TOXICOLOGICAL PROFILE FOR
 

DDT, DDE, and DDD
 

U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES
 

Public Health Service
 

Agency for Toxic Substances and Disease Registry
 

September 2002 
DDT, DDE, and DDD ii 

DISCLAIMER 

The use of company or product name(s) is for identification only and does not imply endorsement by the 

Agency for Toxic Substances and Disease Registry. 
DDT, DDE, and DDD iii 

UPDATE STATEMENT
 

Toxicological profiles are revised and republished as necessary, but no less than once every three years. 

For information regarding the update status of previously released profiles, contact ATSDR at: 

Agency for Toxic Substances and Disease Registry
 

Division of Toxicology/Toxicology Information Branch
 

1600 Clifton Road NE, E-29
 

Atlanta, Georgia 30333
 

V 

FOmWORD 

This toxicological profile is prepared in accordance with guidelines* developed by the Agency for 

Toxic Substances and Disease Registry (ATSDR) and the Environmental Protection Agency (EPA). The 

original guidelines were published in the Federal Register on April 17, 1987. Each profile will be revised 

and republished as necessary. 

The ATSDR toxicological profile succinctly characterizes the toxicologic and adverse health 

effects information for the hazardous substance described therein. Each peer-reviewed profile identifies 

and reviews the key literature that describes a hazardous substance's toxicologic properties. Other 

pertinent literature is also presented, but is described in less detail than the key studies. The profile is not 

intended to be an exhaustive document; however, more comprehensive sources of specialty information 

are referenced. 

The focus of the profiles is on health and toxicologic information; therefore, each toxicological 

profile begins with a public health statement that describes, in nontechnical language, a substance's 

reIevant toxicological properties. Following the public health statement is information concerning levels 

of si,gnificant human exposure and, where known, significant health effects. The adequacy of 

information to determine a substance's health effects is described in a health effects summary. Data 

needs that are of significance to protection of public health are identified by ATSDR and EPA. 

Each profile includes the following: 

The examination, summary, and interpretation of available toxicologic information and 

epidemiologic evaluations on a hazardous substance to ascertain the levels of significant 

human exposure for the substance and the associated acute, subacute, and chronic health 

effects; 

A determination of whether adequate information on the health effects of each substance is 

available or in the process of development to determine levels of exposure that present a 

significant risk to human health of acute, subacute, and chronic health effects; and 

Where appropriate, identification of toxicologic testing needed to identify the types or levels 

of exposure that may present significant risk of adverse health effects in humans. 

principal audiences for the toxicological profiles are health professionals at the federal, state, 

and local levels; interested private sector organizations and groups; and members of the public. 

This profile reflects ATSDRs assessment of all relevant toxicologic testing and information that 

has been peer-reviewed. Staff of the Centers for Disease Control and Prevention and other federal 

scientists have also reviewed the profile. In addition, this profile has been peer-reviewed by a 

nongovernmental panel and was made available for public review. Final responsibility for the contents 

and views expressed in this toxicological profile resides with ATSDR. 

Administrator 

Agency for Toxic Substances and 

Disease Registry 
DDT, DDE, and DDD vi 

*Legislative Background 

The toxicological profiles are developed in response to the Superfund Amendments and 

Reauthorization Act (SARA) of 1986 (Public law 99-499) which amended the Comprehensive 

Environmental Response, Compensation, and Liability Act of 1980 (CERCLA or Superfund). This 

public law directed ATSDR to prepared toxicological profiles for hazardous substances most commonly 

found at facilities on the CERCLA National Priorities List and that pose the most significant potential 

threat to human health, as determined by ATSDR and the EPA. The availability of the revised priority 

list of 275 hazardous substances was announced in the Federal Register on November 17, 1997 (62 FR 

61332). For prior versions of the list of substances, see Federal Register notices dated April 29, 1996 (61 

FR 18744); April 17, 1987 (52 FR 12866); October 20, 1988 (53 FR 41280); October 26, 1989 (54 FR 

43619); October 17, 1990 (55 FR 42067); October 17, 1991 (56 FR 52166); October 28, 1992 (57 FR 

48801); and February 28, 1994 (59 FR 9486). Section 104(i)(3) of CERCLA, as amended, directs the 

Administrator of ATSDR to prepare a toxicological profile for each substance on the list. 
 

 

DDT, DDE, and DDD vii 

QUICK REFERENCE FOR HEALTH CARE PROVIDERS 

Toxicological Profiles are a unique compilation of toxicological information on a given hazardous 

substance. Each profile reflects a comprehensive and extensive evaluation, summary, and interpretation 

of available toxicologic and epidemiologic information on a substance. Health care providers treating 

patients potentially exposed to hazardous substances will find the following information helpful for fast 

answers to often-asked questions. 

Primary Chapters/Sections of Interest 

Chapter 1: Public Health Statement: The Public Health Statement can be a useful tool for educating 

patients about possible exposure to a hazardous substance. It explains a substance’s relevant 

toxicologic properties in a nontechnical, question-and-answer format, and it includes a review of 

the general health effects observed following exposure. 

Chapter 2: Relevance to Public Health: The Relevance to Public Health Section evaluates, interprets, 

and assesses the significance of toxicity data to human health. 

Chapter 3: Health Effects: Specific health effects of a given hazardous compound are reported by type 

of health effect (death, systemic, immunologic, reproductive), by route of exposure, and by length 

of exposure (acute, intermediate, and chronic). In addition, both human and animal studies are 

reported in this section. 

NOTE: Not all health effects reported in this section are necessarily observed in 

the clinical setting. Please refer to the Public Health Statement to identify 

general health effects observed following exposure. 

Pediatrics: Four new sections have been added to each Toxicological Profile to address child health 

issues: 

Section 1.6 How Can (Chemical X) Affect Children?
 

Section 1.7 How Can Families Reduce the Risk of Exposure to (Chemical X)?
 

Section 3.8 Children’s Susceptibility
 

Section 6.6 Exposures of Children
 

Other Sections of Interest: 

Section 3.9 Biomarkers of Exposure and Effect 

Section 3.12 Methods for Reducing Toxic Effects 

ATSDR Information Center 

Phone: 1-888-42-ATSDR or (404) 498-0110 Fax: (404) 498-0057 

E-mail: atsdric@cdc.gov Internet: http://www.atsdr.cdc.gov 

The following additional material can be ordered through the ATSDR Information Center: 

Case Studies in Environmental Medicine: Taking an Exposure History—The importance of taking an 

exposure history and how to conduct one are described, and an example of a thorough exposure 

history is provided. Other case studies of interest include Reproductive and Developmental 

Hazards; Skin Lesions and Environmental Exposures; Cholinesterase-Inhibiting Pesticide 

Toxicity; and numerous chemical-specific case studies. 
DDT, DDE, and DDD viii 

Managing Hazardous Materials Incidents is a three-volume set of recommendations for on-scene 

(prehospital) and hospital medical management of patients exposed during a hazardous materials incident. 

