Mortality

Mating index [(no. with seminal plugs or sperm/no. mated) × 100]

Fertility index [(no. with implants/no. of matings) × 100] 

Gestation length (useful when animals are allowed to deliver pups)

Body weight

 Day 0

During gestation

Day of necropsy

Body weight change 

Throughout gestation

During treatment (including increments of time within treatment period)

Post-treatment to sacrifice

Corrected maternal (body weight change throughout gestation minus gravid

uterine weight or litter weight at sacrifice)

Organ weights (in cases of suspected target organ toxicity and especially when

supported by adverse histopathology findings)

Absolute

Relative to body weight

Relative to brain weight

Food and water consumption (where relevant)

Clinical evaluations 

Types, incidence, degree, and duration of clinical signs

Enzyme markers

Clinical chemistries 

Gross necropsy and histopathology
10

Changes in other endpoints may also be important. For example, changes in relative and

absolute organ weights may be signs of a maternal effect, especially when an agent is suspected of

causing specific organ toxicity and when such findings are supported by adverse histopathologic findings

in those organs. Food and water consumption data are useful, especially if the agent is administered in

the diet or drinking water. The amount ingested (total and relative to body weight) and the dose of the

agent (relative to body weight) can then be calculated, and changes in food and water consumption

related to treatment can be evaluated along with changes in body weight and body weight gain. Data

on food and water consumption also are useful when an agent is suspected of affecting appetite, water

intake, or excretory function.

Clinical evaluations of toxicity also may be used as indicators of maternal toxicity. Daily clinical

observations may be useful in describing the profile of maternal toxicity and alterations in general

homeostasis. Enzyme markers and clinical chemistries may be useful indicators of exposure but must

be interpreted carefully as to whether or not a change constitutes toxicity. Gross necropsy and

histopathology data (when specified in the protocol) may aid in determining toxic dose levels. The

minimum amount of information considered useful for evaluating maternal toxicity [as noted in the

“Proceedings of the Workshop on the Evaluation of Maternal and Developmental Toxicity” (Kimmel et

al., 1987)], includes morbidity or mortality, maternal body weight and body weight gain, clinical signs of

toxicity, food and water consumption (especially if dosing is via food or water), and necropsy for gross

evidence of organ toxicity. In a well-designed study, maternal toxicity is determined in the pregnant

and/or lactating animal over an appropriate part of gestation and/or the neonatal period, and is not

assumed or extrapolated from other adult toxicity studies.

3.1.1.2. Endpoints of Developmental Toxicity: Altered Survival, Growth, and Morphological

 Development

Because the maternal animal, and not the conceptus, is the individual treated during gestation,

data generally are calculated as incidence per litter or as number and percent of litters with particular

endpoints. Table 2 indicates the ways in which offspring and litter endpoints may be expressed.

When treatment of females begins prior to implantation, an increase in preimplantation loss

could indicate an adverse effect on gamete transport, the fertilization process, uterine toxicity, the

developing blastocyst, or on the process of implantation itself. If treatment begins around the time of

implantation (i.e., day 6 of gestation in the mouse, rat, or rabbit), an increase in preimplantation loss

probably reflects variability that is not treatment-related in the animals being used, but the data should

be examined carefully to determine if there is a dose-response
11

Table 2. Endpoints of developmental toxicity

Litters with implants

No. implantation sites/dam

No. corpora lutea (CL)/dama

Percent preimplantation loss

(CL - implantations) × 100a

CL

No. and percent live offspringb/litter

No. and percent resorptions/litter

No. and percent litters with resorptions

No. and percent late fetal deaths/litter

No. and percent nonlive (late fetal deaths + resorptions) implants/litter

No. and percent litters with nonlive implants

No. and percent affected (nonlive + malformed) implants/litter

No. and percent litters with affected implants

No. and percent litters with total resorptions 

No. and percent stillbirths/litter

No. and percent litters with live offspring

Litters with live offspring

No. and percent live offspring/litter

Viability of offspringc

Sex ratio/litter

Mean offspring body weight/litterc

Mean male or female body weight/litterc

No. and percent offspring with external, visceral, or skeletal malformations/litter

No. and percent malformed offspring/litter

No. and percent litters with malformed offspring

No. and percent malformed males or females/litter

No. and percent offspring with external, visceral, or skeletal variations/litter

No. and percent offspring with variations/litter

No. and percent litters having offspring with variations

Types and incidence of individual malformations

Types and incidence of individual variations

Individual offspring and their malformations and variations

(grouped according to litter and dose)

Clinical signs (type, incidence, duration, and degree)

Gross necropsy and histopathology

a

Important when treatment begins prior to implantation. May be difficult to assess in mice.

bOffspring refers both to fetuses observed prior to term and to pups following birth. The endpoints examined

depend on the protocol used for each study.

cMeasured at selected intervals until termination of the study.
12

relationship. If preimplantation loss is related to dose, further studies would be necessary to determine

the mechanism and extent of such effects.

The number and percent of live offspring per litter, based on all litters, may include litters that

have no live implants. The number and percent of resorptions and late fetal deaths give some indication

of when the conceptus died, and the number and percent of nonlive implants per litter (postimplantation

loss) is a combination of these two measures. Expression of data as the number and percent of litters

showing an increased incidence for these endpoints may be less useful than incidence per litter because,

in the former case, a litter is counted whether one or all implants were resorbed, dead, or nonlive.

If a significant increase in postimplantation loss is found after exposure to an agent, the data

may be compared not only with concurrent controls, but also with recent historical control data

(preferably from the same laboratory), since there is considerable interlitter variability in the incidence of

postimplantation loss (Kimmel and Price, 1990). If a given study control group exhibits an unusually

high or low incidence of postimplantation loss compared to historical controls, then scientific judgment

must be used to determine the adequacy of the study for risk assessment purposes.

The endpoint “affected implants” (i.e., the combination of nonlive and malformed conceptuses)

sometimes reflects a better dose-response relationship than does the incidence of nonlive or malformed

offspring taken individually. This is especially true at the high end of the dose-response curve in cases

when the incidence of nonlive implants per litter is greatly increased. In such cases, the malformation

rate may appear to decrease because only unaffected offspring have survived. If the incidence of

prenatal deaths or malformations is unchanged, then the incidence of affected implants will not provide

any additional dose-response information. In studies where maternal animals are allowed to deliver

pups normally, the number of stillbirths per litter should also be noted.

The number of live offspring per litter, based on those litters that have one or more live

offspring, may be unchanged even though the incidence of nonlive in all litters is increased. This could

occur either because of an increase in the number of litters with no live offspring, or an increase in the

number of implants per litter. A decrease in the number of live offspring per litter is generally

accompanied by an increase in the incidence of nonlive implants per litter unless the implant numbers

differ among dose groups. In postnatal studies, the viability of live-born offspring should be determined

at selected intervals until termination of the study.

The sex ratio per litter, as well as the body weights of males and females, can be examined to

determine whether or not one sex is preferentially affected by the agent. However, this is an unusual

occurrence.
13

A change in offspring body weight is a sensitive indicator of developmental toxicity, in part

because it is a continuous variable. In some cases, offspring weight reduction may be the only indicator

of developmental toxicity. While there is always a question as to whether weight reduction is a

permanent or transitory effect, little is known about the long-term consequences of short-term fetal or

neonatal weight changes. Therefore, when significant weight reduction effects are noted, they are used

as a basis to establish the NOAEL. Several other factors should be considered in the evaluation of

fetal or neonatal weight changes; for example, in polytocous animals, fetal and neonatal weights are

usually inversely correlated with litter size, and the upper end of the dose-response curve may be

affected by smaller litters and increased fetal or neonatal weight. Additionally, the average body weight

of males is greater than that of females in the more commonly used laboratory animals.

Live offspring are generally examined for external, visceral, and skeletal malformations and

variations. If only a portion of the litter is examined for one or more endpoints, then random selection

of those pups examined introduces less bias in the data. An increase in the incidence of malformed

offspring may be indicated by a change in one or more of the following endpoints: the incidence of

malformed offspring per litter, the number and percent of litters with malformed offspring, or the number

of offspring or litters with a particular malformation that appears to increase with dose (as indicated by

the incidence of individual types of malformations).

Other ways of examining the data include determining the incidence of external, visceral, and

skeletal malformations and variations that may indicate the organs or organ systems affected. A listing

of individual offspring with their malformations and variations may give an indication of the pattern of

developmental deviations. All of these methods of expressing and examining the data are valid for

determining the effects of an agent on structural development. However, care must be taken to avoid

counting offspring more than once in the evaluation of any single endpoint based on number or percent

of offspring or litters. The incidence of individual types of malformations and variations may indicate

significant changes that are masked if the data on all malformations and/or variations are pooled. 

Appropriate historical control data can be especially helpful in the interpretation of malformations and

variations, particularly those that normally occur at a low incidence and may or may not be related to

dose in an individual study. 

Although a dose-related increase in malformations is interpreted as an adverse developmental

effect of exposure to an agent, the biological significance of an altered incidence of anatomical variations

is more difficult to assess, and must take into account what is known about developmental stage (e.g.,

with skeletal ossification), background incidence of certain variations (e.g., 12 or 13 pairs of ribs in

rabbits), or other strain- or species-specific factors. However, if variations are significantly increased in
14

a dose-related manner, these should also be evaluated as a possible indication of developmental

toxicity.

In addition, although some investigators have considered certain of these effects to simply be

associated with manifestations of maternal toxicity noted at similar dose levels (Khera, 1984, 1985,

1987), such effects are still toxic manifestations and as such are generally considered a reasonable basis

for Agency regulation and/or risk assessment. On a somewhat similar note, the conclusion of

participants in a “Workshop on Reproductive Toxicity Risk Assessment” (Kimmel et al., 1986) was

that dose-related increases in defects that may occur spontaneously are as relevant as dose-related

increases in any other developmental toxicity endpoints.

3.1.1.3. Endpoints of Developmental Toxicity: Functional Deficits

Developmental effects that are induced by exogenous agents are not limited to death, structural

abnormalities, and altered growth. Rather, it has been demonstrated in a number of instances that

alterations in the functional competence of an organ or a variety of organ systems may result from

exposure during critical developmental periods that may occur between conception and sexual

maturation. Sometimes, these functional defects are observed at dose levels below those at which

other indicators of developmental toxicity are evident (Rodier, 1978). Such effects may be transient or

reversible in nature, but generally are considered adverse. Testing for functional developmental toxicity

has not been required routinely by regulatory agencies in the United States, but studies in developmental

neurotoxicity are beginning to be required by the EPA when other information indicates the potential for

adverse functional developmental effects (U.S. EPA, 1986a, 1988a, 1989b, 1991a). Data from

postnatal studies, when available, are considered very useful for further assessment of the relative

importance and severity of findings in the fetus and neonate. Often, the long-term consequences of

adverse developmental outcomes noted at birth are unknown, and further data on postnatal

development and function are necessary to determine the full spectrum of potential developmental

effects. Useful data can also be derived from well-conducted multigeneration studies, although the dose

levels used in these studies may be much lower than in studies with shorter-term exposure.

Much of the early work in functional developmental toxicology was related to behavioral

evaluations, and the term “behavioral teratology” became prominent in the mid-1970s. Recent

advances in this area have been reviewed in several publications (Riley and Vorhees, 1986; Kimmel,

1988; Kimmel et al., 1990a). Several expert groups have focused on the functions that should be

included in a behavioral testing battery (World Health Organization [WHO], 1984; Buelke-Sam et al.,

1985; Leukroth, 1986). These include: sensory systems, neuromotor development, locomotor activity,
15

learning and memory, reactivity and/or habituation, and reproductive behavior. No testing battery has

fully addressed all of these functions, but it is important to include as many as possible, and several

testing batteries have been developed and evaluated for use in testing (Buelke-Sam et al., 1985;

Tanimura, 1986; Elsner et al., 1986). 

