Pesticides in the Diets of Infants and Children, by NRC

Pesticides in the Diets of Infants and Children, by NRC

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Committee on Pesticides in the Diets of Infants and Children
Board on Agriculture and Board on Environmental Studies and Toxicology
Commission on Life Sciences
National Research Council
Suggested Citation: "FRONT MATTER." National Research Council. Pesticides in the Diets of Infants and Children. Washington, DC: The National Academies Press, 1993. doi:10.17226/2126.




o Pesticide Use
o Pesticide Control Legislation
o Approach to the Study
o References
o Growth
o Development
o Conclusions and Recommendations
o References
o Acute Toxicity
o Neurotoxicity
o Immunotoxicity
o Carcinogenesis and Mutagenesis
o Metabolism and Pharmacokinetics
o Scaling and Progression Analysis
o Conclusions and Recommendations
o References
o Current Methods: General Considerations
o Acute Toxicity Studies
o Subchronic Toxicity Studies
o Chronic Toxicity Studies
o Developmental Toxicity Studies
o Reproduction Studies
o Mutagenicity Studies
o General Metabolism Studies
o Neurotoxicity Studies
o Special Testing
o Conclusions and Recommendations
o References
o Food Consumption Surveys
o Survey Methodology
o Survey Design
o Sample Weights
o Sample Size
o Comparisons of Intake Data with Standards
o Validation of Food Consumption Data
o The Strengths and Weaknesses of the Food Consumption Data Bases in Estimating Pesticide Exposure of Children
o Water Intake
o Quantification of Consumption Data
o Age-Related Differences in Dietary Patterns
o Issues Related to the Evaluation of Food Monitoring Data
o Conclusions and Recommendations
o References
o Sources of Data on Usage
o The Occurrence and Fate of Pesticide Residues
o Pesticide Registration and the Development of Analytical Methods
o Methods for Sampling and Analysis
o Monitoring
o Quality Controls
o Limitations of the Data
o Pesticides in Water
o Pesticides in Infant Formula
o Pesticides in Human Milk
o Pesticides in Foods
o Conclusions and Recommendations
o References
o The Use of Food Consumption and Residue Data for Exposure Assessment
o Long-Term Exposure to Benomyl
o Short-Term Exposure to Aldicarb
o Multiple Exposure Assessment: Organophosphate Insecticides
o Nondietary Exposure to Pesticides
o Conclusions and Recommendations
o References
o General Principles of Risk Assessment
o Risk Assessments for Infants and Children
o Conclusions and Recommendations
o References

NOTICE: The project that is the subject of this report was approved by the Governing Board of the National Research Council, whose members are drawn from the councils of the National Academy of Sciences, the National Academy of Engineering, and the Institute of Medicine. The members of the committee responsible for the report were chosen for their special competences and with regard for appropriate balance.

This report has been reviewed by a group other than the authors according to procedures approved by a Report Review Committee consisting of members of the National Academy of Sciences, the National Academy of Engineering, and the Institute of Medicine.

Support for this project was provided by the U.S. Environmental Protection Agency, Contract No. 68D-80101, with contributions from International Life Sciences Institute and Health and Welfare Canada. In addition, support for this project was provided by the Kellogg Endowment Fund of the National Academy of Sciences and the Institute of Medicine.

Copyright 1993 by the National Academy of Sciences. All rights reserved.

Any opinions, findings, conclusions, or recommendations expressed in this publication are those of the author(s) and do not necessarily reflect the view of the organizations or agencies that provided support for this project.


Mount Sinai School of Medicine, New York

Graduate School of Public Health, University of Pittsburgh, Pittsburgh

Rockefeller University and Yeshiva University, New York

Boston University Medical School and Boston City Hospital Pediatrics, Boston

University of Georgia, Athens

University of Medicine and Dentistry of New Jersey, Robert Wood Johnson Medical School, Piscataway

New York State Psychiatric Institute, New York

California State Department of Health Services, Berkeley

Graduate School of Public Health, University of Pittsburgh, Pittsburg

Health and Welfare Canada, Ottawa, Ontario

Gerber Products Company, Fremont, Mich.

U.S. Department of Agriculture, Hyattsville, Md.

University of Nevada, Reno

Michigan State University, East Lansing


FRANCES M. PETER, Project Manager

RICHARD D. THOMAS, Principal Staff Scientist (BEST)

CRAIG A. COX, Senior Staff Officer (BA)

SANDRA S. FITZPATRICK, Senior Program Assistant (BEST)

SHELLEY A. NURSE, Senior Project Assistant (BEST)

RUTH P. DANOFF, Project Assistant (BEST)

Technical Advisers

Environmental Systems International, Vienna, Va.

Health and Welfare Canada, Ottawa, Ontario

Jellinek, Schwartz, Connolly and Freshman, Washington, D.C.

Yale University, New Haven, Conn.

Center for Resource Economics, Washington, D.C.


University of California, Davis

American Association for the Advancement of Science, Washington, D.C.

Montana State University, Bozeman

Cornell University, Ithaca, N.Y.

Delaware State College, Dover

Land O'Lakes, Inc., Minneapolis, Minn.

Natural Resources Consultant, Decorah, Iowa

U.S. Department of Agriculture, Forest Service, Madison, Wis.

Purdue University, West Lafayette, Ind.

University of California, Davis

Cornell University, Ithaca, N.Y.

Colorado State University, Fort Collins

Iowa State University, Ames

The Upjohn Company, Kalamazoo, Mich.

The Trustees of Reservations, Beverly, Mass


SUSAN E. OFFUTT, Executive Director

JAMES E. TAVARES, Associate Executive Director

CARLA CARLSON, Director of Communications


JANET OVERTON, Associate Editor


University of Miami, Oxford, Ohio

Cadwalader, Wickersham & Taft, Washington, D.C.

McGill University School of Medicine, Montreal, Quebec, Canada

University of Michigan, Ann Arbor, Mich.

Virginia Polytechnic Institute and State University, Blacksburg, Va.

Department of Natural Resources and Environmental Control, State of Delaware, Dover, Del.

Lilly Research Laboratories, Greenfield, Ind.

Unionville, Pa.

Fox Chase Cancer Center, Philadelphia, Pa.

Williams College, Williamstown, Mass.

The New York Botanical Garden, Millbrook, N.Y.

Oregon State University, Corvallis, Oreg.

University of Pittsburgh, Pittsburgh, Pa.

Stanford University, Stanford, Calif.

University of Washington, Seattle, Wash.

Vanderbilt University, Nashville, Tenn., and Clemson University, Anderson, S.C.

Hilton Head, S.C.

AFL/CIO, Washington, D.C.

University of Arizona, Tucson, Ariz.

University of Oklahoma, Oklahoma City, Okla.

University of Michigan, Ann Arbor, Mich.


JAMES J. REISA, Director

DAVID J. POLICANSKY, Associate Director and Program Director for Natural Resources and Applied Ecology

RICHARD D. THOMAS, Associate Director and Program Director for Human Toxicology and Risk Assessment

LEE R. PAULSON, Program Director for Information Systems and Statistics

RAYMOND A. WASSEL, Program Director for Environmental Sciences and Engineering


Johns Hopkins Medical School, Baltimore, Md.

University of California, Berkeley, Calif.

Hooper Research Foundation, University of California Medical Center, San Francisco, Calif.

Stanford University School of Medicine, Stanford, Calif.

University of California, Riverside, Calif.

Washington State University, Pullman, Wash.

University of Washington, Seattle, Wash.

University of California, Berkeley, Calif.

University of Oxford, Oxford, United Kingdom

Southwestern Medical Center, Dallas, Tex.

Hoffmann-La Roche Inc., Nutley, N.J.

University of Southern California School of Medicine, Los Angeles, Calif.

Miami University, Oxford, Ohio

University of New Mexico School of Medicine, Albuquerque, N. Mex.

Armonk, N.Y.

University of California, Berkeley, Calif.

University of California at San Diego, La Jolla, Calif.

Merck and Company, Inc., Whitehouse Station, N.J.

Rockefeller University, New York, N.Y.


PAUL GILMAN, Executive Director

The National Academy of Sciences is a private, nonprofit, self-perpetuating society of distinguished scholars engaged in scientific and engineering research, dedicated to the furtherance of science and technology and to their use for the general welfare. Upon the authority of the charter granted to it by the Congress in 1863, the Academy has a mandate that requires it to advise the federal government on scientific and technical matters. Dr. Frank Press is president of the National Academy of Sciences.

The National Academy of Engineering was established in 1964, under the charter of the National Academy of Sciences, as a parallel organization of outstanding engineers. It is autonomous in its administration and in the selection of its members, sharing with the National Academy of Sciences the responsibility for advising the federal government. The National Academy of Engineering also sponsors engineering programs aimed at meeting national needs, encourages education and research, and recognizes the superior achievements of engineers. Dr. Robert M. White is president of the National Academy of Engineering.

The Institute of Medicine was established in 1970 by the National Academy of Sciences to secure the services of eminent members of appropriate professions in the examination of policy matters pertaining to the health of the public. The Institute acts under the responsibility given to the National Academy of Sciences by its congressional charter to be an adviser to the federal government and, upon its own initiative, to identify issues of medical care, research, and education. Dr. Kenneth I. Shine is president of the Institute of Medicine.

The National Research Council was organized by the National Academy of Sciences in 1916 to associate the broad community of science and technology with the Academy's purposes of furthering knowledge and advising the federal government. Functioning in accordance with general policies determined by the Academy, the Council has become the principal operating agency of both the National Academy of Sciences and the National Academy of Engineering in providing services to the government, the public, and the scientific and engineering communities. The Council is administered jointly by both Academies and the Institute of Medicine. Dr. Frank Press and Dr. Robert M. White are chairman and vice-chairman, respectively, of the National Research Council.
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IN 1988, THE U.S. CONGRESS requested that the National Academy of Sciences establish a committee within the National Research Council to study scientific and policy issues concerning pesticides in the diets of infants and children. The Committee on Pesticide Residues in the Diets of Infants and Children appointed to undertake this study was charged with responsibility for examining what is known about exposures to pesticide residues in the diets of infants and children, the adequacy of current risk assessment methods and policies, and toxicological issues of greatest concern. The committee operated under the joint aegis of the Board on Agriculture (BA) and the Board on Environmental Studies and Toxicology (BEST).

The committee first met in October 1988 and held its last meeting in January 1993. Several full committee meetings were held each year, and subgroups of the committee were convened on a number of occasions to address such topics as the physiology of infants and children, the age-specific patterns of children's diets, the measurement of residue levels, and the mathematical modeling of risks. The expertise represented on the committee included pediatrics, toxicology, epidemiology, biostatistics, food science and nutrition, analytical chemistry, and child growth and development. When required, advice was obtained from experts outside the committee on a variety of topics.

Critical assessment of potential risks to health resulting from exposures to toxicants in the environment has been the focus of several recent studies conducted by BEST and BA. Many of the approaches to risk assessment used in this report trace their origins to the reports on Drinking Water and Health developed since 1977. Of particular value was Volume 6 in that series. The committee also found useful Risk Assessment in the Federal Government: Managing the Process (1983), Biologic Markers in Reproductive Toxicology (1989), Biologic Markers in Immunotoxicology (1992), and Environmental Neurotoxicology (1992). The analysis in this volume draws conceptually from the 1987 report from the Board on Agriculture called Regulating Pesticides in Food: The Delaney Paradox—an examination of the process by which levels of pesticide residues in foods are regulated by the U.S. Government.

The Committee on Pesticides in the Diets of Infants and Children was greatly assisted by many individuals and groups who provided information on food consumption patterns and on pesticide residue concentrations in the U.S. diet. The groups include the U.S. Department of Agriculture, the U.S. Food and Drug Administration, the National Food Processors Association, the Gerber Products Company, and the Infant Formula Council. Many other food manufactures as well as pesticide manufacturers also provided useful data to the committee either individually or through various organizations.

The committee is grateful for the assistance of the National Research Council (NRC) staff in the preparation of this report. In particular the committee wishes to acknowledge Frances Peter, project manager; Richard Thomas, principal staff scientist (BEST); Sandi Fitzpatrick, senior program assistant (BEST); James Reisa, director of BEST; and Susan Offutt, executive director of BA. Other staff members who contributed to this effort include Shelley A. Nurse, senior project assistant (BEST); Ruth P. Danoff, project assistant (BEST); Craig Cox, senior staff officer (BA); Mary Lou Sutton, administrative assistant (BA); Carla Carlson, director of communications (BA); Barbara J. Rice, editor (BA); Janet Overton, associate editor (BA); Lee. R. Paulson, program director for information systems and statistics (BEST); Bernidean Williams, information specialist (BEST); and Dawn M. Eichenlaub, production manager, and Richard E. Morris, editor, National Academy Press. Thanks are also due to Richard Wiles and Charles Benbrook, formerly of the BA staff. The interest in this report shown by the Executive Office of the National Research Council, especially by the Deputy Executive Officer Mitchel Wallerstein, is greatly appreciated. These individuals provided invaluable support to the committee throughout its deliberations.

As consultant to the committee, John Wargo of the Yale University School of Forestry and Environmental Studies developed numerous innovative approaches to the analysis of highly complex data. His pellucid presentations permitted clear understanding of issues that previously had been opaque. Valuable assistance was also provided to the committee by Emmanuel Akpanyie, Sheryl Bartlett, and Judy Hauswirth, who served as technical advisers, and Dr. Marcia VanGemert, the EPA project officer.

Last, but by no means least, the work of all the members of the committee is greatly appreciated. We are also grateful to the U. S. Environmental Protection Agency, Health and Welfare Canada, the International Life Sciences Institute, and the Kellogg Endowment Fund of the National Academy of Sciences and the Institute of Medicine, whose financial support made the study possible.

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Executive Summary

PESTICIDES ARE USED WIDELY in agriculture in the United States. Their application has improved crop yields and has increased the quantity of fresh fruits and vegetables in the diet, thereby contributing to improvements in public health.

But pesticides may also cause harm. Some can damage the environment and accumulate in ecosystems. And depending on dose, some pesticides can cause a range of adverse effects on human health, including cancer, acute and chronic injury to the nervous system, lung damage, reproductive dysfunction, and possibly dysfunction of the endocrine and immune systems.

Diet is an important source of exposure to pesticides. The trace quantities of pesticides that are present on or in foodstuffs are termed residues. To minimize exposure of the general population to pesticide residues in food, the U.S. Government has instituted regulatory controls on pesticide use. These are intended to limit exposures to residues while ensuring an abundant and nutritious food supply. The legislative framework for these controls was established by the Congress through the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) and the Federal Food, Drug, and Cosmetic Act (FFDCA). Pesticides are defined broadly in this context to include insecticides, herbicides, and fungicides.

Tolerances constitute the single, most important mechanism by which EPA limits levels of pesticide residues in foods. A tolerance is defined as the legal limit of a pesticide residue allowed in or on a raw agricultural commodity and, in appropriate cases, on processed foods. A tolerance must be established for any pesticide used on any food crop.

Tolerance concentrations are based primarily on the results of field trials conducted by pesticide manufacturers and are designed to reflect the highest residue concentrations likely under normal conditions of agricultural use. Their principal purpose is to ensure compliance with good agricultural practice. Tolerances are not based primarily on health considerations.

This report addresses the question of whether current regulatory approaches for controlling pesticide residues in foods adequately protect infants and children. The exposure of infants and children and their susceptibility to harm from ingesting pesticide residues may differ from that of adults. The current regulatory system does not, however, specifically consider infants and children. It does not examine the wide range of pesticide exposure patterns that appear to exist within the U.S. population. It looks only at the average exposure of the entire population. As a consequence, variations in dietary exposure to pesticides and health risks related to age and to such other factors as geographic region and ethnicity are not addressed in current regulatory practice.

Concern about the potential vulnerability of infants and children to dietary pesticides led to U.S. Congress in 1988 to request that the National Academy of Sciences (NAS) appoint a committee to study this issue through its National Research Council (NRC). In response, the NRC appointed a Committee on Pesticide Residues in the Diets of Infants and Children under the joint aegis of the Board on Agriculture and the Board on Environmental Studies and Toxicology.

The committee was charged with responsibility for examining scientific and policy issues faced by government agencies, particularly EPA, in regulating pesticide residues in foods consumed by infants and children. Specifically, the committee was asked to examine the adequacy of current risk assessment policies and methods; to assess information on the dietary intakes of infants and children; to evaluate data on pesticide residues in the food supply; to identify toxicological issues of greatest concern; and to develop relevant research priorities. Expertise represented on the committee included toxicology, epidemiology, biostatistics, food science and nutrition, analytical chemistry, child growth and development, and pediatrics.

The committee was not asked to consider toxicities resulting from exposures to microorganisms (bacteria and viruses) or from other naturally occurring potential toxins. It was not asked to weigh the benefits and risks to be derived from a plentiful and varied food supply against the potential risks resulting from pesticide exposure. It was not asked to assess the overall safety of the food supply.

In this report, the committee considered the development of children from the beginning of the last trimester of pregnancy (26 weeks) through 18 years of age, the point when all biological systems have essentially matured.


Age-Related Variation in Susceptibility and Toxicity

A fundamental maxim of pediatric medicine is that children are not ''little adults." Profound differences exist between children and adults. Infants and children are growing and developing. Their metabolic rates are more rapid than those of adults. There are differences in their ability to activate, detoxify, and excrete xenobiotic compounds. All these differences can affect the toxicity of pesticides in infants and children, and for these reasons the toxicity of pesticides is frequently different in children and adults. Children may be more sensitive or less sensitive than adults, depending on the pesticide to which they are exposed. Moreover, because these processes can change rapidly and can counteract one another, there is no simple way to predict the kinetics and sensitivity to chemical compounds in infants and children from data derived entirely from adult humans or from toxicity testing in adult or adolescent animals.

The committee found both quantitative and occasionally qualitative differences in toxicity of pesticides between children and adults . Qualitative differences in toxicity are the consequence of exposures during special windows of vulnerability—brief periods early in development when exposure to a toxicant can permanently alter the structure or function of an organ system. Classic examples include chloramphenicol exposure of newborns and vascular collapse (gray baby syndrome), tetracycline and dysplasia of the dental enamel, and lead and altered neurologic development.

Quantitative differences in pesticide toxicity between children and adults are due in part to age-related differences in absorption, metabolism, detoxification, and excretion of xenobiotic compounds, that is, to differences in both pharmacokinetic and pharmacodynamic processes. Differences in size, immaturity of biochemical and physiological functions in major body systems, and variation in body composition (water, fat, protein, and mineral content) all can influence the extent of toxicity. Because newborns are the group most different anatomically and physiologically from adults, they may exhibit the most pronounced quantitative differences in sensitivity to pesticides. The committee found that quantitative differences in toxicity between children and adults are usually less than a factor of approximately 10-fold.

The committee concluded that the mechanism of action of a toxicant—how it causes harm—is generally similar in most species and across age and developmental stages within species. For example, if a substance is cytotoxic in adults, it is usually also cytotoxic in immature individuals.

Lack of data on pesticide toxicity in developing organisms was a recurrent problem encountered by the committee. In particular, little work has been done to identify effects that develop after a long latent period or to investigate the effects of pesticide exposure on neurotoxic, immunotoxic, or endocrine responses in infants and children. The committee therefore had to rely mostly on incomplete information derived from studies in mature animals and on chemicals other than pesticides.

The committee reviewed current EPA requirements for toxicity testing by pesticide manufacturers, as well as testing modifications proposed by the agency. In general, the committee found that current and past studies conducted by pesticide manufacturers are designed primarily to assess pesticide toxicity in sexually mature animals. Only a minority of testing protocols have supported extrapolation to infant and adolescent animals. Current testing protocols do not, for the most part, adequately address the toxicity and metabolism of pesticides in neonates and adolescent animals or the effects of exposure during early developmental stages and their sequelae in later life.

Age-Related Differences in Exposure

Estimation of the exposures of infants and children to pesticide residues requires information on (1) dietary composition and (2) residue concentrations in and on the food and water consumed. The committee found that infants and children differ both qualitatively and quantitatively from adults in their exposure to pesticide residues in foods. Children consume more calories of food per unit of body weight than do adults. But at the same time, infants and children consume far fewer types of foods than do adults. Thus, infants and young children may consume much more of certain foods, especially processed foods, than do adults. And water consumption, both as drinking water and as a food component, is very different between children and adults.

The committee concluded that differences in diet and thus in dietary exposure to pesticide residues account for most of the differences in pesticide-related health risks that were found to exist between children and adults. Differences in exposure were generally a more important source of differences in risk than were age-related differences in toxicologic vulnerability.

Data from various food consumption surveys were made available to the committee. In analyzing these data, the committee found it necessary to create its own computer programs to convert foods as consumed into their component raw agricultural commodities (RACs). This analytic approach facilitated the use of data from different sources and permitted evaluation of total exposure to pesticides in different food commodities. For processed foods, the committee noted that effects of processing on residue concentrations should be considered, but that information on these effects is quite limited. Processing may decrease or increase pesticide residue concentrations. The limited data available suggest that pesticide residues are generally reduced by processing; however, more research is needed to define the direction and magnitude of the changes for specific pesticide-food combinations. The effect of processing is an important consideration in assessing the dietary exposures of infants and young children, who consume large quantities of processed foods, such as fruit juices, baby food, milk, and infant formula.

Although there are several sources of data on pesticide residues in the United States, the data are of variable quality, and there are wide variations in sample selection, reflecting criteria developed for different sampling purposes, and in analytical procedures, reflecting different laboratory capabilities and different levels of quantification between and within laboratories. These differences reflect variations in precision and in the accuracy of methods used and the different approaches to analytical issues, such as variations in limit of quantification. There also are substantial differences in data reporting. These differences are due in part to different record-keeping requirements, such as whether to identify samples with multiple residues, and differences in statistical treatment of laboratory results below the limit of quantification.

Both government and industry data on residue concentrations in foods reflect the current regulatory emphasis on average adult consumption patterns. The committee found that foods eaten by infants and children are underrepresented in surveys of commodity residues. Many of the available residue data were generated for targeted compliance purposes by the Food and Drug Administration (FDA) to find residue concentrations exceeding the legal tolerances established by the EPA under FFDCA.

Survey data on consumption of particular foods are conventionally grouped by broad age categories. The average consumption of a hypothetical "normal" person is then used to represent the age group. However, in relying solely on the average as a measure of consumption, important information on the distribution of consumption patterns is lost. For example, the high levels of consumption within a particular age group are especially relevant when considering foods that might contain residues capable of causing acute toxic effects. Also, geographic, ethnic, and other differences may be overlooked.

To overcome the problems inherent in the current reliance on "average" exposures, the committee used the technique of statistical convolution (i.e., combining various data bases) to merge distributions of food consumption with distributions of residue concentrations. This approach permits examination of the full range of pesticide exposures in the U.S. pediatric population. As is described in the next section, this approach provides an improved basis over the approach now used for assessing risks for infants and children.

A New Approach to Risk Assessment for Infants and Children

To properly characterize risk to infants and children from pesticide residues in the diet, information is required on (1) food consumption patterns of infants and children, (2) concentrations of pesticide residues in foods consumed by infants and children, and (3) toxic effects of pesticides, especially effects that may be unique to infants and children. If suitable data on these three items are available, risk assessment methods based on the technique of statistical convolution can be used to estimate the likelihood that infants and children who experience specific exposure patterns may be at risk. To characterize potential risks to infants and children in this fashion, the committee utilized data on distributions of pesticide exposure that, in turn, were based on distributions of food consumption merged with data on the distribution of pesticide residue concentrations. The committee found that age-related differences in exposure patterns for 1– to 5-year-old children were most accurately illuminated by using 1-year age groupings of data on children's food consumption.

Exposure estimates should be constructed differently depending on whether acute or chronic effects are of concern. Average daily ingestion of pesticide residues is an appropriate measure of exposure for assessing the risk of chronic toxicity. However, actual individual daily ingestion is more appropriate for assessing acute toxicity. Because chronic toxicity is often related to long-term average exposure, the average daily dietary exposure to pesticide residues may be used as the basis for risk assessment when the potential for delayed, irreversible chronic toxic effects exists. Because acute toxicity is more often mediated by peak exposures occurring within a short period (e.g., over the course of a day or even during a single eating occasion), individual daily intakes are of interest. Examining the distribution of individual daily intakes within the population of interest reflects day-to-day variation in pesticide ingestion both for specific individuals and among individuals.

Children may be exposed to multiple pesticides with a common toxic effect, and estimates of exposure and of risk could therefore be improved by accounting for these simultaneous exposures. This can be accomplished by assigning toxicity equivalence factors to each of the compounds having a common mechanism of action. Total residue exposure is then estimated by multiplying the actual level of each pesticide residue by its toxicity equivalence factor and summing the results. This information may be combined with data on consumption to construct a distribution of total exposure to all pesticides having a common mechanism of action. To test this multiple-residue methodology, the committee estimated children's acute health risks resulting from combined exposure to five members of the organophosphate insecticide family. This was accomplished by combining actual food consumption data with data on actual pesticide residue levels.

Through this new analytical procedure, the committee estimated that for some children, total organophosphate exposures may exceed the reference dose. Furthermore, although the data were weak, the committee estimated that for some children exposures could be sufficiently high to produce symptoms of acute organophosphate pesticide poisoning.

Compared to late-in-life exposures, exposures to pesticides early in life can lead to a greater risk of chronic effects that are expressed only after long latency periods have elapsed. Such effects include cancer, neurodevelopmental impairment, and immune dysfunction. The committee developed new risk assessment methods to examine this issue.

Although some risk assessment methods take into account changes in exposure with age, these models are not universally applied in practice. The committee explored the use of newer risk assessment methods that allow for changes in exposure and susceptibility with age. However, the committee found that sufficient data are not currently available to permit wide application of these methods.


On the basis of its findings, the committee recommends that certain changes be made in current regulatory practice. Most importantly, estimates of expected total exposure to pesticide residues should reflect the unique characteristics of the diets of infants and children and should account also for all nondietary intake of pesticides. Estimates of exposure should take into account the fact that not all crops are treated with pesticides that can be legally applied to those crops, and they should consider the effects of food processing and storage. Exposure estimates should recognize that pesticide residues may be present on more than one food commodity consumed by infants and children and that more than one pesticide may be present on one food sample. Lastly, determinations of safe levels of exposure should take into consideration the physiological factors that can place infants and children at greater risk of harm than adults.

• Tolerances. Tolerances for pesticide residues on commodities are currently established by the EPA under FIFRA and FFDCA. A tolerance concentration is defined under FFDCA as the maximum quantity of a pesticide residue allowable on a raw agricultural commodity (RAC) (FFDCA, Section 408) and in processed food when the pesticide concentrates during processing (FFDCA Section 409). Tolerance concentrations on RACs are based on the results of field trials conducted by pesticide manufacturers and are designed to reflect the highest residue concentrations likely under normal agricultural practice. More than 8,500 food tolerances for pesticides are currently listed in the Code of Federal Regulations (CFR). Approximately 8,350 of these tolerances are for residues on raw commodities (promulgated under section 408) and about 150 are for residues known to concentrate in processed foods (promulgated under section 409).