Volumes I and II are planning guides to assist first responders and hospital emergency department 

personnel in planning for incidents that involve hazardous materials. Volume III—Medical Management 

Guidelines for Acute Chemical Exposures—is a guide for health care professionals treating patients 

exposed to hazardous materials. 

Fact Sheets (ToxFAQs) provide answers to frequently asked questions about toxic substances. 

Other Agencies and Organizations 

The National Center for Environmental Health (NCEH) focuses on preventing or controlling disease, 

injury, and disability related to the interactions between people and their environment outside the 

workplace. Contact: NCEH, Mailstop F-29, 4770 Buford Highway, NE, Atlanta, GA 30341­

3724 • Phone: 770-488-7000 • FAX: 770-488-7015. 

The National Institute for Occupational Safety and Health (NIOSH) conducts research on occupational 

diseases and injuries, responds to requests for assistance by investigating problems of health and 

safety in the workplace, recommends standards to the Occupational Safety and Health 

Administration (OSHA) and the Mine Safety and Health Administration (MSHA), and trains 

professionals in occupational safety and health. Contact: NIOSH, 200 Independence Avenue, 

SW, Washington, DC 20201 • Phone: 800-356-4674 or NIOSH Technical Information Branch, 

Robert A. Taft Laboratory, Mailstop C-19, 4676 Columbia Parkway, Cincinnati, OH 45226-1998 

• Phone: 800-35-NIOSH. 

The National Institute of Environmental Health Sciences (NIEHS) is the principal federal agency for 

biomedical research on the effects of chemical, physical, and biologic environmental agents on 

human health and well-being. Contact: NIEHS, PO Box 12233, 104 T.W. Alexander Drive, 

Research Triangle Park, NC 27709 • Phone: 919-541-3212. 

Referrals 

The Association of Occupational and Environmental Clinics (AOEC) has developed a network of clinics 

in the United States to provide expertise in occupational and environmental issues. Contact: 

AOEC, 1010 Vermont Avenue, NW, #513, Washington, DC 20005 • Phone: 202-347-4976 • 

FAX: 202-347-4950 • e-mail: AOEC@AOEC.ORG • Web Page: http://www.aoec.org/. 

The American College of Occupational and Environmental Medicine (ACOEM) is an association of 

physicians and other health care providers specializing in the field of occupational and 

environmental medicine. Contact: ACOEM, 55 West Seegers Road, Arlington Heights, IL 

60005 • Phone: 847-818-1800 • FAX: 847-818-9266. 
DDT, DDE, and DDD ix 

CONTRIBUTORS 

CHEMICAL MANAGER(S)/AUTHORS(S): 

Obaid Faroon, Ph.D. 

M. Olivia Harris, M.A.
 

ATSDR, Division of Toxicology, Atlanta, GA
 

Fernando Llados, Ph.D.
 

Steven Swarts, Ph.D.
 

Gloria Sage, Ph.D.
 

Mario Citra, Ph.D.
 

Daniel Gefell, M.S.
 

Syracuse Research Corporation, North Syracuse, NY
 

THE PROFILE HAS UNDERGONE THE FOLLOWING ATSDR INTERNAL REVIEWS: 

1.	 Health Effects Review. The Health Effects Review Committee examines the health effects 

chapter of each profile for consistency and accuracy in interpreting health effects and classifying 

end points. 

2.	 Minimal Risk Level Review. The Minimal Risk Level Workgroup considers issues relevant to 

substance-specific minimal risk levels (MRLs), reviews the health effects database of each 

profile, and makes recommendations for derivation of MRLs. 

3.	 Data Needs Review. The Research Implementation Branch reviews data needs sections to assure 

consistency across profiles and adherence to instructions in the Guidance. 

DDT, DDE, and DDD xi 

PEER REVIEW
 

A peer review panel was assembled for DDT, DDE, and DDD. The panel consisted of the following 

members: 

1.	 Dr. D. Andrew Crain, Assistant Professor, Maryville College, Maryville, TN; 

2.	 Dr. Donald Michael Fry, Director, Center for Avian Biology, University of California at Davis; 

3.	 Dr. Christopher Metcalfe, Professor and Chair, Environmental and Resource Studies, Trent 

University, Peterborough, Ontario, Canada; and 

4.	 Dr. Mary S. Wolff, Professor of Community Medicine, Mt. Sinai School of Medicine, New York, 

NY. 

These experts collectively have knowledge of DDT, DDE, and DDD's physical and chemical properties, 

toxicokinetics, key health end points, mechanisms of action, human and animal exposure, and 

quantification of risk to humans. All reviewers were selected in conformity with the conditions for peer 

review specified in Section 104(I)(13) of the Comprehensive Environmental Response, Compensation, 

and Liability Act, as amended. 

Scientists from the Agency for Toxic Substances and Disease Registry (ATSDR) have reviewed the peer 

reviewers' comments and determined which comments will be included in the profile. A listing of the 

peer reviewers' comments not incorporated in the profile, with a brief explanation of the rationale for their 

exclusion, exists as part of the administrative record for this compound. A list of databases reviewed and 

a list of unpublished documents cited are also included in the administrative record. 

The citation of the peer review panel should not be understood to imply its approval of the profile's final 

content. The responsibility for the content of this profile lies with the ATSDR. 