The Agency recently has developed a “generic” developmental neurotoxicity test guideline that

can be used for both pesticides and industrial chemicals (U.S. EPA, 1991a). Because of its design, the

developmental neurotoxicity testing protocol may be conducted as a separate study, concurrently with

or as a follow-up to a developmental toxicity (Segment II) study, or be folded into a multigeneration

study in the second generation. Testing is generally conducted in the rat. In the protocol for the

separate study, the test agent is administered orally (other routes may be used on a case-by-case basis)

to at least three treated groups and one concurrent control group of animals on day 6 of gestation

through day 10 postnatally. The highest dose level is selected to induce some overt signs of maternal

toxicity, but not result in more than a 20% reduction in weight gain during gestation and lactation. This

dose also is selected to avoid in utero or neonatal death or malformations sufficient to preclude a

meaningful evaluation of developmental neurotoxicity. At least 20 litters are required per treatment

group. For behavioral tests, one female and one male pup per litter are randomly selected and assigned

to one of the following tests: motor activity, auditory startle, and learning and memory in animals at

weaning and as adults. Neuropathological evaluation and determination of brain weights are conducted

on selected pups at postnatal day 11 and at termination of the study.

Several criteria for selecting agents for developmental neurotoxicity testing have been suggested

(Buelke-Sam et al., 1985; Levine and Butcher, 1990), including: agents that cause central nervous

system malformations, psychoactive drugs and chemicals, agents that cause adult neurotoxicity,

hormonally active agents, and chemicals that are structurally related to others that cause developmental

neurotoxicity or for which widespread exposure and/or release is expected. Data from developmental

neurotoxicity studies should be evaluated in light of the data that may have triggered such testing as well

as all other toxicity data available. 

Less work has been done on other developing functional systems, but the assessment of

postnatal renal morphological and functional development may serve as a model for the use of postnatal

evaluations in the risk assessment process. As an example, standard morphological analyses of the

kidneys of fetal rodents have detected treatment-related changes in the relative growth of the renal

papilla versus the renal cortex, an effect considered in some cases to be a malformation

(hydronephrosis), while in other cases a variation (apparent hydronephrosis, enlarged or dilated renal

pelvis). While some investigators (Woo and Hoar, 1972) have provided data suggesting that the
16

morphological effect represents a transient developmental delay, others have shown that it can persist

well into postnatal life and that physiological function is compromised in the affected individuals

(Kavlock et al., 1987a, 1988; Daston et al., 1988; Couture, 1990). Thus, the biological interpretation

of this effect on the basis of fetal examinations alone is tenuous (U.S. EPA, 1985b). In addition, the

critical period for inducing renal morphological abnormalities extends into the postnatal period

(Couture, 1990), and studies on perinatally induced renal growth retardation (Kavlock et al., 1986,

1987b; Slotkin et al., 1988; Gray et al., 1989; Gray and Kavlock, 1991) have shown that renal

function is generally altered in such conditions, but that manifestation of the dysfunction is not readily

predictable. Thus, both morphological and functional assessment of the kidneys after birth can provide

useful and complementary information on the persistence and biological significance of expressions of

developmental toxicity.

Although not as well studied, data indicate that the cardiovascular, respiratory, immune,

endocrine, reproductive, and digestive systems also are subject to alterations in functional competence

(Kavlock and Grabowski, 1983; Fujii and Adams, 1987) following exposure during development. 

Currently, there are no standard testing procedures for these functional systems; however, when data

are encountered on a chemical under review, they are considered in the risk assessment process.

Direct extrapolation of functional developmental effects to humans is limited in the same way as

for other endpoints of developmental toxicity, i.e., by the lack of knowledge about underlying

toxicological mechanisms and their significance. In evaluations of a limited number of agents known to

cause developmental neurotoxic effects in humans, Adams (1986) concluded that these agents produce

similar developmental neurotoxic effects in animals and humans. This conclusion was strongly

supported by the results of a recent “Workshop on the Qualitative and Quantitative Comparability of

Human and Animal Developmental Neurotoxicity,” sponsored by EPA and the National Institute on

Drug Abuse (NIDA), at which participants critically evaluated and compared the effects of agents

known to cause human developmental neurotoxicity with the effects seen in experimental animal studies

(Kimmel et al., 1990a). The high degree of qualitative correlation between human and experimental

animal data for the agents evaluated lends strong support for the use of experimental animals in

assessing the potential risk for developmental neurotoxicity in humans. Thus, as for other endpoints of

developmental toxicity, the assumption can be made that functional effects in animal studies indicate the

potential for altered development in humans, although the types of developmental effects seen in

experimental animal studies will not necessarily be the same as those that may be produced in humans. 

Thus, when data from functional developmental toxicity studies are encountered for particular agents,

they should be considered in the risk assessment process.
17

Some guidance is provided here concerning important general concepts of study design and

evaluation for functional developmental toxicity studies.

C Several aspects of study design are similar to those important in standard developmental

toxicity studies (e.g., a dose-response approach with the highest dose producing minimal

overt maternal or perinatal toxicity, number of litters large enough for adequate statistical

power, randomization of animals to dose groups and test groups, litter generally considered

the statistical unit, etc.).

C A replicate study design provides added confidence in the interpretation of data.

C A pharmacological/physiological challenge may be valuable in evaluating function and

“unmasking” effects not otherwise detectable, particularly in the case of organ systems that

are endowed with a reasonable degree of functional reserve capacity.

C Functional tests with a moderate degree of background variability may be more sensitive to

the effects of an agent on behavioral endpoints than are tests with low variability that may

be impossible to disrupt without being life-threatening (Butcher et al., 1980).

C A battery of functional tests, in contrast to a single test, is usually needed to evaluate the full

complement of organ function in an animal; tests conducted at several ages may provide

more information about maturational changes and their persistence.

C Critical periods for the disruption of functional competence include both the prenatal and

the postnatal periods to the time of sexual maturation, and the effect is likely to vary

depending on the time and degree of exposure.

C Interpretation of data from studies in which postnatal exposure is included should take into

account possible interaction of the agent with maternal behavior, milk composition, pup

suckling behavior, possible direct exposure of pups via dosed feed or water, etc.

Although interpretation of functional data may be limited at present, it is clear that functional

effects must be evaluated in light of other toxicity data, including other forms of developmental toxicity

(e.g., structural abnormalities, perinatal death, and growth retardation). The level of confidence in an

adverse effect may be as important as the type of change seen, and confidence may be increased by

such factors as replicability of the effect either in another study of the same function or by convergence

of data from tests that purport to measure similar functions. A dose-response relationship is considered

an important measure of chemical effect; in the case of functional effects, both monotonic and biphasic

dose-response curves are likely, depending on the function being tested. 

Finally, there are at least three general ways in which the data from these studies may be useful

for risk assessment purposes: (1) to help elucidate the long-term consequences of fetal and neonatal
18

effects; (2) to indicate the potential for an agent to cause functional alterations and the effective doses

relative to those that produce other forms of toxicity; and (3) for existing environmental agents, to

suggest organ systems to be evaluated in exposed human populations.

3.1.1.4. Overall Evaluation of Maternal and Developmental Toxicity

As discussed previously, individual endpoints of maternal and developmental toxicity are

evaluated in developmental toxicity studies. In order to interpret the data fully, an integrated evaluation

must be performed considering all maternal and developmental endpoints.

Agents that produce developmental toxicity at a dose that is not toxic to the maternal animal are

especially of concern because the developing organism is affected but toxicity is not apparent in the

adult. However, the more common situation is when adverse developmental effects are produced only

at doses that cause minimal maternal toxicity; in these cases, the developmental effects are still

considered to represent developmental toxicity and should not be discounted as being secondary to

maternal toxicity. At doses causing excessive maternal toxicity (that is, significantly greater than the

minimal toxic dose), information on developmental effects may be difficult to interpret and of limited

value. Current information is inadequate to assume that developmental effects at maternally toxic doses

result only from maternal toxicity; rather, when the LOAEL is the same for the adult and developing

organisms, it may simply indicate that both are sensitive to that dose level. Moreover, whether

developmental effects are secondary to maternal toxicity or not, the maternal effects may be reversible

while effects on the offspring may be permanent. These are important considerations for agents to

which humans may be exposed at minimally toxic levels either voluntarily or involuntarily, since several

agents are known to produce adverse developmental effects at minimally toxic doses in adult humans

(e.g., smoking, alcohol, isotretinoin).

Since the final risk assessment not only takes into account the potential hazard of an agent, but

also the nature of the dose-response relationship, it is important that the relationship of maternal and

developmental toxicity be evaluated and described. Then, information from the exposure assessment is

used to determine the likelihood of exposure to levels near the maternally toxic dose for each agent and

the risk for developmental toxicity in humans.

Although the evaluation of developmental toxicity is the primary objective of standard studies

within this area, maternal effects seen within the context of developmental toxicity studies should be

evaluated as part of the overall toxicity profile for a given chemical. Maternal toxicity may be seen in

the absence of or at dose levels lower than those producing developmental toxicity. If the maternal

effect level is lower than that in other evaluations of adult toxicity, this implies that the pregnant female is
19

likely to be more sensitive than the nonpregnant female. Data from reproductive and developmental

toxicity studies on the pregnant female should be used in the overall assessment of risk.

Approaches for ranking agents according to their relative maternal and developmental toxicity

have been proposed; Schardein (1983) has reviewed several of these. Several approaches involve the

calculation of ratios relating an adult toxic dose to a developmentally toxic dose (Johnson, 1981; Fabro

et al., 1982; Johnson and Gabel, 1983; Brown and Freeman, 1984). Such ratios may describe in a

qualitative and roughly quantitative fashion the relationship of maternal (adult) and developmental

toxicity. However, at the U.S. EPA-sponsored “Workshop on the Evaluation of Maternal and

Developmental Toxicity” (Kimmel et al., 1987), there was no agreement as to the validity or utility of

these approaches in other aspects of the risk assessment process. This is due in part to uncertainty

about factors that can affect the ratios. For example, the number and spacing of dose levels,

differences in study design (e.g., route and/or timing of exposure), the relative thoroughness in the

assessment of maternal and developmental endpoints examined, species differences in response, and

differences in the slope of the dose-response curves for maternal and developmental toxicity can all

influence the maternal and developmental effects observed and the resulting ratios (Kimmel et al., 1987;

U.S. EPA, 1985b). Also, maternal and developmental endpoints used in the ratios need to be better

defined to permit cross-species comparison. Until such information is available, the applicability of

these approaches in risk assessment is not justified.

3.1.1.5. Short-Term Testing in Developmental Toxicity

The need for short-term tests for developmental toxicity has arisen from the need to establish

testing priorities for the large number of agents in or entering the environment, the interest in reducing

the number of animals used for routine testing, and the expense of testing. These approaches may be

useful in making preliminary evaluations of potential developmental toxicity, for evaluating structureactivity relationships, and for assigning priorities for further, more extensive testing. Furthermore, as the

risk assessment process begins to incorporate more pharmacokinetic and mechanistic data, short-term

tests should be particularly useful. Kimmel (1990) has recently discussed the potential application of in

vitro systems in risk assessment in a context that is broader than chemical screening. However, the

Agency currently considers a short-term test as “insufficient” by itself to carry out a risk assessment

(see Section 3.3).