The determination of what might be a safe level of residue exposure is made by considering the results of toxicological studies of the pesticide's effects on animals and, when data are available, on humans. Both acute and chronic effects, including cancer, are considered, although acute effects are treated separately. These data are used to establish human exposure guidelines (i.e., a reference dose, RfD) against which one can compare the expected exposure. Exposure is a function of the amount and kind of foods consumed and the amount and identity of the residues in the foods (i.e., Theoretical Maximum Residue Contributions, TMRCs). If the TMRCs exceed the RfD, then anticipated residues are calculated for comparison with the proposed tolerance. The percent of crop acreage treated is also considered. If the anticipated residues exceed the RfD, then the proposed tolerance is rejected, and the manufacturer may recommend a new tolerance level.

Although tolerances establish enforceable legal limits for pesticide residues in food, they are not based primarily on health considerations, and they do not provide a good basis for inference about actual exposures of infants and children to pesticide residues in or on foods.

Tolerances constitute the only tool that EPA has under the law for controlling pesticide residues in food. To ensure that infants and children are not exposed to unsafe levels of pesticide residues, the committee recommends that EPA modify its decision-making process for setting tolerances so that it is based more on health considerations than on agricultural practices. These changes should incorporate the use of improved estimates of exposure and more relevant toxicology, along with continued consideration of the requirements of agricultural production. As a result, human health considerations would be more fully reflected in tolerance levels. Children should be able to eat a healthful diet containing legal residues without encroaching on safety margins. This goal should be kept clear.

• Toxicity testing. The committee believes it is essential to develop toxicity testing procedures that specifically evaluate the vulnerability of infants and children. Testing must be performed during the developmental period in appropriate animal models, and the adverse effects that may become evident must be monitored over a lifetime. Of particular importance are tests for neurotoxicity and toxicity to the developing immune and reproductive systems. Extrapolation of toxicity data from adult and adolescent laboratory animals to young humans may be inaccurate. Careful attention to interspecies differences in pharmacokinetics and metabolism of pesticides and the relative ages at which organ systems mature is essential. It is also important to enhance understanding of developmental toxicity, especially in humans, during critical periods of postnatal development, including infancy and puberty.

• Uncertainty factors. For toxic effects other than cancer or heritable mutation, uncertainty factors are widely used to establish guidelines for human exposure on the basis of animal testing results. This is often done by dividing the no-observed-effect level (NOEL) found in animal tests by an uncertainty factor of 100-fold. This factor comprises two separate factors of 10-fold each: one allows for uncertainty in extrapolating data from animals to humans; the other accommodates variation within the human population. Although the committee believes that the latter uncertainty factor generally provides adequate protection for infants and children, this population subgroup may be uniquely susceptible to chemical exposures at particularly sensitive stages of development.

At present, to provide added protection during early development, a third uncertainty factor of 10 is applied to the NOEL to develop the RfD. This third 10-fold factor has been applied by the EPA and FDA whenever toxicity studies and metabolic/disposition studies have shown fetal developmental effects.

Because there exist specific periods of vulnerability during postnatal development, the committee recommends that an uncertainty factor up to the 10-fold factor traditionally used by EPA and FDA for fetal developmental toxicity should also be considered when there is evidence of postnatal developmental toxicity and when data from toxicity testing relative to children are incomplete. The committee wishes to emphasize that this is not a new, additional uncertainty factor but, rather, an extended application of a uncertainty factor now routinely used by the agencies for a narrower purpose.

In the absence of data to the contrary, there should be a presumption of greater toxicity to infants and children. To validate this presumption, the sensitivity of mature and immature individuals should be studied systematically to expand the current limited data base on relative sensitivity.

• Food consumption data. The committee recommends that additional data on the food consumption patterns of infants and children be collected within narrow age groups. The available data indicate that infants and children consume much more of certain foods on a body weight basis than do adults. Because higher exposures can lead to higher risks, it is important to have accurate data on food consumption patterns for infants and children. At present, data are derived from relatively small samples and broad age groupings, making it difficult to draw conclusions about the food consumption patterns of infants and children. Because the composition of a child's diet changes dramatically from birth through childhood and adolescence to maturity, "market basket" food consumption surveys should include adequate samples of food consumption by children at 1-year intervals up to age 5, by children between the ages of 5 and 10 years, and by children between 11 and 18 years. Food consumption surveys should be conducted periodically to ascertain changes in consumption patterns over time.

• Pesticide residue data. To maximize the utility of pesticide residue data collected by various laboratories, the committee recommends the use of comparable analytical methods and standardized reporting procedures and the establishment of a computerized data base to collate data on pesticide residues generated by different laboratories. Reports on pesticide residue testing should describe the food commodity analyzed (whether processed or raw), the analytical methods used, the compounds for which tests were conducted, quality assurance and control procedures, and the limit of quantification of the tests. All findings should be reported, whether or not the residue sought is found.

• In its surveillance of pesticide residues, FDA should increase the frequency of sampling of the commodities most likely to be consumed by infants and children. The residue testing program should include all toxic forms of the pesticide, for example, its metabolites and degradation products.

• Food residue monitoring should target a special "market basket" survey focused toward the diets of infants and children.

• Pesticide field trials currently conducted by pesticide manufacturers in support of registration provide data on variation in residue concentrations associated with different rates and methods of application. Such data should be consulted to provide a basis for estimating potential maximum residue levels.

• More complete information is needed on the effects of food processing on levels of pesticides—both the parent compound and its metabolites—in specific food-chemical combinations potentially present in the diets of infants and children.

• Risk assessment. All exposures to pesticides—dietary and nondietary—need to be considered when evaluating the potential risks to infants and children. Nondietary environmental sources of exposure include air, dirt, indoor surfaces, lawns, and pets.

• Estimates of total dietary exposure should be refined to consider intake of multiple pesticides with a common toxic effect. Converting residues for each pesticide with a common mechanism of action to toxicity equivalence factors for one of the compounds would provide one approach to estimating total residue levels in toxicologically equivalent units.

• Consumption of pesticide residues in water is an important potential route of exposure. Risk assessment should include estimates of exposure to pesticides in drinking water and in water as a component of processed foods.

Given adequate data on food consumption and residues, the committee recommends the use of exposure distributions rather than single point data to characterize the likelihood of exposure to different concentrations of pesticide residues. The distribution of average daily exposure of individuals in the population of interest is most relevant for use in chronic toxicity risk assessment, and the distribution of individual daily intakes is recommended for evaluating acute toxicity. Ultimately, the collection of suitable data on the distribution of exposures to pesticides will permit an assessment of the proportion of the population that may be at risk.

Although the committee considers the use of exposure distributions to be more informative than point estimates of typical exposures, the data available to the committee did not always permit the distribution of exposures to be well characterized. Existing food consumption surveys generally involve relatively small numbers of infants and children, and food consumption data are collected for only a few days for each individual surveyed. Depending on the purpose for which they were originally collected, residue data may not reflect the actual distribution of pesticide residues in the food supply. Since residue data are not developed and reported in a consistent fashion, it is generally not possible to pool data sets derived from different surveys. Consequently, the committee recommends that guidelines be developed for consumption and residue data permitting characterization of distributions of dietary exposure to pesticides.

The committee identified important differences in susceptibility to the toxic effects of pesticides and exposure to pesticides in the diet with age. For carcinogenic effects, the committee proposed new methods of cancer risk assessment designed to take such differences into account. Preliminary analyses conducted by the committee suggest that consideration of such differences can lead to lifetime estimates of cancer risk that can be higher or lower than estimates derived with methods based on constant exposure. However, underestimation of risk assuming constant exposure was limited to a factor of about 3- to 5-fold in all cases considered by the committee. Because these results are based on limited data and specific assumptions about the mechanisms by which carcinogenic effects are induced, the applicability of these conclusions under other conditions should be established.

Currently, most long-term laboratory studies of carcinogenesis and other chronic end points are based on protocols in which the level of exposure is held constant during the course of the study. To facilitate the application of risk assessment methods that allow for changes in exposure and susceptibility with age, it would be desirable to develop bioassay protocols that provide direct information on the relative contribution of exposures at different ages to lifetime risks. Although the committee does consider it necessary to develop special bioassay protocols for mandatory application in the regulation of pesticides, it would be useful to design special studies to provide information on the relative effects of exposures at different ages on lifetime cancer and other risks with selected chemical carcinogens.

In addition to pharmacodynamic models for cancer risk assessment, the committee recommends the development and application of physiologically based pharmacokinetic models that describe the unique features of infants and children. For example, differences in relative organ weights with age can be easily described in physiologic pharmacokinetic models; special compartments for the developing fetus may also be incorporated. Physiologically based pharmacokinetic models can be used to predict the dose of the proximate toxicant reaching target tissues, and may lead to more accurate estimates of risk.

In summary, better data on dietary exposure to pesticide residues should be combined with improved information on the potentially harmful effects of pesticides on infants and children. Risk assessment methods that enhance the ability to estimate the magnitude of these effects should be developed, along with appropriate toxicological tests for perinatal and childhood toxicity. The committee's recommendations support the need to improve methods for estimating exposure and for setting tolerances to safeguard the health of infants and children.
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1. Background and Approach to the Study

PESTICIDES ARE USED widely in agriculture in the United States. When effectively applied, pesticides can kill or control pests, including weeds, insects, fungi, bacteria, and rodents. Chemical pest control has contributed to dramatic increases in yields for most major fruit and vegetable crops. Its use has led to substantial improvements over the past 40 years in the quantity and variety of the U.S. diet and thus in the health of the public (see, for example, Block et al., 1992).

On the negative side, many pesticides are harmful to the environment and are known or suspected to be toxic to humans. They can produce a wide range of adverse effects on human health that include acute neurologic toxicity, chronic neurodevelopmental impairment, cancer, reproductive dysfunction, and possibly dysfunction of the immune and endocrine systems.

The diet is an important source of exposure to pesticides. The trace quantities of pesticides and their breakdown products that are present on or in foodstuffs are termed residues. Residue levels reflect the amount of pesticide applied to a crop, the time that has elapsed since application, and the rate of pesticide dissipation and evaporation. Pesticide residues are widespread in the U.S. diet. They are consumed regularly by most Americans, including infants and children.

To protect the U.S. public against dietary pesticides and their potentially harmful effects, the U.S. Congress has enacted legislation to regulate residue exposures and to ensure that the food supply is safe as well abundant and nutritious. The two principal components of the legislative framework—the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) and the Federal Food, Drug, and Cosmetic Act (FFDCA)—have provided the foundation for a comprehensive regulatory system.

Concern has arisen in recent years that the current pesticide regulatory system, which is intended to minimize health risk to the general population, may not adequately protect the health of infants and children. The traditional system assesses dietary pesticide risk on the basis of the average exposure of the entire U.S. population. However, it does not consider the range of exposures that exists within the population, nor does it specifically consider exposures of infants and children. The exposure of infants and children and their susceptibility to harm from ingesting pesticide residues may differ considerably from that of adults.

Concern about this uncertainty led the U.S. Congress in 1988 to request that the National Academy of Sciences (NAS) appoint a committee to study scientific and policy issues concerning pesticides in the diets of infants and children through its National Research Council (NRC). The committee was specifically charged with examining.

• what is known about exposures to pesticide residues in the diets of infants and children;

• the adequacy of current risk assessment methods and policies; and

• toxicological issues of greatest concern and in greatest need of further research.


A pesticide is defined under FIFRA as ''any substance or mixture of substances intended for preventing, destroying, repelling, or mitigating any insects, rodents, nematodes, fungi, or weeds, or any other forms of life declared to be pests, and any substance or mixture of substances intended for use as a plant regulator, defoliant, or desiccant."

Pesticides have been used by humankind for centuries. Their use was recorded as early as the eighth century BC when the application of fungicides was documented in Homeric poems (Mason, 1928; McCallan, 1967). From the until the present, numerous mixtures have been developed to control fungi, insects, weeds, and other pests.

In the 19th century, sulfur compounds were developed as fungicides, and arsenicals were used to control insects attacking fruits and vegetables. Those compounds were highly toxic and consequently were replaced by chlorinated organic pesticides such as DDT and benzenehexachloride (BHC), which were developed during the 1930s and became widely used in the 1950s and 1960s. Chlorinated hydrocarbon insecticides such as DDT, BHC, dieldrin, aldrin, and toxaphene were enthusiastically adopted by farmers who hoped to control previously uncontrolled insects with what were believed to be relatively safe compounds with long environmental persistence. These chemicals were also used widely in the control of malaria and other insectborne diseases. By 1955, more than 90% of all pest control chemicals used in U.S. agriculture were synthetic organic compounds, and in 1961 DDT was registered for use on 334 crops. Phenoxy herbicides such as 2,4-dichlorophenoxyacetic acid (2,4-D), 2,4,5-trichlorophenoxyacetic acid (2,4,5-T), and ethylenebisdithiocarbamates (EBDCs) and dicarboximide fungicides also gained widespread use during that time.

Beginning in the late 1960s, the potential of the chlorinated hydrocarbons for bioaccumulation and long-term toxicity became widely recognized. Also, pest resistance to chlorinated pesticides became increasingly evident and problematic throughout the 1960s, leading many farmers to substitute organophosphates and carbamates for DDT and other chlorinated compounds. Public pressure to end the use of chlorinated pesticides contributed to the creation of the Environmental Protection Agency (EPA) in 1970 and the ultimate administrative revocation in 1972 of the use of DDT on all food sources in the United States. By the end of the 1980s, most food uses of chlorinated compounds were discontinued in this country, although heavy application continues in other nations.

Since the late 1960s, a decline has occurred in insecticide use on major commodities such as corn, soybeans, cotton, and wheat. This decrease was primarily the result of pest management programs, which led to an approximately 50% reduction in pesticide application to cotton crops nationwide. Another important factor was the development and widespread adoption of synthetic pyrethroid compounds, which are applied in gram quantities rather than pounds per acre. During this period, fungicide use on peanuts and wheat declined, but because of the continued application of fungicides to fruits and vegetables and the increasing acreage of those crops under cultivation, the overall volume of fungicides used has remained steady.

In contrast, the use of herbicides has increased dramatically. In 1955 approximately 3% of all acreage planted with corn and soybean crops were treated with a herbicide; by 1985 that figure had increased to more than 95%, primarily because of the development of effective herbicides that were applied before the crop was planted. Herbicides now account for approximately 66% of all agricultural pesticides, but for a lower percentage of dietary exposure than is attributed to fungicides and insecticides, which are applied directly to the food closer to, or even after, its harvest. More than 90% of all herbicides are applied to just four crops: corn, soybeans, cotton, and wheat.

Today, most pesticides are synthetically produced organic and inorganic chemicals or microbial agents. Some of these pesticides have been found naturally and have been synthetically reproduced for commercial use. The variety and amounts of pesticides now used are far greater than at any previous time in human history. Approximately 600 pesticides are currently registered with the EPA (P. Fenner-Crisp, EPA, personal commun., 1993).

The most common food-use pesticides fall into three classes: insecticides, herbicides, and fungicides. In 1991, an estimated 817 million pounds of active pesticide ingredients were used for agricultural application in the United States. Of this total, herbicides accounted for 495 million pounds; insecticides, 175 million pounds, fungicides, 75 million pounds; and other pesticides, 72 million pounds; (EPA, 1992). "Other" pesticides were defined as rodenticides, fumigants, and molluscicides but do not include wood preservatives, disinfectants, and sulfur.

• Insecticides. Insecticides control insects that damage crops through a variety of modes. Some work as nerve poisons, muscle poisons, desiccants, sterilants, or pheromones; others exert their effects by physical means such as by clogging air passages. The classes of insecticides most commonly used today are chlorinated hydrocarbons, organophosphates, and carbamates, and of these, the organophosphates are the most widely used. Typically they are very acutely toxic, but they do not persist in the environment. Well-known organophosphate pesticides include parathion, dichlorvos, malathion, chlorpyrifos, and azinphos-methyl. The toxicity to humans resulting from exposure to these compounds can differ markedly from chemical to chemical.

The carbamate insecticides are also very widely used in the United States today. They too are highly toxic, e.g., aldicarb. Other insecticides such as synthetic pyrethroids, e.g., permethrin, are valued because of their fast action and relatively low toxicity to mammals.

• Herbicides. Herbicides are used to control weeds, which compete with crop plants for water, nutrients, space and sunlight. By reducing the weed population, the need for farm labor is decreased and crop quality is enhanced. Herbicides work through a variety of modes of action. Some damage leaf cells and desiccate the plant; others alter nutrient uptake or photosynthesis. Some herbicides inhibit seed germination or seedling growth. Others are applied to foliage and kill on contact, thereby destroying leaf and stem tissues. Some of the most widely used herbicides are 2,4-D [(2,4-dichlorophenoxy) acetic acid], atrazine, simazine, dacthal, alachlor, metolachlor, and glyphosate.

• Fungicides. Fungicides control plant molds and other diseases. They include compounds of metals and sulfur as well as numerous synthetics. Some fungicides act by inhibiting the metabolic processes of fungal organisms and can be used on plants that have already been invaded and damaged by the organism. Other fungicides protect plants from fungal infections and retard fungal growth before damage to plants can occur. Fungicides frequently provide direct benefit to humans by retarding or eliminating fungal infections that can produce toxicants such as aflatoxins. Fungicides that have been used heavily over the years include benomyl, captan, and the EBDC family of fungicides such as mancozeb.

In addition to their agricultural applications, pesticides are also used for many nonagricultural purposes, e.g., in homes and public buildings to kill termites and other pests; on lawns and ornamental plantings to kill weeds, insects, and fungi; and on ponds, lakes, and rivers to control insects and weeds. Therefore, humans are exposed to pesticides from a variety of sources other than the diet, for example, through the skin or by inhalation. Some of these exposures are especially important when considering total exposures of infants and children.


The societal response to the dual nature of pesticides—to their combination of benefits and toxicity—has been to develop a comprehensive regulatory system that seeks to make possible the beneficial use of pesticides while minimizing their hazards to public health and the environment. This regulatory system originated with the enactment of FIFRA in 1947. The legislation regulating pesticides in the United States now consists of FIFRA, its comprehensive amendments of 1972, 1975, 1978, 1980, and 1988, and certain provisions of the FFDCA, which was enacted in 1954 and later amended.

FIFRA is intended by Congress to be a "balancing" or risk-benefit statute. It states that a pesticide when used for its intended purpose must not cause "unreasonable adverse effects on the environment." This balancing process must take into account "the economic, social, and environmental costs as well as the potential benefits of the use of any pesticide" [7 USC 136(a) (1978)]. Wilkinson (1990, p. 11) has commented: ''While use of the term 'unreasonable risk' implies that some risks will be tolerated under FIFRA, it is clearly expected that the anticipated benefits will outweigh the potential risks when the pesticide is used according to commonly recognized, good agricultural practices."

Under FIFRA, pesticide use is controlled through a registration process. This process is administered by EPA. A given pesticide may have several different uses, and each use is required to have its own registration. EPA registration of a pesticide use and approval of a label detailing the legally binding instructions for that use are required before a pesticide can be legally sold.

For a pesticide to be registered, manufacturers must submit to EPA the data needed to support the product's registration, including substantiation of its usefulness and disclosure of its chemical and toxic properties, its likely distribution in the environment, and its possible effects on wildlife and plants.

Pesticides that are to be registered for use on food crops must be granted a tolerance by EPA. These tolerances constitute the principal mechanism by which EPA limits levels of pesticides residues in foods. A tolerance concentration is defined under FFDCA as the maximum quantity of a pesticide residue allowable on a raw agricultural commodity (RAC) (FFDCA, Section 408) and in processed food when the pesticide has concentrated during processing (FFDCA, Section 409). A tolerance must be defined for any pesticide used on food crops. Tolerance concentrations on RACs are based on the result of field trials conducted by pesticide manufacturers and are designed to reflect the highest residue concentrations likely under normal agricultural practice. Thus, tolerances are based on good agricultural practice rather than on considerations of human health.

The determination of what might be a safe level of residue exposure is made by considering the results of toxicological studies of the pesticide's effects on animals and, when data are available, on humans. Both acute and chronic effects, including cancer, are considered, although currently, acute effects are treated separately. These data are used to establish human exposure guidelines (i.e., reference dose, RfD) against which one can compare the expected exposure. Exposure is a function of the amount and kind of foods consumed and the amount and identity of residues in the foods (i.e., Theoretical Maximum Residue Contributions, TMRCs). If the TMRCs exceed the RfD, then anticipated residues are calculated and compared with the proposed tolerance. The percent of crop acreage treated is also considered. If the anticipated residues exceed the RfD, then the proposed tolerance is rejected, and the manufacturer may recommend a new level.

Tolerances are the single most important tool by which the U.S. Government regulates pesticide residues in food. More than 8,500 food tolerances for all pesticides are currently listed in the Code of Federal Regulations (CFR). Approximately 8,350 of these tolerances are for residues on raw commodities (promulgated under section 408) and about 150 are for residues known to concentrate in processed foods (promulgated under Section 409). Table 1-1 shows the number of tolerances established for insecticides, herbicides, and fungicides in the mid-1980s for purposes of comparison.


Infants and children are unique. They are undergoing growth and development. Their metabolic rates are rapid. Their diets and their patterns of dietary exposure to pesticide residues are quite different from those of adults.

TABLE 1-1 Food Tolerances Established Under Sections 408 and 409 of the Federal Food, Drug, and Cosmetic Act

Number of Tolerances Under:

Type of Pesticide / Section 408 / Section 409

Insecticides / 3,654 / 63

Herbicides / 2,462 / 39

Fungicides / 1,256 / 20

Total / 7,372 / 122

NOTE: This table does not include feed-additive tolerances listed in the CFR.
SOURCE: NRC, 1987.

To determine whether the current regulatory system in the United States adequately protects infants and children against dietary residues of pesticides, the committee considered two main issues—susceptibility and exposure:

• Susceptibility: Are infants and children more or less susceptible (sensitive) than adults to the toxic effects of pesticides? Is there a uniform and predictable difference in susceptibility, or must each pesticide (and each toxic response) be considered separately? Does susceptibility increase during periods of rapid growth and development? Does high metabolic activity lead to more rapid excretion of xenobiotic compounds and thus to reduced susceptibility? Is the ability to repair damaged tissues and organs greater in childhood, thus leading to apparently lower sensitivity? In what fashion does the potentially long life span of infants and children affect their susceptibility to diseases with long latent periods?

• Exposure: What foods do infants and children eat? How much of these foods do they eat? How much variation in diet is there among children in the United States? How much, and what residues are found in or on the food eaten by infants and children? What are the nonfood sources of pesticide exposure? How important are they? What data are available on exposure? Are there adequate, frequently collected food consumption data categorized by age, sex, and race that can serve as a basis for computations of intakes by potentially more sensitive subgroups in the population? What are the proper measures of exposure?

The committee examined current procedures for toxicity testing of pesticides to learn whether these approaches provide sufficient information on toxicity in the young. Specific questions posed by the committee included: How are toxic effects identified? If they are determined by experiments in laboratory animals, what problems exist in transferring the results to humans? To infants and children? What information on toxicity is needed? For example, is information on mechanisms of action needed to establish risks to children? Are animal studies on weanlings and older animals adequate to estimate toxicity in infants and children at relatively earlier stages of development? Are there toxicities unique to some species of laboratory animals? To humans? How can exposures of animals to toxicants late in life predict responses in humans exposed early in life?

The committee reviewed approaches to pesticide risk assessment to assess whether these approaches adequately consider the effects of exposure in young age groups. Specific issues included: How is exposure to pesticide residues associated with response? If special consideration needs to be given to childhood exposures that result in risk, how can laboratory data from lifetime animal studies be used to develop meaningful estimates? Does risk accumulate faster during the early years of life? When exposure to a pesticide leads to more than one toxic responses, how can, or should, the total toxicity be described or evaluated?

Two final issues that the committee considered were:

• How can the lifetime risks associated with exposures to pesticides and other chemicals during infancy and childhood be assessed?

• How can methods for assessing and controlling these risks be improved?

In this report, the committee considers the development of children from the last trimester of gestation (26 weeks) through adolescence—approximately 18 years of age. Twenty-six weeks of gestation is considered the beginning of infancy because this age coincides closely with the earliest point at which an infant can survive outside the uterus. All major organ systems can function independently at that point, and the lungs have developed to the degree that reasonable exchanges of oxygen and carbon dioxide can take place.

Chapters 2, 3, and 4 of the report consider the susceptibility of infants and children to pesticides. Chapter 2 examines current evidence on the impact of children's exposures to pesticides and other toxicants in light of the special demands imposed by their rapid development, their special nutritional requirements, and their rapid metabolism. Chapter 3 explores current data on perinatal and pediatric toxicity. In Chapter 4, the committee reviews EPA's current and proposed toxicity testing requirements for pesticide registration and tolerance setting. Chapters 5, 6, and 7 assess the dietary exposure of infants and children to pesticides. The committee began this examination by reviewing in Chapter 5 the food consumption patterns of this age group and exploring the ways that the patterns differ from those of adults—not only in the types and amounts of food and water consumed, but also in the proportion of the diet comprising certain foods. Then in Chapter 6 the committee reviews the data available on pesticide residues in food and gives particular attention to sampling of the foods consumed most by infants and children. In Chapter 7, the committee ties together the information on dietary patterns and residue levels from the two preceding chapters and provides examples for estimating the dietary pesticide exposures of infants and children. This linking of the data on dietary patterns of infants and children with data on pesticide residue levels was accomplished by applying a computer-based technology that enabled the committee to examine and quantify the full range of dietary pesticide exposures. This methodologic innovation obviates the need to study the average exposure of the hypothetical "normal" child and focuses instead on the full distribution of exposures.

In Chapter 8, the committee focuses on risk assessment. Using the data developed in Chapters 5, 6 and 7 on exposure levels, the committee presents a new method that can be used by government regulatory agencies to assess the health risks to infants and children resulting from exposures to pesticide residues in the diet. Like the exposure assessment method developed in Chapter 7, the risk assessment method permits examination of the full range of risks across the entire pediatric population.

This report embodies three unique features:

• It is the first assessment of dietary exposures to pesticides that has focused specifically on infants and children. It makes the case that children are different from the rest of the population, both in their vulnerability to toxicants as well as in their patterns of dietary exposure to pesticide residues. Children therefore deserve specific attention in the risk assessment and regulatory processes.

• It considers the total distribution of dietary exposures to pesticides among infants and children. It does not focus merely on average exposure, nor does it simply use summary statistics to examine the pesticide exposures of a hypothetical "average" child. Instead, through the use of newly applied statistical techniques, the committee was able to examine and quantify the entire range of exposures confronting the pediatric population of the United States. In this way, the committee was able to develop improved estimates of the numbers of children with high levels of dietary exposure to pesticides. This approach should be of considerable value to the government regulatory agencies, especially EPA, as they continue their efforts to use risk assessment methodologies to safeguard the health of the U.S. population.