DDT, DDE, and DDD xiii 

CONTENTS
 

FOREWORD ....................................................................... v 
 

QUICK REFERENCE FOR HEALTH CARE PROVIDERS ................................. vii 
 

CONTRIBUTORS ................................................................... ix 
 

PEER REVIEW ..................................................................... xi 
 

LIST OF FIGURES ................................................................ xvii 
 

LIST OF TABLES .................................................................. xix 
 

1. PUBLIC HEALTH STATEMENT .................................................... 1 
 

1.1	 WHAT ARE DDT, DDE, AND DDD? ...................................... 2 
 

1.2	 WHAT HAPPENS TO DDT, DDE, AND DDD WHEN THEY ENTER THE
 

ENVIRONMENT? ..................................................... 2 
 

1.3	 HOW MIGHT I BE EXPOSED TO DDT, DDE, AND DDD? .................... 4 
 

1.4	 HOW CAN DDT, DDE, AND DDD ENTER AND LEAVE MY BODY? .......... 5 
 

1.5	 HOW CAN DDT, DDE, AND DDD AFFECT MY HEALTH? ................... 6 
 

1.6	 HOW CAN DDT, DDE, AND DDD AFFECT CHILDREN? .................... 8 
 

1.7	 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO DDT, DDE, AND
 

DDD? ............................................................... 10 
 

1.8	 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN
 

EXPOSED TO DDT, DDE, AND DDD? ................................... 11 
 

1.9	 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO
 

PROTECT HUMAN HEALTH? .......................................... 11 
 

1.10 	 WHERE CAN I GET MORE INFORMATION? ............................. 13 
 

2. RELEVANCE TO PUBLIC HEALTH ............................................... 15 
 

2.1	 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO DDT/DDE/DDD IN
 

THE UNITED STATES ................................................ 15 
 

2.2	 SUMMARY OF HEALTH EFFECTS ..................................... 19 
 

2.3 	 MINIMAL RISK LEVELS .............................................. 27 
 

3. HEALTH EFFECTS .............................................................. 33 
 

3.1	 INTRODUCTION ..................................................... 33 
 

3.2	 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE ............ 34 
 

3.2.1	 Inhalation Exposure ............................................. 35 
 

3.2.1.1 Death .................................................. 36 
 

3.2.1.2 Systemic Effects ......................................... 36 
 

3.2.1.3 Immunological and Lymphoreticular Effects ................... 37 
 

3.2.1.4 Neurological Effects ...................................... 37 
 

3.2.1.5 Reproductive Effects ...................................... 37 
 

3.2.1.6 Developmental Effects ..................................... 37 
 

3.2.1.7 Cancer ................................................. 37 
 

3.2.2	 Oral Exposure .................................................. 37 
 
 

 

 

DDT, DDE, and DDD xiv 

3.2.2.1 Death .................................................. 72 
 

3.2.2.2 Systemic Effects ......................................... 74 
 

3.2.2.3 Immunological and Lymphoreticular Effects ................... 86
 

3.2.2.4 Neurological Effects ...................................... 90 
 

3.2.2.5 Reproductive Effects ...................................... 93 
 

3.2.2.6 Developmental Effects .................................... . 99
 

3.2.2.7 Cancer ................................................ 107
 

3.2.3 Dermal Exposure .............................................. 126
 

3.2.3.1 Death ................................................. 127
 

3.2.3.2 Systemic Effects ........................................ 127
 

3.2.3.3 Immunological and Lymphoreticular Effects .................. 128
 

3.2.3.4 Neurological Effects ..................................... 129
 

3.2.3.5 Reproductive Effects ..................................... 129
 

3.2.3.6 Developmental Effects .................................... 129
 

3.2.3.7 Cancer ................................................ 129
 

3.3	 GENOTOXICITY .................................................... 129
 

3.4	 HEALTH EFFECTS IN WILDLIFE POTENTIALLY RELEVANT TO HUMAN
 

HEALTH ........................................................... 133
 

3.5	 TOXICOKINETICS .................................................. 138
 

3.5.1 Absorption ................................................... 139
 

3.5.1.1 Inhalation Exposure ...................................... 139
 

3.5.1.2 Oral Exposure .......................................... 139
 

3.5.1.3 Dermal Exposure ........................................ 140
 

3.5.2	 Distribution ................................................... 140
 

3.5.3	 Metabolism ................................................... 143
 

3.5.4	 Elimination and Excretion ....................................... 149
 

3.5.5	 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) 

Models ...................................................... 151
 

3.6	 MECHANISMS OF ACTION ........................................... 160
 

3.6.1	 Pharmacokinetic Mechanisms ..................................... 160
 

3.6.2	 Mechanisms of Toxicity ......................................... 161
 

3.6.3	 Animal-to-Human Extrapolations .................................. 176
 

3.7	 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS ....... 176
 

3.8	 CHILDREN’S SUSCEPTIBILITY ....................................... 180
 

3.9	 BIOMARKERS OF EXPOSURE AND EFFECT ............................ 187
 

3.9.1	 Biomarkers Used to Identify or Quantify Exposure to DDT, DDE, and 

DDD ........................................................ 188
 

3.9.2	 Biomarkers Used to Characterize Effects Caused by DDT, DDE, and 

DDD ........................................................ 190
 

3.10 INTERACTIONS WITH OTHER CHEMICALS ............................ 191
 

3.11 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE ................. 193
 

3.12 METHODS FOR REDUCING TOXIC EFFECTS ........................... 193
 

3.12.1 Reducing Peak Absorption Following Exposure ...................... 194
 

3.12.2 Reducing Body Burden .......................................... 194
 

3.12.3 Interfering with the Mechanism of Action for Toxic Effects ............. 195
 

3.13 ADEQUACY OF THE DATABASE ..................................... 196
 

3.13.1 Existing Information on Health Effects of DDT, DDE, and DDD ......... 196
 

3.13.2 Identification of Data Needs ...................................... 199
 

3.13.3 Ongoing Studies ............................................... 210
 

4. CHEMICAL AND PHYSICAL INFORMATION ...................................... 215
 
 

 

 

DDT, DDE, and DDD xv 

4.1 CHEMICAL IDENTITY ............................................... 215
 

4.2 PHYSICAL AND CHEMICAL PROPERTIES ............................. 215
 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL ............................ 221
 