Although short-term tests for developmental toxicity are not routinely required, such data are

encountered in the review of chemicals. Two approaches are considered here in terms of their

contribution to the overall testing process: an in vivo mammalian screen and in vitro test systems.
20

3.1.1.5.1. In vivo mammalian developmental toxicity tests. The most widely studied in vivo

short-term approach is that developed by Chernoff and Kavlock (1982). This approach is based on

the hypothesis that a prenatal injury, which results in altered development, will be manifested postnatally

as reduced viability and/or impaired growth. When originally proposed, the test substance was

administered to mice over the period of major organogenesis at a single dose level that would elicit

some degree of maternal toxicity. At the NIOSH “Workshop on the Evaluation of the

Chernoff/Kavlock Test for Developmental Toxicity” (Hardin, 1987), use of a second lower dose level

was encouraged to potentially reduce the chances of false positive results, and the recording of

implantation sites was recommended to provide a more precise estimate of postimplantation loss

(Kavlock et al., 1987c). In this approach, the pups are counted and weighed shortly after birth, and

again after 3-4 days. Endpoints that are considered in the evaluation include: general maternal toxicity

(including survival and weight gain), litter size, pup viability and weight, and gross malformations in the

offspring. Several schemes have been proposed for ranking the results as a means of prioritizing agents

for further testing (Chernoff and Kavlock, 1982; Brown, 1984; Schuler et al., 1984).

The mouse was chosen originally for this test because of its low cost, but the procedure has

been applied to the rat as well (Wickramaratne, 1987). The test can predict the potential for

developmental toxicity of an agent in the species used while extrapolation of risk to other species,

including humans, has the same limitations as for other testing protocols. The EPA Office of Toxic

Substances has developed testing guidelines for this procedure (U.S. EPA, 1985c), and the Office of

Pesticide Programs has applied similar protocols on a case-by-case basis (U.S. EPA, 1985b). The

National Toxicology Program also has developed a protocol that incorporates aspects of a rangefinding study, with the intent of providing information on appropriate exposure levels should a standard

developmental toxicity study be required (Morrissey et al., 1989). Although testing guidelines are

available, such procedures are required on a case-by-case basis. Application of this procedure in the

risk assessment process within the Office of Toxic Substances has been described (Francis and

Farland, 1987), and the experiences of a number of laboratories are detailed in the proceedings of a

NIOSH-sponsored workshop (Hardin, 1987).

Recently, the OECD developed a screening protocol to be used for prioritizing existing

chemicals for further testing (draft as of March 22, 1990). This protocol is similar to the design of the

Chernoff-Kavlock test except that it involves exposure of male and female rats 2 weeks prior to

mating, throughout mating and gestation, and postnatally to day 4. Male animals are exposed following

mating for a period corresponding to that of the females. Adult animals are evaluated for general

toxicity and effects on reproductive organs. Pups are counted, weighed, and examined for any gross
21

physical or behavioral abnormalities at birth and on postnatal day 4. This protocol permits evaluation of

reproductive and developmental toxicity following repeated dosing with an agent, provides an indication

for the need to conduct additional studies, and provides guidance in the design of further studies. 

Currently, this study design is insufficient by itself to make an estimate of human risk without further

studies to confirm and extend the observations. 

3.1.1.5.2. In vitro developmental toxicity screens. Test systems that fall under the general heading

of “in vitro” developmental toxicity screens include any system that employs a test subject other than the

intact pregnant mammal. Examples of such systems include isolated whole mammalian embryos in

culture, tissue/organ culture, cell culture, and developing nonmammalian organisms. These systems

have long been used to assess events associated with normal and abnormal development, but more

recently they have been considered for their potential as screens in testing (Wilson, 1978; Kimmel et

al., 1982b; Brown and Fabro, 1982). Many of these systems are now being evaluated for their ability

to predict the developmental toxicity of various agents in intact mammalian systems. This validation

process requires certain considerations in study design, including defined endpoints for toxicity and an

understanding of the system’s ability to handle various test agents (Kimmel et al., 1982a; Kimmel,

1985; FDA, 1987; Brown, 1987).

While in vitro test systems can provide significant information, they are considered insufficient,

by themselves, for carrying out a risk assessment (see Section 3.3). In part, this is due to limitations in

the application of the data to the whole-animal situation. But it is also due to the lack of assays that

have been fully validated, as has been noted in several reviews of available in vitro systems (FDA,

1987; Brown, 1987; Faustman, 1988) and at a recent workshop on in vitro teratology (Morrissey et

al., 1991).

3.1.1.6. Statistical Considerations

In the assessment of developmental toxicity data, statistical considerations require special

attention. Since the litter is generally considered the experimental unit in most developmental toxicity

studies, and fetuses or pups within litters do not respond independently, the statistical analyses are

generally designed to analyze the relevant databased on incidence per litter or on the number of litters

with a particular endpoint. The analytical procedures used and the results, as well as an indication of

the variance in each endpoint, should be evaluated carefully when reviewing data for risk assessment

purposes. Analysis of variance (ANOVA) techniques, with litter nested within dose in the model, take

the litter variable into account while allowing use of individual offspring data and an evaluation of both
22

within and between litter variance as well as dose effects. Nonparametric and categorical procedures

have also been widely used for binomial or incidence data. In addition, tests for dose-response trends

can be applied. Although a single statistical approach has not been agreed upon, a number of factors

important in the analysis of developmental toxicity data have been discussed (Haseman and Kupper,

1979; Kimmel et al., 1986).

Studies that employ a replicate experimental design (e.g., two or three replicates with 10 litters

per dose per replicate rather than a single experiment with 20 to 30 litters per dose group) allow

broader interpretation of study results since the variability between replicates can be accounted for

using ANOVA techniques. Replication of effects due to a given agent within a study, as well as among

studies or laboratories, provides added strength in the use of data for the estimation of risk.

An important factor to consider in evaluating data is the power of a study (i.e., the probability

that a study will demonstrate a true effect), which is limited by the sample size used in the study, the

background incidence of the endpoint observed, the variability in the incidence of the endpoint, and the

analysis method. As an example, Nelson and Holson (1978) have shown that the number of litters

needed to detect a 5% or 10% change was dramatically lower for fetal weight (a continuous variable

with low variability) than for resorptions (a binomial response with high variability). With the current

recommendation in testing protocols being 20 rodents per dose group (U.S. EPA, 1982b, 1985a), the

minimum change detectable is an increased incidence of malformations 5 to 12 times above control

levels, an increase 3 to 6 times the in utero death rate, and a decrease 0.15 to 0.25 times the fetal

weight. Thus, even within the same study, the ability to detect a change in fetal weight is much greater

than for the other endpoints measured. Consequently, for statistical reasons only, changes in fetal

weight are often observed at doses below those producing other signs of developmental toxicity. Any

risk assessment should present the detection sensitivity for the study design used and for the endpoint(s)

evaluated.

Although statistical analyses are important in determining the effects of a particular agent, the

biological significance of data is most relevant. It is important to be aware that with the number of

endpoints that can be observed in standard protocols for developmental toxicity studies, a few

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. This may be true especially for the incidence of malformations or in utero death because of the

low power of standard study designs in which a relatively large difference is required to be statistically

significant. It should be apparent from this discussion that a great deal of scientific judgment, based on
23

experience with developmental toxicity data and with principles of experimental design and statistical

analysis, may be required to adequately evaluate such data.

3.1.2. Human Studies

In principle, human data are preferred for risk assessment. However, the complexities of

obtaining sufficient human data are such that these data are not available for many potential toxicants. 

The following describes the methods of generation of human data, their evaluation, and the weight they

should be given in risk assessments.

The category of “human studies” includes both epidemiologic studies and other reports of

individual cases or clusters of events. Greatest weight should be given to carefully designed

epidemiologic studies with more precise measures of exposure, since they can best evaluate exposureresponse relationships (see Section 4). 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 to

qualitative risk assessments, but are of limited use for quantitative risk assessments because of the

generally broad categorical groupings. 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 give added support to

associations suggested by other human or animal data, but cannot stand by themselves in risk

assessments. Risk assessors should seek the assistance of professionals trained in epidemiology when

conducting a detailed analysis. 

3.1.2.1. Epidemiologic Studies

Good epidemiologic studies provide the most relevant information for assessing human risk. As

there are many different designs for epidemiologic studies, simple rules for their evaluation do not exist. 

3.1.2.1.1. 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; U.S. EPA, 1981; Wilcox, 1983; Sever and Hessol, 1984; Axelson, 1985; Tilley et al., 1985;

Kimmel et al., 1986). Some of the more prominent factors are as follows:

(a) 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 in order to have a high probability of detecting a modest increase
24

in risk (e.g., 133 in both exposed and unexposed groups to detect a doubling of background; alpha =

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 in both

exposed and unexposed groups) (Bloom, 1981; Selevan, 1981; Sever and Hessol, 1984; Selevan,

1985; Stein et al., 1985; Kimmel et al., 1986). In case-referent studies, study sizes are dependent on

the frequency of exposure within the source population. The confidence one has in the results of a

study without positive findings is related to the power of the study to detect meaningful differences in the

endpoints studied.

Power may be enhanced by combining populations from several studies using a meta-analysis

(Greenland, 1987). The combined analysis would increase confidence in the absence of risk for agents

with negative findings. However, care must be exercised in the combination of potentially dissimilar

study groups.

A posteriori determination of power of the actual study may be useful in evaluating

contradictory studies in risk assessment. Absence of positive findings in a study of low power would be

given less weight than either a positive study or a null study (one with no significant differences) with

high power. Positive findings from very small studies are open to question due to the instability of the

risk estimates and the potential for highly selected study groups.

(b) Potential bias in data collection: Sources of bias may include selection bias and information

bias (Rothman, 1986). Selection bias may occur when an individual’s willingness to participate varies

with certain characteristics relating to the exposure status or health status of that individual. In addition,

selection bias may operate in the identification of subjects for study. For example, in studies of

embryonic loss, use of hospital records to identify embryonic or early fetal loss will underascertain

events, because women are not always hospitalized for these outcomes. More weight might be given in

a risk assessment to a study in which a more complete list of pregnancies is obtained by, for example,

collecting biological data [e.g., human chorionic gonadotropin (hCG) measurements] on pregnancy

status from study members. These studies may also be affected by bias. 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; however, this strategy carries with it the potential for recall bias, discussed in further

detail below. A second example of different levels of ascertainment of events is the use of hospital

records to study congenital malformations. Hospital records contain more complete data on

malformations than do birth certificates (Mackeprang et al., 1972). Consequently, birth defects

registries that are based on searches of hospital records are more complete than those based on vital
25

records (Selevan, 1986). Thus, a study using hospital records to identify congenital malformations

would be given more emphasis in a risk assessment than one using birth certificates.

Studies of working women present the potential for additional bias since some factors that

influence employment status may also be associated with 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). 

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 will reduce the likelihood of such bias. 

Studies with lower likelihood of the above-listed biases 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 considerably more

complete and valid than those of the husbands (Selevan, 1980). A more recent study (Schnatter,

1990) found small, nonsignificant improvements in reporting of birth weights by mothers compared to

fathers, and that males who provide early fetal loss data with the aid of their wives give better data

(borderline significance). Studies based on interview data from the appropriate respondent(s) would

carry more weight than those from proxy respondents (e.g., the specific individual when examining

exposure history and the woman or both partners when examining pregnancy history).

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 indicate 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 and Hatch, 1987).

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).

(c) Collection of data on other risk factors, effect modifiers, and confounders: Risk factors for

reproductive and developmental toxicity include such characteristics as age, smoking, alcohol
26

consumption, drug use, and past reproductive history. Additionally, occupational and environmental

exposures are potential risk factors for reproductive and developmental effects. Known and potential

risk factors should be examined to identify those that may be effect modifiers or confounders. An effect

modifier is a factor that produces different exposure-response relationships at different levels of that

factor. For example, maternal age would be an effect modifier if the risk associated with a given

exposure increased with the mother’s age. A confounder is a variable that is a risk factor for the

disease 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, socioeconomic status might be a

confounder in a study of the association of smoking and fertility, since socioeconomic status may be

associated with both.