• It proposes new cancer risk assessment methods that take into account temporal patterns of exposure to pesticide residues in the diet of infants and children, as well as tissue growth and changes in cell kinetics with age. Because of their greater consumption of certain foods relative to body weight, children may be at greater risk than adults from pesticides with carcinogenic potential. Infants and children are subject to rapid tissue growth and development, which will have an impact on cancer risk.

This report indicates how such variations in exposure with age can be accommodated in the Moolgavkar-Venzon-Knudson model of carcinogenesis (Moolgavkar et al., 1988), along with data on tissue growth and changes in cell kinetics. The methods proposed here can be adapted and extended, based on the availability of appropriate data on dietary exposure to pesticides and on tissue growth and cell kinetics, to arrive at improved estimates of lifetime cancer risks that may be posed by dietary exposure to pesticides.


Block, G., B. Patterson, and A. Subar. 1992. Fruit, vegetables, and cancer prevention: A review of the epidemiological evidence. Nutr. Cancer 18:1–29.

EPA (U.S. Environmental Protection Agency). 1992. Pesticide Industry Sales and Usage. 1990 and 1991 Market Estimates. Office of Pesticide Programs. Washington, D.C.: U.S. Environmental Protection Agency.

Mason, A.F. 1928. Spraying, Dusting, and Fumigating of Plants. New York: Macmillan.

McCallan, S.E.A. 1967. History of fungicides. In Fungicides: An Advanced Treatise, Vol. 1, D.C. Torgeson, ed. New York: Academic.

Moolgavkar, S.H., A. Dewanji, and D.J. Venzon. 1988. A stochastic two-stage model for cancer risk assessment. I. The hazard function and the probability of tumor. Risk Anal. 8:383–392.

NRC (National Research Council). 1987. Regulating Pesticides in Food: The Delaney Paradox . Washington, D.C.: National Academy Press. 288 pp.

Wilkinson, C.F. 1990. Introduction and overview. Pp. 5–33 in Advances in Modern Environmental Toxicology: The Effects of Pesticides on Human Health, Vol. XVIII, S.R. Baker and C.F. Wilkinson, eds. Princeton, N.J.: Princeton Scientific. 438 pp.
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Postby admin » Sun Mar 13, 2016 2:36 am

2. Special Characteristics of Children

BECAUSE THEY ARE growing and developing, infants and children are different from adults in composition and metabolism as well as in physiological and biochemical processes. In a period of 26 weeks, or about 6 months, the human concepts grows from microscopic size to recognizable human form weighing almost 500 g (1 pound). At that time, its organs and body systems (cardiovascular, pulmonary, genitourinary, gastrointestinal, neurological, hematological, immunologic, endocrine, and musculoskeletal) are sufficiently mature that extrauterine existence is possible—but survival is very risky. After 3 more months of intrauterine growth (38 weeks of gestation), the average fetal weight increases to 3.5 kg (7.5 pounds), and the organs and body systems become mature enough that adaptation to life outside the uterus is relatively assured. From birth through adolescence, physical growth and functional maturation of the body continue. The rates of physical growth and functional development vary from system to system, organ to organ, and tissue to tissue during this time. Thus, not only do infants and children differ from adults, but at any point during maturation, the individual differs in structure and function from herself or himself at any other age.

This chapter summarizes what is known about these differences. Several aspects of physical growth (structure) and development (functional maturation) are considered in terms of the implications they may have for evaluating the impacts of pesticides on children's health.

Physical development of the body (overall growth), nervous and digestive systems, liver and kidneys, and the proportions of body water and body fat are of special concern in the study of developmental toxicology. Prior to full maturation, damage to an organ or organ system, such as the central nervous system, could permanently prevent normal physical maturation. Also of concern are the physical properties of the toxic substances. For example, water-soluble compounds will be more diluted when the proportion of water in the body is higher, as in infancy, and lipid-soluble substances will be more concentrated in fatty or adipose tissue when the proportion of fat in the body is lower, as may also occur in infancy.

Functional development involves changes in the operational reactions that constitute the process of living, such as the ability to digest and absorb substances in the gastrointestinal tract, to alter their composition in the liver, to develop new metabolic pathways, and to excrete chemicals in the urine. These functions also develop at differing rates, so the organism may respond to chemicals in different ways at different ages. For example, because the filtering function of the kidney develops at a rate different from that of the reabsorptive and secretory functions, the kidney's overall ability to rid the body of toxic substances and the relative ability to excrete substances that are partially reabsorbed or are secreted will vary in a nonlinear fashion with increasing age.

The development of the functions of digestion, absorption, distribution, metabolic alteration, and elimination is of major importance in studies of developmental toxicology. This importance results in part from the possibility that toxicity to a functional system (e.g., glucose homeostasis) prior to its full development may permanently affect that system, resulting in altered function (e.g., glucose metabolism) in the mature animal, and in part from differences in the rates of absorption, metabolism, and excretion and therefore differences in susceptibility to toxicity at different ages prior to maturation.

Behavioral development includes the maturational changes in physical and mental activities associated with the relation of the individual to the environment. Behavioral development has four interrelated aspects: (a) gross motor and fine motor activities; (b) cognitive ability; (c) emotional development; and (d) social development. Alteration in one of these domains can affect the development of each of the other three.

Because of the dependence of behavioral development on physical and functional development, toxic effects occurring before maturation may permanently alter behavioral development. The most commonly encountered and well-known toxicants that can permanently change all four of the components of behavioral development are bilirubin toxicity in the newborn and lead toxicity in the infant or young child. All four aspects of behavioral development are important in studies of developmental toxicology, but much more attention has been given to the first two because they are easier to measure.

The committee's conclusions regarding how infants and children differ from adults with respect to susceptibility to toxicity are based on an extensive literature review. The few available epidemiological studies of chemical toxicity in humans, usually adults, are supplemented by experimental studies in animals. However, much less experimental work has been done on the relationship between body systems and pesticide toxicity in immature organisms than in mature ones. Therefore, data on the relationship between differences in structure and function of the young as they pertain to pesticide toxicity are supplemented with data on the effects of other toxic substances on immature organisms.

Although infant and adult nonhuman animals differ in much the same way that human infants and adults differ, there are substantial interspecies differences among the young. For example, the newborn mouse or rat more nearly resembles the human fetus in the third trimester of gestation than the human infant at birth. On the other hand, the rate of maturation and growth of the mouse or rat after birth is relatively more rapid than that of the human. Thus, cross-species comparisons of potential toxicity for pesticides in the very young animal, although helpful, cannot be used in the same manner that cross-species comparisons are used with adult animals because of differences in developmental patterns. At birth, the dog's brain reaches approximately 8% of its mature weight, the hamster 8%, the rat 15%, the mouse 22%, humans 24%, and the monkey 60% (Himwich, 1973). Subsequent rates of growth are shown in Figure 2-1. Examining the data in another way, as percent of mature weight, human brain weight at 15 to 20 months of age is similar to rat brain weight at 13 to 17 days of age. The rat brain from birth to 26 days is like the hamster brain from 3 to 17 days. For a different variable, the development of γ-aminobutyric acid (GABA) in the cat between the fetal ages of 40 and 44 days is like that of the dog between birth and 10 days of postnatal age (Himwich, 1973).

In some parts of this report, the term development is used to include both functional and behavioral development. Thus when reference is made to "growth and development," the term growth refers to the physical or structural changes associated with the process of maturing, and development refers to the functional and/or behavioral changes that occur during maturation.


Physical growth is a regulated process that represents the sum of the processes of growth of individual cells, tissues, organs, and body systems. These components do not grow at the same rate, but each component has its own rate characteristics. Thus, it is possible to predict the composition of the body from one time to the next; however, the overall composition


FIGURE 2-1 Differences in growth rate of the brain in several animal species. The relative-age scale depends on the age for mature brain size in each species. At birth, for example, the monkey brain will be 60% of its mature size, the human and mouse brains will be 24 and 22% of their mature size, and the hamster and dog brains will be less than 10% of mature size. Body weight does not reach 50% of its adult value until after 10 years of age, but by about 6 months of age brain weight is half of adult brain weight. Skeletal muscle grows more slowly in mass than total body weight before adolescence but increases in weight more rapidly than does body weight after adolescence. The combined weight of liver, heart, and kidneys increases more slowly than brain weight and is about 50% of that of the brain during early childhood. Liver, heart, and kidney weight reaches 50% of its adult value before adolescence and exceeds brain weight by early adolescence. B, birth. SOURCE: Based on data from Himwich, 1973. Age at brain maturity: monkey, 4 yrs.; human, 12.5 yrs.; mouse, 90 days; rat, 120 days; hamster, 60 days; dog, 120 wks.

is never the same from moment to moment until growth processes are complete.

Normal Human Growth

After birth of the full-term human infant, growth occurs at an average rate of 800 g/month or at an incremental rate of 25% of total body weight per month. Because this rate soon slows, the actual doubling of birth weight takes 5 to 6 months. At 1 year, the infant reaches almost three times its birth weight (Table 2-1). After the first year, growth proceeds at about 200 g/month. A peak gain is reached at adolescence: 500 to 600 g/month for boys and somewhat less for girls (Cheek, 1968a).


TABLE 2-1 Average Changes in Various Body Constituents, Birth to 17 Years

Growth in body proportions also changes dramatically during childhood as it does in utero. By 2 years of age, at four times birth weight, the toddler achieves about 20% of adult weight and about 50% of adult height; skull circumference and brain size will already be near their adult values. At birth, the length of the head is 25% of body length; by adulthood the head is only about 14% of the length of the body. Thus, growth proceeds from the head downward. The midpoint of the newborn's body is at midabdomen, but by adulthood it is at the junction of the legs with the trunk.

During infancy and adolescence, children are growing and adding new tissue more rapidly than during any other period in their postuterine life, but their various organs, tissues, and metabolic processes are maturing at different rates. For example, the neuronal cell population is relatively complete by 2 years of age, but full myelination of neuronal tissue is not complete until their 18th year. The brain achieves 50% of its adult weight by 6 months of age, whereas approximately 50% of adult stature is not reached until 2 years of age. In contrast, 50% of the adult weight of the liver, heart, and kidneys is not reached until the children are about 9 years old, and the same point in growth of skeletal muscle and total body weight is not attained until approximately the 11th year of age. Thus, as the child grows, his or her body consists of differing proportions of various tissues and organs that comprise the body. Various tissues—brain, skeletal muscle, liver, heart, and kidney—have different metabolic rates and biochemical pathways, and their changing physical proportions will alter the disposition of xenobiotic compounds over time.

The major periods of rapid growth include infancy and puberty. Growth is most rapid in the combined period of in utero development plus infancy and puberty. From birth to 4 years of age, the rate of growth decreases from 50 cm/year to the childhood value of about 10 cm/year. During puberty, the rate of linear growth increases to approximately 12 cm/year. These periods of rapid growth are believed to be susceptible to adverse influence of toxicants (Karlberg, 1989).

The hypothalamus, pituitary glands, gonads, and other tissues are involved in the control and expression of growth during puberty. Growth hormone (GH), both directly and through stimulation of the somatomedins, plays a major role in growth through endocrine, autocrine, and paracrine mechanisms. GH deficiency is associated with impaired growth, and GH excess with increased growth. The androgens and estrogens, through anabolic and other endocrine mechanisms, also modulate growth at puberty. High levels of estrogens (or estrogen-like molecules such as some pesticides) can, through effects on the growing bone, decrease the optimum height attained in adulthood or rate of linear growth during puberty.

Human Compared to Animal Infants

The newborn rabbit, rat, mouse, and hamster can double their birth weights in less than 1 week, much faster than the human infant can (Altman and Dittmer, 1962). These different growth velocities may alter the toxicity of pesticides and other chemicals among different species of infant animals. In various theories of carcinogenesis, it has been postulated that the rapidity of DNA synthesis and cell proliferation affects on carcinogenicity or other toxic manifestations of chemicals. Therefore one might expect the biological effect (either positive or negative) to be more pronounced in the more rapidly growing animal—e.g., the dog, rabbit, or rat—than in the human infant. The effects on the mouse or hamster should be less pronounced than those on the dog, rabbit, or rat but still more pronounced than those on the human infant. However, if toxicity is related to absolute rates of growth, the order of increased sensitivity would be the hamster (36 days for a 10-fold increase in birth weight), the mouse (54 days), the rabbit (80 days), the rat (160 days), the dog (165 days), the monkey (1,000 days), and humans (5,000 days). This issue is complex, but in the absence of other factors, direct carcinogens are more potent in rapidly growing animals (Cohen and Ellwein, 1991; Weinstein, 1991).

Toxicologic Implications of Growth in Cell Numbers and Size

The rapid growth from conception through infancy in both animals and humans is achieved primarily by an absolute increase in the number of cells in the body (hyperplasia). After infancy, the bulk of growth is the result of increase in cell size (hypertrophy). The major exception is in the growth related to increased secretion of a series of hormones in adolescence. The change from hyperplasia to hypertrophy has major implications for adult form and size. Because of the possibility of a permanent reduction in total cell number, factors that alter growth prior to 2 years of age, even if only transient, are much more likely to result in diminished adult size than similar transient effects that occur after 2 years of age. An example of this phenomenon is the permanent reduction in body size that results from congenital rubella.

During the period of hyperplasia, which varies in duration from one organ to another, the increased rate of DNA/RNA replication and of protein synthesis may have different implications for toxicity. In a study of the age-related sensitivity of 3-, 7-, 11-, and 29-week-old rats to the induction of anemia, phenylhydrazine (PHZ) produced more marked anemia in the older animals on a milligram-per-kilogram basis (Vesell, 1982). The author postulated that the ''resistance" of the younger animals could have been secondary to their increased rate of red blood cell proliferation. When a dosage based on surface area was used, the results continued to demonstrate a proportionate age-related sensitivity to PHZ.

Increased rates of cell proliferation can be associated with negative as well as positive outcomes. In the mammary gland, the proliferation of cells is greatest at the time of menarche—the first menstrual period—and the greatest risk for carcinogenesis from radiation is when the radiation occurs immediately before and after that event (Norman, 1982; Miller et al., 1989). On the other hand, radiation during infancy for treatment of an enlarged thymus also increased the risk of breast cancer 30 to 40 years later (Hildreth et al., 1989). The role of cell proliferation in carcinogenesis remains unclear and may depend on a variety of other factors (Moolgavkar and Venzon, 1979; Moolgavkar and Knudson, 1981; Finhorn, 1982; Rajewsky, 1985; Moolgavkar, 1988; Moolgavkar et al., 1988, 1990.a,b; Cohen and Ellwein, 1990, 1991; Moolgavkar and Luebeck, 1990; Cohen et al., 1991).

In one study, 3- and 14-month-old rats were given intravenous doses of N-nitrosomethylurea (NMU), a known carcinogen (Anisimov, 1981). The 3-month-old rats received doses of 200 or 100 mg/kg of body weight (bw) (30 or 15 mg per rat) (groups 1 and 2, respectively), and the 14-month-old animals received 100 mg/kg bw (30 mg per rat) (group 3). The animals were observed throughout their life span, and tumor incidence was compared with that in a nontreated control group (group 4). Mammary adenocarcinomas occurred only in the rats treated at 3 months of age: 71% of group 1 and 32% of group 2 animals. Cervicovaginal sarcomas and carcinomas were found only in the group 3 animals injected at 14 months of age. Kidney tumors occurred in groups 1, 2, and 3 but were most common (54%) in the group 2 animals (the 3-month-old rats receiving the lower dose of NMU). The incidence of kidney tumors in group 1 was 10% and in group 3, 17%; there were none in the control group. Neoplasia of the hematopoietic system was found equally in groups 1, 2, and 3 (33%, 32%, and 25%, respectively, compared with 9% in the controls). The author postulated that the younger rats developed mammary and renal tumors more readily than the older animals because the proliferation of mammary gland cells and renal epithelium had decreased with age. The lack of mammary tumors in the animals treated at 14 months may reflect the insufficient number of remaining months of life to allow for development of the tumors. However, the growth of mammary tissue generally precedes the growth of vaginal/uterine tissue in most species. The older animals developed more uterine, cervical, and vaginal tumors. The higher level of estrogens at their age (14 months) was believed to be associated with more rapid proliferative activity of the target tissues (Anisimov, 1981).

In a study by Drew et al. (1983) of the effect of age and duration of exposure on the carcinogenicity of vinyl chloride (VC) in female rats, hamsters, and mice, younger animals were generally more affected than older ones. Beginning at about 2 months of age, animals were exposed to VC for 6, 12, 18, and 24 months. Other animals were exposed for 6 or 12 months beginning 6, 12, or 18 months later than the original groups. For all animals, the life span was shortened when exposure began at 2 months of age, regardless of duration of exposure (except for rats exposed for 6 months). Exposures beginning 6, 12, or 18 months later did not shorten the life span except in B6C3F1 mice, which survived about 315 days from the onset of exposure regardless of the age at which exposure was initiated.

The incidence of tumors was also directly related to age of onset of exposure and duration of exposure, but age of onset appeared to be the dominant factor. One exception to this finding was that hepatocellular carcinomas were more prevalent in rats whose 6-month exposure began 6 months later than that of the group treated at 8 weeks. The incidences of hemangiosarcomas and mammary gland adenocarcinomas, as well as neoplastic nodules, were all related to the earlier age of exposure. The greatest incidence of hemangiosarcomas and stomach adenomas was observed in hamsters whose 6-month exposures began at the earliest ages. Longer exposures did not change the incidence. Similar findings were noted in the mice. However, the incidence of hemangiosarcomas was not affected by age in the B6C3F1 mice, but was related to age in the Swiss mice.

In general, exposures beyond 12 months did not change the incidence of tumors in rats, and 6-month exposures of hamsters and mice were sufficient to achieve maximum tumor incidence with only two exceptions. Of note is the finding that the incidence of hemangiosarcomas and stomach adenomas in hamsters was reduced by exposures greater than 6 months. However, Drew et al. (1983) postulated that the shortened life span resulting from the longer exposures could have caused the animals whose exposure began at 2 months to die before their tumors could be expressed. The authors pointed out that animals exposed early in life must be allowed to live out their full life span if appropriate tumor incidence is to be ascertained. When this was done, it was quite apparent that the animals exposed at the youngest age had, with an occasional exception, the highest incidence of a variety of tumors. These studies lend support to the concept that the rate of cell proliferation (greatest in the youngest animals) is a contributing factor to the carcinogenic effect of VC. Similar conclusions were drawn by Swenberg et al. (1992).

Toxicologic Implications of the Growth and Development of Organs

The pattern of the rate growth of the various organs in the human infant varies from the pattern of the overall body growth rate. Figure 2-2A illustrates the growth rate for several organs in humans. The thymus grows most rapidly initially, and throughout most of childhood exceeds its adult size. The brain approaches adult size early in childhood, although behavioral development continues for many years. The uterus and testes grow slowly until adolescence, and the growth of the ovaries is similar to that of the kidney and spleen and to increase in total body weight. Figures 2-2B and 2-2C present growth curves for some of the same organs in the rat and mouse. Each of these species has a different pattern of specific organ growth, and these differences complicate comparative studies between species. Much less is known about the rate of functional development of various organs. The age period in which specific organs or tissues undergo their most rapid rate of development and the age at which development is completed have major implications for studies of toxicity to those organs in growing animals. Toxicity that is dependent on rates of cell proliferation (DNA/RNA replication and protein synthesis) might affect different tissues or organs at various stages of animal growth. Furthermore, whether the effect generated permanently alters the pattern of adult function, depresses adult function, or has no permanent effect may depend on whether the particular organ or tissue is still evolving or has reached its adult capacity at the time of exposure. Knowledge of the developmental pattern for various tissues and organs in animals used for studying pesticide toxicity is important in determining potential target sites and toxic end points in humans.


FIGURE 2-2 Organ development and stature or body weight as percentage of adult values by age in the human (A), rat (B), and mouse (C).

SOURCE: Based on data from Altman and Dittmer, 1962.



On another level, cellular functions within the body and within organs change with the period of overall growth. From early gestation to early infancy, blood cells are formed in the liver and spleen, but by late infancy only the bone marrow functions in the initial development of blood cells. However, "the volume of bone marrow in newborns is in fact so large that it is nearly equal to the marrow space occupied by hematopoietic cells in adults" (Nathan, 1989). For compounds that are potentially toxic to the hematopoietic system, the anatomic location and the relative size of the system may determine the degree of toxicity. Substances that accumulate in the liver or spleen in the young infant could have a direct effect on hematopoiesis—an effect that might be absent in the older child or adult. Similarly, the relatively large volume of hematopoietic tissue in the young could contribute to a more profound effect for substances that are widely distributed in the body but could reduce toxicity through dilution of compounds that concentrate in such tissue.

In several portions of the brain, the nerve cells (or neurons) that will ultimately be near the surface of the brain (in the cortex) must migrate from more central locations to the cortex during late fetal and early infant life (Rakic, 1970, 1972). This is only one of many developmental processes taking place in the brain that could be disrupted by substances toxic to the central nervous system (Langman et al., 1972). Although the developmental sequences of tissue organization and cellular maturation are similar in various animal species, the specific rate of each sequence relative to other sequences and across species differs significantly (Rodier, 1980). The myelination of nerve tracts in the spinal cord and peripheral nerves is a process that continues throughout childhood. Incomplete myelination of nerve fibers could alter their response to xenobiotic agents. The cumulative dose of cisplatin "that causes peripheral neuropathy tends to be higher in children and younger patients than in the elderly" (Legha, 1990). Whether the decreased sensitivity in the child is related to incomplete myelination or to other metabolic differences is unknown. Thus, the impact of toxic products can produce quite different outcomes that vary both with time and with species.

Toxicologic Implications of Changes in Body Composition

The human infant is compositionally immature at birth. The infant's body is unlike the adult's in terms of body water and the relation of skeletal mass to other lean body tissue. Table 2-1 provides values for various body constituents at ages that represent significant developmental periods in the life of a child. The bottom row presents the value ratios for 17 year olds compared with newborns. From birth to age 17, body weight increases about 18-fold, stature slightly more than 3-fold, and surface area about 8-fold. Total body fat increases about 36-fold in the female and about 17-fold in the male. Body water increases about 15-fold, but extracellular water only 10-fold. Total body protein increases almost 20-fold, whereas bone mineral increases by 25-fold and total body potassium by about the same amount as total body protein because protein is located primarily in cells and potassium is primarily an intracellular ion.

The rate at which various body constituents increase and their relative proportions in the body shift during childhood (see Table 2-1). For example, total body fat accumulates most rapidly in infancy and again in adolescence, especially in the female. Bone mineral also increases markedly in adolescence—especially during the male's major growth spurt in late adolescence. Another relationship that has been known for many years is the correspondence between the extracellular fluid volume and the surface area of the body. Both of these sets of values also have a linear relationship to metabolic rate for most of childhood.

The implications of these changes in body composition for pesticide toxicity are not well defined experimentally, but one can hypothesize a number of possibilities. The relatively larger extracellular fluid volume of the infant would result in a somewhat greater dilution if the infant were to ingest equivalent amounts (on the basis of body weight) of water soluble substances. Alternatively, lipid-soluble substances given in equivalent amounts on a body weight basis would be more concentrated in the fat of the young child because of the lower amount of fat per kilogram of body weight. Similarly, compounds that have an impact on bone growth would be more concentrated per unit of bone mass because of the bone's relatively smaller size per unit of body weight in the infant and child.

Premature and full-term infants differ from older children and adults in their body composition. Whereas the premature infant may be 85% water, the lean body mass of the full-term infant is 82% water, compared with 72% water in the lean (fat-free or non-fat-containing) body mass of the adult. Most of the "excess" water in the infant is extracellular (Forbes, 1968). Thus, the overall amount of organ tissue per unit of whole body mass is less in the infant. One exception is the liver, which in the child is relatively larger per unit of body weight than in the adult. Several investigators have postulated that this relatively larger size of the liver could play a role in the capacity of the young child to metabolize drugs such as phenylbutazone and antipyrine more rapidly than the adult (Coppoletta and Wolbach, 1933; Alvares et al., 1975; Vesell, 1982). In addition, the organs themselves contain more water than do the organs of adults (Dickerson and Widdowson, 1960; Widdowson and Dickerson, 1960, 1964; Dickerson, 1962). Brain water decreases from 90% in the infant to 77% in the adult (Altman and Dittmer, 1973). Liver water decreases from 78% to 71% and kidney water from 84% to 81% (Widdowson, 1968). Although the total water content of skeletal muscle is relatively unchanged, the extracellular water of muscle decreases from 35% to 18%, but the intracellular water increases from 45% to 61%. The change in intracellular water is to a large extent related to an increase in cell size.

The changes in body water compartment sizes with increasing age may bear some relationship to changes in pesticide toxicity, depending on the distribution of the toxic compound. Those agents that are water soluble and distributed extracellularly will be more diluted in the youngest animals, or in human infants and small children, for comparable exposures on a milligram-per-kilogram basis. Thus one might expect toxicity to vary with age for water-soluble compounds. If the compound or its toxic metabolite is distributed intracellularly, the relatively decreased volume of intracellular water in the young would lead to toxicity varying indirectly with age. These age-related differences in toxicity emphasize the importance of ascertaining the distribution of water-soluble toxic materials within the specific tissue water compartments, extracellular or intracellular.

Cell size in infancy is relatively small. Thus, there is proportionately more cell membrane per unit of cell mass in the infant than in the adult (Cheek, 1968b). The effect of the relatively greater cell membrane and, therefore, of cell surface area compared to cell mass in the infant and young child has not been explored, but one might postulate that these proportions could increase the sensitivity of the cells of the young to compounds that act primarily at the level of the cell membrane.


Development, as used in this context, refers to the functional maturation of various cells, tissues, organs, organ systems, and the organism as a whole. As noted previously, the changes in function follow patterns similar to but not identical with physical growth. Functional development, like physical growth, is established to a large extent by genetic mechanisms. However, alterations in the patterns of functional development may be more readily modified by external (environmental) factors than the patterns of physical growth. In addition, there may well be external factors that modify growth but do not affect development and vice versa.