5.1 PRODUCTION ...................................................... 221
 

5.2 IMPORT/EXPORT ................................................... 222
 

5.3 USE ............................................................... 222
 

5.4 DISPOSAL ......................................................... 223
 

6. POTENTIAL FOR HUMAN EXPOSURE ........................................... 225
 

6.1 OVERVIEW ........................................................ 225
 

6.2 RELEASES TO THE ENVIRONMENT ................................... 230
 

6.2.1 Air .......................................................... 230
 

6.2.2 Water ........................................................ 230
 

6.2.3 Soil ......................................................... 231
 

6.3 ENVIRONMENTAL FATE ............................................ 232
 

6.3.1 Transport and Partitioning ....................................... 232
 

6.3.2 Transformation and Degradation .................................. 237
 

6.3.2.1 Air ................................................... 237
 

6.3.2.2 Water ................................................. 237
 

6.3.2.3 Sediment and Soil ....................................... 238
 

6.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT .......... 241
 

6.4.1 Air .......................................................... 241
 

6.4.2 Water ........................................................ 244
 

6.4.3 Sediment and Soil .............................................. 246
 

6.4.4 Other Environmental Media ...................................... 251
 

6.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE ............. 263
 

6.6 EXPOSURES OF CHILDREN .......................................... 278
 

6.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES ................ 279
 

6.8 ADEQUACY OF THE DATABASE ..................................... 281
 

6.8.1 Identification of Data Needs ...................................... 281
 

6.8.2 Ongoing Studies ............................................... 284
 

7. ANALYTICAL METHODS ....................................................... 287
 

7.1 BIOLOGICAL MATERIALS ........................................... 287
 

7.2 ENVIRONMENTAL SAMPLES ........................................ 291
 

7.3 ADEQUACY OF THE DATABASE ..................................... 296
 

7.3.1 Identification of Data Needs ...................................... 297
 

7.3.2 Ongoing Studies ............................................... 298
 

8. REGULATIONS AND ADVISORIES .............................................. 299
 

9. REFERENCES ................................................................. 309
 

10. GLOSSARY .................................................................. 397
 

APPENDICES 

A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETS .......................... A-1 
 
DDT, DDE, and DDD	 xvi 

B. 	 USER’S GUIDE ............................................................ B-1 
 

C.	 ACRONYMS, ABBREVIATIONS, AND SYMBOLS .............................. C-1 
 

D.	 HEALTH EFFECTS IN WILDLIFE POTENTIALLY RELEVANT TO HUMAN 

HEALTH .................................................................. D-1 
 

E. 	 INDEX ................................................................... E-1 
 
DDT, DDE, and DDD xvii 

LIST OF FIGURES 

3-1. Levels of Significant Exposure to DDT, DDE, and DDD (except o,p’ isomers) - Oral ......... 59 
 

3-2. Levels of Significant Exposure to o,p’-DDT, DDE, and DDD - Oral ...................... 69 
 

3-3a. Model I Metabolic Scheme for DDT .............................................. 147 
 

3-3b. Model II Metabolic Scheme for DDT ............................................. 148 
 

3-4. Proposed Metabolic Pathway for the Conversion of p,p’-DDE to its Methylsulfone 

Derivative ................................................................... 152 
 

3-5. Conceptual Representation of a Physiologically Based Pharmacokinetic (PBPK) Model for a
 

Hypothetical Chemical Substance ................................................ 158 
 

3-6. Diagrammatic Representation of the PBPK Model for Gestation ......................... 159 
 

3-7. Diagrammatic Representation of the PBPK Model for the Lactating Dam and Nursing Pup .... 160 
 

3-8. Existing Information of Health Effects of DDT, DDE, and DDD ......................... 201 
 

6-1. Frequency of NPL Sites with DDT, DDE, and DDD Contamination ...................... 233 
 

DDT, DDE, and DDD xix 

LIST OF TABLES 

3-1. Levels of Significant Exposure to DDT, DDE, or DDD (except o,p’ isomers) - Oral .......... 38 
 

3-2. Levels of Significant Exposure to o,p’-DDT, DDE, and DDD - Oral ...................... 65 
 

3-3. Genotoxicity of DDT, DDE, and DDD In Vivo ....................................... 132 
 

3-4. Genotoxicity of DDT, DDE, and DDD In Vitro ...................................... 133 
 

3-5. Tissue:Blood Partition Coefficients and Pharmacokinetic Constants for Modeling DDE 

Disposition in the Pregnant Rat .................................................. 156 
 

3-6. Physiological Constants Used in the PBPK Model for the Lactating Dam and the Nursing 

Pup ........................................................................ 157 
 

3-7. Ongoing Studies on the Health Effects of DDT and DDT Analogues ..................... 215 
 

4-1. Chemical Identity of p,p'- and o,p'-DDT, DDE, and DDD .............................. 220 
 

4-2. Physical and Chemical Properties of p,p'- and o,p'-DDT, DDE, and DDD .................. 222 
 

6-1. Concentrations of DDT and Metabolites in Biota ..................................... 260 
 

6-2. Average Residues in Food Groups and Average Daily Intake from U.S. FDA Total Diet 

Studies ...................................................................... 265 
 

6-3. Mean Daily Intake of DDT Per Unit Body Weight (μg/kg body weight/day) for Various Age 

Groups in the United States ..................................................... 269 
 

6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent Studies .......... 273 
 

7-1. Analytical Methods for Determining DDT, DDE, and DDD in Biological Samples .......... 292 
 

7-2. Analytical Methods for Determining DDT, DDE, and DDD in Environmental Samples ....... 296 
 

8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD ............................ 304 
 

DDT, DDE, and DDD 1 

1. PUBLIC HEALTH STATEMENT
 

This public health statement tells you about DDT, DDE, and DDD and the effects of exposure. 

The Environmental Protection Agency (EPA) identifies the most serious hazardous waste sites in 

the nation. These sites make up the National Priorities List (NPL) and are the sites targeted for 

long-term federal cleanup activities. DDT, DDE, and DDD have been found in at least 442 of 

the 1,613 current or former NPL sites. However, the total number of NPL sites evaluated for 

these substances is not known. As more sites are evaluated, the sites at which DDT, DDE, and 

DDD are found may increase. This information is important because exposure to these 

substances may harm you and because these sites may be sources of exposure. 

When a substance is released from a large area, such as an industrial plant, or from a container, 

such as a drum or bottle, it enters the environment. This release does not always lead to 

exposure. You are exposed to a substance only when you come in contact with it. You may be 

exposed by breathing, eating, or drinking the substance, or by skin contact. 

If you are exposed to DDT, DDE, and DDD, many factors determine whether you’ll be harmed. 

These factors include the dose (how much), the duration (how long), and how you come in 

contact with them. You must also consider the other chemicals you’re exposed to and your age, 

sex, diet, family traits, lifestyle, and state of health. 

While this document is specifically focused on the primary forms or isomers of DDT, DDE, and 

DDD (namely p,p’-DDT, p,p’-DDE, and p,p’-DDD), other isomers of these compounds will be 

discussed when appropriate. In some cases, the term DDT will be used to refer to the collection 

of all forms of DDT, DDE, and DDD. Should this not be clear from the context, the term ΣDDT 

(Σ is used to mean sum of) will be used. 
DDT, DDE, and DDD 2 

1. PUBLIC HEALTH STATEMENT 

1.1	 WHAT ARE DDT, DDE, AND DDD? 

DDT (1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane) is a pesticide that was once widely used to 

control insects on agricultural crops and insects that carry diseases like malaria and typhus, but is 

now used in only a few countries to control malaria. Technical-grade DDT is a mixture of three 

forms, p,p’-DDT (85%), o,p’-DDT (15%), and o,o’-DDT (trace amounts). All of these are 

white, crystalline, tasteless, and almost odorless solids. Technical grade DDT may also contain 

DDE (1,1-dichloro-2,2-bis(p-chlorophenyl)ethylene) and DDD (1,1-dichloro-2,2-bis(p-chlorophenyl)ethane) as contaminants. DDD was also used to kill pests, but to a far lesser extent than 

DDT. One form of DDD (o,p’-DDD) has been used medically to treat cancer of the adrenal 

gland. Both DDE and DDD are breakdown products of DDT. 