Studies that fail to account for effect modifiers and confounders should be given less weight in a

risk assessment. Both of these important factors 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, 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).

(d) Statistical factors: As in animal studies, pregnancies experienced by the same woman are

not independent events (Kissling, 1981; Selevan, 1985). Women who have had embryo/fetal loss are

reported to be more likely to have subsequent losses (Leridon, 1977). In animal studies, the litter is

generally used as the unit of measure to deal with nonindependence of events. In studies of humans,

pregnancies are sequential with the risk factors changing for different pregnancies, making analyses

considering nonindependence of events very difficult (Epidemiology Workgroup, 1981; Kissling,

1981). If more than one pregnancy per woman is included, as is often necessary due to 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). Biased estimates of risk might also result if family size confounds the relationship between

exposure and outcome. Some approaches to deal with these issues have been suggested (Kissling,

1981; Stiratelli et al., 1984; Selevan, 1985). At this point in time, a generally accepted solution to this

problem has not been developed.

3.1.2.1.2. Selection of outcomes for study. As already discussed, a number of endpoints can be

considered in the evaluation of adverse developmental effects. However, some of the outcomes are not
27

easily observed in humans, such as early embryonic loss and reproductive capacity of the offspring. 

Currently, the most feasible endpoints for epidemiologic studies are reproductive history studies of

some pregnancy outcomes (e.g., embryo/fetal loss, birth weight, sex ratio, congenital malformations,

postnatal function, and neonatal growth and survival) and measures of fertility/infertility, which would

include indirect evaluations of very early embryonic loss. Postnatal outcomes for examination could

include physical growth and development, organ or system function, and behavioral effects of exposure. 

Factors requiring control in the design or analysis (such as effect modifiers and confounders) may vary

depending on the specific outcomes selected for study.

The developmental outcomes available for epidemiologic examination are limited by a number

of factors, including the relative magnitude of the exposure, because differing spectra of outcomes may

occur at different exposure levels, different size and demographic characteristics of the population, and

different ability to observe the developmental outcome in humans. Improved methods for identifying

some outcomes such as very early embryonic loss using new hCG assays may change the spectrum of

outcomes available for study (Wilcox et al., 1985; Sweeney et al., 1988).

Demographic characteristics of the population, such as marital status, age distribution,

education, socioeconomic status (SES), and prior reproductive history are associated with the

probability of whether couples will attempt to have children. Differences in the use of birth control

would also affect the number of outcomes available for study. In addition, women with live births are

more likely to terminate employment than are those with other outcomes, such as infertility or early

embryonic loss. Thus, retrospective studies of female exposure that do not include terminated women

workers may be of limited use in risk assessment because the level of risk for these outcomes is likely

to be overestimated (Lemasters and Pinney, 1989). 

In addition to the above-mentioned factors, developmental endpoints may be envisioned as

effects recognized at various points in a continuum, starting at conception through death of the offspring. 

Thus, a malformed stillbirth would not be included in a study of defects observed at live birth, even

though the etiology could be identical (Stein et al., 1975; Bloom, 1981). A shift in the patterns of

outcomes could result from differences in timing or in level of exposure (Selevan and LeMasters,

1987). 

3.1.2.1.3. Reproductive history studies. (a) Measures of fertility: Normally, studies of sub- or

infertility would not be included in an evaluation of developmental effects. However, in humans it is

difficult to identify very early embryonic loss, and distinguish it from sub- or infertility. Thus, studies that

examine sub- or infertility indirectly examine loss very early in the gestational period. Infertility or
28

subfertility may be thought of as a nonevent: a couple is unable to have children within a specific time

frame. Therefore, the epidemiologic measurement of reduced fertility is typically indirect, and is

accomplished by comparing birth rates or time intervals between births or pregnancies. In these evaluations, the couple’s joint ability to procreate is estimated. One method, the Standardized Birth Ratio

(SBR; also referred to as the Standardized Fertility Ratio), compares the number of births observed to

those expected based on the person-years of observation stratified by factors such as time period, age,

race, marital status, parity, contraceptive use, etc. (Wong et al., 1979; Levine et al., 1980, 1981;

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).

Analysis of the time period between recognized pregnancies or live births has been suggested

as another indirect measure of fertility (Dobbins et al., 1978; Baird et al., 1986; Weinberg and Gladen,

1986). Because the time interval 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.

Fertility may also be affected by alterations in sexual behavior. However, limited data are

available linking toxic exposures to these alterations in humans. Moreover, such data are not easily

obtained in epidemiology studies. More information on this subject is available in the proposed male

and female reproductive risk assessment guidelines (U.S. EPA, 1988b, 1988c).

(b) Pregnancy outcomes: Pregnancy outcomes examined in human studies of parental

exposures may include embryo/fetal loss, congenital malformations, birth weight, sex ratio at birth, and

postnatal effects (e.g., physical growth and development, organ or system function, and behavioral

effects of exposure). Postnatal effects are discussed in more detail in the next section. As mentioned

previously, epidemiologic studies that focus on only one type of pregnancy outcome may miss a true

effect of exposure because of the continuum of outcomes. Examination of individual outcomes could

mask a true effect due to reduced power resulting from fewer events for study. Studies that examine

multiple endpoints could yield more information, but the results may be difficult to interpret.

Evidence of a dose-response relationship is usually an important criterion in the assessment of a

toxic exposure. However, traditional dose-response relationships may not always be observed for

some endpoints. For example, with increasing dose, a pregnancy might end in a 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
29

description, see Section 3.1.2.1.5). Therefore, a risk assessment should, when possible, attempt to

look at the interrelationship of different reproductive endpoints and patterns of exposure.

(c) Postnatal developmental effects: These effects may include changes in growth, behavior,

organ or system function, or cancer. Studies of neurological and reproductive function are discussed

here as examples. Postnatal behavioral and functional effects in humans have been examined for a small

number of environmental and occupational agents (e.g., lead, PCBs, methyl mercury, alcohol). For

some agents (e.g., lead and PCBs), subtle changes have been observed in groups of children at lower

exposures than for other developmental effects (e.g., Bellinger et al., 1987; Needleman, 1988; Davis et

al., 1990; Tilson et al., 1990). This may not be true for all toxic agents. These subtle differences would

be difficult to identify in individuals, but could result in an overall shifting of mean values when

comparing groups of exposed and unexposed children. Some postnatal studies have examined infants

or young children using standard developmental scales (e.g., Brazelton Neonatal Behavioral

Assessment Scale, Bayley Scales of Infant Development, Stanford Binet IV, and Wechsler Scales) and

some biologic measure of exposure (e.g., blood lead levels). These tests are designed to examine

certain endpoints and have been developed to cover certain age ranges. Certain tests examine specific

aspects of development. For example, the Bayley Scales look at motor and language development, but

do not examine sensory function. Batteries of tests are important for a proper evaluation because of the

possibility of interrelated effects, e.g., hearing deficits and language development. Thus, batteries of

tests will give a clearer indication of direct effects of exposure resulting in postnatal developmental

deficits.

Factors that may influence the examination of these effects include parental education, SES,

obstetrical history, and health characteristics independent of exposure that may affect functional

measurement (e.g., injuries and infections). Many social and lifestyle factors may also affect scoring on

these scales (e.g., neonatal-maternal interactions, SES, home environment).

Studies of premature infants carry special problems. For proper comparisons, tests keyed to

age in very young children (less than 2.5 years of age) need to “correct” the age for premature infants to

the age they would have been had they been born at term. In addition, premature infants or those with

low birth weight for their gestational age may have problems resulting from the birth process not directly

related to exposure (e.g., intraventricular hemorrhage in the brain which can then cause developmental

problems). Thus, the developmental effects resulting from exposure may have their own sequelae. 

Other studies may examine effects occurring at a later age (e.g., in utero exposure and cancer in

young women). This long time interval typically carries with it the need for retrospective studies, with

the inherent limitations in accurate determination of exposure, effect modifiers, and confounders. Risk
30

assessment methods for cancer are described in the “Guidelines for Carcinogen Risk Assessment”

(U.S. EPA, 1986b).

Reproductive effects may result from developmental exposures. For example, environmental

exposures may result in oocyte toxicity, in which a loss of primordial oocytes irreversibly affects a

woman’s fertility. The exposures of importance may occur during both the prenatal period and after

birth. 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). Risk assessment methods for female reproductive effects are

described in the “Proposed Guidelines for Assessing Female Reproductive Risk” (U.S. EPA, 1988c).

Developmental exposures to males could affect their reproductive function (e.g., deplete stem

or Sertoli cells potentially affecting sperm production) (Zenick and Clegg, 1989). If stem cell death

occurs with exposure at any age, recovery is possible as long as some stem cells survive. The same is

true for Sertoli cells, except that they cease multiplication before puberty. Thus, cell replication cannot

compensate for Sertoli cell death after puberty. Human studies of stem and Sertoli cells would be

difficult due to the invasiveness of the measure. Less direct measures, e.g., sperm count, morphology,

and motility, could be evaluated, but this would not indicate what cells or stage of spermatogenesis had

been affected. Risk assessment methods for male reproductive effects are described in the “Proposed

Guidelines for Assessing Male Reproductive Risk” (U.S. EPA, 1988b).

In addition to the above effects, genetic damage to germ cells may result from developmental

exposures. Outcomes resulting from germ-cell mutations could include reduced probability of

conception as well as increased probability of embryo/fetal loss and other developmental effects. 

These endpoints could be studied using the approaches described above. However, a human germ-cell

mutagen has not yet been demonstrated (U.S. EPA, 1986c). Based on animal studies, critical

exposures are to germ cells or early zygotes. Germ-cell mutagenicity could also be expressed as

genetic diseases in future generations. Unfortunately, these studies would be very difficult to conduct in

human populations because of the long time lag between exposure and outcome. For more

information, refer to the “Guidelines for Mutagenicity Risk Assessment” (U.S. EPA, 1986c).

3.1.2.1.4. Community studies/surveillance programs. Epidemiologic studies may also be based

on broad populations such as a community, a nationwide probability sample, or surveillance programs

(such as birth defects registries). Other studies have examined environmental exposures, such as toxic

agents in the water system, and adverse pregnancy outcome (Swan et al., 1989; Deane et al., 1989). 

Unfortunately, in these studies maternally mediated effects may be difficult to distinguish from paternally
31

mediated effects. In addition, the presumably lower exposure levels (compared to industrial settings)

may require very large groups for 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), birth defects (Hemminki et al., 1980; Kwa and Fine, 1980;

Papier, 1985), and childhood cancer (Kwa and Fine, 1980; Zack et al., 1980; Hemminki et al., 1981;

Peters et al., 1981). In these reports, jobs are typically classified into broad categories based on the

probability of exposure to certain classes or levels of exposure (e.g., Kwa and Fine, 1980). 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, such studies are not typically useful for risk assessment

of a particular agent.

Surveillance programs may also exist in occupational settings. In this case, reproductive

histories and/or clinical evaluations could be followed to monitor for reproductive effects of exposures. 

Both could yield very useful data for risk assessment; however, a clinical evaluation program would be

costly to maintain, and there are numerous impediments to the collection of reliable and valid

information in the workplace. These might include concerns similar to those previously discussed plus

potentially low participation rates due to employee sensitivities and confidentiality concerns.

3.1.2.1.5. Identification of exposures important for developmental effects. For all examinations

of the relationship between developmental effects and potentially toxic exposures, the identification of

the appropriate exposure is crucial. Preconceptional exposures to either parent and in utero exposures

have been associated with the more commonly examined outcomes (e.g., fetal loss, malformations, birth

weight, and measures of infertility). These exposures, plus postnatal exposure from breast milk, food,

and the general environment, may be associated with postnatal developmental effects (e.g., changes in

behavioral and cognitive function, or growth). The magnitude of exposure may affect the spectrum of

outcomes observed. This issue is discussed in more detail in Sections 3.1.1.2 and 3.2.