Genetics, Development, and the Environment

Understanding infant and child development in relation to the toxicity of pesticides requires an understanding of the constantly evolving interaction between a person's genetic endowment, developmental processes, and the environment. The infant or small child's phenotype (characteristics produced through interaction between genetic properties and environment) more closely resembles its genotype (genetic properties) than is true for the adult. At conception, each individual is genetically unique. From the time of conception, environmental factors may alter the genotype to produce a different phenotype. As the phenotype changes, it may alter the environment and be further altered by the environment. Specific enzyme systems may be enhanced, delayed, or altered permanently in their development by environmental factors. In addition, environmental exposures may impair, alter, or delay the development of some biochemical or physiological systems. Thus, not only will the programmed development of enzyme function alter the responses to xenobiotic compounds in immature, compared with mature, organisms, but environmental exposure to such substances before maturity may further modify response at a later time, or in some cases, throughout the life span. The specific stage in tissue or organ development when environmental factors can modify the cells to produce an effect apparent only in later life are termed critical periods of development. For example, giving insulin or glucosamine to newborn animals may permanently alter the mature animal's insulin levels and blood glucose values (Csaba and Dobozy, 1977; Csaba et al., 1979). Another example is the elevation of serum bilirubin in the human neonate, which produces altered brain function that becomes evident as the child matures. Similar elevations of serum bilirubin after infancy fail to produce these changes. Some of the damage to the central nervous system resulting from exposure to low levels of lead may not be apparent until the development of more mature functions in test animals (Csaba et al., 1979) or of such skills as reading and arithmetic in children. Similarly, neonatal exposure to diethylstilbestrol may produce effects later in life in the reproductive system (adenocarcinoma of the vagina and impaired function of the reproductive and immune systems) (Kalland, 1982).

As a general rule, compounds that interact with some genetic component of an individual are likely to be more active and cause greater impact on the young, whereas those compounds that depend for their activity on the development of acquired characteristics—such as elevated blood pressure, atherosclerosis, and loss of renal function—will be more active in the elderly. For example, the red blood cells of older mice are more easily damaged by oxidants, possibly because of acquired changes in their cell membrane (Tyan, 1982) and impaired ability to withstand oxidant stress. Similarly, the increasing presence of chromosome aberrations in the cells of the older person, perhaps because of longer exposure to toxic substances in the environment, may contribute to the increased carcinogenic susceptibility of the elderly (Singh et al., 1986).


The metabolic rate, a measure of the total energy expenditure of an organism, increases as the size of the organism increases. However, when energy expenditure or energy intake is evaluated per unit of mass, it decreases with increasing body size. Thus, the metabolism of exogenous substances administered in terms of body size will generally be more rapid in the immature than in the mature animal. This may have important implications for toxicity of pesticides—decreasing toxicity when the parent compound is the toxic agent and increasing toxicity when a breakdown product is the toxic substance. However, there is a variety of other aspects of changes in metabolism that can also act to modify toxicity.

The infant's body surface area per unit of body mass is two to three times greater than the adult's. This relatively large surface area per unit of mass is associated with a metabolic rate that is more than twice the adult's on a weight basis but closer to the adult value when compared per unit of surface area. Because food and water intake is related to metabolic rate, the infant's consumption will be greater than the adult's on a weight basis, but total energy consumption will be more nearly comparable to the adult's if surface area is the basis of comparison. Thus, infants differ from adults, not only in specific metabolic and pharmacokinetic measures, but also in overall metabolic rate.

Because of differences in the types of food consumed by the young (see Chapter 5), the amount of a food consumed, with its related additives, preservatives, and contaminants, may be 1 order of magnitude higher or lower for children than for adults per unit of body weight. The percent absorption varies with the development of the gastrointestinal system. The volume of distribution also depends on the degree of maturity as well as on the solubility characteristics of the substance. Because of developmental changes in enzyme activity, rates of deactivation or activation are related to the stages of maturation. The number of cell membrane receptors and the degree of protein binding are similarly variable with age. Rates of excretion by the liver and kidneys depend on both the overall organ development and the differential rates of maturation of specific organ components. For example, the development of the filtration function of the kidney is slower than the development of the reabsorption and secretory functions. Finally, the cellular response itself is dependent on intracellular maturation and differentiation. Thus, the human infant or infant animal may respond quite differently from the adult of its species to many xenobiotic substances. There is no reason to believe that these differences in response would not be equally applicable to pesticides.

Developmental Toxicity Studies

Because so many bodily functions are at various stages of development throughout infancy and early childhood, toxic effects of chemical agents during these age periods not only produce the same sorts of direct injuries to established organ tissues and functions seen in adults, but also have the potential to affect the later development of anatomic, physiologic, and metabolic processes.

During organogenesis, functional integrity does not necessarily coincide with morphological maturity, and relative organ size varies as development proceeds. Because of the evolving development of various organs and tissues, the effect of exposure to toxic substances will vary in a complex way with the age at exposure. Substances that are toxic to adults may have minimal effects at one stage of development, but at another stage these same substances may produce permanent damage to the organism or may be lethal. Such effects are particularly prominent for the central nervous system. For example, radiation treatment for medulloblastoma resulted in major cognitive problems at a later age for children less than 4 years old at the time of radiation, minimal problems later if treatment occurred at 5 to 7 years of age, and no residual cognitive difficulties when radiation was administered at 8 years and older (Chin and Maruyama, 1984).

The infant kidney is immature at birth, and the relatively poor glomerular filtration rate leads to delayed drug excretion and, therefore, an increased likelihood of toxicity (Kleinman, 1982). In the liver, the capacity to detoxify drugs by the process of conjugation develops slowly and is a major factor contributing to the toxicity of chloramphenicol in infants (Vesell, 1982).

The complexity of age-related effects of exposure for a specific tissue was demonstrated by Fleisch (1980), who identified the following factors that could change with age and thus alter the age-related sensitivity of blood vessels to various drugs:


FIGURE 2-3 Changes in the arterial smooth muscle ß-adrenergic system with age. The overall decrease in arterial smooth muscle relaxation with isoproterenol stimulation as the organism ages can be related to three separate changes in the model. SOURCE: Based on data from Fleisch, 1980.

• the number of impinging nerve cells,

• the number of drug receptors per cell,

• the relative proportion of one type of receptor to others,

• the cell membrane composition,

• the second messenger systems, and

• the concentration of reactive proteins.

The complexity of the age-related changes in vascular responsiveness is illustrated by analysis of the changes in contractile function of the arterial smooth muscle in the rat. The responsiveness of the venous smooth muscle to the stimulant isoproterenol does not change with age, whereas the responsiveness of the arterial smooth muscle increases in the postnatal period and then declines with increasing maturity. In the simplified model used to exemplify the response, isoproterenol combines with its receptor in the cell membrane and then stimulates a chemical messenger such as cyclic adenosine monophosphate (AMP), which in turn reduces the contractility of the structural protein (see Figure 2-3). In this model the affinity of isoproterenol for the receptor does not change with age, but the number of receptors per cell does decrease with age. The sensitivity of the chemical messenger to receptor stimulation does not change with maturity, but the sensitivity of the contractile protein decreases in the older rats. In addition, the overall contractile property of the structural protein shows a decline with maturation. Incidentally, there are no comparable unambiguous data on the changes in alternative stimulants with age, although age-related changes in function have been demonstrated.

Without detailed knowledge of all such age-related physiological changes and their potential interactions, it is impossible to extrapolate the impact of xenobiotic effects from mature to young animals. It is therefore necessary to examine such effects in immature animals of various ages.

A further complication is that developmental aspects of organs, tissues, and cells in the young may increase sensitivity to some drugs and decrease sensitivity to others. In fact, for the same compound one aspect of the immature organism may increase sensitivity and another aspect may decrease sensitivity. For example, the immaturity of some metabolic functions may lead toprolongation of the half-life of a toxic substance, but the larger volume of extracellular water may reduce its concentration. Reduced numbers of binding sites on target cells could also reduce the toxicity of the substance.

After reviewing a variety of studies of developmental toxicity, Calabrese (1986) concluded that there is no systematic way to predict differences in toxicity related to age. Because of differences in mechanisms of toxicity, rates of metabolism, enzyme development, stages of organ system development (some of which are more resistant to damage than others), and relative proportions of organ mass, it is unlikely that a predictive model of toxicity and development will evolve in the near future (Calabrese, 1986).

In 1983, Anisimov examined the age of animals as it relates to their sensitivity to carcinogens. He discussed four factors that may alter sensitivity on the basis of age:

• changing activity of carcinogen-metabolizing enzymes,

• changes in binding affinity,

• age-related alteration in the accuracy of DNA repair, and

• variation in proliferative activity of target tissues as a function of age and of secondary modifiers of proliferation such as hormones and tissue-specific growth factor.

Because epidemiological data provide contradictory relationships between age at exposure and the development of cancer, and because such relationships vary with specific tumors, Anisimov examined 52 studies of chemical-induced carcinogenesis in animals of different age groups. In a tabulation of these studies for the effect of aging on latency, incidence, and tumor size, he noted that aging increased chemical carcinogenesis in 28 studies, decreased it in 19 studies, and had no effect in 5 studies. Analysis by animal species, target tissue, or chemical agent did not reveal any specific pattern (Anisimov, 1983). In his own studies of N-nitrosomethylurea in rats of various ages, Anisimov (1983) made several observations associated with increasing age at exposure: a decrease in tissue sensitivity to the action of the carcinogen in the mammary gland and kidney; an increased sensitivity in the cervix, uterus, and vagina; and no influence in the hematopoietic system (Anisimov, 1981).

If one were to examine the entire life span at many points, however, it is likely that sensitivity to pesticides and other toxic agents would not follow a regular progression in toxicity from infancy to senescence. One example is a study of the toxicity of 14 pesticides in mallard ducks at 1.5, 7, 30, and 180 days of age. In no case was there a regular progression of toxicity relative to age (Hudson et al., 1972). In all cases but two, there was either an initial decrease in toxicity followed by an increase (four cases) or an initial increase in toxicity in the earliest ages followed by a decrease (eight cases).


Infants and children from the beginning of the third trimester of pregnancy until 18 years of age are the subject of this report. This age span was selected to cover the period of development from the onset of possible extrauterine existence to the age of maturation of essentially all biological systems. Reproductive toxicology and teratology are not a focus of this study.

The chapter discusses maturation in terms of physical growth, physiological and behavioral development, and the impact of the changing patterns of anatomical structure and function on the toxicity of chemicals in the immature individual. The changes do not begin at similar times or proceed at similar rates. At present, toxicity data on mature animals are insufficient for extrapolation to immature animals.


• Human infants and children differ from human adults not only in size but also, and more importantly, in the relative immaturity of biochemical and physiological functions in major body systems; body composition in terms of proportions of water, fat, protein, and mineral mass, as well as the chemical constituents of these body components; the anatomic structure of organs; and the relative proportions of muscle, bone, solid organs, and brain. These structural and functional differences between neonates and adults can potentially influence the toxicity of pesticides, due to qualitative and quantitative alterations in the magnitude of systemic absorption, distribution, binding, metabolism, interaction of the chemical with cellular components of target organs, and excretion. Without detailed knowledge of all age-related physiological changes and their potential interactions, it is not possible to extrapolate the impact of xenobiotic effects from mature to young animals.

• Whatever the target tissue or organ may be, some responses to pesticides may alter physiological processes that influence the uptake and metabolism of food constituents needed for optimal growth and development. Interference with such processes could result in disorders of overall growth or functional disorders of specific organ systems. It is often not known which developmental stage of individual biochemical systems, tissues, or organs will enhance, diminish, or not alter the infant's or child's sensitivity to the toxic effects of specific pesticides. In the experimental setting, therefore, the stage of growth and development of laboratory animals is a critical variable in evaluating the toxicity of pesticides.

• There are, however, specific periods in development when toxicity can permanently alter the function of a system at maturity. These special windows of vulnerability (critical periods) are often found in the early months of human pregnancies; however, some systems (e.g., the central nervous, immunologic, reproductive, and endocrine systems) continue to mature and may demonstrate particular sensitivity during the postnatal period.

• If a compound's toxicity is age related in one species, it is reasonable to assume that there will be an age relationship in other species in which the compound is toxic.


• Care must be taken when selecting an appropriate animal model for investigating the toxic effects of pesticides in infants and children, interpreting the data, and extrapolating the data from young animals to young humans. Toxicity in young rodents can vary substantially over a period of days, since maturation occurs so rapidly in these animals.

• In the evaluation of pesticide toxicity for immature animals, overall growth should be evaluated by measurement of growth rates and adult size at maturity.

• Studies should be conducted to examine age-related physiological changes and their potential interactions in immature animals of various ages.

• Because of the variable rates of organ development within and between species, specific organ system development should be evaluated by functional measures specific to each organ system. For example, if a chemical interfered with glomerular development in the kidney to the extent that the glomerular filtration rate were reduced by 50%, such an effect would not be likely to become apparent until maturity, and only then would the effect be documented by actual measurement of that specific function in the mature animal.


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Part 1 of 5

3. Perinatal and Pediatric Toxicity

BOTH ACUTE AND CHRONIC toxic reactions in the young are often considered together under the title of developmental toxicity. Such toxicity can be further subdivided by the organ system involved or by whether the toxic effect occurred before or after birth. The developmental purview of the committee extends from the beginning of the third trimester through 18 years of age; however, no single theoretical framework or unifying set of principles readily applies to so broad a developmental span. Teratology, the study of congenital malformations, has traditionally focused on the process of organogenesis, the sensitive period in prenatal development when birth defects can be induced by exposure to either endogenous (e.g., endocrine) or exogenous (e.g., xenobiotic) agents. One view of teratogenesis is that this type of abnormal development represents a special form of embryotoxicity.

Developmental toxicology includes the study of chemically induced alterations of the normal sequence of developmental processes. It both encompasses and expands the domain of abnormal development beyond that implied by teratology. Although the term denotes adverse chemical effects on development, its end points are not restricted to gross anatomical defects but encompass multiple expressions of abnormal outcome. This research specialty combines basic principles, concepts, and working assumptions from several disciplines, including developmental and cellular biology, pharmacology, and toxicology. A major objective is to understand how exogenous agents interfere with the normal progression of developmental events to produce phenotypically abnormal cells, tissues, organs, and function. Since this report's focus begins with the third trimester, the committee does not directly consider the teratogenicity of pesticides, i.e., their potential to produce gross structural malformations. Rather, the focus is on processes that occur after the completion of organogenesis and continue well into the postnatal period. However, the origins of this broader concern with peri- and postnatal toxicology are inextricably rooted in experimental teratology.

Studies of the toxicity of xenobiotic compounds in children have demonstrated the potential for either acute or chronic exposure to result in serious malfunctions at a later age. This potential exists because of the developmental character of the physiologic/biochemical/molecular function of the young individual. While a biologic system is developing, a toxic event can alter one aspect of that development so that all subsequent reactions are altered or modified. For example, transient elevations of serum bilirubin during the newborn period may produce changes in the basal ganglia of the brain that may not become apparent until several years later but are then permanent in nature.


In this section, the committee discusses and summarizes the relative sensitivity of infants, children, and adults to the acute toxicity of chemicals. Acute toxicity here is defined as toxicity resulting from a single exposure to a chemical. The injury may be immediate or delayed in onset. Both lethality and target organ injury will be considered as toxic end points. A limited number of findings from studies of laboratory animals are summarized where data on humans are inadequate. Because of the meager data base on age-dependent acute toxicity of pesticides, some examples of pharmacologic effects and adverse effects of therapeutic agents in pediatric and adult populations are described. Attention is focused, in turn, on age-related differences in the lethality of pesticides and other chemicals, differential effects of cholinesterase inhibitors in immature and mature subjects, and age-related effects of toxic and pharmacologic actions of selected therapeutic agents.

Data on age-related susceptibility to the lethal effects of chemicals are largely limited to acute LD50 studies in laboratory animals. Done (1964) was one of the first investigators to compile the results of LD50s and other measures of lethality of a variety of chemicals in immature and mature animals. Immature animals were more sensitive to 34 chemicals, whereas mature animals were more sensitive to 24 compounds. Thiourea was 50 to 400 times more toxic (i.e., lethal) in adult than in infant rats. Conversely, chloramphenicol was 5 to 16 times more toxic in 1- to 3-day-old rats. Thus, Done (1964) concluded that immaturity does not necessarily entail greater sensitivity and that age-dependent toxicity is chemical dependent. Goldenthal (1971) tabulated LD50 values for newborn and neonatal animals as compared to adult animals primarily from data submitted by pharmaceutical firms in drug applications. Approximately 225 of these compounds were more acutely toxic (lethal) to neonates, whereas about 45 were more toxic to adult animals. Almost all the age-related differences in LD50s in the reports of Goldenthal (1971) and Done (1964) were less than 1 order of magnitude; indeed, most varied no more than two- to threefold.

As discussed in Chapter 2, there are important differences between immature laboratory animals and humans. Nonprimate species are generally less mature at birth than are humans. Newborn mice and rats are among the most immature of commonly used test species, so it is not surprising that they often differ markedly from adult animals in sensitivity to chemicals. This phenomenon is particularly evident in the paper by Goldenthal (1971), who reported five times as many chemicals to be more acutely toxic to newborn than to adult animals. Since full-term human newborns are more mature, such pronounced age-dependent differences in toxicity would not be anticipated. Maturation in rodents is very rapid, so that even a few days of age can result in a marked disparity in test results (Done, 1964). Furthermore, organs and their associated functions mature at different rates in different species. Uncertainty in extrapolating findings among different species of mature animals is appreciable. When the additional variable of interspecies maturation patterns is introduced, the choice of an appropriate animal model for pesticide toxicity of neonates, infants, and children becomes even more complex.

The relative acute lethality of pesticides to immature and mature animals has been the subject of a number of studies. Goldenthal (1971), in his extensive compilation of LD50 values for newborn and adult animals, included several fungicides, herbicides, and the insecticide heptachlor. Each of these compounds was more toxic to newborn than to adult rats. Gaines and Linder (1986) more recently contrasted the acute toxicity of 36 pesticides given orally to weanling (4 to 6 weeks old) and to young adult Sherman rats. Age-related differences, where they existed, were usually no more than two- to threefold. Weanlings were more sensitive than adults to only 4 of the 36 compounds. Lu et al. (1965) observed that 14- to 16-day-old rats were intermediate between newborns (most sensitive) and adults (least sensitive) in their susceptibility to malathion poisoning. Such findings are in agreement with the observation that physiological and biochemical processes, which govern the pharmacodynamics of pesticides, mature quite rapidly in rodents. Indeed, metabolism and renal clearance of xenobiotic compounds and their metabolites soon approach and may exceed adult capacities in rodents within 2 to 3 weeks. This same phenomenon occurs in humans, albeit at a somewhat slower pace (i.e., within the first weeks to months of life). Higher metabolism may confer protection against pesticides or increased susceptibility to injury, depending on the relative toxicity (and rate of elimination) of the parent compound compared to its metabolites. The findings of Lu et al. (1965) are a good case in point. These investigators contrasted acute oral LD50 values for newborn, 14- to 16-day-old, and young adult Wistar rats. The adult animals were the most resistant to malathion, as would be anticipated, since adult rats most efficiently metabolize organophosphates and organophosphates are metabolically inactivated (Benke and Murphy, 1975). Conversely, the older rats of Lu et al. (1965) were the most sensitive to the acute toxicity of dieldrin. Thus, susceptibility to acute pesticide toxicity appears to be a function of age, species, and chemical.

Limitations of acute lethality data should be recognized. Acute doses of chemicals high enough to cause death may damage organ systems by mechanisms that are quite different from those that produce biological effects from chronic exposures to lower levels. MacPhail et al. (1987) examined age-related effects of a number of pesticides on lethality, serum chemistry, and motor activity in weanling and adult male rats. Although age was generally not an important determinant of toxicity for most of the pesticides, there were age-related differences in the effects of carbaryl and diazinon on motor activity. These results could not have been predicted on the basis of LD50 values for the two groups, leading MacPhail et al. (1987) to conclude that mortality may be a poor predictor of morbidity and that nonlethal end points should be used to assess the age-dependency of the neurobehavioral toxicity of pesticides. More sensitive indices should also be used to monitor other potentially vulnerable systems in infants and children, including the hormonal and reproductive systems, the immune system, the nervous system, developmental effects, and carcinogenesis/mutagenesis. Unfortunately, relatively few well-controlled studies have been conducted, particularly in humans, in which sensitive end points are used to assess the relative toxicity of comparable doses of pesticides or other chemicals in pediatric and adult populations.

Cholinesterase inhibition, a mechanism by which organophosphate and carbamate insecticides produce excessive cholinergic effects, is a sensitive end point that can be monitored in humans and other mammals. Brodeur and DuBois (1963) reported that weanling (23-day-old) rats were more susceptible than adults to the acute toxicity of 14 of 15 organophosphates tested. The greater toxicity of parathion in weanling rats was tentatively attributed to deficient hepatic detoxification of parathion and its bioactive oxygen analogue, paraoxon (Gagne and Brodeur, 1972). A comprehensive investigation was reported by Benke and Murphy (1975) in five age groups of male and female Holtzman rats: 1, 12 to 13, 23 to 24, 35 to 40, and 56 to 63 days old. There was a progressive decrease in susceptibility to poisoning by parathion and parathion-methyl with increasing age up to 35 to 40 days for both sexes. Detailed experiments were conducted to determine the influence of aging on metabolic activation of the two compounds, as well as on detoxification systems (e.g., aryl esterase-catalyzed hydrolysis, glutathione-dependent dearylation and dealkylation, and binding in the liver and plasma). Benke and Murphy (1975) concluded that increased detoxification of the active oxygen analogues of parathion and parathion-methyl was largely responsible for the lower acute toxicity of the two insecticides in adult animals. Murphy (1982) subsequently pointed to two other factors that contributed to the lower sensitivity of adult rats to organophosphates: greater binding to noncritical tissue constituents and more rapid catabolism of the parent compounds.

The limited information available suggests that immature humans also experience greater susceptibility to organophosphate- and carbamate-induced cholinesterase inhibition and related effects. In 1976 in Jamaica, 79 people were acutely poisoned as a result of eating parathion-contaminated flour (Diggory et al., 1977). Seventeen of the patients died. Case-fatality ratios were highest (i.e., 40%) among children ranging from newborns to 4 years of age. Zwiener and Ginsburg (1988) presented the clinical histories of 37 infants and children exhibiting moderate to severe organophosphate and carbamate toxicity. Although most of these patients ingested the pesticides, six became intoxicated after playing on sprayed surfaces. Zwiener and Ginsburg (1988) noted that 76% of their subjects were younger than 3 years old. The investigators found there was a paucity of information in the literature on the toxicity of cholinesterase inhibitors in infants and children.

Parathion contamination of stored foodstuffs (Diggory et al., 1977) and aldicarb contamination of crops (Goldman et al., 1990) have resulted in the most widespread outbreaks of foodborne pesticide toxicity in North America. Goldman and co-workers investigated more than 1,000 cases of illness caused by consumption of aldicarb-contaminated watermelons and cucumbers. Unfortunately, infants and children were not studied as a subpopulation at risk. The investigators did calculate doses of aldicarb sulfoxide that produced illness in the general population and estimated that a 10-kg child could readily consume enough of the pesticide on watermelons to experience toxicity. The U.S. Environmental Protection Agency (EPA, 1988) concluded that infants and children are at the greatest risk of acute aldicarb toxicity. This conclusion was based on dietary consumption and contamination patterns, however, rather than on the greater sensitivity of infants and children to this potent cholinesterase inhibitor.

Although immature humans appear to be more susceptible than adults to the acute effects of cholinesterase inhibitors, the age-dependency of this phenomenon is not entirely clear. Some of the most applicable information has been provided by a study of the perinatal development of human blood esterases (Ecobichon and Stephens, 1973). Erythrocyte acetylcholinesterase and plasma pseudocholinesterase and arylesterase activities were measured in premature newborns of varying gestational age as well as in full-term newborns, children of different ages, and adults. Apparent Km values for the three enzymes did not vary significantly with age for a variety of substrates, indicating that the enzyme properties were similar in all age groups. Enzymatic activity, however, did vary significantly with age. Levels of all three enzymes progressively increased during gestation, then rose markedly during the first year of life. Thereafter, erythrocyte cholinesterase and pseudocholinesterase activities increased gradually to adult levels. If one were to assume that one of these peripheral enzymes (e.g., erythrocyte cholinesterase) reflects brain acetylcholinesterase levels, then the most pronounced effects of cholinesterase inhibitors may be expected to occur in newborns, neonates, and infants, since a chemically induced depression of enzymatic activity may be more apparent when baseline cholinesterase levels are relatively low. Ecobichon and Stephens (1973) provided evidence of another mechanism of increased susceptibility of newborns—namely, diminished detoxification capacity (i.e., significantly lower plasma arylesterase and paraoxon hydrolysis activities). Children 2 to 8 years old had slightly lower activities than adults, suggesting that younger children may be somewhat more susceptible to cholinesterase inhibitors. The consequences of brain acetylcholinesterase inhibition on nervous system development and postnatal function remain largely unexplored.

Because of the paucity of data on the age-dependency of acute toxicity of pesticides in humans, the remainder of this section focuses on relative effects of therapeutic agents in pediatric and adult populations. Substantially more information should be available on drugs, due to their common use in all age groups and stringent requirements by the Food and Drug Administration (FDA) for demonstration of safety and efficacy. Data from well-controlled, parallel studies in infants, children, and adults, however, are quite limited for most drugs.

Done et al. (1977) reported what was termed a therapeutic orphan problem—namely, that safety and efficacy for children had not been proved for 78% of new drugs then marketed in the United States. A 1990 survey by the American Academy of Pediatrics revealed that the labeling of 80% of new drugs approved by the FDA between 1984 and 1989 did not include information on pediatric use. The FDA's policy has allowed the marketing of drugs that have been approved for adults but not studied in children, as long as labeling included disclaimers and no instructions about pediatric use. Without adequate information, physicians commonly prescribe such medications for children, possibly placing pediatric populations at increased risk of uncertain efficacy or adverse reactions. The FDA (1992) proposed to amend labeling requirements for prescription drugs to promote their safe and effective use in children. Misunderstandings and concern about legal and ethical implications have limited clinical research in pediatric populations. The newly proposed guidelines provide alternative ways to assess effectiveness and safety in children without necessarily having to conduct comprehensive studies. Results from well-controlled studies in adults can be extrapolated to children under some circumstances, although separate pharmacokinetic studies are needed to establish appropriate pediatric dosage regimens. The intent of the proposed amendment is to provide more complete information on labeling of prescription drugs concerning use and possible hazards for children.