DDT does not occur naturally in the environment. After 1972, the use of DDT was no longer 

permitted in the United States except in cases of a public health emergency. It is, however, still 

used in some other areas of the world, most notably for controlling malaria. The use of DDD to 

kill pests has also been banned in the United States. 

You will find further information on the physical properties and uses of DDT, DDE, and DDD in 

Chapters 4 and 5 of this profile. 

1.2	 WHAT HAPPENS TO DDT, DDE, AND DDD WHEN THEY ENTER THE 

ENVIRONMENT? 

Before 1973 when it was banned, DDT entered the air, water, and soil during its production and 

use as an insecticide. DDT is present at many waste sites, including NPL sites; releases from 

these sites might continue to contaminate the environment. Most DDT in the environment is a 

result of past use; DDD was also used as a pesticide to a limited extent in the past. DDT still 

enters the environment because of its current use in other areas of the world. DDE is only found 

in the environment as a result of contamination or breakdown of DDT. DDD also enters the 

environment during the breakdown of DDT. 
DDT, DDE, and DDD 3 

1. PUBLIC HEALTH STATEMENT 

Large amounts of DDT were released into the air and on soil or water when it was sprayed on 

crops and forests to control insects. DDT was also sprayed in the environment to control 

mosquitos. Although the use of DDT is no longer permitted in the United States, DDT may be 

released into the atmosphere in other countries where it is still manufactured and used, including 

Mexico. DDT, DDE and DDD may also enter the air when they evaporate from contaminated 

water and soil. DDT, DDE, and DDD in the air will then be deposited on land or surface water. 

This cycle of evaporation and deposition may be repeated many times. As a result, DDT, DDE, 

and DDD can be carried long distances in the atmosphere. These chemicals have been found in 

bogs, snow, and animals in the Arctic and Antarctic regions, far from where they were ever used. 

Some DDT may have entered the soil from waste sites. DDT, DDE, and DDD may occur in the 

atmosphere as a vapor or be attached to solids in air. Vapor phase DDT, DDE, and DDD may 

break down in the atmosphere due to reactions caused by the sun. The half-life of these 

chemicals in the atmosphere as vapors (the time it takes for one-half of the chemical to turn into 

something else) has been calculated to be approximately 1.5–3 days. However, in reality, this 

half-life estimate is too short to account for the ability of DDT, DDE, and DDD to be carried 

long distances in the atmosphere. 

DDT, DDE, and DDD last in the soil for a very long time, potentially for hundreds of years. 

Most DDT breaks down slowly into DDE and DDD, generally by the action of microorganisms. 

These chemicals may also evaporate into the air and be deposited in other places. They stick 

strongly to soil, and therefore generally remain in the surface layers of soil. Some soil particles 

with attached DDT, DDE, or DDD may get into rivers and lakes in runoff. Only a very small 

amount, if any, will seep into the ground and get into groundwater. The length of time that DDT 

will last in soil depends on many factors including temperature, type of soil, and whether the soil 

is wet. DDT lasts for a much shorter time in the tropics where the chemical evaporates faster 

and where microorganisms degrade it faster. DDT disappears faster when the soil is flooded or 

wet than when it is dry. DDT disappears faster when it initially enters the soil. Later on, 

evaporation slows down and some DDT moves into spaces in the soil that are so small that 

microorganisms cannot reach the DDT to break it down efficiently. In tropical areas, ΣDDT 

may disappear in much less than a year. In temperate areas, half of the ΣDDT initially present 
DDT, DDE, and DDD 4 

1. PUBLIC HEALTH STATEMENT 

usually disappears in about 5 years. However, in some cases, half of the ΣDDT initially present 

will remain for 20, 30, or more years. 

In surface water, DDT will bind to particles in the water, settle, and be deposited in the sediment. 

DDT is taken up by small organisms and fish in the water. It accumulates to high levels in fish 

and marine mammals (such as seals and whales), reaching levels many thousands of times higher 

than in water. In these animals, the highest levels of DDT are found in their adipose tissue. 

DDT in soil can also be absorbed by some plants and by the animals or people who eat those 

crops. 

More information about what happens to DDT, DDE, and DDD in the environment can be found 

in Chapter 6. 

1.3 HOW MIGHT I BE EXPOSED TO DDT, DDE, AND DDD? 

People in the United States are exposed to DDT, DDE, and DDD mainly by eating foods 

containing small amounts of these compounds. Although not common today, exposure to DDT 

could also occur through inhalation or absorption through the skin during the handling or 

application of DDT. Even though DDT has not been used in this country since 1972, soil may 

still contain some DDT that may be taken up by plants and eaten by animals and people. DDT 

from contaminated water and sediment may be taken up by fish. The amount of DDT in food 

has greatly decreased since DDT was banned and should continue to decline. In the years 1986 

to 1991, the average adult in the United States consumed an average of 0.8 micrograms (a 

microgram is a millionth of a gram) of DDT a day. Adults consumed slightly different amounts 

based on their age and sex. The largest fraction of DDT in a person’s diet comes from meat, 

poultry, dairy products, and fish, including the consumption of sport fish. Leafy vegetables 

generally contain more DDT than other vegetables, possibly because DDT in the air is deposited 

on the leaves. Infants may be exposed by drinking breast milk. 

DDT or its breakdown products are still present in some air, water, and soil samples. However, 

levels in most air and water samples are presently so low that exposure is of little concern. DDT 
DDT, DDE, and DDD 5 

1. PUBLIC HEALTH STATEMENT 

levels in air have declined to such low levels that it often cannot be detected. In cases where 

DDT has been detected in air, it is associated with air masses coming from regions where DDT 

is still used or from the evaporated DDT from contaminated water or soil. p,p’-DDT and p,p’­

DDE concentrations measured in air in the Great Lakes region in 1990 reached maximum levels 

of 0.035 and 0.119 nanograms (a nanogram is a billionth of a gram) of chemical per cubic meter 

of air (ng/m3), respectively. Levels were generally much lower, especially during the winter 

months. In 1995–1996, soils in the corn belt, where DDT was heavily used in the past, 

contained on the average about 10 nanograms of DDT in a gram of soil. In recent years, most 

surface water has not contained detectable amounts of DDT. 