Infants and young children may receive disproportionate levels of exposure due to their

tendency to “put everything” in their mouths (pica) and the greater time they spend on the floor. 

Carpets may serve as a reservoir for toxic agents (e.g., pesticides and lead dust), and the air nearer the

floor may have greater levels of certain airborne toxicants (e.g., mercury from latex paints).

Exposures in environmental settings are frequently lower than in industrial and 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, pesticides usage). Larger

populations are necessary in settings with lower exposures (Lemasters and Selevan, 1984). Other
32

factors affect the identification of reproductive or developmental events with various levels of exposure. 

Exposed individuals may move in and out of areas with differing levels and types of exposures, affecting

the number of exposed and comparison events for study. Thus, exposures can be short-term or

chronic.

Data on exposure from human studies are frequently qualitative, such as employment or

residence histories. More quantitative data may be difficult to obtain due to the nature of certain study

designs (e.g., retrospective studies) and historical limitations in exposure measurements. Many

developmental outcomes result from exposures during certain critical times. The appropriate exposure

classification depends on the outcome(s) studied, the biologic mechanism affected by exposure, and the

biologic half-life of the agent. The biologic half-life, in combination with the patterns of exposure (e.g.,

continuous or intermittent) affect the individual’s body burden and consequently the “true” 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 et al., 1986). As more prospective studies are done, better estimates of

exposure will be developed.

3.1.2.2. Examination of Clusters or Case Reports/Series

The identification of cases or clusters of adverse pregnancy outcomes is generally limited to

those identified by the women involved, or clinically by their physicians. Examples of outcomes more

easily identified include mid-to-late fetal loss or congenital malformations. Identification of other effects,

such as very early embryonic loss, may be difficult to separate from the study of sub- or infertility. Such

“nonevents” (e.g., lack of pregnancies or children) are much harder to recognize than are

developmental effects such as malformations resulting from in utero exposure. While case reports have

been important in the recognition of some agents that cause developmental toxicity, they may be of

greatest use in suggesting topics for further investigation (Hogue, 1985). 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 animals also occur in humans. Previous discussion of the use of human data should

be taken into account wherever possible.

3.1.3. Other Considerations

Several other types of information may be considered in the evaluation and interpretation of

human and animal data. Information on pharmacokinetics and structure-activity relationships may be

very useful, but is often lacking for developmental toxicity risk assessments.
33

3.1.3.1. Pharmacokinetics

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 and/or peak plasma concentrations, placental

metabolism and transfer, excretion in breast milk, comparative metabolism, and concentrations of the

parent compound and metabolites may be useful in predicting risk for developmental toxicity. Such

data may also be helpful in defining the dose-response curve, developing a more accurate comparison

of species sensitivity (Wilson et al., 1975, 1977), determining dosimetry at target sites, and comparing

pharmacokinetic profiles for various dosing regimens or routes of exposure. Pharmacokinetic studies in

developmental toxicology are most useful if conducted in animals at the stage when developmental

insults occur. The correlation of pharmacokinetic parameters and developmental toxicity data may be

useful in determining the contribution of specific pharmacokinetic parameters to the effects observed

(Kimmel and Young, 1983).

While human pharmacokinetic data are often lacking, absorption data in laboratory animals for

studies conducted by any relevant route of exposure may assist in the interpretation of the

developmental toxicity studies in the animal models for the purposes of risk assessment. Specific

guidance regarding both the development and application of pharmacokinetic data was agreed upon by

the participants at the “Workshop on the Acceptability and Interpretation of Dermal Developmental

Toxicity Studies” (Kimmel and Francis, 1990). It was concluded that absorption data are needed both

when a dermal developmental toxicity study shows no developmental effects and when developmental

effects are seen. The results of a dermal developmental toxicity study showing no adverse

developmental effects and without blood level data (as evidence of dermal absorption) are potentially

misleading and would be insufficient for risk assessment, especially if interpreted as a “negative” study. 

In studies where developmental toxicity is detected, regardless of the route of exposure, absorption

data can be used to establish the internal dose in maternal animals for risk extrapolation purposes.

3.1.3.2. Comparisons of Molecular Structure

Comparisons of the chemical or physical properties of an agent with those known to cause

developmental toxicity may indicate a potential for developmental 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 have not been well studied in developmental

toxicology, although data are available that suggest structure-activity relationships for certain classes of
34

chemicals (e.g., glycol ethers, steroids, retinoids). Under certain circumstances (e.g., in the case of new

chemicals), this is one of several procedures used to evaluate the potential for toxicity when little or no

data are available.

3.2. DOSE-RESPONSE EVALUATION

The evaluation of dose-response relationships for developmental toxicity includes the evaluation

of data from both human and animal studies. When quantitative dose-response data are available in

humans and with sufficient range of exposure, dose-response relationships may be examined. Since

data on human dose-response relationships have been available infrequently, the dose-response

evaluation is usually based on the assessment of data from tests performed in laboratory animals.

Evidence for a dose-response relationship is an important criterion in the assessment of

developmental toxicity, which is usually based on limited data from standard studies using three dose

groups and a control group. Most agents causing developmental toxicity in humans alter development

at doses within a narrow range near the lowest maternally toxic dose (Kimmel et al., 1984). Therefore,

for most agents, the exposure situations of concern will be those that are potentially near the maternally

toxic dose range. For those few agents that produce developmental effects at much lower levels than

maternal effects, the potential for exposing the conceptus to damaging doses is much greater than when

the maternal and developmental toxic doses are similar. As mentioned previously (Section 3.1.1.2),

however, traditional dose-response relationships may not always be observed for some endpoints. For

example, as exposure increases, embryolethal levels may be reached, resulting in an observed decrease

in malformations with increasing dose (Wilson, 1973; Selevan and LeMasters, 1987). The potential for

this response pattern indicates that dose-response relationships of individual endpoints as well as

combinations of endpoints (e.g., dead and malformed combined) must be carefully examined and

interpreted.

The evaluation of dose-response relationships includes the identification of effective dose levels

as well as doses that are associated with no increased incidence of adverse effects when compared

with controls. Much of the focus is on the identification of the critical effect(s) (i.e., the adverse

effect(s) observed at the lowest dose level) and the LOAEL and NOAEL associated with that

developmental effect, which may be any of the four manifestations of developmental toxicity. The

NOAEL is defined as the highest dose at which there is no statistically or biologically significant

increase in the frequency of an adverse effect in any of the possible manifestations of developmental

toxicity when compared with the appropriate control group in a data base characterized as having

sufficient evidence for use in a risk assessment (see Section 3.3). The LOAEL is the lowest dose at
35

which there is a statistically or biologically significant increase in the frequency of adverse developmental

effects when compared with the appropriate control group in a database characterized as having

sufficient evidence. Although a threshold is assumed for developmental effects, the existence of a

NOAEL in an animal study does not prove or disprove the existence or level of a biological threshold;

it only defines the highest level of exposure under the conditions of the study that is not associated with

a significant increase in adverse effects.

Several limitations in the use of the NOAEL have been described (Gaylor, 1983; Crump,

1984; Kimmel and Gaylor, 1988; Gaylor, 1989; Brown and Erdreich, 1989, Kimmel, 1990): (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) Since 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 is 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 influence the dose chosen for the NOAEL. (5) Since the NOAEL is defined as a dose

that does not produce an observed increase 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). Crump (1984) and Gaylor (1989) have estimated the upper confidence limit on risk at

the NOAEL to be 2% to 6% for specific developmental endpoints from several data sets. 

Because of the limitations associated with the use of the NOAEL (Kimmel and Gaylor, 1988;

Gaylor, 1989; Kimmel, 1990), the Agency is evaluating the use of an additional approach for more

quantitative dose-response evaluation when sufficient data are available, i.e., the benchmark dose

(Crump, 1984). The benchmark dose is based on a model-derived estimate of a particular incidence

level, such as 10% incidence. More specifically, the benchmark dose (BD) is derived by modeling the

data in the observed range, selecting an incidence level within or near the observed range (e.g., the

effective dose to produce a 10% increased incidence of response, the ED10), and determining the upper

confidence limit on the model. The upper confidence value corresponding to, for example, a 10%

excess in response is used to derive the BD, which is the lower confidence limit on dose for that level of

excess response, in this case the LED10 (see Figure 1).

Various mathematical approaches have been proposed for deriving the benchmark dose for

developmental toxicity data (e.g., Crump, 1984; Rai and Van Ryzin, 1985; Kimmel and Gaylor, 1988;

Faustman et al., 1989; Chen and Kodell, 1989; Kodell et al., 1991). Such models may be used to

calculate the benchmark dose, and the particular model used may be less critical since estimation of the
36

benchmark dose is limited to the observed dose range. Since the model is only used to fit the observed

data, the assumptions about the existence or nonexistence of a threshold are not as pertinent. Thus,

models that fit the empirical data may well provide a reasonable estimate of the benchmark dose,

although biological factors known to influence data should be incorporated into the model [e.g.,

intralitter correlations, correlations among endpoints (Ryan et al., 1991)]. The Agency is currently

conducting studies to evaluate the application of several models to actual data sets for calculating the

benchmark dose, to determine the minimum data required for modeling, and to develop methods for

application to continuous data. In addition, information from these studies will be used to develop

guidance for application of the benchmark dose approach to the calculation of the RfDDT or the RfCDT,

since the Agency has limited experience with this approach (see Section 3.4 for a discussion of the

RfDDT and RfCDT).

Using the benchmark dose approach, an LED10 can be calculated for each effect of an agent

for which there is a database with sufficient evidence to conduct a risk assessment. In some cases, the

data may be sufficient to also estimate the ED05 or ED01, which should be closer
35

Figure 1. This graphical illustration of the benchmark dose approach is based on Crump (1984) and Kimmel and Gaylor (1988). The

benchmark dose (BD) is derived by modeling the data in the observed range, selecting an incidence level within or near the observed range

(e.g., the effective dose to produce a 10% increased incidence of response, the ED10), and determining the upper confidence limit on the

model. The upper confidence value corresponding to, for example, a 10% excess in response is used to derive the BD, which is the lower

confidence limit on dose for that level of excess response, in this case the LED10. The RfDDT or RfCDT estimated by applying uncertainty

factors (UF) to the BD would be greater than or equal to the BD/UF.
38

to a true no-effect dose. A level between the ED01 and the ED10 usually corresponds to the lowest

level of risk that can be estimated for binomial endpoints from standard developmental toxicity studies

Certain principles are especially applicable for determining the NOAEL, LOAEL, and

benchmark dose for developmental toxicity studies. First, the NOAEL, LOAEL, or benchmark dose

are identified for both developmental and maternal or adult toxicity, based on the information available

from studies in which developmental toxicity has been evaluated. The NOAEL, LOAEL, or

benchmark dose for maternal or adult toxicity should be compared with the corresponding values from

other adult toxicity data to determine if the pregnant or lactating female or the paternal animal (if

exposure is prior to mating) may be more sensitive to an agent than adult males or nonpregnant females

in other toxicity studies that generally involve longer exposure times.