Several instances of severe adverse effect from pharmaceutical agents in pediatric populations have attracted widespread attention. During the 1950s, chloramphenicol produced a pallid cyanosis, which progressed to circulatory collapse and death in some newborns (Sutherland, 1959). This so-called gray baby syndrome has been attributed to the diminished hepatic glucuronide conjugation and renal secretory capacities of newborns. Weiss et al. (1960) reported blood half-lives of 26, 10, and 4 hours for chloramphenicol at birth, at 10 to 16 days of age, and in children 4 to 5 years old, respectively. Thus, there is a substantial increase in chloramphenicol metabolism and excretion capacity during the first days and weeks of life. Decreased metabolic and excretory capacities of newborns and neonates have been associated with exaggerated toxicity of a number of other chemicals, including benzyl alcohol (Gershanik et al., 1982), hexachlorophene (Tyrala et al., 1977), and diazepam (Nau et al., 1984). The hexachlorophene poisonings appeared to be associated with increased percutaneous absorption as well as deficient metabolism in newborns. Floppy infant syndrome in babies born to mothers given diazepam is apparently the result of a number of age-dependent factors, including a smaller volume of distribution and thus greater target organ concentrations of the lipophilic drug due to a smaller adipose tissue volume in newborns, increased amounts of free diazepam due to displacement of the drug from plasma protein binding sites by elevated free fatty acid levels, and a prolonged half-life as a result of diminished oxidative and conjugative metabolism (Warner, 1986). As discussed in Chapter 2, most physiological processes that govern the kinetics of drugs and other chemicals mature during the first year after birth. Indeed, profound changes in some processes (e.g., phase I and II metabolism) occur during the first days and weeks of life (Morselli, 1989). Thus, the most pronounced differences from adults in susceptibility to drug toxicity would be expected in newborns, neonates, and infants; the youngest are most likely to experience the most aberrant responses.

The net effect of immature physiological and biochemical processes on drug efficacy and toxicity is difficult to predict. The various processes mature of different rates and may enhance or offset one another. Local anesthetics provide a good illustration. These drugs are commonly administered to the mother during labor and delivery and may readily enter the maternal circulation and cross the placenta (Tucker and Mather, 1979). Cardiovascular depression and respiratory depression in newborns have occasionally been reported, although subtle neurophysiological impairment and behavioral changes are probably more common consequences (Dodson, 1976; Ostheimer, 1979). Premature and full-term newborns exhibit lower plasma protein binding of local anesthetics. This should result in increased amounts of free drug and a more pronounced pharmacologic response, but the greater volume of distribution in newborns reduces the concentration of drug at sites of action. Rates of hepatic microsomal metabolism and plasma pseudocholinesterase-catalyzed hydrolysis of anesthetics such as procaine are quite low in newborns. This deficit in metabolism, coupled with the larger distribution volume that must be cleared of drug, accounts for the prolonged half-life and long duration of action of lidocaine and its analogues in neonates (Morselli, et al., 1980).

Hepatic metabolism and renal clearance of xenobiotic compounds change dramatically during the first year of life. Phase I metabolic reactions (e.g., oxidation) may rise from one-fifth to one-third of the adult rate during the first 2 to 3 postnatal weeks to two to six times the adult rate (Neims et al., 1976; Morselli, 1989). Different isozymes and enzymes mature at different ages. Certain phase II (e.g., glucuronidation) reactions do not reach adult levels for months, while maturation of alcohol dehydrogenase activity may take as long as 5 years (Kearns and Reed, 1989). The majority of xenobiotics, however, are metabolized most rapidly by individuals between 2 to 4 months and about 3 years of age. Thereafter, drug metabolism gradually declines to adult levels (Warner, 1986). Development of renal function displays a similar age-dependency. Glomerular filtration increases dramatically during the first week of life, approaching and exceeding adult values within 3 to 5 months. Renal tubular secretory and absorptive processes mature more slowly (Kearns and Reed, 1989). Older infants and children, therefore, may be less susceptible than adults to drugs that are metabolized to less toxic, more readily excretable metabolites. Spielberg (1992) noted that clearance of nearly all anticonvulsant drugs is quite limited in newborns, especially premature newborns. Conversely, clearance of such drugs (e.g., phenytoin, phenobarbital, carbamazepine, and diazepam) in infants and children, when calculated on a milligram-per-kilogram-of-body-weight basis, was well above that in adults until around puberty. Thus, children are less likely than adults to exhibit toxicity and require higher doses (on a milligram-per-kilogram-of-body-weight basis) of anticonvulsants to achieve therapeutic levels. In contrast, infants and children may be at greater risk from other drugs and chemicals that undergo metabolic activation (i.e., conversion to bioactive or cytotoxic metabolites). Unfortunately, there is lack of information on such agents in humans in the published literature.

There was concern that acetaminophen (Tylenol), a drug that undergoes metabolic activation to hepatocytotoxic metabolite(s) via a P-450-mediated mixed-function oxidase (MFO) pathway, would cause increased morbidity and mortality in young children. This concern was never realized, however, since hepatotoxicity in young children was found to be less severe than in adults, and has rarely (Rumack, 1984). Acetaminophen is metabolized by several parallel pathways. The two major detoxification pathways involve conjugation of the parent compound with sulfate or glucuronide. Thus only a small fraction of the drug remains to be oxidized by the P-450-mediated pathway to a reactive intermediate (N-acetyl-p-benzoquinonimine). This metabolite is conjugated with glutathione to produce nontoxic products or can bind covalently to cell proteins and nucleic acids, causing cellular injury (Hinson et al., 1990). Although prepubescent children have relatively high hepatic MFO activity, they also exhibit a greater capacity than adults to detoxify acetaminophen by phase II metabolic reactions, primarily sulfate conjugation (Miller et al., 1977). Also, higher glutathione levels in the young may contribute to protection from hepatotoxicity. Thus, the lower susceptibility of children to acetaminophen poisoning is due to their greater capacity to eliminate the drug by nontoxic pathways (Kauffman, 1992).

Clinical trials in infants and children are relatively infrequent for most classes of drugs, but this is not the case for many antineoplastic agents. Although some types of childhood cancer are refractory to chemotheraphy, others have excellent cure rates (Petros and Evans, 1992). Therefore, phase I clinical trials are frequently conducted in both adult and pediatric populations to define the maximum tolerated dose (MTD) for appropriate dosage schedules in phase II trials. Antineoplastic agents include a wide variety of different types of chemicals that act by diverse mechanisms. Thus, results of phase I studies of anticancer drugs afford scientists some of the most comprehensive data sets for contrasting toxic effects of chemicals in children and adults. The investigations typically involve repetitive dosage regimens lasting days or weeks, however, rather than single, acute exposures.

Comparable clinical trials of antineoplastic agents in pediatric and adult patient populations have revealed toxic effects that are often similar qualitatively but different quantitatively (Glaubiger et al., 1982; Marsoni et al., 1985; Evans et al., 1989). In compilation of data on 16 compounds for which there had been comparable phase I trials in adults and children, the types of toxic effects that limited further dosage escalation were generally the same (Glaubiger et al., 1982). As shown in Table 3-1, the MTD for children was higher than that for adults for 13 of the compounds. Similar findings were reported by Marsoni et al. (1985). These investigators compared the MTDs and recommended phase II doses in children and adults for 14 drugs in patients with solid tumors and 8 drugs in patients with acute leukemia. Children with solid tumors exhibited a greater dose tolerance for 12 of the 14 drugs. Children with leukemia appeared to have tolerances similar to those of adults.


TABLE 3-1 Maximum Tolerated Dose (MTD) of Some Anticancer Drugs in Children and Adults

SOURCE: Glaubiger et al., 1982.

Data on daunomycin in relation to the incidence of congestive heart failure in children and adults have been compiled. Children seem to be more sensitive than adults to this complication at comparable doses, even through the MTD is approximately 20% higher in children than in adults.

The greater tolerance of children to many anticancer drugs may be attributable to higher rates of metabolic or renal clearance. Both Glaubiger et al. (1982) and Marsoni et al. (1985) expressed MTDs on a milligram-per-square-meter rather than a milligram-per-kilogram-of body-weight basis. Had the relative doses been calculated as milligram per kilogram, the interage differences should have been even more pronounced. Pinkel (1958) observed that pediatric patients tolerated more methotrexate on a milligram-per-kilogram basis than did adults, but the MTDs were similar when calculated on the basis of body surface area. Methotrexate is eliminated primarily by glomerular filtration and active renal tubular secretion of the parent compound. It is not surprising, therefore, that children with relatively high renal function exhibit greater rates of plasma elimination than do adults (Wang et al., 1979). In a study of 47 patients (3 to 39 years old) receiving methotrexate, Bleyer (1977) found a significantly higher incidence of neurotoxicity in the adults. Conversely, young infants have diminished renal function and exhibit lower systemic clearance and a greater potential for injury than do children (McLeod et al., 1992).

As maturation of xenobiotic metabolism and renal function generally parallel one another during the first year of life, it is not surprising that neonates and young infants may be at increased risk of injury from anticancer drugs that undergo metabolic inactivation. Vincristine is one such drug. It is detoxified in the liver and eliminated primarily via biliary excretion. Woods et al. (1981) reported a significantly higher incidence of neurotoxicity and hepatotoxicity in small infants than in children receiving vincristine. On the other hand, compounds that undergo metabolic activation may place children at greater risk than neonates or adults, since children have a higher metabolic capacity. Marsoni et al. (1985) observed that indicine N-oxide was one of the few anticancer drugs tested to have a lower MTD in children than in adults. Indicine N-oxide is believed to be converted to the toxic metabolite dehydroindicine by the liver. Cyclophosphamide is another drug that undergoes metabolic activation to cytotoxic metabolites. Certain of its metabolic pathways, however, also involve inactivation/detoxification. The half-life of cyclophosphamide is shorter in children (1 to 6.5 hours) than in adults (4 to 10 hours) (Crom et al., 1987). Although metabolic activation of drugs such as cyclophosphamide may be highest in children, the operability of concurrent detoxification pathways and inactivation of the reactive metabolites, coupled with rapid urinary excretion of the metabolites, apparently combine to hasten the elimination and thereby to negate expression of greater toxicity in children.

Because of the rapid increase in human immunodeficiency virus (HIV) positive children and the significant morbidity and mortality of the resultant disease, drugs for HIV treatment are being tested in both pediatric and adult populations. One of the most widely tested anti-HIV drugs is azidothymidine (AZT, Restrovir). McKinney et al. (1991) studied the effects of AZT in 88 children (mean age, 3.9 years; range, 4 months to 11 years). Maha (1992) reported that the efficacy and incidence of side effects (e.g., hematological abnormalities, primarily neutropenia) were similar in both adults and children but noted that the mean duration of therapy was much longer in the cohort of children, suggesting that they tolerated AZT somewhat better than did adults.


Postnatal Effects of Neurotoxicants

Studies in animals suggest that the nature of an injury is determined by the stage of brain development at the time of exposure rather than by the relationship of the insult to the time of the birth event. Measures of brain development (e.g., gross brain weight and measures of biochemical change, physiologic function, and microanatomic structure) indicate that the processes and timing of brain development relative to birth differ among species (Himwich, 1973). These considerations are important in evaluating and comparing neurodevelopmental toxicology data from laboratory animals and human epidemiologic studies, especially when exposures occurred during the prenatal and weanling stages, reflecting different stages of brain development in different species. In humans, significant brain development and structural alteration occur until at least 4 to 6 years of age. It is plausible, therefore, that effects could result from exposures occurring several years after birth.

Studies evaluating microanatomic development of the brain indicate that the numerous brain structures have differing peak periods of growth. Therefore, toxic exposures at a particular time would differentially affect the structures undergoing peak development. Studies in animals indicate that exposures at different stages of brain development have differing effects on brain and behavioral function (Rodier, 1980). These critical periods or windows of vulnerability must be seriously considered when evaluating neurotoxic effects.

Because human brain development continues for years after birth, it can be hypothesized that postnatal exposure to xenobiotic compounds would alter the structure or function of the human nervous system. If this hypothesis is correct, there should be evidence of children suffering measurable effects from neurotoxic exposures at levels that do not affect adults. An alternative hypothesis suggests that children are less vulnerable because of the increased plasticity of the developing brain. In this case, children could be less vulnerable to insult. Unfortunately, the epidemiologic literature on childhood effects of neurotoxins is extremely difficult to evaluate because of the complex nature brain function and because of the multiple factors that affect brain development and confound evaluation.

The data on prenatal and early childhood exposure to lead indicate that effects occur at levels well below those that are toxic to adults (Bellinger et al. 1987). Irradiation studies also suggest vulnerability of the developing brain. Studies on fetal alcohol syndrome (FAS) and on neonatal drug addiction are based on less accurate dose data than are the lead studies, but the occurrence of permanent changes in brain capacity from fetal exposure is strongly suggestive of special vulnerability of the fetus. Damage from a given level of oxygen deprivation (anoxia) is generally more severe for the developing brain than for the mature brain (Menkes, 1981). In certain cases, vulnerability of the infant to neurotoxins may be related not only to the stage of neurologic development but also to the immaturity or failure of various other protective barriers. For example, the vulnerability of the neonatal brain to bilirubin exposure resulting in kernicterus may be related to the immaturity of the so-called blood-brain barrier. Bilirubin concentrations in the 40s (mg/dl) appear to cause no adverse effects in adults, but are not tolerated in children.

The data strongly suggest that exposure to neurotoxic compounds at levels believed to be safe for adults could result in permanent loss of brain function if it occurred during the prenatal and early childhood period of brain development. This information is of particular relevance to dietary exposure to pesticides, since policies that established safe levels of exposures to neurotoxic pesticides for adults could not be assumed to adequately protect a child less than 4 years of age. Knowledge of the degree of variations in neurotoxic dose levels between children and adults is necessary for establishing risk of exposure to the developing brain. Unfortunately, only minimal data are available on the effects of exposure at levels likely to occur in the food supply. The expansion of the knowledge base, particularly the refinement of animal models, is an important first step.

Measuring Neurotoxic Effects in Humans

Techniques for measuring neurotoxic effects attempt to match the various types of neurologic functions (Bondy, 1986; Triebig et al., 1987; Weiss, 1988). Acute severe clinical effects such as seizure, coma, or death are clear, measurable and points, whereas more subtle effects that occur at low exposures must be measured with more sensitive techniques.

Effects involving the peripheral nervous system can be assessed with the use of nerve conduction tests. Stimulus-response times are measured in animals to evaluate more complete reflex arcs. Specific sensory function may be quantified by using vibratory sensitivity measures (Singer et al., 1982; Wu et al., 1985). Specific sensory pathways may be measured by using evoked brain responses for auditory or visual signals (Otto, 1986; Weiss, 1988). Neurotoxic effects can be measured with electroencephalogram (EEG) technology (Dyer and Boyes, 1983; Dyer, 1985) and with biochemical measurements of neurotransmitter and neuroendocrine levels (Healy et al., 1984; Rosecrans et al., 1982; Finkelstein et al., 1988).

Cognitive and behavioral processes can be measured by testing a multiplicity of pathways and functions with methods that evaluate altered behavior in animals or psychological testing in humans. Unfortunately, testing is complicated by the fact that cognitive and behavioral outcomes can be influenced by many factors other than exposure to neurotoxins. Rigorous experimental or statistical designs are necessary to control for such confounding variables (see, for example, Weiss, 1983; Tilson and Mitchell, 1984; Weiss, 1988; Annau, 1990) Behavioral and developmental assessments have been conducted in children and in adults to identify age-related vulnerabilities to neurotoxins (Pearson and Dietrich, 1985). Animal models for behavioral and developmental studies are being evaluated (Buelke-Sam and Mactutus, 1990; Stanton and Spear, 1990; Tyl and Sette, 1990).

Vorhees (1986) attempted to define the areas of behavioral dysfunction that could be affected by prenatal brain damage, stating that behavioral teratogenesis could be expressed as impairment of several categories of neurobehavioral functions (e.g., sensory, cognitive, motor), delayed behavioral maturation of these functions, or other indices of compromised behavioral competence. He further noted that ''the behavioral effects of some teratogens, even if concomitant with physical defects, may be the most significant devastating and noncorrectable of all the effects observed within the syndrome (associated with the teratogen)" (Vorhees, 1986, p.43).

The Lead Model

The most extensive body of data describing the effects of a neurotoxin on the postnatal developing brain pertains to childhood exposure to lead. Progress in this area is reviewed below to illustrate the issues involved in evaluating the developmental effects of neurotoxins.

In the 1970s, researchers found that lead had measurable effects on the behavioral and cognitive function of children (Perino and Erinhart, 1974; Needleman et al., 1979; Graef, 1980) at blood-lead levels (20 to 40 µg/dl) considerably lower than the threshold previously considered to cause clinical lead disease or biochemical effects in adults.

During the 1980s, neurobehavioral and neurotoxic effects of lead exposure were found in children and in the human fetus at progressively lower levels of exposure (Moore et al., 1982; Needleman, 1983; Winneke et al., 1985; Bellinger et al., 1986; Mayer-Popken et al., 1986; Dietrich et al., 1987; Ruff and Bijur, 1989). Exposure levels resulting in blood-lead levels than 20 µg/dl were implicated. Measures of neurophysiologic and neurochemical disturbances (Otto and Reiter, 1984; Alfano and Petit, 1985; Otto et al., 1985; Moore et al., 1986) have supported the findings of toxic effects from exposure to low lead levels in humans (<30 µg/dl) and in animals. By testing for subtle neurologic, cognitive, and behavioral effects in children, these investigators elucidated neurodevelopmental toxic effects of lead.

The research that produced these findings was characterized by a variety of methodological approaches:

• Multifaceted approaches included a range of methods for biochemical and neurophysiologic measurements in animals and (where ethically possible) humans.

• Studies to evaluate subtle neurologic and developmental effects of lead included innovative methods that combined extensive batteries of different psychologic tests and rigorous statistical design focusing on various constructs to control confounding variables.

• Investigators looking for lead toxicity did not assume safe levels of exposure or protective effects. When biochemical alteration in function was found at what was considered to be subclinical levels of exposure, researchers looked for methods that would measure subtle functional changes.

Pesticides as Neurotoxicants

Many classes of compounds are used as pesticides. Some of them are known neurotoxicants. Important subclasses of the substances in use are known to have neurotoxic effects. Organophosphates and carbamates are used for demonstration purposes in this section of the extensive data—not because they present greater potential risk than other compounds.

Data suggest that in addition to short-term effects, there are other neurologic effects of a long-term nature in adult humans. For example, symptoms of organophosphate-induced delayed neurotoxicity have been found several weeks after acute exposure and have continued for many months (Whorton and Obrinsky, 1983; Vasilesque et al., 1984; Cherniak, 1988). An intermediate syndrome starts several days after acute exposure and involves paralytic symptoms for many days (Senanayake and Karalliedde, 1987). Abnormal nerve conduction velocities have also been observed in some settings involving low-level, long-term exposure (Misra et al., 1988). Neurobehavioral and psychiatric effects have been reported in some epidemiologic studies of adult populations (Maizlish et al., 1987) and in studies of adult animals (Overstreet, 1984).

The evidence on chronic effects, particularly neurobehavioral effects of organophosphate and carbamate exposure, is less well established, but is strongly suggestive. Similar to the data on lead, there is strong evidence that acute, high-level exposure results in severe systemic disease caused by biochemical mechanisms that affect the nervous system directly. In addition, the data suggest that more long-term effects may result from exposure and that low-level exposure may have subtle, but measurable, effects on neurologic function.

The emerging data suggest that neurotoxic and behavioral effects may result from low-level chronic exposure to some organophosphate and carbamate pesticides. Sophisticated methods will be required to pursue this line of research. For many other pesticides, the data are far less complete. However, when animal studies have shown that a pesticide functions by disrupting neurologic cellular function and when systemic toxic effects are known to occur after high-level acute exposures, the possibility of low-level chronic neurotoxic and behavioral effects must be considered.

Effects of Pesticides in Children

In reviewing the data on the effects of pesticides, two questions must be addressed: Is there evidence that pesticides cause neurotoxic effects in children after acute exposure to high doses? Is there reason to suspect low-level, long-term developmental effects different from effects in adults?

Acute exposure of children to pesticides and resultant disease similar to neurotoxic effects in adults has been described for a range of pesticides, including organophosphate, carbamates, and organochlorines (Hayes 1970; Mortenson, 1986). Pediatric cases involving neurotoxic effects due to acute exposure continue to be reported for other pesticides (e.g., Roland et al., 1985, who reported on exposure to insect repellents and encephalopathy). Data on children as segments of larger exposed populations have also been reported (e.g., CDC, 1986).

Very few pesticides have been well studied for effects on neurologic development in humans and animals. Studies on polychlorinated biphenyls (PBBs) and polychlorinated biphenyls (PCBs) strongly suggest developmental effects from low-level exposures similar to the effects found for lead.

Data on exposure of humans were generated following a 1973–1974 exposure to PBBs in Michigan. Neuropsychological and developmental data were collected on children who were exposed in utero and during infancy. Physiological testing showed significant differences that were related to measures of body dose (Weil et al., 1981; Seagull, 1983).

In Taiwan, children exposed in utero to PCBs in contaminated cooking oil experienced deficits in developmental testing and abnormalities in behavioral assessment (Rogan et al., 1988). This study did not include good body burden measures, but sample sizes were large, permitting elucidation of more subtle effects.

Data on most compounds are not as extensive as those on PBBs, PCBs, and lead. Nevertheless, the pattern shown in the data on those compounds generate concern about the vulnerability of the developing human brain to any neurotoxic pesticides.

Levels of Pesticides Affecting Children

Although the vulnerability of the developing brain to neurotoxic exposure is of serious concern, it is entirely unclear from the data available whether exposures at levels consistent with usual dietary exposures would pose a substantial risk to the long-term neurologic development of children in general or to particular subgroups of children that are neurologically vulnerable.

It is theoretically possible that certain children with preexisting neurologic conditions such as hyperactivity might be more vulnerable to certain low-level neurotoxic exposures. There has been a scientific controversy surrounding the effects of "food additives" (i.e., dyes, flavors, and sugar) on children diagnosed as hyperactive. Responses vary with the study methodology, but even studies that do show effects do not show that all children in the hyperactive subpopulation are affected. These studies do not quantify effects of trace pesticide exposures, but they do raise the question, What would the dose curve for neurodevelopmental toxicants look like, and would all children be similarly vulnerable?

Comparability of Neurotoxicity Effects in Laboratory Animals

An evaluation of the accuracy with which adverse effects are detected across species (Stanton and Spear, 1990) was included in the proceedings of a workshop on "Similarities and Differences Between Children and Adults: Implications for Risk Assessment," sponsored by the International Life Sciences Institute (Kimmel et al., 1990). Species were subdivided into rodents, nonhuman primates, and humans and compared across several categories of neurobehavioral function (sensory, motivational/arousal, cognitive, motor, social). Such an analysis is extremely complex, and required a meticulously detailed comparison of hundreds of research reports for the seven toxicants considered. Overall, the investigators concluded that despite wide species differences in neurobehavioral functional categories, there was close agreement across species for the neurotoxic agents reviewed. Agents that produced cognitive, motor, and sensory deficits in humans generally resulted in corresponding deficits in laboratory animals. Although this relationship held up well at higher doses, comparability across species at lower doses was more difficult to assess. When the outcome measures were operationally similar, however, effects across species were observed with a high degree of reliability. This observation provides an essential basis for adequately predicting and formulating risk assessment guidelines for agents with potential developmental neurotoxicity.


The primary function of the immune system is to provide resistance to pathogenic agents and surveillance against neoplastic cells. These functions are accomplished by both specific antibodies and cellular components of the immune system. Environmental agents may exert an influence on the immune system by altering cellular function or communication or by serving as a foreign structure and inducing a specific immune response. Altered immune function can result in impaired health by predisposing individuals to infectious disease, malignancy, or autoimmune disease. Because the immune system is not fully developed until adolescence, immunotoxic effects of environmental exposure in children and adults may differ.
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Part 2 of 5

Effects of Environmental Agents on the Immune System

Environmental agents may affect the immune system in a variety of ways. The potential outcomes can be summarized as follows:

• immunosuppression, or depressed function of the immune system;

• altered host resistance against infections or neoplastic agents;

• hypersensitivity, or autoimmune reactivity; and

• uncontrolled proliferation of immune components, such as lymphoma or leukemia (see section, "Carcinogenesis and Mutagenesis," below).

Animal Studies

Most of the studies investigating the effects of pesticides on the immune system have been conducted in animals and have focused on immunosuppression or impaired host resistance following subchronic exposure. For example, host resistance was evaluated in adult Swiss-Webster and B6C3F1 mice following exposure to aldicarb (0.1 to 1,000 ppb) in drinking water (Thomas and Ratajczak, 1988; Thomas et al., 1990). After daily consumption for 34 days, no effect was noted on host resistance to infectious viral challenge, the functional ability of interferon-induced splenic NK cells to lyse YAC-1 lymphoma target cells, or cytotoxic T-cell function. In addition, there was no change in production of splenic antibody resulting from immunization with sheep erythrocytes, no effect on spleen lymphocyte blastogenesis to B- and T-cell mitogens, and no effect on the mixed lymphocyte culture response, blood counts, differential leukocyte counts, body weight, or relative lymphoid organ weights. The studies concluded that no exposure-related immunologic effects resulted from environmentally relevant concentrations of aldicarb.

The immunotoxic effect of sublethal exposure to dieldrin and aminocarb has also been examined (Fournier et al., 1988). Mice were exposed to the pesticides by gavage or intraperitoneal injection of sublethal (<LD50) doses in corn oil or dimethyl sulfoxide on two occasions, then subsequently infected with mouse hepatitis virus (MHV3). Resistance to the viral infection indicated the status of cell-mediated immunity. Dieldrin increased the cumulative mortality of animals, whereas aminocarb did not. In addition, splenic lymphocytes from the dieldrin-treated mice were found to be functionally suppressed, as evidenced by their reduced ability to respond in a mixed lymphocyte culture. Aminocarb-exposed lymphocytes were not affected. These data indicate that cell-mediated immunity may be affected by pesticide exposure.

The immunotoxicity of captan was evaluated in rats and mice following oral administration (LaFarge-Frayssinet and Declöitre, 1982). Animals were fed a diet with or without 0.3% (wt/wt) captan [cis-N-(trichloromethylthio)-4-cyclohexene-1,2-dicarboximide] for 7, 14, 21, and 42 days. After 14 days of treatment, antibody formation was found to be depressed by about 70% in both species. The effect waned by day 42. Other effects noted on day 14 were reduced splenic T- and B-cell proliferation to mitogens. These responses also improved by day 42.

The effects of lindane, malathion, and dichlorophos on the immunocompetence of rabbits were assessed (Dési et al., 1978). Doses of 1/2.5 to 1/40 of the LD50 were given orally, in capsules, five times per week for 5 to 6 weeks. Animals were intravenously immunized weekly with Salmonella typhi, and antibody titers were assessed. Each of the pesticides caused a decreased antibody titer. Depression of red blood cell cholinesterase activity correlated with the immune suppression to show dose response.

Oral ingestion of lindane- and carbaryl-containing food increased antibody production in response to the antigenic stimulus, sheep red blood cells, in mice. However, decreased resistance to infection was noted following feeding of lindane. Duration of giardiasis was increased in mice, although nonreaginic antibody levels to the parasite were elevated (André et al., 1983).

Studies in mice with the organophosphorus pesticide O,O,S-trimethyl phosphorothioate (an impurity in malathion) demonstrated the ability of this chemical to block both generation of cytotoxic T lymphocytes and antibody responses at doses that did not affect body weight or splenic lymphocyte number (Rodgers et al., 1986). The macrophage appeared to be the affected splenic cell type. The suppression was reversible. Recovery time was dependent on the dose administered. A dose of 1 mg/kg was immunotoxic.