People who work or live around NPL sites or work with contaminated soil or sediment would 

most likely be exposed by accidentally swallowing soil, having skin contact with the soil, 

inhaling DDT vapor, or breathing in DDT in dust. 

You can find more information on exposure to DDT, DDE, and DDD in Chapter 6 of this profile. 

1.4 HOW CAN DDT, DDE, AND DDD ENTER AND LEAVE MY BODY? 

Today in the United States, DDT, DDE, or DDD enters the body mainly when a person eats 

contaminated food. The actual amounts of DDT, DDE, and DDD absorbed from foods depends 

on both the concentration of chemical in the food and the amount of food eaten. Small amounts 

of DDT, DDE, and DDD may also be breathed in and absorbed into the body. DDT, DDE, and 

DDD are often attached to particles too large to pass very far into the lungs after air containing 

them is breathed. These particles are more likely to be carried upward in the mucus of the air 

passages and swallowed than for the DDT to be absorbed in the lungs. DDT, DDE, and DDD do 

not enter the body through the skin very easily. 

Once inside the body, DDT can break down to DDE or DDD. DDE and DDD, in turn, break 

down to other substances (called metabolites). DDT, DDE, and DDD are stored most readily in 

fatty tissue, especially DDE. Some of these stored amounts leave the body very slowly. Levels 

in fatty tissues may either remain relatively the same over time or even increase with continued 
DDT, DDE, and DDD 6 

1. PUBLIC HEALTH STATEMENT 

exposure. However, as exposure decreases, the amount of DDT in the body also decreases. 

DDT metabolites leave the body mostly in urine, but may also leave by breast milk and pass 

directly to nursing infants. See Chapter 3 for more information on how DDT, DDE, and DDD 

enter and leave the body. 

1.5 HOW CAN DDT, DDE, AND DDD AFFECT MY HEALTH? 

Eating food with large amounts (grams) of DDT over a short time would most likely affect the 

nervous system. People who swallowed large amounts of DDT became excitable and had 

tremors and seizures. They also experienced sweating, headache, nausea, vomiting, and 

dizziness. These effects on the nervous system went away once exposure stopped. The same 

type of effects would be expected by breathing DDT particles in the air or by contact of the skin 

with high amounts of DDT. Tests in laboratory animals confirm the effect of DDT on the 

nervous system. 

No effects have been reported in adults given small daily doses of DDT by capsule for 

18 months (up to 35 milligrams [mg] every day). People exposed for a long time to small 

amounts of DDT (less than 20 mg per day), such as people who worked in factories where DDT 

was made, had some minor changes in the levels of liver enzymes in the blood. A study in 

humans showed that increasing concentrations of p,p’-DDE in human breast milk were 

associated with reductions in the duration of lactation. An additional study in humans found that 

as the DDE levels in the blood of pregnant women increased, the chances of having a pre-term 

baby also increased. It should be mentioned, however, that the levels of DDE in the blood at 

which this was noticed were higher than those currently found in women from the general 

population in the United States, but not higher than those that may be found in women in 

countries where DDT is still being used. 

To protect the public from the harmful effects of toxic chemicals and to find ways to treat people 

who have been harmed, scientists use many tests. 
DDT, DDE, and DDD 7 

1. PUBLIC HEALTH STATEMENT 

One way to see if a chemical will hurt people is to learn how the chemical is absorbed, used, and 

released by the body; for some chemicals, animal testing may be necessary. Animal testing may 

also be used to identify health effects such as cancer or birth defects. Without laboratory 

animals, scientists would lose a basic method to get information needed to make wise decisions 

to protect public health. Scientists have the responsibility to treat research animals with care and 

compassion. Laws today protect the welfare of research animals, and scientists must comply 

with strict animal care guidelines. 

Animal studies show that long-term exposure to moderate amounts of DDT (20–50 mg per 

kilogram [kg] of body weight every day) may affect the liver. Tests in animals also suggest that 

short-term exposure to DDT and metabolites in food may have a harmful effect on reproduction. 

In addition, we know that some breakdown products of DDT can cause harmful effects on the 

adrenal gland. This gland is situated near the kidney and produces hormones (substances 

produced by organs and released to the bloodstream to regulate the function of other organs). 

Studies in animals have shown that oral exposure to DDT can cause liver cancer. Studies of 

DDT-exposed workers did not show increases in deaths or cancers. Based on all of the evidence 

available, the Department of Health and Human Services has determined that DDT is reasonably 

anticipated to be a human carcinogen. Similarly, the International Agency for Research on 

Cancer (IARC) has determined that DDT is possibly carcinogenic to humans. EPA has 

determined that DDT, DDE, and DDD are probable human carcinogens. See Chapter 3 for more 

information on the health effects associated with exposure to DDT, DDE, and DDD. 
DDT, DDE, and DDD 8 

1. PUBLIC HEALTH STATEMENT 

1.6 HOW CAN DDT, DDE, AND DDD AFFECT CHILDREN? 

This section discusses potential health effects from exposures during the period from conception 

to maturity at 18 years of age in humans. 

Children can be exposed to DDT, DDE, or DDD by eating food or drinking breast milk 

contaminated with these compounds. DDT is a pesticide, and even though it has not been used 

in this country since 1972, soil has small amounts and, under certain conditions, contaminated 

soil transfers DDT to crops. Children can be exposed also by eating food imported from 

countries where DDT is still being used. Because of their smaller weight, intake of an equivalent 

amount of DDT by children and adults would result in a higher dose (amount of DDT ingested 

per kilogram of body weight) in children than in adults. In the United States between 1985 and 

1991, the average 81⁄2-month-old infant consumed 4 times as much DDT for each pound of body 

weight than the average adult. However, the amounts of DDT consumed were much smaller 

than the amounts that have been tested in studies in animals. 

DDT from the mother can enter her unborn baby through the placenta. DDT has been found in 

amniotic fluid, human placentas, fetuses, and umbilical cord blood. DDT has been measured in 

human milk; therefore, nursing infants are also exposed to DDT. In most cases, however, the 

benefits of breast-feeding outweigh any risks from exposure to DDT in mother’s milk. 

Nevertheless, women with unusually high amounts of DDT or metabolites in their bodies 

(compared to background amounts measured in the general population) should be informed of 

the potential exposure of the fetus if they become pregnant and the potential risks of breastfeeding. 