Second, for developmental toxic effects, a primary assumption is that a single exposure at a

critical time in development may produce an adverse developmental effect, i.e., repeated exposure is

not a necessary prerequisite for developmental toxicity to be manifested. In most cases, however, the

data available for developmental toxicity risk assessment are from studies using exposures over several

days of development, and the NOAEL, LOAEL, and/or benchmark dose is most often based on a

daily dose, e.g., mg/kg-day. Usually, the daily dose is not adjusted for duration of exposure because

appropriate pharmacokinetic data are not available. In cases where such data are available,

adjustments may be made to provide an estimate of equal average concentration at the site of action for

the human exposure scenario of concern. For example, inhalation studies often use 6 hr/day exposures

during development. If the human exposure scenario is continuous and pharmacokinetic data indicate

an accumulation with continuous exposure, appropriate adjustments can be made. If, on the other

hand, the human exposure scenario of concern is very brief or intermittent, pharmacokinetic data

indicating a long half-life may also require adjustment of dose. When quantitative absorption data by

any route of exposure are available, the NOAEL may be adjusted accordingly; e.g., absorption of 50%

of administered dose could result in a 50% reduction in the NOAEL. If absorption in the experimental

species has been determined, but human absorption is not known, human absorption is generally

assumed to be the same as that for the species with the greatest degree of absorption. NOAELs from

inhalation exposure studies are adjusted to derive a human equivalent concentration (HEC) by taking

into account known anatomical and physiological species differences (e.g., minute volume, respiratory

rate, etc.) (U.S. EPA, 1991b).

In summary, the dose-response evaluation identifies the NOAEL, LOAEL, or benchmark

dose, defines the range of doses for a given agent that are effective in producing developmental and

maternal toxicity; the route, timing, and duration of exposure; species specificity of effects, and any
39

pharmacokinetic or other considerations that might influence the comparison with human exposure

scenarios. This information should always accompany the characterization of the health-related

database (discussed in the next section).

3.3. CHARACTERIZATION OF THE HEALTH-RELATED DATABASE

This section describes the process for evaluating the health-related database as a whole on a

particular agent and provides criteria for characterizing the evidence for judging a potential

developmental hazard in humans within the context of expected exposure or dose. This determination

provides the basis for judging whether or not there are sufficient data for proceeding further in the risk

assessment process. This section does not address the nature and magnitude of human health risks,

which are discussed as part of the final characterization of risk along with estimates of potential human

exposure and the relevancy of available data for estimating human risk. Characterization of hazard

potential within the context of exposure or dose should assist the risk assessor in clarifying the strengths

and uncertainties associated with a particular database. Because a complex interrelationship exists

among study design, statistical analysis, and biological significance of the data, a great deal of scientific

judgment, based on experience with developmental toxicity data and with the principles of study design

and statistical analysis, may be required to adequately evaluate the database. Scientific judgment is

always necessary, and in many cases, interaction with scientists in specific disciplines (e.g.,

developmental toxicology, epidemiology, statistics) is recommended.

A categorization scheme for characterizing the evidence for developmental toxicity is presented

in Table 3. The categorization scheme contains two broad categories, sufficient evidence and

insufficient evidence, which are defined in the table. Data from all available studies, whether indicative

of potential hazard or not, must be evaluated and factored into a judgment as to the strength of evidence

available to support a complete risk assessment for developmental toxicity. The primary considerations

are the human data, if available, and the experimental animal data. The judgment of whether the data

are sufficient or insufficient should consider quality of the data, power of the studies, number and types

of endpoints examined, replication of effects, relevance of the test species to humans, relevance of route

and timing of exposure for both human and animal studies, appropriateness of the dose selection in

animal studies, and number of species examined. In addition, pharmacokinetic data and structureactivity considerations, data from other toxicity studies, as well as other factors that may affect the

strength of the evidence, should be taken into account. 

In general, the categorization is based on criteria that define the minimum evidence necessary to

conduct a hazard identification/dose-response evaluation. Establishing the
40

Table 3. Categorization of the health-related database for hazard identification/doseresponse evaluation

SUFFICIENT EVIDENCE

The sufficient evidence category includes data that collectively provide enough information to

judge whether or not a human developmental hazard could exist within the context of dose, duration,

timing, and route of exposure. This category includes both human and experimental animal evidence. 

Sufficient Human Evidence: This category includes data from epidemiologic studies (e.g., case control

and cohort) that provide convincing evidence for the scientific community to judge that a causal

relationship is or is not supported. A case series in conjunction with strong supporting evidence may

also be used. Supporting animal data may or may not be available.

Sufficient Experimental Animal Evidence/Limited Human Data: This category includes data from

experimental animal studies and/or limited human data that provide convincing evidence for the scientific

community to judge if the potential for developmental toxicity exists. The minimum evidence necessary

to judge that a potential hazard exists generally would be data demonstrating an adverse developmental

effect in a single, appropriate, well-conducted study in a single experimental animal species. The

minimum evidence needed to judge that a potential hazard does not exist would include data from

appropriate, well-conducted laboratory animal studies in several species (at least two) which evaluated

a variety of the potential manifestations of developmental toxicity and showed no developmental effects

at doses that were minimally toxic to the adult.

INSUFFICIENT EVIDENCE

This category includes situations for which there is less than the minimum sufficient evidence

necessary for assessing the potential for developmental toxicity, such as when no data are available on

developmental toxicity, as well as for databases from studies in animals or humans that have a limited

study design (e.g., small numbers, inappropriate dose selection/exposure information, other

uncontrolled factors), or data from a single species reported to have no adverse developmental effects,

or databases limited to information on structure/activity relationships, short-term tests,

pharmacokinetics, or metabolic precursors.
41

minimum sufficient human evidence necessary to do a hazard identification/dose-response evaluation is

difficult, since there are often 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, there are currently few agents that can be classified in this category. In the

case of animal data, agents that have been tested adequately in laboratory animals according to current

test guidelines generally would be included in the “Sufficient Experimental Animal Evidence/Limited

Human Data” category. The strength of evidence for a database increases with replication of the

findings and with additional animal species tested. Information on pharmacokinetics or mechanisms, or

on more than one route of exposure may reduce uncertainties in extrapolation to the human. 

More evidence is necessary to judge that an agent is unlikely to pose a hazard for

developmental toxicity than that required to judge a potential hazard. This is because it is more difficult,

both biologically and statistically, to support a finding of no apparent adverse effect than a finding of an

adverse effect. For example, to judge that a hazard for developmental toxicity could exist for a given

agent, the minimum evidence necessary would be data from a single, appropriate, well-executed study

in a single experimental animal species that demonstrate developmental toxicity, and/or suggestive

evidence from adequately conducted clinical/epidemiologic studies. On the other hand, to judge that an

agent is unlikely to pose a hazard for developmental toxicity, the minimum evidence would include data

from appropriate, well-executed laboratory animal studies in several species (at least two) which

evaluated a variety of the potential manifestations of developmental toxicity and showed no adverse

developmental effects at doses that were minimally toxic to the adult animal. In addition, there may be

human data from appropriate studies supportive of no adverse developmental effects. 

If a database on a particular agent includes less than the minimum sufficient evidence (as defined

in the “Insufficient Evidence” category) necessary for a risk assessment, but some data are available,

this information could be used to determine the need for additional testing. 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 Evidence” criteria. 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 in an attempt to define the strength of evidence of the database for

assessing the potential for developmental toxicity.

Judging that the health-related database is sufficient to indicate a potential developmental hazard

does not mean that the agent will be a hazard at every exposure level (because of the assumption of a
42

threshold) or in every situation (e.g., hazard may vary significantly depending on route and timing of

exposure). In the final risk characterization, the characterization of the health-related database should

always be presented with information on the dose-response evaluation (e.g., LOAEL, NOAEL, and/or

benchmark dose), exposure route, timing and duration of exposure, and with the human exposure

estimate.

3.4. DETERMINATION OF THE REFERENCE DOSE (RfDDT) OR REFERENCE

CONCENTRATION (RfCDT) FOR DEVELOPMENTAL TOXICITY

The RfDDT or RfCDT is an estimate of a daily exposure to the human population that is assumed

to be without appreciable risk of deleterious developmental effects. The use of the subscript DT is

intended to distinguish these terms from the reference dose (RfD) for oral or dermal exposure or the

reference concentration (RfC) for inhalation exposure, terms that refer primarily to chronic exposure

situations (U.S. EPA, 1991b). The RfDDT or RfCDT is derived by applying uncertainty factors to the

NOAEL (or the LOAEL, if a NOAEL is not available), or the benchmark dose. To date, the Agency

has applied uncertainty factors only to the NOAEL or LOAEL to derive an RfDDT or RfCDT. The

Agency is planning eventually to use the benchmark dose approach as the basis for derivation of the

RfDDT or RfCDT and will develop guidance as information is acquired and analyzed from ongoing

Agency studies.

The most sensitive developmental effect (i.e., the critical effect) from the most appropriate

and/or sensitive mammalian species is used for determining the NOAEL, LOAEL, or the benchmark

dose in deriving the RfDDT or RfCDT (Section 3.2). Uncertainty factors (UFs) for developmental and

maternal toxicity applied to the NOAEL generally include a 10-fold factor for interspecies variation and

a 10-fold factor for intraspecies variation. In general, an uncertainty factor is not applied to account for

duration of exposure.

Additional factors may be applied to account for other uncertainties or additional information

that may exist in the database. For example, the standard study design for a developmental toxicity

study calls for a low dose that demonstrates a NOAEL, but in some cases, the lowest dose

administered may cause significant adverse effect(s), and thus be identified as the LOAEL. 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. In addition, if a benchmark dose has been calculated, it

may be used to help interpret how close the LOAEL is to a level that would not be detectable from

controls (equivalent to the NOAEL), and thus the size of the uncertainty factor to be applied. Other
43

modifying factors (MFs) may be used depending on the characterization of the database (Section 3.3),

data on pharmacokinetics, or other considerations that may alter the level of confidence in the data

(U.S. EPA, 1991b). The total size of the uncertainty factor will vary from agent to agent and will

require the exercise of scientific judgment, taking into account interspecies differences, variability within

species, the slope of the dose-response curve, the background incidence of the effects, the route of

administration, and pharmacokinetic data.

As stated above, there is little experience with the application of uncertainty factors to the

benchmark dose approach for calculating the RfDDT or RfCDT, and there are several issues that must

be addressed prior to its use for this purpose. For example, which benchmark dose (e.g., LED01,

LED05, LED10) should be used for calculating the RfDDT or RfCDT, and what are the appropriate

uncertainty factors that should be applied to the benchmark dose for deriving the RfDDT or RfCDT? 

That is, should the uncertainty factor applied to an LED10 be similar to that applied to a LOAEL, or

should the uncertainty factor applied to an LED01 be equal to or less than that applied to a NOAEL? 

These and other questions are being addressed in ongoing Agency studies on the calculation of the

RfDDT or RfCDT using the benchmark dose approach. As results become available, and as further

guidance is developed, this information will be published as a supplement to these Guidelines.

The total uncertainty factor selected is divided into the NOAEL or LOAEL (or the benchmark

dose) for the critical effect in the most appropriate and/or sensitive mammalian species to determine the

RfDDT or RfCDT. If the NOAEL, LOAEL, or benchmark dose for maternal toxicity is lower than that

for developmental toxicity, this should be noted in the risk characterization, and this value compared

with data from other studies in which adult animals are exposed.

The modeling approaches that have been proposed for developmental toxicity are, for the most

part, statistical probability models that do not take into account underlying biological processes or

mechanisms (e.g., Crump, 1984; Rai and Van Ryzin, 1985; Kimmel and Gaylor, 1988; Faustman et al.,

1989; Chen and Kodell, 1989; Kodell et al., 1991). These models can be applied to derive doseresponse curves for data in the observed dose range, but may or may not accurately predict risk at low

levels of exposure. It has generally been assumed that there is a biological threshold for developmental

toxicity; however, a threshold for a population of individuals may or may not exist because of other

endogenous or exogenous factors that may increase the sensitivity of some individuals in the population. 

Thus, the addition of a toxicant may result in an increased risk for the population, but not necessarily for

all individuals in the population.