By contrast, another malathion impurity, O,S,S-trimethyl phosphorodithioate, was immunostimulatory (Rodgers et al., 1987). At nontoxic doses, mice demonstrated elevated cytotoxic T lymphocyte responses and heightened humoral immune responses.

The immunotoxic effects of the herbicide 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD) have been studied extensively. In laboratory animals, the immune system appears to be a sensitive target organ. Immunosuppression is characterized by depressed cell-mediated immunity, which is most evident after perinatal exposure during the period of thymic organogenesis. The mechanism of immunosuppression in mice appears to be a defect in T-cell regulation, because nude mice (which lack T-cell populations) were more resistant than their normal littermates (Kerkvliet and Brauner, 1987). Exposure of adult animals to a TCDD concentration of 2.7 µg/kg resulted in depressed humoral immunity (Exon, 1984). In animals, the response is dependent on Ah locus, suggesting a genetic basis for susceptibility.

In the rat, the developing immune system has been shown to be more susceptible than the immune system of the adult to the immunotoxic effects of TCDD (Vos and Moore, 1974; Faith and Moore, 1977). Fetal and neonatal rats were exposed to TCDD through maternal dosing (5 µg/kg). The doses were administered by gavage on day 18 of gestation and on days 0, 7, and 14 of postnatal life. At this concentration, TCDD suppressed the developing immune system but not the immune system of the adult (Faith and Moore, 1977). In mice treated only at 1 month of age (not during the fetal or neonatal periods), there was reduced spleen cell response to phytohemagglutinin (PHA), which was not observed in mice treated at 4 months (Kerkvliet and Brauner, 1987). However, this effect was noted only at a toxic level of TCDD.

Few studies have examined the development of hypersensitivity following exposure to pesticides in laboratory animals. Localized dermal sensitivity has been reported for some pesticides such as naled, malathion, captan, Difulatan, DDT, and Omite (Ercegovich, 1973).

Studies in Humans

No studies have been conducted to examine the immunotoxic effects of pesticides on infants or children. Immunologic effects of chronic exposure to aldicarb in adults were investigated as a result of groundwater contamination by this carbamate pesticide in Wisconsin from 1981 to 1985 (Thomas et al., 1990). Levels of >1 to <61 ppb had been measured (enforcement standard for groundwater is 10 ppb). The average aldicarb level in the groundwater was 16.1 ppb. Adult women from 18 to 70 years of age were examined for immune status in 1985. The 23 women who consumed the contaminated groundwater were compared for health status, immune function, and fluid intake with 27 who consumed water with no known contamination. Aldicarb levels in the groundwater samples averaged 16.1 ppb. Results suggested an association between consumption of aldicarb and T-cell subset abnormalities, elevated response to Candida stimulation, increased number of T8 cells, and increased percentage of T8 to T4 cells. The T-cell analyses were repeated on three more occasions and gave reproducible results. Dose-response data indicated a statistically significant association between aldicarb levels (using well-water values from individual households) and T4:T8 abnormalities as well as Candida stimulation results. However, although the stimulation results differed between groups, values for both groups were within normal limits. In addition, there was no self-reported clinical evidence of adverse health effects in the study groups (Thomas et al., 1990).

Health effects in humans from TCDD exposure were examined. In 1971 TCDD-contaminated sludge waste was mixed waste oil and sprayed for dust control on residential, recreational, and commercial areas in eastern Missouri (Hoffman et al., 1986). Some reduction in activities in these areas was recommended in 1982. As a consequence, the longest period of exposure was 11 years. Individuals were exposed at nine residential sites. At least 1 ppb TCDD was found in all soil samples. Levels as high as 2,200 ppb were found in some samples.

The study involved 155 unexposed persons and 154 people exposed for 6 or more months. The exposed group had increased frequencies of abnormal T-cell subsets (10.4% compared with 6.8%). The T4:T8 ratio was less than 1 (8.1% compared with 6.4%). The exposed group had an increased frequency of anergy (11.8% compared with 1.1%) and relative anergy (35.3% compared with 11.8%). Anergy was correlated with the length of time the individual lived in the area. Chloracne was not observed. These results suggest an effect of TCDD exposure on the T-cell component of the immune system; however, the effect did not produce any clinical illness (Hoffman et al., 1986).

Hypersensitivity to pesticides has been examined. Few problems of dermatitis were noted after exposure to DDT and lindane, which were applied to the skin and clothing of individuals to control disease vectors (Ercegovich, 1973). Furthermore, there are no documented reports of sensitization to pesticides as a result of food or environmental exposure, nor are there reports of antibodies in sera from individuals exposed to pesticides, as would be expected if pesticides functioned as haptens and induced allergic responses.


Carcinogenesis is a multistage or multistep process by which a normal cell loses its ability to control its rate of proliferation and differentiation and becomes a cell from which a tumor may arise. These alterations may occur as a result of mutagenesis, which involves direct alteration of the structure of DNA, or as a result of nongenotoxic mechanisms that alter the expression of DNA or indirectly lead to mutagenesis. An increased rate of cell proliferation is an example of an indirect mechanism that can lead to carcinogenesis by increasing the likelihood that spontaneous mutation will occur or by decreasing the time available to repair DNA damage. Children may be more susceptible than adults to carcinogenesis or mutagenesis because as developing organisms, their rates of growth and thus of cell proferation are much greater. Experimental and epidemiologic observations do not always support this, however.

Carcinogenesis in the Developing Organism

Animal Studies

Comparisons of tumor incidence observe in rodents at the same age and at the same dose rate but after different exposure durations indicate that tumor incidence is not solely a function of total accumulated lifetime dose but may depend on age at first exposure as well (Gaylor, 1988). This conclusion is supported by the observations of Toth (1968) and Rice (1979), who reported that in comparison to older animals, newborn and young animals are generally more susceptible to chemically induced tumor induction at some sites (including lung and liver) but are often more resistant to tumors at other sites (such as skin and breast). For example, intraperitoneal injections of the solvent urethane in mice produced a sixfold higher rate of leukemia when treatment was begun shortly after birth than when it was begun at about 45 days of age (Berenblum et al., 1966). Sensitivity to the induction of preneoplastic cells in the pancreas by the antibiotic azaserine is maximal in postnatal rats when the level of pancreatic DNA synthesis is high, whereas treatment is less effective in weanlings and ineffective in adults (Longnecker et al., 1977). When perinatal administration of ethylenethiourea was combined with 2 years of dietary administration to rats and mice, the incidence of thyroid tumors was slightly enhanced as compared to that obtained in the absence of perinatal exposure (NTP, 1992). By contrast, a number of studies do not support the conclusion that younger animals are more susceptible carcinogenesis or mutagenesis than older animals. For example, Greenman (1987) failed to demonstrate an effect of age on 2-acetylaminofluorene-induced bladder cancer in mice and found that younger animals were more resistant to histopathologic changes in both the bladder and the liver. Singh et al. (1986) treated both young and old mice with ethylnitrosourea and observed that genetic alterations in bone marrow cells occurred with a greater frequency among older animals. Methylcholanthrene did not produce skin tumors when applied to new born mice but did produce tumors in 42% of the mice treated as adults (Toth, 1968).

Anisimov (1983) surveyed the literature to determine the effects of aging on tumor latency, incidence, and size at different sites in different species for a variety of chemicals. Although these are not pesticides, the studies provide further evidence of end organ changes with age that may be applicable in the study of pesticide toxicity (Table 3-2). It is apparent from the table that results are contradictory and generelizations are impossible. Skin painting experiments with the dermal carcinogen dimethylbenz-[a]anthracene, for example, showed that younger mice are both more susceptible (Lee and Peto, 1970) and less susceptible (Stenbäck et al., 1981) than older mice to skin tumors. Increasing the age at which the carcinogen diethylnitrosamine was administered to rodents both increased (mice; Clapp et al., 1977) and decreased (rats; Reuber, 1976) the number of esophageal and forestomach tumors observed. The experiments that have been performed in animals to evaluate the effects of aging on susceptibility to chemical carcinogenesis clearly demonstrate that age may be an important factor but do not support the conclusion that younger animals are always more susceptible than older animals.

Cancer risk can thus be a function of age at first exposure, although increasing the age at first exposure does not necessarily decrease susceptibility. One explanation for this inconsistency is that as the number of cells in a target tissue increases with age, the total number of cell divisions may also increase, even if the mitotic rate decreases. There are likely to be a multitude of factors in addition to age and rates of cell proliferation that modulate carcinogenesis.


TABLE 3-2 Effect of Aging on Latency, Incidence, and Size of Tumors at Different Sites

NOTE: MC, 3-methylcholanthrene; BP, benzo[a]pyrene; TC, tobacco smoke condensate; DMBA, 7,12-dimethylbenz[a]anthracene; DBA, 1,2,5,6-dibenzanthracene; MNU, N-nitrosomethylurea; FBAA, N-4(fluorobiphenyl) acetamide; CCI4 carbon tetrachhloride; DMH, 1,2-dimethylhydrazine; AFB1 aflatoxin B1; BMNNG, N-methyl-N'-nitro- N-nitrosoguanidne; DENA, N-nitrosodiethylamine; DMNA, N-nitrosodimethylamine; DMNA, dimethylnitramine; PMS, pregnant mare serum.

SOURCE: Anisimov, 1983.

Increased susceptibility to carcinogenesis at younger ages, when it occurs, may be attributable to two factors: increased rates of cell proliferation and differing metabolic capabilities. The many roles that cell proliferation may play in carcinogenesis are described above; overall, increased rates of cell proliferation can contribute to an increased likelihood of carcinogenesis. For example, polycyclic aromatic hydrocarbons and aflatoxin B1, produce liver tumors when administered to newborn rodents but not when administered to older animals, presumably because the liver proliferates rapidly in the developing organism but slowly in older animals. Differing metabolic capabilities may contribute to greater susceptibility if the developing organism has less competent detoxifying or conjugating abilities than the adult. Conversely, less competent activating enzymes may protect the developing animal from chemicals that require metabolic activation to their reactive forms to elicit effects. Ethylnitrosourea, which does not require metabolic activation, is very effective as a carcinogen in neonatal rodents as compared to adults, whereas diethylnitrosamine, which requires activation, is not (Vesselinovitch et al., 1979). In addition, there may be age-related differences in DNA repair abilities and in the fidelity of DNA replication.

Human Studies

Epidemiologic studies of the effects of age on susceptibility to carcinogenesis are conflicting. The risk of bladder cancer associated with employment in a ''hazardous occupation" (e.g., an industry believed to be associated with an increased risk of bladder cancer, such as the rubber or leather industries, or work with dyestuffs, paint, and other organic chemicals) was greater in younger people (Hoover and Cole, 1973), whereas the risk of nasal cancer among nickel workers increased in proportion to age at beginning of exposure (Doll et al., 1970). Tucker et al. (1987) demonstrated that chemotherapeutic treatment of children with cancer using alkylating agents, which can form adducts with DNA and induce mutations, resulted in a significantly elevated risk of secondary leukemia. No study has been performed to determine whether similar treatment of adults has the same outcome, however, so it is not possible to conclude that children are more susceptible to chemically induced carcinogenesis on the basis of these limited data. Evidence from epidemiologic studies is thus inadequate to demonstrate a consistent increased susceptibility to carcinogenesis among children, nor would one assume that children would regularly be more susceptible to toxic end points in pesticide toxicity. These data emphasize the need to evaluate each pesticide specifically for age-related toxicity. The incidence of most cancers in humans increases with age, with the exception of certain tumor types that are associated with childhood and that are suspected to result from inborn genetic alterations or prenatal genetic damage. An example of a childhood tumor is retinoblastoma, in which a mutation occurs in the retinoblast population resulting from genetic damage either before or after conception, creating a population of altered retinal cells that is very susceptible to malignant transformation.

From 1973 through 1989, the incidence of cancer among children of all races from 0 to 14 years old increased 7.6%. The greatest increases were observed for acute lympocytic leukemia (23.7%), brain and nervous system cancers (28.6%), and cancers of the kidney and renal pelvis (26.9%). The incidence of several other childhood cancers decreased (bones and joints, -15.1%; Hodgkins disease, -1.5%; non-Hodgkins lymphomas, -0.9%). During the same period, total cancer incidence for the entire U.S. population increased approximately 16.1% (Miller et al., 1992).


Data on pharmacokinetics are basic to considerations of the relative risks of toxic injury from pesticides in both children and adults. The fundamental goal of pharmacokinetic studies is to delineate the uptake and disposition of pesticides, drugs, and other chemicals in the body. A basic tenet of toxicology is that toxic responses are a function of the concentration of the active chemical in target tissues. Thus the degree and duration of a toxic effect depend on the quantity of the reactive form of a chemical that reaches its target site and the length of time the agent remains there. These factors in turn depend on the magnitude of systemic absorption, binding, distribution, metabolism, interaction with cellular components, and elimination of the chemical from the tissue and body. The important structural and functional differences between infants and adults can have an impact on one or more of these pharmacokinetic processes, which in turn may result in different effects of chemicals on the two age groups.

This section focuses on age-related factors that influence the pharmacokinetics of pesticides, drugs, and other chemicals in humans. The study subjects are grouped as follows: premature newborns, full-term newborns, neonates (birth to 4 weeks), infants (4 weeks to 1 year), young children (1 to 5 years), older children (6 to 12 years), and adolescents (13 to 18 years). Consideration in this section is largely limited to information from studies in humans, since there are major difficulties in extrapolating from immature animals to immature humans. Nonprimate species are less mature in many respects than humans at birth. Maturation in most lower animals, however, is quite rapid; some adult-like characteristics and functions are attained in as little as 14 to 21 days in rodents. A difference of only a few days in exposure age can thus have a marked effect on the handling of a chemical and its ensuing effects in such species (Done, 1964; Neims et al., 1976). Various body structures and associated functions mature at different rates in different species. Utmost care must be exercised in selecting an appropriate animal model for developmental pharmacokinetic and toxicology studies, in interpreting the data, and in extrapolating the data to humans. Animal studies are presented here when data on humans are inadequate or when findings in animals elucidate ontogenetic mechanisms.

For infants and children, exposure to pesticides occurs primarily through ingestion, inhalation, and through the skin. The newborn may have previously encountered chemical agents in utero, but an in-depth examination of in-utero exposure is beyond the prescribed scope of this report. The major emphasis in this section is the absorption and disposition of ingested chemicals. Because children put all kinds of things into their mouths, they are at risk of ingesting pesticides from nonfood sources, including contaminated household objects, ornamental plants, sod, and paint. In certain situations, significant exposure may result from inhalation of pesticides or skin contact with contaminated surfaces (see Chapter 7). Dermal and inhalation exposures are also addressed because they may contribute to the total systemic dose and need to be considered when establishing prudent levels of dietary intake for infants and children.

Dermal and Pulmonary Exposure

The skin area of the infant per unit of body weight is double that of the adult, whereas the permeability of the infant's skin, except for those born prematurely, appears to be similar to that of the adult. These are important factors to remember when considering dermal absorption or penetration of xenobiotic compounds. The stratum corneum (the outer layer of the skin, which serves as the barrier to penetration by chemicals) is fully developed in the human newborn. Studies of the bacteria-inhibiting agent hexachlorophene in premature and full-term infants, the hormone testosterone in infant and adult monkeys (Wester et al., 1977), and alcohols in premature and full-term infants and human adults have shown no differences in penetration, but differences in absorption have been shown for fatty acids (Wester and Maibach, 1982).

There is little evidence to suggest that percutaneous absorption of chemicals varies greatly with age during the preadolescent period, since the overall thickness of the stratum corneum remains relatively constant throughout postnatal development (Rasmussen, 1979). There is a paucity of information, however, from well-controlled studies on percutaneous absorption of chemicals in this age group. McCormack et al. (1982) observed no difference in the rate of penetration of a series of alcohols through premature, full-term newborn, and adult skin specimens in vitro. They did find differences in penetration of a series of fatty acids, which the investigators attributed to differences in solubilization of the fatty acids in epidermal lipids. Wester et al. (1977) reported that the percutaneous absorption of testosterone was similar in the newborn and adult rhesus monkey.

Some studies have been conducted to assess the age-dependency of dermal absorption of pesticides in rodents. Solomon et al. (1977) found no significant differences between newborn and adult guinea pigs in blood or brain concentrations of the insecticide γ-benzene hexachloride following its topical application. Knaak et al. (1984), however, reported that the fungicide triadimefon was more rapidly absorbed through the skin of young rats than through the skin of adult rats. Shah et al. (1987) contrasted the percutaneous absorption of 14 pesticides in young (33-day-old) versus adult (82-day-old) female rats. The pesticides studied were structurally diverse, in that they included organophosphates, carbamates, organometallics, chlorinated hydrocarbons, biological insecticides, and a triazine compound. No clear age-related pattern of absorption was found. At least four of the compounds (atrazine, carbaryl, chlordecone, and chlorpyrifos) were absorbed more extensively by the younger animals. Six compounds, however, were better absorbed by the adults, and the others seemed to be equally well absorbed by both age groups. No one class of compounds was better absorbed in one particular age group, other than the organometallics, which were more extensively absorbed by the adult rats. Skin penetration was not well correlated with the octanol-water partition coefficients of this diverse group of chemicals (Shah et al., 1987).

The premature human newborn may be a special case. Studies have shown that formation of the stratum corneum is incomplete until just before birth at full term (Singer et al., 1971). Greaves et al. (1975) reported that premature newborns bathed in an antibacterial hexachlorophene solution had considerably higher blood levels of hexachlorophene than did full-term newborns monitored by Curley et al. (1971). Tyrala et al. (1977) similarly observed an inverse relationship between body weight and post-conceptional age versus blood hexachlorophene concentrations in a group of 54 premature and full-term newborns. Subjects with large areas of abraded skin exhibited particularly high blood levels. Thus diminished effectiveness of the epidermal barrier to absorption was apparently a major determinant of the elevated blood levels of hexachlorophene. Another important factor was the reduced capacity of premature and full-term newborns to metabolize and eliminate hexachlorophene. Dermal hexachlorophene exposures have resulted in a number of cases of brain damage in newborns (Powell et al., 1973; Shuman et al., 1974), suggesting that this very early age group could be at increased risk of toxicity from direct skin contact with pesticides.

Several factors may contribute to increased percutaneous absorption and toxicity of pesticides in neonates and infants. Dermal absorption of a variety of chemicals is markedly increased under diapers and rubber pants. These materials retard the evaporation of volatile chemicals and enhance the hydration and temperature of the skin, thereby increasing penetration by water-soluble chemicals. Damage to the stratum corneum, as in diaper rash, circumvents this barrier layer. In addition, some skin surfaces such as the male scrotum and the face are more absorbent than skin in other areas of the body.

The ratio of surface area to body weight in newborns and infants is approximately 2.5-fold greater than that of adults. Thus, if the exposed area of skin and percutaneous absorption rate in a neonate and adult were equivalent, the neonate would receive almost three times the systemic dose, on a kilogram-of-body-weight basis (Wester and Maibach, 1982). Nevertheless, there are few data to indicate which types of chemicals may be more extensively absorbed through the skin of neonates and infants.

The alveolar epithelium is another potential portal of entry into the body for pesticides. Most xenobiotics are absorbed through the alveolar epithelium into the pulmonary (blood) circulation by simple passive diffusion, rather than specific active transport processes (Schanker, 1978). Therefore, changes in several parameters with age may be of consequence in pulmonary absorption of pesticides, including alveolar surface area, thickness of alveolar membranes, porosity and other properties of the membranes, pulmonary blood flow, and respiratory volume. The normal respiratory volume of the resting infant is approximately twice that of the resting adult, when expressed per unit of body weight. The structural development of the human lung is known to continue postnatally (Hislop and Reid, 1981; Langston, 1983). There is a marked increase in alveolar surface area for the first 18 to 24 months of life. Thereafter, pulmonary structures continue to increase in size, and alveolar surface area increases gradually throughout childhood (Thurlbeck, 1982). There is a progressive increase up to 18 years of age in collagenous elastic fiber bundles in the alveolar walls. Some of these factors may counteract others in terms of their influence on the pulmonary absorption of chemicals. For example, the effect of increased respiratory volume may be offset by the infant's smaller surface area for absorption. Unfortunately, little information is available on the pulmonary absorption and bioavailability of xenobiotic compounds in infants and children.

One group of investigators has studied the pulmonary absorption of nonvolatile drugs in neonatal and adult animals. Hemberger and Schanker (1978) injected measured doses of a series of drugs into the tracheas of neonatal (3 to 27 days of age) and adult rats. Systemic absorption was determined by assay of the quantity of drug remaining in the lungs and trachea after 2 hours. Lipid-soluble compounds (procainamide and sulfisoxazole) were absorbed at similar rates by neonates and adults, indicating that the properties of the alveolar epithelium do not change substantially with age. Lipid-insoluble compounds (p-aminohippuric acid, mannitol, and tetraethylammonium bromide), however, were absorbed about twice as readily by the 3- to 12-day-old rats as by rats 18 days old or older. The lipid-soluble drugs were absorbed much more rapidly than the lipid-insoluble drugs, but the amount of all drugs absorbed per unit of time was directly proportional to the administered concentration, leading Hemberger and Schanker (1978) to conclude that there was no evidence for an absorption mechanism other than simple diffusion in either the neonate or the adult rats. Lipid-insoluble substances are believed to be absorbed in the lung primarily by diffusion through aqueous channels, or pores (Schanker, 1978). Because the alveolar membrane is reported to become thinner in neonatal rats as they age (Burri et al., 1974), Hemberger and Schanker (1978) concluded that greater porosity must account for the greater absorption of lipid-insoluble drugs in neonatal rats.
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Part 3 of 5

Oral Exposure

The gastrointestinal tract is the major portal of entry of pesticides into the body. Absorption depends on the physical and chemical properties of the pesticide, as well as on conditions within the gastrointestinal tract itself. Some of the more important conditions, or factors, include gastric emptying and intestinal motility, gut flora, acid and enzyme secretory activity, mucosal structure and surface area, cellular transport systems, and gastrointestinal blood supply. All these change with postnatal development. A change in one may tend to enhance chemical absorption, whereas a change in another may have the opposite effect. The complexity of the system, with its multiple, interrelating factors, makes it difficult to predict the net effect of maturation on absorption of pesticides and other chemicals.


The majority of information on gastrointestinal absorption of chemicals in infants and children has come from nutrition and drug studies. The musculature of the gastrointestinal tract is relatively thin and quiescent before birth but then must rapidly make the transition from placental to intestinal nutrition (Balistreri, 1988). Ingested nutrients entering the gut and regulatory hormones secreted in response to food are important in the adaptation to extrauterine life. For example, plasma levels of the peptide YY increase substantially after consumption of milk during the first 2 weeks of life (Adrian et al., 1986). Peptide YY reduces the rate of gastric emptying and slows intestinal transit, thereby increasing the efficiency of absorption of nutrients. The rate and extent of absorption of orally administered drugs are often quite variable in newborns, largely because of irregular, unpredictable gastric emptying and peristalsis (Warner, 1986). Some drugs (such as digoxin and diazepam) appear to be absorbed as efficiently by full-term newborns and neonates as by adults. Premature and full-term newborns, however, exhibit slow, incomplete oral absorption of other drugs, including phenobarbital, chloramphenicol, rifampin, valproate, and phenytoin (Morselli et al., 1980; Morselli, 1989). Although poor bioavailability of phenytoin is generally seen during the first month of life, infants absorb the drug well. Similarly, infants absorb valproate at a rate comparable to that of adults.

Absorption of Proteins and Related Large Molecules. The microvillus surface of the adult small intestine serves as a barrier to penetration by many substances, particularly large, charged molecules such as proteins. There is evidence from human and animal studies, however, that the immature intestine can allow passage of intact macromolecules. Antibodies, including immunoglobulin, bovine serum albumin, α-lactalbumin, and other antigens, are reported to be absorbed intact by newborns (Grand et al., 1976). Acquisition of both allergies and passive immunity has been attributed to gastrointestinal absorption of intact proteins by neonates (see section on "Immunotoxicity," above). The period of macromolecular uptake varies in laboratory animals, from as short as 1 to 2 days after birth in guinea pigs to as long as 23 to 24 days in rabbits (Hoffmann, 1982). Selective absorption of intact γ-globulins from colostrum occurs in ruminant species for only a few days, but absorption is retained by suckling rats for up to 20 days. Absorption of α-globulins and other proteins such as cholera toxin has been shown to involve high-affinity binding to specific receptors on the microvillus surface of intestinal mucosal cells in rodents. The binding is followed by invagination of the membrane to form protein-filled vesicles, which migrate through the cytoplasm to the basolateral surface of the cell, where the vesicle contents are extruded (Walker and Isselbacher, 1974). Because this energy-dependent, endocytotic process involves receptor binding of specific proteins, it is unlikely that pesticides would be absorbed in significant quantities in human neonates by this mechanism. And, indeed, most data indicate that uptake of proteins in humans is limited and nonselective, rather than a specific, receptor-dependent transport process as is found in the rat (Walker and Isselbacher, 1974). Because macromolecular uptake in human newborns appears to be nonselective, there could be a transitory period during which compounds that are normally excluded by the gastrointestinal tract of children and adults are absorbed.

Age-Related Changes in Absorption. Although there is a paucity of data on absorption of drugs as related to age in human neonates and infants, some definitive investigations have been conducted using laboratory animals. Hoffmann (1982) summarized the results of a number of animal studies. The rate of penetration of four model compounds (antipyrine, sodium salicylate, tetraethylammonium bromide, and phenosulfonphthaline) from everted intestinal sacs of rats was evaluated in an in vitro experiment. Penetration of all compounds was two to three times greater for 10- than for 30-day-oldrats, which in turn exhibited somewhat greater penetration rates than did adult animals. In an in vivo study, disappearance of the compounds from the duodenum of anesthetized rats was examined. Antipyrine, sodium salicylate, and tetraethylammonium bromide were each absorbed more rapidly by 10-day-old rats than by adult rats. These differences in the rate of absorption were not reflected by differences in blood levels, apparently because of a greater volume of distribution in the neonatal rats. Large synthetic molecules such as polyvinylpyrrolidinone have also been found to be well absorbed in neonatal and suckling mammals, as have heavy metals (Hoffmann, 1982). Closure, or decrease to the low-level uptake characteristic of adult animals, generally has been found to occur at the time of weaning.