We do not know whether children differ from adults in their susceptibility to health effects from 

DDT. There are few studies of young children exposed to DDT. A child who drank DDT in 

kerosene only once vomited and had tremors and convulsions and eventually died; however, we 

do not know how much of this was caused by the kerosene. Adults who swallowed DDT in 

much greater amounts than those found in the environment had effects on their nervous systems. 

The same harmful effects will probably happen to young children if they eat food or drink 
DDT, DDE, and DDD 9 

1. PUBLIC HEALTH STATEMENT 

liquids with large amounts of DDT. However, because DDT is no longer used or made in the 

United States, such exposure is not likely to happen. Two studies have shown a higher dose of 

DDT is needed to kill newborn and young rats than adult rats. In one study, when the dose was 

divided up and given over 4 days, the same dose of DDT killed rats of all ages. 

There is no evidence that exposure to DDT at levels found in the environment causes birth 

defects in people. One study in U.S. children 12 to 14 years of age found that boys whose 

mothers had higher DDE levels in their bodies when they were pregnant were taller than those 

whose mothers had lower DDE levels. A study of German children found that girls with higher 

DDE in the blood at 8 years of age were shorter than those with lower DDE levels. The reason 

for the discrepancy between the two studies is unknown. Studies in animals have shown that 

DDT given during pregnancy can slow the growth of the fetus. Exposure to DDT or its 

metabolites during development may change how the reproductive and nervous systems work. 

This seems to be caused by the property of DDT or its metabolites to mimic the action of natural 

hormones. Male rats exposed to the DDT breakdown product, p,p’-DDE, as fetuses or while 

nursing, showed changes in the development of their reproductive system. One study found that 

the beginning of puberty is delayed in male rats given relatively high amounts of p,p’-DDE as 

juveniles. Also, one study showed that exposure of mice to DDT during the first weeks of life 

resulted in neurobehavioral problems when tests were done later in life. These studies raise 

concerns that exposure to DDT early in life might cause harmful effects that remain or begin 

long after exposure has stopped. 

More information regarding children’s health and DDT and related compounds can be found in 

Section 3.8. 
DDT, DDE, and DDD 10 

1. PUBLIC HEALTH STATEMENT 

1.7	 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO DDT, DDE, AND 

DDD? 

If your doctor finds that you have been exposed to significant amounts of DDT, DDE, and DDD, 

ask whether your children might also be exposed. Your doctor might need to ask your state 

health department to investigate. 

At this time, most people are exposed to DDT and its breakdown products as a result of eating 

foods or drinking liquids that may be contaminated with small amounts of DDT. DDT is a 

pesticide, but it was banned in the United States in 1972. However, because of its chemical 

characteristics, it has stayed in the environment and low levels of DDT may be present in foods 

(i.e., fruits, vegetables, meat, and fish) for many years. Studies have shown that cooking will 

reduce the amount of DDT in fish. Many other countries still use DDT; therefore, food brought 

into the United States from these countries may contain DDT. The Food and Drug 

Administration (FDA) analyzes a wide variety of imported food items (coffee, tropical fruits, 

etc.) as well as domestic products to insure that pesticide residues are below FDA tolerances. 

DDT has been found in both root and leafy vegetables. DDT attaches to the roots of plants, but 

it does not easily move to other parts of the plants. DDT in the air can be deposited on to the 

surfaces of plants. Washing fruits and vegetables before eating them is a healthful practice. 

You and your children may be exposed to DDT by eating certain types of fish or wildlife caught 

from certain locations. Some states, Native American tribes, and U.S. territories have issued fish 

and wildlife advisories to warn people about DDT-contaminated fish and turtles. Each state, 

Native American tribe, or U.S. territory sets its own criteria for issuing fish and wildlife 

advisories. A fish advisory will specify which bodies of water have restrictions. The advisory 

will tell you what type and sizes of fish are of concern. The advisory may completely ban eating 

fish or tell you to limit your meals of a certain fish type. For example, an advisory may tell you 

to eat a certain type of fish no more than once a month. The advisory may tell you only to eat 

certain parts of the fish or turtle and how to prepare or cook the fish or turtle to decrease your 

exposure to DDT. The fish or wildlife advisory may be stricter to protect pregnant women, 

nursing mothers, and young children. To reduce your children’s exposure to DDT, obey fish and 
DDT, DDE, and DDD 11 

1. PUBLIC HEALTH STATEMENT 

wildlife advisories. Information on fish and wildlife advisories in your state is available from 

your state health or natural resources department. Signs may also be posted in certain fishing 

areas. 

More information regarding exposure to DDT can be found in Sections 6.5, 6.6, and 6.7. 

1.8	 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN 

EXPOSED TO DDT, DDE, AND DDD? 

DDT, DDE, and DDD can be measured in fat, blood, urine, semen, and breast milk. Samples of 

blood and urine are easy to get, and levels in these samples may help show the amount of 

exposure. These tests are not readily available at your doctor's office, but your doctor can tell 

you where they can be done. Tests may show low, moderate, or excessive exposure to these 

compounds. However, such tests cannot show the exact amount of DDT, DDE, or DDD to 

which a person was exposed, or predict the chance of health effects in the person. See 

Chapters 3 and 7 for more information on tests to detect these compounds in the body. 

1.9	 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO 

PROTECT HUMAN HEALTH? 

The federal government develops regulations and recommendations to protect public health. 

Regulations can be enforced by law. Federal agencies that develop regulations for toxic 

substances include the Environmental Protection Agency (EPA), the Occupational Safety and 

Health Administration (OSHA), and the Food and Drug Administration (FDA). 

Recommendations provide valuable guidelines to protect public health but cannot be enforced by 

law. Federal organizations that develop recommendations for toxic substances include the 

Agency for Toxic Substances and Disease Registry (ATSDR) and the National Institute for 

Occupational Safety and Health (NIOSH). 

Regulations and recommendations can be expressed in not-to-exceed levels in air, water, soil, or 

food that are usually based on levels that affect animals; then they are adjusted to help protect 
DDT, DDE, and DDD 12 

1. PUBLIC HEALTH STATEMENT 

people. Sometimes these not-to-exceed levels differ among federal organizations because of 

different exposure times (an 8-hour workday or a 24-hour day), the use of different animal 

studies, or other factors. 

Recommendations and regulations are also periodically updated as more information becomes 

available. For the most current information, check with the federal agency or organization that 

provides it. Some regulations and recommendations for DDT, DDE, and DDD include the 

following: 

All uses of DDT were banned by EPA in 1972, except in cases of public health emergencies. 