Models that are more biologically based should provide a more accurate estimation of lowdose risk to humans. The development of biologically based dose-response models in developmental
44

toxicology has been limited by a number of factors, including a lack of understanding of the biological

mechanisms underlying developmental toxicity, intra/interspecies differences in the types of

developmental events, appropriate pharmacokinetic data, and the influence of maternal effects on the

dose-response curve. The Agency is currently supporting several major research efforts to develop

biologically based dose-response models for developmental toxicity risk assessment that include the

consideration of threshold under its Research to Improve Health Risk Assessment program.

3.5. SUMMARY

In summary, the hazard identification/dose-response evaluation of developmental toxicity data is

used as part of the final characterization of risk along with information on estimates of human exposure. 

This analysis depends on scientific judgment as to the accuracy and sufficiency of the health-related

data, biological relevance of significant effects, the conditions of human exposure, and other

considerations important in the extrapolation of data from animals to humans. Scientific judgment is

always necessary, and in many cases, interaction with scientists in specific disciplines (e.g.,

developmental toxicology, epidemiology, statistics) is recommended.
45

4. EXPOSURE ASSESSMENT

In order to obtain quantitative estimates of risk for human populations, estimates of human

exposure are required. This discussion is not intended to provide definitive guidance on exposure

assessment; the “Guidelines for Estimating Exposures” have been published separately (U.S. EPA,

1986d) and will not be discussed in detail here. Rather, the issues important to developmental toxicity

risk assessment are addressed. In general, the exposure assessment describes the magnitude, duration,

frequency, and route(s) of exposure. This information is usually developed from monitoring data and

from estimates based on various scenarios of environmental exposures.

There are several exposure considerations that are unique for developmental toxicity. For

example, exposure to developing individuals is often secondary via placental transfer or through breast

milk. Thus, exposure to the embryo/fetus or child may not be the same as for the pregnant or lactating

mother, and measurements of an agent in maternal or cord blood and in breast milk may provide a

better estimate of developmental exposure. Direct exposure of neonates and children may also occur

via environmental media such as water, air and soil, and thus may require estimates of exposure from

multiple sources. Duration and period of exposure also must be related to stage of development, if

possible (e.g., first, second, or third trimester of pregnancy, infancy, early, middle, and late childhood,

adolescence, etc.). These stages of development may have different sensitivities to agents, and

exposure estimates should be derived for as many as possible. In addition, exposure to either parent

prior to conception must be considered in relation to adverse developmental effects. 

There is also a possibility that a single exposure may be sufficient to produce adverse

developmental effects (i.e., 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 sufficiently long to produce an increasing body burden

over time). Therefore, it is assumed that, in most cases, a single exposure at any of several

developmental stages may be sufficient to produce an adverse developmental effect. Most of the data

available for risk assessment involve exposures over several days of development. Thus, human

exposure estimates used to calculate margins of exposure (MOE, see following section) or to compare

with the RfDDT or RfCDT are usually based on a daily dose that is not adjusted for duration or pattern of

exposure. For example, it would be inappropriate in developmental toxicity risk assessments 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-hr inhalation exposure to account for a 24-hr exposure

scenario), unless pharmacokinetic data were available to indicate an accumulation with continuous
46

exposure. In the case of intermittent exposures, examination of the peak exposure(s), as well as the

average exposure over the time period of exposure, would be important.

It should be recognized that, based on the definition used in these Guidelines for developmental

toxicity, exposure of almost any segment of the human population may lead to risk to the developing

organism. This would include fertile men and women, the developing embryo and fetus, and children up

to the age of sexual maturation. Although some effects of developmental exposures may be manifested

while the exposure is occurring (e.g., spontaneous abortion, structural abnormality present at birth,

childhood mental retardation), some effects may not be detectable until later in life, long after exposure

has ceased (e.g., perinatally induced carcinogenesis, impaired reproductive function, shortened

lifespan).
47

5. RISK CHARACTERIZATION

5.1. OVERVIEW

Risk characterization is the culmination of the risk assessment process. In this final step, risk

characterization involves integration of the toxicity information from the hazard identification/doseresponse evaluation 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 the

results of the risk assessment to a risk manager. The risk manager can then use the risk assessment,

along with other risk management elements, to make public health decisions. The following sections

describe these three 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 developmental toxicity.

5.2. INTEGRATION OF THE HAZARD IDENTIFICATION/DOSE-RESPONSE

EVALUATION AND EXPOSURE ASSESSMENT

In developing the hazard identification/dose-response and exposure portions of the risk

assessment, the risk assessor makes many judgments concerning human relevance of the toxicity data,

including the appropriateness of the various animal models for which data are available, the route,

timing, and duration of exposure relative to expected human exposure, etc. These judgments should be

summarized at each stage of the risk assessment process (e.g., the biological relevance of anatomical

variations may be made in the hazard identification process, or species differences in metabolic patterns

in the dose-response evaluation). When data are not available to make such judgments, as is often the

case, the background information and assumptions discussed in the Introduction (Section 1) provide a

default position. The risk assessor 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 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 the

analysis of these elements. Each of these elements should have its own characterization as a part of it. 

Interpretation of data should be explained, and the risk manager should be given a clear picture of

consensus or lack of consensus that exists about significant aspects of the assessment. Whenever more

than one view is supported by the data and choosing between them is difficult, both views should be
48

presented. If one has been selected over another, the rationale should be given; if not, then both should

be presented as plausible alternative results.

The risk characterization should not only examine the judgments, but also explain the constraints

of available data and the state of knowledge about the phenomena studied in making them, including:

C the qualitative conclusions about the likelihood that the agent 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 to use in a risk assessment;

C a discussion of the dose-response patterns for the critical effect(s), data such as the shapes

and slopes of the dose-response curves for the various endpoints, the rationale behind the

determination of the NOAEL, LOAEL, and/or calculation of the benchmark dose, and the

assumptions underlying the estimation of the RfDDT or RfCDT; and

C the estimates of the magnitude of human exposure, the route, duration, and pattern of the

exposure, relevant pharmacokinetics, and the size and characteristics of the populations

exposed.

The risk characterization of an agent should be based on data from the most appropriate

species, or, if such information is not available, on the most sensitive species tested. It should also be

based on the most sensitive indicator of toxicity, whether maternal, paternal, or developmental, when

such data are available, and should be considered in relationship to other forms of toxicity.

If data used in characterizing risk 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, 1984, 1985b).

The level of confidence in the hazard identification/dose-response evaluation should be stated to

the extent possible, including determination of the appropriate category regarding sufficiency of the

health-related data. A comprehensive risk assessment ideally includes information on a variety of

endpoints that provide insight into the full spectrum of developmental responses. A profile that

integrates both human and test species data and incorporates a broad range of developmental effects

provides more confidence in a risk assessment for a given agent. 

The ability to describe the nature of human exposure is important for prediction of specific

outcomes and the likelihood of permanence or reversibility of the effect. An important part of this effort

is a description of the nature of the exposed populations. For example, the consequences of exposure

to the developing individual versus the adult can differ markedly and again can influence whether the
49

effects are transient or permanent. Other considerations relative to human exposures might include

potential synergistic effects, increased susceptibility resulting from concurrent exposures to other agents,

concurrent disease, and nutritional status.

5.3. DESCRIPTORS OF DEVELOPMENTAL TOXICITY RISK

There are a number of ways to describe risks. These include:

5.3.1. Estimation of the Number of Individuals Exposed to Levels of Concern

The RfDDT or RfCDT is assumed to be a level at or below which no significant risk occurs. 

Therefore, information on the populations at or below the RfDDT or RfCDT (“not likely to be at risk”)

and above the RfDDT or RfCDT (“may be at risk”) may be useful information for risk managers.

This method is particularly useful to a risk manager considering possible actions to ameliorate

risk for a population. If the number of persons in the “at risk” category can be estimated, then the

number of persons potentially removed from the “at risk” category after a contemplated action is taken

can be used as an indication of the efficacy of that action.

5.3.2. Presenting Specific Scenarios

Presenting specific scenarios in the form of “what if?” questions is particularly useful to give

perspective to the risk manager, especially where criteria, tolerance limits, or media quality limits are

being set. The question being asked in these cases is, “At this proposed limit, what would be the

resulting risk for developmental toxicity above the RfDDT?”

5.3.3. Risk Characterization for Highly Exposed Individuals

This measure and the next are examples of specific scenarios. The purpose of this measure is

to describe the upper end of the exposure distribution. This allows risk managers to evaluate whether

certain individuals are at disproportionately high or unacceptably high risk.

The objective of looking at the upper end of the exposure distribution is to derive a realistic

estimate of a relatively highly exposed individual(s), for example, by identifying a specified upper

percentile of exposure in the population and/or by estimating the exposure of the most highly exposed

individual(s). Whenever possible, it is important to express the number of individuals who comprise the

highly exposed group and discuss the potential for exposure at still higher levels.

If population data are absent, it will often be possible to describe a scenario representing highend exposures using upper percentile or judgment-based values for exposure variables. In these
50

instances, caution should be taken not to overestimate the high-end values if a “reasonable” exposure

estimate is to be achieved.

5.3.4. Risk Characterization for Highly Sensitive or Susceptible Individuals

The purpose of this measure is to quantify exposure to identified sensitive or susceptible

populations to the effect of concern. Sensitive or susceptible individuals are those within the exposed

population at increased risk of expressing the adverse effect. All stages of development might be

considered highly sensitive or susceptible, but certain subpopulations can sometimes be identified

because of critical periods for exposure; for example, pregnant or lactating women, infants, children,

adolescents.

In general, not enough is understood about the mechanisms of toxicity to identify sensitive

subgroups for all agents, although factors such as nutrition, personal habits (e.g., smoking, alcohol

consumption, illicit drug abuse), or pre-existing disease (e.g., diabetes) may predispose some

individuals to be more sensitive to the developmental effects of various agents.

5.3.5. Other Risk Descriptors

In risk characterization, dose-response information and the human exposure estimates may be

combined either by comparing the RfDDT or RfCDT and the human exposure estimate or by calculating

the margin of exposure (MOE). The MOE is the ratio of the NOAEL from the most appropriate or

sensitive species to the estimated human exposure level from all potential sources (U.S. EPA, 1985b). 

If a NOAEL is not available, a LOAEL may be used in the calculation of the MOE, but considerations

for the acceptability would be different from those when a NOAEL is used. Considerations for the

acceptability of the MOE are similar to that for the uncertainty factor applied to the LOAEL, NOAEL,

or the benchmark dose. 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 dose-response, route, timing, and duration information. The RfDDT or RfCDT

comparison with the human exposure estimate and the calculation of the MOE are conceptually similar

but are used in different regulatory situations. If the MOE is equal to or more than the uncertainty

factor used as a basis for an RfDDT or RfCDT, then the need for regulatory concern is likely to be

reduced. 

The choice of approach is dependent upon several factors, including the statute involved, the

situation being addressed, the database used, and the needs of the decision maker. While these

methods of describing risk do not actually estimate risks per se, they give the risk manager some sense
51

of how close the exposures are to levels of concern. The RfDDT, RfCDT, and/or the MOE are

considered along with other risk assessment and risk management issues in making risk management

decisions, and the scientific issues that must be taken into account in establishing them have been

addressed here.

5.4. COMMUNICATING RESULTS

Once the risk characterization is completed, the focus turns to communicating results to the risk

manager. The risk manager uses the results of the risk characterization, other technological factors, and

nontechnological social and economic considerations in reaching a regulatory decision. Because of the

way in which these risk management factors may impact different cases, consistent but not necessarily

identical risk management decisions must be made on a case-by-case basis. Consequently, it is entirely

possible and appropriate that an agent with a specific risk characterization may be regulated differently

under different statutes. These Guidelines are not intended to give guidance on the nonscientific aspects

of risk management decisions.
52

6. SUMMARY AND RESEARCH NEEDS

These Guidelines summarize the procedures that the U.S. Environmental Protection Agency

uses in evaluating the potential for agents to cause developmental toxicity. While these are the first

amendments to the developmental toxicity guidelines issued in 1986, further revisions and updates will

be made as advances occur in the field. These Guidelines discuss the assumptions that should be made

in risk assessment for developmental toxicity because of gaps in our knowledge about underlying

biological processes and how these compare across species. 