Closure has been associated with structural and functional maturation of intestinal epithelial cells. It has been suggested that a marked reduction in pinocytotic activity is largely responsible for this phenomenon. Pinocytosis is believed to be the primary mechanism for nonselective absorption of macromolecules by the neonatal intestinal epithelium in mammals (Lecce, 1972). Bierring et al. (1964) observed large numbers of pinocytotic vacuoles associated with phagosomes at the base of microvilli in the intestinal epithelial cells of human fetuses. Udall and Walker (1982) saw such vacuoles associated with an extensive apical tubular network system, as well as pseudopodlike cytoplasmic extensions projecting through the lamina propria in the intestine of 1-week-old rabbits. The intestinal epithelium of the adult rabbit showed a marked decrease in the tubular network and pseudopods, which accompanied cessation of systemic uptake of bovine serum albumin. Pang et al. (1983) found that the membrane of newborn rabbits had a significantly higher lipid-to-protein ratio than did that of adult animals. Electron spin resonance spectra revealed that the membranes from the newborn rabbits were more disorganized and fluid, which could account for the more efficient penetration and diffusion of macromolecules during the perinatal period.

Absorption and Retention of Lead and Other Heavy Metals Although lead and other heavy metals are not commonly used now as pesticides by themselves, they have been used in the past. Lead, in particular, has been widely used and studied. Lead is still of major concern as a health hazard for a number of reasons, including its prevalence in the environment and the increased sensitivity of children to its adverse effects. Barltrop (1965) reported one of the earliest investigations indicating that blood lead concentrations were age related. In a study group of 470 London children, lead levels in blood increased to a maximum in 3 year olds and then progressively decreased during the next 4 to 5 years of life. Subsequent studies demonstrated that young children had higher lead blood levels than adults living with the children (Barltrop et al., 1974; McNeil et al., 1975).

Research findings for both laboratory animals and humans indicate that increased absorption and retention of lead are important factors in the relatively high incidence of lead toxicity in the young. Suckling rats were shown to absorb a greater percentage of ingested lead than did older rats (Kostial et al., 1971). The increased absorption of lead and other divalent cations, including iron, calcium, cadmium, mercury, and manganese, typically lasted until the animals were weaned (Kostial et al., 1978). Ziegler et al. (1978) conducted metabolic balance studies in 14- to 746-day-old human subjects consuming cow's milk, infant formula, fruit juice, and strained fruits or vegetables that contained known small amounts of lead. Absorption and retention of lead increased with increasing dietary lead intake. Net absorption and retention averaged approximately 42% and 32%, respectively, when intake exceeded 5 µg/kg bw/day in 61 balance studies conducted by Ziegler et al. (1978). Data of other investigators indicate that human adults absorb only about 10% of ingested lead (Hursh and Suomela, 1968; Rabinowitz et al., 1976). Thus, greater gastrointestinal absorption, in concert with the increased retention and target organ deposition, appears to contribute significantly to the higher risk of lead poisoning in infants and young children.

Increased lead absorption in the very young is commonly attributed to pronounced pinocytotic activity in the gastrointestinal epithelium, but there appear to be other determinants as well. Because absorption decreases substantially when animals are weaned, Kostial et al. (1978) studied the effect of cow's milk on lead uptake. Inclusion of milk in the diet resulted in a substantially greater absorption not only of lead but also of cadmium, mercury, and manganese in 6- or 18-week-old rats. Still, absorption of the metals by the milk-fed animals was not as great as in 1-week-old suckling rats. The mechanism of milk's effect is unknown, although it is possible that the metals may bind to some milk constituent, and this binding would facilitate their penetration of the gastrointestinal mucosal barrier. Barltrop (1982) reported that raising dietary fat content from 5% to 10% enhanced lead absorption by 80% in rats and proposed that certain protein deficiencies could result in enhanced absorption because protein diets with a high sulfhydryl content impaired gastrointestinal mucosal binding and uptake of lead. Trace element deficiencies may also play a role. Six and Goyer (1972) demonstrated that iron deficiency resulted in increased lead deposition and toxicity in young rats, possibly because of increased lead absorption. Iron deficiency has been commonly associated with lead poisoning in children. Low dietary intake of calcium has also been shown to lead to greater lead uptake in immature animals and humans (Ziegler et al., 1978).

Other Factors Affecting Oral Absorption

There are additional physiological and morphological processes undergoing continuous maturational changes after birth that can affect gastrointestinal absorption of metals and other chemicals. The microflora of the gut changes considerably during the neonatal and infancy period (Long and Swenson, 1977). The fecal flora in milk-fed infants exhibit negligible demethylating ability, in contrast to that of weaned children and adults. This difference could be important for a chemical such as methylmercury, which is absorbed much more readily than inorganic mercury. Infants would be expected to absorb more of an ingested dose of methylmercury because much of it would not be demethylated in the gut (Rowland et al., 1983). The same would be true for the methylmercury that reenters the gastrointestinal tract via the bile.

Gastric pH varies considerably, falling during the initial hours after birth but returning to neutrality for 10 to 15 days; thereafter, it declines gradually, not reaching adult levels until about 2 years of age (Morselli et al., 1980). Agunod et al. (1969) observed that chemically stimulated hydrochloric acid secretion was low in the gastric juice of neonates but increased until it approached the lower limit of the normal range for adults after 3 months.

Achlorhydria may result in diminished absorption and bioavailability of acidic compounds. The converse should be true for basic compounds. The relatively small surface area of the neonatal intestine proportionally reduces absorption of all chemicals. Although villi and microvilli are present in the intestinal epithelium of newborns, cell proliferation is apparently quite slow (Grand et al., 1976). Autoradiographic experiments in rats demonstrated clearly that the villi of the small intestine of sucklings were shorter and that epithelial cell migration proceeded at a rate of 20%, or less than that, of weaned animals (Koldovsky et al., 1966). Varga and Csaky (1976) found that the blood supply to the gastrointestinal tract of rats changed with age. Fractional blood flow to the total gastrointestinal tract decreased from 20% in 20-day-old rats to 8% to 12% in adult animals.

The net effect of the aforementioned factors on pesticide absorption in the immature individual is hard to predict because they sometimes oppose one another, change at different rates in the maturing organism, and often are ill-defined in humans.

Distribution and Uptake of Chemicals

Distribution of a chemical to sites of action in different tissues, following systemic absorption, is governed by a number of factors. These include plasma protein binding, extracellular fluid volume, adipose tissue mass, organ blood flow, tissue uptake, and tissue binding. The factors exert their influence concurrently and may compete with one another. They may change to varying degrees at different rates during postnatal development. Thus, the net effect of maturation on the quantity of a particular chemical reaching a target tissue is difficult to ascertain. There are few data on binding and distribution of pesticides in infants and children, but a variety of drugs have been relatively well studied (Kearns and Reed, 1989). These are used to illustrate how chemical distribution and uptake can vary with age in the developing individual.

Protein Binding Numerous chemicals, including a variety of pesticides, bind reversibly to plasma proteins. As long as the compounds are bound, they are not able to leave the bloodstream and reach sites of action (i.e., produce biological effects) in extravascular tissues. Although increased plasma protein binding thereby generally reduces the maximum bioactivity of chemicals, binding can prolong their effects by slowly releasing them to sites of action as well as to sites of inactivation or elimination. Much of our current knowledge of how altered plasma protein binding affects the health of infants and children has been gained from clinical studies of therapeutic agents.

Many drugs exhibit significantly lower plasma protein binding in premature and full-term newborns than in adults (Table 3-3). Investigators typically take blood samples from the umbilical cord at the time of delivery and determine the percentage of free, or unbound, drug in the sample. A diverse group of therapeutic agents, including both acidic and basic compounds, exhibit considerably reduced plasma protein binding in perinatal subjects (Morselli et al., 1980; Morselli, 1989). This group includes such common drugs as phenobarbital, digoxin, theophylline, phenytoin, lidocaine, imipramine, and diazepam. Though data for certain age groups are lacking for a number of these drugs, it appears that the age at which binding reaches adult levels is compound-specific. Rane et al. (1971), in a comprehensive study of the age-dependency of the plasma protein binding of phenytoin (a weak acid), found that adult values were approached in infants and young children 3 months to 2 years old. The percent of unbound phenytoin diminished little thereafter in succeeding age groups. Values comparable to those of adults for binding of acidic drugs are often reached during the second to third year of life, whereas γ-globulins, which are believed to be important in binding of nonacidic compounds, may not attain adult levels until ages 7 to 12 years (Morselli et al.,1980). Reduced binding of imipramine has been reported in children younger than 10 years (Windorfer et al., 1974).

TABLE 3-3 Protein Binding of Some Drugs in Cord Plasma in Relation to Adult Plasma

Lower Binding / Higher Binding

Acid Drugs / --

Ampicillin / Valporic acida (indirect evidence)

Benzylpenicillin / --

Nafcillin / Salicyclic acida

Naproxen / Sulfisoxazolea

Salicylates / Cloxacillina

Phenytoin (same or lower binding) / Flucloxacillina

Phenylbutazone / --

Phenobarbitone / --

Pentobarbitone / --

Cloxacillin / --

Flucloxacillin / --

Sulfamethoxypyrazine / --

Sulfaphenazole / --

Sulfadimethoxine / --

Sulfamethoxydiazine / --

Neutral drugs / --

Digoxin / --

Dexamethasone / --

Basic drugs / --

Diazepam / Diazepama

Imipramine / --

Desmethylimipramine / --

Bupivacaine / --

Lidocaine / --

Propranolol / --

Metocurinea / --

D-Tubocurarinea / --

a As compared with maternal plasma.

SOURCE: Adapted from Rane, 1992.

Quantitative and qualitative differences in circulating plasma proteins have been well documented (Morselli, 1976), and it is known that during the perinatal period there is a decrease in plasma protein binding. There appear to be four primary reasons for this decrease. First, the concentration of albumin, the most important binding protein for most compounds, is reduced. Second, there is a persistence of fetal albumin, which has a lower affinity for many drugs. Third, levels, of γ-globulins and lipoproteins are also low postnatally. Fourth, a transient hyperbilirubinemia is universally present during the first few days after birth (Done, 1964; Nau et al., 1984). Bilirubin competes with drugs, particularly acidic ones, for albumin-binding sites. Rane et al. (1971), for example, demonstrated a correlation between the unbound fraction of phenytoin in umbilical cord blood and the total concentration of bilirubin in plasma of neonates. Conversely, large doses of highly bound drugs can displace bilirubin from plasma proteins, resulting in jaundice. Shortly after birth, lipolysis occurs, resulting in an elevation in free fatty acids in the blood. Free fatty acids compete with and can displace some drugs from plasma protein-binding sites. Nau et al. (1984) conducted a comprehensive study in which they measured binding of diazepam and N-desmethyldiazepam, its major active metabolite, in the serum of adults, fetuses, and neonates 1 to 11 days old. The free fatty acid concentration and free fraction of diazepam and its metabolite were highest on day 1. During the next 10 days of life, there were progressive, parallel decreases in serum free fatty acids, albumin, and free fraction of drug and metabolite. Bilirubin seemed to play a less important role in diazepam binding because it was relatively low on day 1 (the day the free fraction was highest) and peaked on day 3 during the time the free fraction was diminishing. Nau et al. (1984) concluded that increased plasma free fatty acids and albumin and decreased bilirubin levels, coupled with deficient metabolism and elimination of diazepam, predisposed the newborn to excessive, potentially adverse effects of the drug.

The extent of plasma protein binding can have a major impact on the magnitude and duration of chemical action and toxicity. Diminished binding, as mentioned previously, results in higher concentrations of free drug available for diffusion from the blood to sites of action in target tissues and sites of metabolism and/or excretion. Thus, diminished binding typically results in an increased intensity of pharmacological effects and potential for toxicity but a shorter duration of action. Clearance is directly proportional to the free fraction of chemical for compounds for which elimination is dependent on diffusion across cell membranes (e.g., into metabolizing liver cells) or glomerular filtration (and urinary excretion). Decreased binding will therefore normally result in increased elimination of a chemical. The neonate, however, often exhibits compromised hepatic metabolic and renal clearance capacities, so the duration of biological effects may be longer than anticipated.

Distribution Volumes The distribution of xenobiotics and many natural compounds in the body is known to change with age. As mentioned previously, infants have a higher percentage of water in lean body tissues than do adults. The additional water is primarily extracellular, so that the volume of extracellular fluid per unit of body weight in infants is about twice that of adults (Widdowson and Dickerson, 1964). As a result, water-soluble chemicals have a greater volume (per unit of body weight) in which to distribute. The newborn is exceptionally resistant to the skeletal muscle relaxants succinylcholine and decamethonium. This resistance can be attributed to the distribution of these small, highly ionized molecules in the relatively large extracellular fluid volume, which in effect reduces their concentration and their resulting pharmacologic action. Penicillins have a higher volume of distribution* in neonates because of lower plasma protein binding and higher extracellular water content. Although decreased plasma protein binding of a drug such as lidocaine would be expected to result in an increased amount of free drug and thus an exaggerated pharmacologic response in newborns, its greater volume of distribution reduces its concentration at the site of action. One age-related difference negates the other. Morselli et al. (1980) pointed out that newborns need twice as long as adults to eliminate lidocaine because of the large distribution volume that must be cleared of the drug. In the newborn, the decreased metabolism of lidocaine is apparently offset by increased renal clearance; a low glomerular filtration rate is offset by diminished tubular reabsorption of the drug. Therefore, despite a number of age-dependent differences that affect pharmacokinetics, in this instance the differences nullify one another such that lidocaine's total body clearance and pharmacologic potency are comparable in newborns and adults.

Most drugs have a larger volume of distribution during infancy and early childhood, although the reverse is true for some other compounds (Done et al., 1977). The decreased binding and increased extracellular fluid volume typically seen postnatally result in a greater volume of distribution for relatively polar chemicals. Many drugs, however, have volume of distribution values in excess of the extracellular fluid volume or total body water as a result of their solubility in body fat. Adipose tissue usually makes up a smaller percentage of body weight in newborns and infants than in adults. Thus diazepam, a lipophilic drug, has a somewhat smaller volume of distribution in neonates and infants than in adults (Morselli et al.,1980). Distribution volume differences, which may be observed for the first 10 years of life, appear to disappear more slowly than most other age-dependent differences that alter pharmacokinetics.

* Volume of distribution is an apparent volume based on the dose administered divided by the concentration in the plasma water (e.g., dose = 100 mg/kg; plasma water concentrations = 0.5 mg/ml; volume of distribution = 100 mg per kg/0.5 mg per ml = 200 ml/kg).

Barriers to Distribution Barriers to tissue uptake, as well as distribution within organs, may vary with age, in conjunction with morphological and functional maturation. In most regions of the body only the vascular endothelium serves as a barrier to diffusion of chemicals from the bloodstream into surrounding tissue. Generally, diffusion is limited to the un-ionized, more lipid-soluble form of chemicals. In some organs (such as the liver), there are pores in the endothelium and gaps between adjacent endothelial cells, which facilitate passage of large, charged molecules. The vascular endothelium of certain organs (such as the brain and testes), however, is devoid of pores and pinocytotic vesicles, has tight cell junctions, and is encased in specialized pericapillary cells. The blood-brain barrier limits entry into the central nervous system to un-ionized, lipophilic compounds (Benet et al., 1990).

There is speculation that neonates and infants may be more susceptible to chemically induced neurotoxicity, in part because of the immaturity of their blood-brain barrier. Watanabe et al. (1990) point out that the central nervous system in developing individuals is potentially vulnerable to chemicals for a protracted period because the central nervous system requires longer than most other organ systems for cellular differentiation, growth, and functional organization. Therefore, any increase in accessibility to cytotoxic agents because of delayed maturation of the blood-brain barrier could have serious consequences.

One of the most commonly cited examples of this phenomenon is lead poisoning in infants and children. It is argued that children exhibit neurological disturbances at lower blood lead concentrations than adults, suggesting that lead enters the central nervous system of children in larger amounts (Barltrop, 1982). Although data on humans are lacking, laboratory studies show that heavy metals accumulate in the brain of immature animals in much greater amounts than in adults (Jugo, 1977; Kostial et al., 1978). These investigators found substantially higher levels of lead, mercury, and manganese in the brains of 1- and 2-week-old rats than in older animals given the agents by intravenous or intraperitoneal injection. The greater toxicity of morphine in immature rats was associated with higher brain-to-blood ratios (Kupferberg and Way, 1963). Pylkko and Woodbury (1961) attributed changes in the convulsant effects of strychnine and brucine in rats to maturation of the blood-brain barrier. The time of maturation of the blood-brain barrier in the rat appears to vary, ranging from about 1 week for 5,7-dihydroxytryptamine (Sachs and Jonsson, 1975) to as long as 3 weeks for cadmium (Wong and Klaassen, 1980). It is not known when this barrier becomes fully functional in humans.


Immature humans and laboratory animals typically exhibit greater systemic retention of heavy metals than do adults. Ziegler et al. (1978) found that retention of lead in human infants (14 to 746 days old) was dose dependent. Although urinary and fecal excretion of lead increased with dose, they apparently could not compensate for lead intake at higher doses. Investigations have shown greater whole body retention, elevated blood levels, and higher target organ concentrations in young animals than in older animals given heavy metals by injection (Kostial et al., 1978; Wong and Klaassen, 1980). Possible explanations for these age-related differences include diminished excretory capacity, high growth rates and high rate of protein synthesis, altered binding to proteins and other ligands in tissues, higher extracellular fluid volume, and greater permeability of tissue barriers in the immature organism. The brain and testes, two organs believed to have effective barriers to ionized molecules in the adult, exhibited substantially higher levels of lead, mercury (Kostial et al., 1978), and cadmium (Wong and Klaassen, 1980) in young rats. Conversely, the kidneys, another target organ, contained smaller amounts of these compounds in the young animals. Such differences may have significant toxicological implications. Data on tissue distribution as it relates to age are generally lacking for most other classes of chemicals.

Metabolism of Xenobiotic Compounds

The human newborn exhibits decreased capacity to metabolize a variety of drugs and other xenobiotic compounds (Warner, 1986; Reed and Besunder, 1989). The premature newborn is usually more deficient than the full-term newborn, although metabolic functions increase rapidly in both during the initial days after birth. A deficiency in MFO activity does not necessarily entail greater susceptibility to toxicity; indeed, it may have the opposite effect (Done, 1964). Inefficient metabolism would make the newborn more susceptible to the action of compounds that are converted to less active, more readily excreted metabolites. Conversely, inefficient metabolism should confer protection against the compounds that are metabolically activated (i.e., converted to reactive, cytotoxic metabolites). Human fetuses and newborns have higher MFO activity, in relation to adult values, than do nonprimates (Neims et al., 1976). Thus, human newborns should be more sensitive to chemicals requiring metabolic activation, whereas most laboratory animals should be more sensitive to chemicals that are detoxified and eliminated via the MFO system.


TABLE 3-4 Risk Assessment for Infants and Children: Pharmacokinetic Factors

SOURCE: Adapted from Rane, 1992.

It is difficult to generalize about age-dependent deficiencies in the metabolism of xenobiotic compounds because different enzymatic pathways seem to exhibit dissimilar maturational patterns (Neims, 1982). There are a number of forms of cytochrome P-450 in the human liver that appear to have distinctive substrate specificity and unique developmental patterns. Pelkonen et al. (1973) found that fetal liver was much more deficient in aryl hydroxylase activity than in other hepatic microsomal monooxygenases. Metabolism of caffeine and theophylline, which initially undergo N-demethylation, is particularly slow in the human newborn. Neonates exhibit a plasma half-life for caffeine of approximately 4 days, as compared to 4 hours for adults (Aldridge et al., 1979). These investigators observed that adult levels and patterns of caffeine metabolism were reached at 7 to 9 months of age. Glucuronidation is one of the most inefficient pathways for metabolism during early development and may take the longest to mature (Done et al., 1977). In contrast, mixed-function-catalyzed oxidation of a number of other drugs increases rapidly during the first days of life, soon approaching and exceeding adult values (Neims, 1982). Thus, the ontogeny of metabolism of the xenobiotic compounds, and its implications in toxicology, is quite compound-specific.

The time course of postnatal development of MFOs has been delineated for several drugs (see Table 3-4). Loughnan et al. (1977) measured the plasma half-life of phenytoin in 2-day- to 96-week-old subjects given the drug intravenously. Although the half-life was prolonged and variable during the first week of life for full-term newborns, the premature newborn exhibited even longer and more inconsistent values. Adult values seemed to be reached by 7 weeks of age, but the number of study subjects more than 1 week old was too small to be definitive. Neims et al. (1976) used the data of several investigators to estimate phenytoin half-life values as a function of age. A marked increase in phenytoin metabolism (i.e., a decrease in half-life) was observed during the first few days postnatally, followed by a progressive increase to levels two- to threefold greater than those of adults in neonates and infants more than 2 weeks old. The large interindividual variability seen in newborns diminished with age. A similar, rapid increase in metabolism and a decrease in variability with age were seen when data for phenobarbital were evaluated (Neims et al., 1976). Both drugs are metabolized by aromatic hydroxylation. Clearance of many drugs, when normalized to unit body weight, is two- to fourfold greater in infants and young children than in older children and adults. As illustrated in Figure 3-1, plasma levels and the dose of theophylline required to maintain therapeutic levels vary substantially with age. Such increases in metabolic clearance, evident for many drugs from the age of 2 to 3 months to the age of 2 to 3 years, tend to decline gradually during childhood until adult values are reached (Morselli et al., 1980).

Metabolism of xenobiotic compounds in the newborn and neonate may be both qualitatively and quantitatively different from that in the adult. Although chloramphenicol is primarily metabolized by hydrolysis and by glucuronidation, a glycolic acid derivative not found in adults has been identified in newborns (Morselli et al., 1980). Premature newborns, unlike infants and adults, exhibit substantial N-methylase activity. This enzyme, acting in the presence of N-demethylase deficiency, can convert theophylline to caffeine. This process is the opposite of what occurs in adults. Aldridge et al. (1979) evaluated the pattern of caffeine metabolites as affected by age in neonates and infants. The researchers found that the proportion of individual metabolites varied until an adultlike metabolite pattern was reached at 7 to 9 months. The toxicological implications of age-dependent qualitative differences in the metabolism of xenobiotic compounds are for the most part unknown.

Enzyme Development. The ontogeny of metabolism of xenobiotic compounds in humans remains largely unexplored. The number of investigations undertaken during fetal and neonatal periods is largely dictated by the availability of tissue. Accordingly, there have been a considerable number of studies of fetal hepatic enzymic differentiation during the first and second trimesters, but relatively few studies during the third trimester and postnatally (Rane and Sjoqvist, 1972; Neims et al., 1976). The major components of the monooxygenase systems are present in fetal liver during midgestation. Smooth endoplasmic reticulum, the principal site of localization of the MFO system in the hepatocyte, is also present at this time (Gillette and Stripp, 1975). Levels of cytochrome P-450 and other monooxygenase components appear to remain relatively constant through parturition. Measurements at term have shown that P-450 levels and NADPH-cytochrome c reductase activity are each about 50% of adult values (Aranda et al., 1974). Postnatal increases in enzyme activity could conceivably result from any one or a combination of the following: increased synthesis of enzyme, conversion of inactive to active enzyme, decreased catabolism of enzyme, disappearance of an endogenous inhibitor, or appearance of an activating substance. Studies of the differentiation of a variety of liver enzymes during the late fetal and postnatal periods indicate that de novo synthesis is largely responsible for increased enzymatic activity (Greengard, 1977). This synthesis is attributed to gene expression as a result of natural (hormones and other endogenous substrates) and unnatural (drugs and other chemicals) stimuli encountered in the extrauterine environment. It is possible that relative proportions of different forms, or isozymes, of P-450 change with development of the individual. Warner (1986) pointed out that the period of most rapid MFO metabolism (from 2 or 3 months to 3 years) coincides with the period when concentrations of endogenous substrates (such as steroid hormones) are low. The gradual decline in metabolism to adult levels at puberty parallels the increase in sex steroids accompanying maturation.


FIGURE 3-1 Theophylline dose requirements and plasma concentrations. (Top) Estimated dose requirements of theophylline (mg/kg/day) to maintain a steady-state plasma concentration Cpss of 10 mg/liter. (Lower) Estimated plasma concentrations of theophylline at steady state if dose is kept at 20 mg/kg/day. Shaded areas indicate tentative therapeutic level for bronchodilation and antiapneic activity.

SOURCE: Aranda, 1984.

Liver Development The liver undergoes a series of integrated morphological and functional changes perinatally, including, differentiation of hepatocytes and emergence of constitutive enzymes. One important function, which requires coordinated maturation of a series of processes, is enterohepatic circulation. Bile flow depends on the adequate synthesis, conjugation, secretion, and recirculation of bile acids (Balistreri et al., 1983). Suchy et al. (1981) found that serum levels of cholylglycine and chenodeoxycholate, the two major bile acids, became markedly elevated in normal neonates during the first 4 days of life. The levels then gradually declined over the next 4 to 6 months to values typical of children and adults. The initial increase, which the investigators termed physiologic cholestasis, was attributed to impared intestinal reabsorption and hepatic transport processes. They speculated that the transport deficit could be caused by functional immaturity in hepatocellular uptake, binding, conjugation, or secretion of bile acids.

Two important possible consequences of the physiologic cholestasis during the first months of life are inefficient intestinal fat digestion and inhibition of biliary excretion. Impaired fat digestion could be of toxicologic importance when lipophilic chemicals are ingested with oils. Studies in rats have demonstrated that halogenated hydrocarbons such as carbon tetrachloride are more poorly absorbed and less acutely toxic when given in oils than when given undiluted or in aqueous vehicles (Kim et al., 1990). The oils serve as a reservoir in the gut to retard systemic absorption of lipophilic chemicals. Biliary excretion is one of the two major pathways for elimination of chemicals from the body. Bilirubin, a product of hemoglobin catabolism, can accumulate in the body if its excretion in the bile is unduly hindered. A transient physiological hyperbilirubinemia is commonly seen in newborns, but it usually lasts only a few days, rather than months as does physiological cholestasis (Suchy et al., 1981). The physiologic hyperbilirubinemia is generally attributed primarily to the newborn's diminished glucuronidation capacity, which in turn is believed to result from low hepatic microsomal uridine diphosphoglucuronyltransferase activity (Kawade and Onishi, 1981). Glucuronide conjugation is depressed to a greater extent and for a longer time during perinatal development than sulfate or glycine conjugation (Dutton, 1978). There have been a number of reports of toxicities associated with neonates' decreased ability to conjugate and eliminate chemicals in the bile and urine. Chemicals implicated in such cases include chloramphenicol (Sutherland, 1959; Weiss et al., 1960), hexachlorophene (Tyrala et al., 1977), benzyl alcohol (Gershanik et al., 1982), and diazepam (Nau et al., 1984).


Renal excretion is the principal pathway for elimination of most chemicals from the body. Although volatile parent compounds and metabolites can be exhaled, this route of elimination is of little quantitative significance for most pesticides in current use. As described above, biliary excretion may play a role in elimination of some parent compounds and metabolites, notably conjugates formed in phase II-type reactions of liver metabolism. The conjugates may be eliminated concurrently in both bile and urine.

The kidneys are anatomically and functionally immature at birth. Although nephrogenesis is complete in the full-term human newborn, anatomic immaturity is still manifest for several weeks (Lorenz and Kleinman, 1988). The final phase of postnatal anatomical kidney development is that of increase in nephron size. Renal blood flow is relatively low in the neonate as a result of receiving a smaller percentage of total cardiac output and high intrarenal vascular resistance. Developmental changes in glomerular filtration rate parallel changes in renal blood flow (Lorenz and Kleinman, 1988). Glomerular filtration rate in premature newborns may be as low as 5.0% of that in adults, whereas in full-term newborns it is typically 30% to 40% of adult values. The glomerular filtration rate increases rapidly in the infant, becoming equivalent to that of the adult per unit of surface area within 10 to 20 weeks (West et al., 1948). Glomerular function appears to be more advanced at birth than renal tubular function (Weil, 1955). Tubular functions include both active and passive reabsorptive and secretory processes for specific agents. The deficient transport processes and reduced glomerular filtration rate produce smaller medullary solute gradients than are produced in adults, which in turn result in a diminished capacity of the neonate to concentrate urine. Maturation of tubular transport systems is relatively slow, in that maximum capacity may not be reached until about 8 months of age (West et al., 1948; Calcagno and Rubin, 1963).

Excretion of chemicals by the kidneys depends primarily on glomerular filtration and tubular secretion and reabsorption. A decrease in one or the other in neonates can result in delayed clearance of a chemical from the bloodstream and the body. Under such circumstances, the chemical may have a prolonged duration of action and there may be an increased risk of toxicity. Aminoglycoside antibiotics, such as kanamycin and gentamicin, are examples of compounds that are not metabolically degraded or bound to plasma proteins to a significant extent. They are primarily eliminated by glomerular filtration. Total body clearance values for premature newborns are only about 5% of adult values, whereas full-term newborn values are about 10% to 30% of the values of adults. Increase in aminoglycoside clearance during the neonatal period is closely related to maturation of glomerular function and correlates well with increased creatinine clearance (McCracken et al., 1971). Penicillins, on the other hand, form a class of compounds eliminated primarily by the kidneys via tubular secretion. Newborns and neonates given penicillins typically exhibit prolonged blood half-lives. Morselli et al. (1980) note that glomerular filtration and tubular secretion mature more rapidly than does tubular reabsorption, so clearance values for some compounds may be quite high during the first 2 to 24 months of life.


Scaling is the mathematical process used to adjust the dosage of therapeutic drugs or toxic substances to achieve comparable effects between animals of different size in one species or between different species of animals of markedly different sizes. In this section, body weight, surface area, metabolic rate, and regression analysis will be used as variables to relate dosages of compounds and body size. The choice of an appropriate scaling variable is important because of the need to compare toxicities in newborn animals weighing hundreds of grams and then extrapolating these findings to human infants weighing several kilograms and to human adults weighing 50 to 100 kg.

Although there are sound theoretical bases for choosing any one of the variables that can bridge great differences in size, there is no current consensus on which is most appropriate. Part of the difficulty in using a single approach is that the rate of development of the processes involved in absorption, distribution, metabolism, and excretion of xenobiotic compounds vary with age and across species. Thus, although some agreement is necessary to establish one system of scaling so that one set of studies can be compared with another, no scaling method will resolve all the differences between animals of different ages and of different species.

Body Weight

Traditionally, the therapeutic effectiveness or the toxicity of xenobiotic compounds in animals of different sizes has been evaluated by comparing dosages based on weight and expressing these as, for example, milligrams of compound per kilogram of body Weight or micrograms of compound per 100 g of body weight. In comparing animals that differ in size as much as mice and rats or rats and human adults do, or immature and mature animals, the values per unit of weight have often been found to be unpredictable for equivalent levels of therapeutic efficacy or toxicity. For example, the therapeutic dose of the anticancer drug methotrexate in the mouse is 1.5 mg/kg; in the rat, 0.5 mg/kg; and in humans, 0.07 mg/kg (Pinkel, 1958). Within one species, the vitamin K analog synkavite at a dosage of 0.16 mg/kg produces a bilirubin level of more than 30 mg/dl in the newborn rat, but only 3 mg/dl in the adult rat. In other words, a dose of 0.64 mg/kg in the adult rat is required to increase bilirubin levels by the same amount that is produced by a dose of slightly less than 0.04 mg/kg in the newborn rat (Wynn, 1963). In pesticide studies, the ratio of the LD50 of adult to newborn rats was 2.4 for parathion (3.6 mg/kg for adults and 1.5 mg/kg for 23-day-old rats), whereas for octamethyl pyrophosphate the LD50 ratio was reversed (0.2; the adult LD50 was 10 mg/kg and for the 23-day-old rat LD50 was 49 mg/kg) (Brodeur and DuBois, 1963).

Other Effects of Body Size

Another way to scale dosages of xenobiotic compounds between large and small animals is on the basis of some function of body size other than weight. The growth rate varies from organ to organ or component to component in relation to increases in body size. Examination of this phenomenon led to the use of the allometric expression in the form

y = aBx,

in which y is the weight of an organ or component of the body, B is total body weight, a is a constant, and x is an exponent of body weight. Transforming this equation yields.

log y = x log B + log a,

and then a plot of log y versus log B yields a simple linear graph with x as the slope and log a as the intercept. This type of representation indicates that the growth of a particular organ bears some constant relationship to the growth of the animal, but that relationship may vary from structure to structure. In physiology, it became apparent that a variety of physiological functions also increased allometrically with body size. These relationships were extensively reviewed by Boxenbaum (1982) and Boxenbaum and D'Souza (1990).

When these allometric concepts were applied to metabolic rate, the equation would be

metabolic rate = a·wtx,

where a is a constant that depends on the units used for metabolic rate and weight (wt), and x would be a fraction that was less than 1.0. Brody (1945) recognized that the expression for metabolic rate was similar to the expression for the surface area of the body. He further made the assumption that since the rate of cooling of a solid body was proportioned to its surface area, by analogy the metabolic rate should be directly proportional to surface area. Thus if surface are = a·wt2/3, then metabolic rate = a·wt2/3. (For an extensive discussion of these relationships, see Calabrese [1986].)

When the potency or toxicity of a substance is related to the persistence of some concentration of the compound in the body, the effect will be related to the rate at which the original substance is metabolized to less potent or to more potent compounds. In many cases, the rate at which a xenobiotic compound is metabolized will be related to the overall metabolic rate of the animal. Because smaller animals, in general, have greater metabolic rates per unit of body mass than do larger animals, the rate of metabolism of many substances per unit of body weight will be more rapid in smaller animals than in larger ones.

Except for the energy required for growth (a small fraction of energy consumption), energy intake and expenditure are essentially equivalent. Total energy expenditure involves the energy required for resting metabolism, for diet-induced thermogenesis, and for physical activity. The total expenditure of energy is the overall metabolic rate, and this can be closely approximated by the total energy intake when body weight is relatively stable. The advantage of this is that energy intake is generally more easily measured than is metabolic rate.

Surface Area

A variety of methods related to metabolic rate have been developed to equalize the values for energy expenditure (and consumption) between individuals of different sizes. The two principal approaches have been to use a power function of weight, such as weight2/3 or weight3/4, or to use calculated surface area. Although surface area is reasonably approximated by weight2/3, it is more precisely a power function of weight with an additional factor based on a power function of height or length. The surface area methodology has been the more widely used of these two approaches, although weight2/3 and surface area have often been used interchangeably despite the fact that they are not actually equivalent (Brody, 1945; Lindstedt, 1987). Values for usual energy intakes by human individuals of different sizes and the relationship between kilocalories per kilogram of body weight and kilocalories per square meter of body surface area are shown in Table 3-5. Because of the comparability of energy requirements throughout infancy and childhood when calculated on a surface area basis (˜2,000 kcal/m2), it has became common to assume that the metabolism of therapeutic drugs would also be comparable on this basis. Therefore, many medication dosages for children are calculated in terms of milligrams per square meter.

Although these assumptions are reasonably valid for many drugs in individuals of a comparable age group, they are not necessarily valid across all age groups or for all drugs (Lamanna and Hart, 1968). Almost all the comparisons between animals of different sizes (within and across species) in which surface area has been used to compare physiological functions, metabolism of drugs, or other body parameters have used comparisons between mature animals. Because of differences in the rate of maturation of individual organs and their functions, within and between species, as the animal proceeds from birth to maturity, it is not likely that the simple surface area relationship will be true for the comparison of immature and mature animals. Infants, children, and adults, just as immature and mature animals of other species, differ from each other in stages of functional development of individual organ systems, the processes of growth, and maturation of enzyme systems involved in the metabolism of xenobiotic compounds, as well as in their metabolic rates expressed on a weight or surface-area basis (see Table 3-5). Compounds such as chloramphenicol, which is poorly detoxified in the newborn (Calabrese, 1978); acetaminophen, which is excreted primarily as a sulfate conjugate in the young and as a glucuronide in the adult (Sonawane, 1982); isoniazid, which is acetylated at a reduced rate in the newborn (Nyhan, 1961); and caffeine, which is poorly biotransformed in the infant (Neims, 1982), are examples of drugs that differ in metabolism and toxicity in relation to age, whether doses are compared in terms of weight or surface area.
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TABLE 3-5 Representative Values of Weight, Length (Height), Surface Area, and Caloric Intake for Individuals at Various Ages

a Kilocalories per kilogram are given to indicate the three-to fourfold differences in metabolic rate when calculated on a weight basis.

b Kilocalories per square meter are shown to demonstrate the relative stability of this measure throughout childhood and also to emphasize the differences between children and adults, even on this basis.

SOURCE: Based on data from Hamill et al., 1979; NRC, 1989.


TABLE 3-6 Comparison of Doses Across Age Groups, Based on Weight, Surface Area, and Metabolic Rate

NOTE: To illustrate the relative differences in dosage on these three bases, the dosage for the adult male is held constant in absolute figures (7,000). By dividing that dose by the adult male's weight, surface area, and metabolic rate, one obtains the figures of 100 mg/kg (column 2), 3,784 mg/m2 (column 3), and 2,333 mg/1,000 kcal (column 4). These figures are then used to calculate the absolute dosages for each individual on each basis. These data demonstrate that the three bases for equating drug dosage between individuals of different ages and weights will result in absolute dosages that may vary several fold for the same age and size person.

a These values per square meter (m2) and per 1,000 kcal were used to provide equivalent adult dosages.

Metabolic Rate

Another method for comparing children of all sizes with adults in order to establish equivalent dosages of drugs is to relate the dose directly to metabolic rate (Calabrese, 1986). Dosage is represented as the amount per unit of metabolic rate, e.g., milligrams per 1,000 kcal per 24 hours. This method has the theoretical advantage of directly relating dosage to the variable that is commonly responsible for differences in drug metabolism between individuals of different sizes and ages. An example of dosages based on weight, surface area, and metabolic rate is given in Table 3-6, in which the adult dose is made constant. With the exception of the very-low-birth-weight infant, dosages based on kilocalories consumed (or metabolized) yield higher values for children than the other approaches for equivalent adult dosage. Even this approach can yield erroneous results, as illustrated by Schmidt-Nielsen (1972), whose calculations of the dosage of lysergic acid needed to produce the condition called musth in the male elephant were as follows:

Based on body weight, elephant versus cat / = 297 mg

Based on metabolic rate, elephant versus cat / = 80 mg

Based on body weight, elephant versus human / = 8 mg

Based on metabolic rate, elephant versus human / = 3 mg

Based on brain weight, elephant versus human / = 0.4 mg

Another problem arises when one considers whether compounds are metabolized to less toxic or to more toxic substances in the liver. This difference is of particular importance in extrapolation from one animal species to another, especially to the human species. For example, the extent of the postnatal development of cytochrome P-450-dependent oxidations is contingent on both species and substrate as well as on the sex of the animal. In the hydroxylation of aromatic compounds, maximum activity is reached at about 30 days in male and female rats but decreases after that in the female and is constant or slowly increasing in males for another 30 days (Klinger, 1982). In humans, there appears to be a relatively consistent difference; humans have a significantly lower rate of hepatic metabolism than found in other mammalian species. Boxenbaum and D'Souza (1990) discussed this difference between human and other animal species in detail and used the term neoteny to explain these differences. Neoteny is the ''retention of formerly juvenile characteristics by adult descendants … produced by retardation of somatic development" (Gould, 1977).

The significance of this concept for pesticide toxicology is that the hepatic metabolism of many xenobiotics in the mature human subject will occur at a reduced rate when compared to the rates in other mature animals even when corrected for scaling difference. It remains uncertain at this time how the problem of neoteny applies to immature animals. This is an area that deserves investigation although the studies will be complex because of the difficulty in identifying comparable stages of maturity in young animals. Further description of the maturation of enzyme systems is provided in the monograph by Calabrese (1986) entitled Age and Susceptibility to Toxic Substances and in the detailed review by Klinger (1982).

Regression Analysis

An alternative approach to scaling that has overcome some of the problems mentioned is that of regression equations that use the allometric concept without a predetermined power value. In this approach, the log of dose for a specific outcome is plotted against the log of body weight. For many substances and many animal species a linear relationship exists, but the slopes and intercepts are specific for each compound and each species. Krasovskii (1976) explained this process in detail, and after examining the action of several hundred compounds stated, "[I]t was shown that the regularities of the comparative sensitivity of the animals to 80–85% of the substances can be characterized by a straight line equation." When this approach was used to evaluate human toxicity of 107 substances based on data from four to six animal species, the results indicated there was a transfer error that did not exceed a value of 300% to 400% between calculated and observed values.

For risk assessment purposes, the problem becomes simpler as long as both exposure and toxicity are expressed on the same basis, e.g., micrograms per kilogram, micrograms per square meter, or micrograms per 1,000 kilocalories. To clarify this concept, assume that pesticide X is used on food A. The residue at the point of consumption is found to be 100 µg/100 g (1 µg/g) of food A. If the average adult consumption of food A is found to be 210 g with a total energy intake of 3,000 kcal/day (Table 3-5) and the average 2-year-old male infant consumes the same proportion of his total calories—1,200—as food A (Table 3-5), the following data apply:

Adult: 210 g of food A in a 3,000 kcal total intake

2 year old: (1,200 kcal ÷ 3,000 kcal) × 210 g food A = 84 g food A

Adult exposure = 210 g × 1 µg/g = 210 µg of pesticide X

2-year-old exposure = 84 µg × 1 µg/g = 84 µg of pesticide X

If these intakes are then scaled by weight, surface area, or calories for both adult and child, the value of intake of pesticide X becomes 70 µg/1,000 kcal for both (see Table 3-7).

If toxicity is evaluated on the same basis as exposure is assessed (i.e., on the basis of weight, surface area, or calories metabolized) using the same ratio of dosages, the results will be identical whatever basis is used. For example, if the no-observed-effect level (NOEL) were 6 µg/kg for the adult and 13 µg/kg for the child, then the NOEL expressed on the basis of surface area would be 228 µg/m2 in the adult and 300 µg/m2 in the child. On a metabolic rate basis, the NOEL for adult and child would both be 140 µg/1,000 kcal. Thus, on whatever basis one calculated toxicity, both the infant and adult would be consuming X at half the NOELs for their age group. Similar calculations would be needed to translate these ratios from the human species to the species to be used for toxicity testing.

Because the diets of small children are limited in diversity, it would be reasonable for the 2 year old to consume twice as much of food A as the adult as a proportion of his total energy intake. In this situation, the amount of residue ingested by the infant would be 13 µg/kg, 300 µg/m2, and 140 µg/1,000 kcal. Because of the relative increase in intake by the child compared to the adult, the child would be at the child's NOEL whatever the basis of the calculation.

If the efficacy or toxicity of a compound is not related to its rate of metabolism, using energy consumption as a basis for relating dosages probably would not provide equivalent levels of toxicity. In acute toxicity, for example, if the toxicity depended on peak concentration, comparable dosage would be on a simple weight basis because volumes of distribution based on body water compartments are more closely related directly to weight than to surface area or metabolic rate. Even under such circumstances, differences in rates of absorption, extent of protein binding, plasma half-life, or concentration at the receptor site could modify the dosage generated on a weight basis.

TABLE 3-7 Expression of Exposure Values for the Adult and Child


In chronic exposures, the differential rate of development of metabolizing enzymes such as glucuronyltransferase and the P-450-dependent mixed-function oxidases can have an impact on toxicity that is independent of overall metabolic rate. If the parent compound were the toxic substance, delayed enzymatic degradation would enhance toxicity. However, if a metabolite were the source of toxicity, the slower metabolism would result in reduced toxicity.

If the metabolism of a xenobiotic material is fully elucidated, it would be appropriate to use a reference base or denominator (weight, surface area, or metabolic rate) that best reflects the pharmacokinetics of that compound. When the developmental pharmacokinetics of a substance are not well delineated, as is often the case for pesticides, it has been demonstrated empirically that on a weight basis, the toxicity difference between immature and mature animals, based on comparative LD50s, is usually less than a factor of 5 and has only occasionally been reported to exceed a factor of 10 for any pesticide studied to date (Calabrese, 1986).



In this chapter, the committee summarized the mechanisms of toxicity and explored the potentially vulnerable organ systems or functional systems in the developing animal. In addition, it reviewed data on toxicants that provide the basis for these conclusions. Although the committee was interested in drawing conclusions from toxicity testing of pesticides, in many cases there were no data on developing animals. To illustrate the principles of toxicity as they pertain to developing animals, the committee therefore utilized information from testing of other toxicants, including drugs.

As described in Chapter 2, the nervous system, the immune system, and reproductive systems continue to develop after birth. This observation heightens concern that toxicity during these postnatal developmental stages or periods may have lasting consequences throughout adult life.

• Differences in toxicity between young and mature animals may be in either direction but are generally modest. The younger animal may be more sensitive or may be less sensitive than the older animal to comparable levels of exposure of toxic agents. The direction of these differences appears to be compound specific as well as age specific because toxicity may not vary linearly with age. In those instances where such measures as LD50s are significantly different, the differences are usually less than 1 order of magnitude and often substantially less.

• Data on age-dependent pharmacokinetics of pesticides are lacking for most animals, and the data base on pharmacokinetics and metabolism of drugs in immature humans is also limited. Available information reveals that some functional immaturities offset or cancel one another, whereas others tend to be additive. For example, inefficient gastrointestinal absorption of the anti-inflammatory drug indomethacin is offset by decreased biliary and renal clearance. Maturation of most biochemical and physiological processes occurs within the first 2 years; indeed, substantial changes occur within the first days and weeks of life. Compared to adults, therefore, neonates and infants can be anticipated to have the greatest differences in pharmacokinetics and susceptibility to pesticide toxicity—the youngest being the most likely to exhibit aberrant responses. Metabolic and renal clearance of many xenobiotics reaches and exceeds adult levels (when expressed on a body weight basis) during the first year. Therefore, older infants and young children may metabolize pesticides more extensively and eliminate them more rapidly than adults. This rapid elimination may confer increased resistance or susceptibility to toxicity, depending on the nature of the compound to which the individual is exposed.

• On the basis of our understanding of mechanisms of action of toxicants in mature animals, including the human adult, it is generally possible to predict that similar mechanisms of action will occur in immature animals, including the human infant, child, or adolescent (i.e., biochemical mechanisms of toxicity are similar across age and developmental stages). For example, if a toxicant is cytotoxic (causes cell death) in the adult, it should cause cell death in the immature animal by the same mechanism. This principle suggests that mechanisms of action should also be comparable across species.

• Studies of the toxicity of xenobiotic compounds in developing mammals—both laboratory animals and humans—demonstrate the potential for acute and chronic toxicity. Toxicity in the perinatal and pediatric periods is of special concern, since systems and structures under development at those times are important for survival over the lifetime of the mammal.

• Studies in laboratory animals have demonstrated an age-related difference in acute toxicity; however, the direction of the difference is dependent on the chemical, and the magnitude of the effect is usually no more than 1 order of magnitude and often is considerably less. A developing animal may be more, less, or equally sensitive to a given chemical than is an adult.

• Studies of the influence of age on the toxicity of xenobiotic compounds in laboratory animals may provide an incomplete or inaccurate picture of similar effects in humans. Rodents are less mature at birth than humans. Thus, more pronounced age-related differences would be anticipated in rodents than in humans during the neonatal period. Rodents and most other commonly used test animals mature very rapidly during this time, so that differences of even a few days in age can profoundly affect susceptibility to the toxicity of xenobiotic compounds.

• Very few data were found on the relative toxicity of pesticides in immature and mature humans. There is, however, a limited data base on pharmaceutical agents. As in animals, age dependency of humans to therapeutic and side effects is drug and age dependent. Since premature and full-term newborns are the most different anatomically and physiologically from adults, it follows that they typically exhibit the most pronounced differences from adults in sensitivity to drugs.

• Pharmacokinetic processes control the amount of bioactive chemicals in target organs and, in turn, the magnitude and duration of toxic effects. The net effect of immaturity on the various processes that affect chemical disposition is difficult to predict. The situation is complex in a number of respects. Most laboratory animals are less mature than humans at birth, although maturation in animals is more rapid. In addition, various body structures and associated functions mature at different rates in different species.

• Comparison of exposures between immature and mature animals and across species is complex, and no single mathematical expression is universally applicable. The toxicity of a pesticide varies with its rate-limiting pharmacokinetic processes.

• Data from studies in humans (e.g., in children and adults treated with cytotoxic chemotheraupetic agents) show that toxic effects are similar qualitatively but may differ quantitatively. That is, the types of toxic effects that limit treatment were similar in children and adults; however, the dose at which treatment limitations were reached was different. Indeed, for some drugs, the maximum tolerated dose is greater in developing than in human adults (i.e., such drugs are less toxic to infants and children).

• Studies of the ontogeny of plasma esterases in humans suggest that immature individuals may be at greater risk for acute toxic effects of pesticides that are cholinesterase inhibitors. The increase in sensitivity appears to be greatest during the first months of life, apparently because of relatively low levels of esterases and a diminished capacity to detoxify such pesticides.

• Although the principles of developmental toxicity following in utero exposure have received considerable attention, there has been little attention to the principles of developmental toxicity following exposure to pesticides in the postnatal and perinatal periods.

• The central nervous system (CNS) continues to develop during the second and third trimester and during postnatal life. These developmental processes include many—if not all—the developmental processes that occur during prenatal development. As expected, alteration of the development of the nervous system can be blatant or silent—until the function is needed.

• The immune system is responsible for modulating host defenses against a range of human diseases. The successful development and functioning of the immune system require recognition and response to a range of cellular and circulating signals acting by endocrine, autocrine, and paracrine mechanisms. These complex control systems offer multiple opportunities for disruption by environmental chemicals—such as agricultural pesticides.

• Assessment of the effects of pesticides on the developing human nervous system is difficult because the methodology for such assessment is complex and poorly delineated. Development of the CNS is characterized by exacting architectural complexity and localization of function occurring over a prolonged period postnatally. The effects of altered neurologic development may be measured either as anatomic or behavioral and cognitive outcomes.

• The scientific consensus on appropriate neurodevelopmental outcome measures for evaluation of exposures (in animal models and human epidemiologic studies) is still evolving (NRC, 1992). Measurement of these end points is complicated, not only because of the elaborate nature of the end points that are measured but also because the timing of the insult may change the outcome and the functional end points may not be manifested until long after the exposure.

• Despite the difficulties in measuring effects, exposure to xenobiotic compounds has been found to alter CNS development at the anatomic and functional level. The alteration in development can be irreversible, thus resulting in permanent loss of function. These damaging effects of xenobiotic compounds on the CNS of the developing organism can occur at exposure levels that are safe for the adult.

• Certain classes of pesticides, including organophosphates, carbamates, and organochlorines, are known to have neurotoxic effects, especially as the result of high-dose acute exposures. Generally, data are insufficient for evaluation and determination of the neurodevelopmental effects of low-dose exposure to these broad classes of agents, and risk assessment for low-level exposures is not possible using current data. Nevertheless, the biochemical (neurotoxic) mode of action of these classes of compounds and the distinctive qualitative vulnerability of the child's developing nervous system makes the evaluation of low-dose neurodevelopmental effects a concern.


• The establishment of a standard developmental assessment model or protocol would allow one to interpret toxicity studies in immature animals in a systematic manner. The committee recommends that a standard protocol for evaluation of immature animals be established and required as part of the basic assessment of pesticides for toxicity to immature animals.

• Given the potential for lifelong effects following perinatal or pediatric toxicity, it is essential to develop toxicity testing procedures that specifically evaluate in appropriate animal models vulnerability during the developmental period and the adverse effects, if any, over the life of the animal.

• Given that toxicity is generally age related, consideration of this phenomenon must be included in regulatory action, testing methodologies, and public health policies.

• In general, it is possible to extrapolate data on the end points of toxic concern from adults to developing humans, although the dose that produces the toxicity is likely to be different. The developing human frequently seems to be more resistant than the adult to cytotoxicity from the anticancer and anti-AIDS drugs tested.

• Extrapolation of toxicity data from laboratory animals to humans—especially those for developing organisms—may be inaccurate. Careful attention to species differences in disposition and metabolism as well as in stages of maturation of organ systems is essential for accurate policy development and public health protection.

• Greater attention is needed to develop a broader understanding of the principles guiding developmental toxicity of organisms, especially humans, following birth and during critical periods of postnatal development, including infancy and puberty.

• Neurodevelopmental effects must be part of the battery of end points evaluated for toxicants, including pesticides and agricultural chemicals.

• Analysis of the impact or toxicity of agricultural chemicals on the immune system is essential. Regulatory development of a battery of consensus tests is critical to protect the developing immune system. At present, there is a paucity of information on the effects of many chemicals on the developing and indeed on the mature immune system.

• Although it is extremely difficult to assess neurodevelopmental effects, the CNS may be peculiarly vulnerable during a prolonged period of development, even if the exposure is at a level known to be safe for adults. Thus, a feasible, streamlined, and publicly credible method of assessment must be developed. Effectiveness of animal model and epidemiologic evaluation must be considered. Regulatory development of a battery of consensus tests will be difficult but necessary to ensure public confidence. Agencies involved should actively support research and innovation in this area of assessment.

• Since the kinetics of a variety of chemicals can be profoundly different in immature and mature subjects, the influence of immaturity on pesticide kinetics and toxicity is complex and must therefore be assessed on a case-by-case and chemical-by-chemical basis.

• When the pharmacokinetics of a specific compound are understood well enough to indicate that the metabolism (and elimination) of the substance is, in fact, proportional to the animal's rate of metabolism, as is often the case, then comparisons on a metabolic rate basis, on a surface area basis, or weight2/3 basis would be reasonable. If surface area is used for adjusting from animals to humans, it should also be used for adjusting from infants to adults. In the many situations when such data are not available, the use of a simple body weight relationship for toxicity testing may be used as long as potential exposure is calculated on the same basis. Since most dietary pesticide exposure data are based on body weight, this is an added reason to use body weight as the basis for examining toxicity. In any case, scaling methods should be consistent.
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