Research to improve the risk assessment process is needed in a number of areas. For example,

research is needed to delineate the mechanisms of developmental toxicity and pathogenesis, provide

comparative pharmacokinetic data, examine the validity of short-term in vivo and in vitro tests, elucidate

possible functional alterations and their critical periods of exposure to toxic agents, develop improved

animal models to examine the developmental effects of exposure during the premating and early

postmating periods and in neonates, further evaluate the relationship between maternal and

developmental toxicity, provide insight into the concept of threshold, develop approaches for improved

mathematical modeling of adverse developmental effects, and improve animal models for examining the

effects of agents given by various routes of exposure. Epidemiologic studies with quantitative measures

of exposure are also strongly encouraged. Such research will aid in the evaluation and interpretation of

data on developmental toxicity, and should provide methods to more precisely assess risk.
53

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PART B: RESPONSE TO PUBLIC AND SCIENCE ADVISORY BOARD COMMENTS

1. INTRODUCTION

This section summarizes the major issues raised in the public and Science Advisory Board

(SAB) comments on the Proposed Amendments to the Guidelines for the Health Assessment of

Suspect Developmental Toxicants published March 6, 1989 [54 FR 9386-9403]. Comments were

received from 25 individuals or organizations. The Agency’s initial summary of the public comments

and proposed responses were presented to the Environmental Health Committee of the SAB on

October 27, 1989. The report of the SAB Committee was provided to the Agency on April 23, 1990.

The SAB and public comments were diverse and addressed issues from a variety of

perspectives. The majority of the comments were favorable and in support of the Proposed

Amendments to the Guidelines. Many praised the Agency’s efforts as being timely and well-justified. 

Most commentors also gave specific comments or criticisms for further consideration, clarification, or

re-evaluation. For example, there was concern expressed about the Guidelines imposing further testing

requirements, particularly functional testing, and many commentors felt that the Proposed Amendments

discounted the role of maternal toxicity in developmental toxicity. In addition, there was concern that

the proposed weight-of-evidence scheme would promote labeling of agents as causing developmental

toxicity before the entire risk assessment process was completed.

The SAB Committee also indicated that the proposed revisions were adequately founded in

developmental toxicology and represented a step forward for the Agency. They suggested that the

Agency revisit the weight-of-evidence scheme to avoid confusion with more commonly applied uses of

such classifications, and to develop a more powerful conceptual approach. Further, the SAB

Committee urged that the Agency begin to move away from the current use of the no-observedadverse-effect level (NOAEL) and lowest-observed-adverse-effect level (LOAEL) basis for

calculating the reference dose for developmental toxicity to a benchmark dose and confidence limit

approach tied to empirical models of dose-response relationships.

In response to the comments, the Agency has modified or clarified many sections of the

Guidelines. For the purposes of this discussion, the major issues reflected by the public and SAB

comments are discussed. Several minor recommendations, which are not discussed specifically here,

also were considered by the Agency in the revision of these Guidelines.
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2. INTENT OF THE GUIDELINES

Many of the public comments indicated some misunderstanding of the intent of the Guidelines,

apparently assuming that the risk assessment guidelines impose testing requirements. In particular,

some commentors suggested that because the Agency was providing guidance on the interpretation of

tests not required in the EPA testing guidelines, the Agency was suggesting that these tests be required

in the future.

The 1986 Guidelines and the 1989 Proposed Amendments clearly state that these guidelines

are not Agency testing guidelines, but rather are intended to ensure uniform interpretation of all existing,

relevant data. However, to avoid any confusion, the discussion of study designs has been changed to

avoid the impression that these Guidelines set testing requirements. In the evaluation of data on an

agent for risk assessment, relevant data are often encountered that have been generated from

nontraditional tests. In such cases, it is imperative that the Agency provide guidance so that all data

considered to be relevant are included in the risk assessment and are interpreted uniformly.

3. BASIC ASSUMPTIONS

In the 1986 Guidelines, several assumptions were implicit in the approach to risk assessment,

but were not explicitly stated. These assumptions were detailed in the 1989 Proposed Amendments. 

Comments received from the public and the SAB favored presentation of these assumptions and

generally agreed with the wording, except for the fourth assumption, which concerns the use of the most

relevant or most sensitive species. The 1989 Proposed Amendments stated that “it is assumed that the

most sensitive species should be used to estimate human risk. When data are available (e.g.,

pharmacokinetic, metabolic) to suggest the most appropriate species, that species will be used for

extrapolation.” The SAB recommended that, for this assumption, the basic position of the Agency

should be to use data from the most relevant species, and that use of data from the most sensitive

species should be the default position. In addition, the SAB recommended that the threshold

assumption be considered carefully in the dose-response assessment of any agent, and that the Agency

develop more comprehensive approaches to risk assessment as discussed further in the following

sections.

Changes have been made in the statement of the basic assumptions in line with the SAB and

public comments that clarify, but do not alter, the intent of the assumptions.
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4. MATERNAL/DEVELOPMENTAL TOXICITY

The 1989 Proposed Amendments stated that “when adverse developmental effects are

produced only at maternally toxic doses, they are still considered to represent developmental toxicity

and should not be discounted as being secondary to maternal toxicity.” This statement and others

concerning the interpretation of developmental toxicity in the presence of maternal toxicity were the

subject of a considerable number of public comments and were also addressed by the SAB. In

general, commentors were divided in their opinions on whether they supported the Agency’s statements

or felt that they discounted the role of maternal toxicity in developmental toxicity, but in general, the

recommended changes did not significantly alter the intent of the statements. The SAB endorsed the

proposed revision, and suggested that the Agency retain the statement that was made in the Proposed

Amendments.

In these Guidelines, the position is further clarified by indicating that when maternal toxicity is

significantly greater than the minimal maternally toxic dose, developmental effects at that dose may be

difficult to interpret. This statement is added to clarify, but not to change, the intent or meaning of the

statements regarding the relationship between maternal and developmental toxicity. From a risk

assessment point of view, whether a developmental effect is or is not secondary to maternal toxicity

does not impact on the selection of the NOAEL or other dose-response methodology.

5. FUNCTIONAL DEVELOPMENTAL TOXICITY

The 1989 Proposed Amendments provided information on the state-of-the-art in the evaluation

of functional effects resulting from developmental exposures. Several commentors voiced strong

objection to this section because they perceived it as indicating an imminent requirement for testing. 

Several indicated there are no standard methods for functional testing, some felt that functional

endpoints should not be used to establish the NOAEL, and others voiced concern about the problems

with using postnatal exposures in animal studies. 

The final Guidelines further update this section to include a discussion of the latest changes in

the requirements for functional developmental toxicity testing by the Agency, and reflect the current

approach to interpretation of such data, with incorporation of information from the EPA/NIDAsponsored “Workshop on the Qualitative and Quantitative Comparability of Human and Animal

Developmental Neurotoxicity” (1990). The intent of these Guidelines, as stated above, is not to change

testing requirements but to give guidance when these types of data are encountered in the risk
69

assessment process. The Guidelines also indicate that functional developmental toxicity endpoints will

be used for establishing the NOAEL when they are found to be the adverse effect occurring at the

lowest dose in appropriate, well-conducted studies. Interpretation of postnatal exposure data is a

concern, and must take into consideration effects on the mother, her offspring, and possible

interactions; a statement to this effect has been added. Further interpretation of data will be discussed

in the guidance being developed by the Agency on neurotoxicity risk assessment.

6. WEIGHT-OF-EVIDENCE SCHEME

The 1989 Proposed Amendments described important considerations in determining the

relative weight of various kinds of data in estimating the risk of developmental toxicity in humans. The

intent of the proposed weight-of-evidence (WOE) scheme was that it not be used in isolation, but be

used as the first step in the risk assessment process, to be integrated with dose-response information

and the exposure assessment. 

The WOE scheme was the subject of a considerable number of public comments, and was one

of the major concerns of the SAB. The concern of public commentors was that the reference to human

developmental toxicity in this scheme suggested that a chemical could be prematurely designated, and

perhaps labeled, as causing developmental toxicity in humans prior to the completion of the risk

assessment process. The SAB suggested that the intended use of this scheme was not consistent with

the use of the term “weight of evidence” in other contexts, since WOE is usually thought of as an

evaluation of the total composite of information available to make a judgment about risk. In addition,

the SAB Committee proposed that the Agency consider development of a more conceptual approach

using decision analytical techniques to predict the relationships among various outcomes. 

In the final Guidelines, the terminology used in the WOE scheme has been completely changed

and retitled “Characterization of the Health-Related Database.” The intended purpose of the scheme is

to provide a framework and criteria for making a decision on whether or not sufficient data are

available to conduct a risk assessment. This decision is based on the available data, whether animal or

human, and does not necessarily imply human hazard. This decision process is part of, but not the

complete, WOE evaluation, which also takes into account the RfDDT or RfCDT and the human

exposure information, culminating in risk characterization. 

The final Guidelines also place strong emphasis on the integration of the dose-response

evaluation with hazard information in characterizing the sufficiency of the health-related database. In

line with this approach, the Guidelines have been reorganized to combine hazard identification and
70

dose-response evaluation. Finally, the SAB comments on developing a conceptual matrix provide an

interesting challenge, but current data indicate that the relationships among endpoints of developmental

toxicity are not consistent across chemicals or species. The Agency is currently supporting modeling

efforts to further explore the relationship among various development toxicity endpoints and the

development of biologically based dose-response models that consider multiple effects.

7. APPLICABILITY OF THE RfDDT CONCEPT AND 

THE BENCHMARK DOSE APPROACH

The 1989 Proposed Amendments introduced the term “reference dose for developmental

toxicity - RfDDT,” based on short-term exposure, to distinguish it from the reference dose (RfD), which

is used for chronic exposure situations. The public comments received generally supported the RfDDT

approach. The SAB also agreed with the concept of the RfDDT for developmental toxicity risk

assessment, based on short-term exposure. In addition, the SAB urged the Agency to consider

strengthening the RfD approach by moving to more quantitative alternatives to the NOAEL. In

particular, the use of a benchmark dose approach to replace the NOAEL was strongly suggested. 

The final Guidelines have incorporated many of the SAB Committee’s suggestions concerning

the development of more quantitative approaches to the RfD, and state that the Agency is beginning to

use the benchmark dose approach for comparison with and interpretation of the NOAEL. That is,

benchmark dose calculations may allow better interpretation of dose-response data and, in particular,

what level of risk may be associated with the NOAEL. The Agency also has developed the concept of

an inhalation reference concentration (RfC), and the RfC DT is being calculated for inhalation

concentrations based on developmental toxicity. Guidance for use of the benchmark dose in the

calculation of the RfDDT or RfCDT is not included in the final Guidelines, because of the limited

experience of the Agency with this approach. There are several issues that must be addressed prior to

its use for this purpose; for example, which benchmark dose (e.g., LED01, LED05, LED10) should be

used for calculating the RfDDT or RfCDT, and what are the appropriate uncertainty factors that should

be applied to the benchmark dose for deriving the RfDDT or RfCDT? Should the uncertainty factor

applied to an LED10 be similar to that applied to a LOAEL, or should the uncertainty factor applied to

an LED01 be equal to or less than that applied to a NOAEL? These and other questions are being

addressed in ongoing Agency studies on the calculation of the RfDDT or RfCDT using the benchmark

dose approach. As results become available, and as further guidance is developed, this information will

be published as a supplement to these Guidelines.
71