DDT was banned because the chemical was building up in the environment and possibly hurting 

wildlife. Also, some cancer tests in laboratory animals showed positive results. Although DDT 

is no longer used in the United States, federal regulations still control the amounts of DDT 

allowed in food and water. 

OSHA states that workers may not be exposed to amounts of DDT greater than 1 milligram of 

DDT per cubic meter of air (1 mg/m3) for an 8-hour workday, 40-hour work week. EPA 

estimates that drinking 2 liters of water per day containing 0.59 nanograms of DDT per liter of 

water (1 nanogram is one billionth of a gram) and eating 6.5 grams of fish and shellfish per day 

(from waters containing 0.59 nanograms DDT per liter) would be associated with an increased 

lifetime cancer risk of one in one million. Fish and shellfish tend to concentrate DDT from the 

surrounding water in their tissues. FDA has set action levels for DDT/DDE/DDD; these are 

limits at or above which FDA will take legal action to remove products from the market. Action 

levels for a variety of products are listed in Table 8-1. See Chapter 8 for more information on 

regulations. 
DDT, DDE, and DDD 13 

1. PUBLIC HEALTH STATEMENT 

1.10 WHERE CAN I GET MORE INFORMATION? 

If you have any more questions or concerns, please contact your community or state health or 

environmental quality department or 

Agency for Toxic Substances and Disease Registry
 

Division of Toxicology
 

1600 Clifton Road NE, Mailstop E-29
 

Atlanta, GA 30333
 

Web site: http://www.atsdr.cdc.gov
 

* Information line and technical assistance 

Phone: 1-888-42-ATSDR (1-888-422-8737)
 

Fax: 1-404-498-0057
 

ATSDR can also tell you the location of occupational and environmental health clinics. These 

clinics specialize in recognizing, evaluating, and treating illnesses resulting from exposure to 

hazardous substances. 

* To order toxicological profiles, contact 

National Technical Information Service
 

5285 Port Royal Road
 

Springfield, VA 22161
 

Phone: 1-800-553-6847 or 1-703-605-6000
 

DDT, DDE, and DDD 15 

2. RELEVANCE TO PUBLIC HEALTH
 

2.1	 BACKGROUND AND ENVIRONMENTAL EXPOSURES TO DDT/DDE/DDD IN THE 

UNITED STATES 

DDT is an organochlorine insecticide that has found a broad range of agricultural and nonagricultural 

applications in the United States and worldwide beginning in 1939. In 1972, DDT use was banned in the 

United States and in many parts of the world, except for use in controlling emergency public health 

problems. DDT is still used in certain parts of the world to control vector-borne diseases, such as malaria. 

The release of DDT into the environment occurs primarily through spraying applications onto agricultural 

crops, forest lands, other nonagricultural land, and homes (for the control of disease-bearing vectors). 

Exposures in the home occurred through the use of DDT as a mothproofing agent, to control lice and, in 

some parts of the world, to control mosquitoes and other disease-bearing vectors. Both DDD and DDE 

are degradation products of DDT. DDD was also manufactured and used as an insecticide, but to a much 

lesser extent than DDT. DDE has no commercial use, but is commonly detected along with DDT at 

concentrations in the environment that often exceed those measured for DDT. 

Upon introduction into the environment, DDT will enter soil, water, or air. DDT and its metabolites are 

essentially immobile in soil, becoming strongly absorbed onto the surface layer of soils. Likewise, as a 

consequence of their extremely low water solubilities, DDT and its metabolites become absorbed onto 

particulates in water and settle into sediments. Because of its chemical characteristics, DDT can undergo 

long-range transport through the atmosphere in a process known as “global distillation” where DDT 

migrates from warmer regions to colder regions through repeated cycles of volatilization from soil and 

water surfaces followed by deposition of DDT onto surfaces through dry and wet deposition processes. 

This long-range transport of DDT results in the wide dispersion of DDT and its metabolites throughout 

the world, even into remote areas, such as the Arctic or Antarctic regions. The rate and extent of 

disappearance of DDT may result from transport processes as well as from degradation and 

transformation. Both the slow degradation of DDT and its metabolites and the ban on DDT use in the 

early 1970s in the United States and most of the world have contributed to a decrease in the levels of 

these compounds in the environment over the past 30 years. DDT can be degraded through atmospheric 

photooxidation in air or photolysis on the surface of water or soil. DDT can undergo slow biodegradation 

through reductive dechlorination to form DDE and DDD, and then be further degraded to form other 

metabolites. The persistence of DDT and its metabolites, in combination with their high lipophilicity, 

have contributed to the bioaccumulation (increasing concentration of a chemical in an organism which 
DDT, DDE, and DDD 16 

2. RELEVANCE TO PUBLIC HEALTH 

exceeds that in its environment) and biomagnification (increasing concentration of a chemical in an 

organism as a function of trophic level) of DDT and its degradation products in the environment. DDT, 

DDE, and DDD accumulate in fatty tissues, with tissue concentrations typically increasing with the 

trophic level of the organism. In addition to DDT, DDE, and DDD, methylsulfonyl derivatives of DDE 

can be formed in higher organisms, such as seals, polar bears, beluga whales, and humans, and are found 

predominantly in the liver and fatty tissues. 

Exposure of the general public to DDT, DDE, and DDD has been declining since the ban on the use of 

DDT, at rates that depend on geographic location and environmental conditions. The predominant route 

of exposure of the general public to DDT and its metabolites is through the diet. Although DDT and its 

metabolites are ubiquitous in the atmosphere, they are present in such low concentrations that exposures 

through inhalation or dermal contact are considered to be negligible. From the standpoint of dietary 

exposures, the main exposure route is through the consumption of foods either obtained from areas of the 

world where DDT is still used or that have the potential to contain bioaccumulated residues of DDT and 

its metabolites (e.g., meat, fish, poultry, dairy products). Exposure to DDT in drinking water is 

considered negligible because of the extremely low water solubility of DDT and the efficiency of 

standard drinking water methods. With the ban on the use of DDT, occupational exposures that result 

from formulation, packaging, and application activities should be negligible, except in areas where DDT 

use remains. Activities that result in the mobilization of DDT (e.g., site remediation) may increase 

exposure of workers to DDT and its metabolites. With the discontinued use of DDT, exposure of 

populations living near areas of heavy DDT use or deposition may not be much greater than for the 

general public, except in areas of the world where DDT is still used to control disease-bearing vectors or 

under conditions where DDT and metabolites become mobilized (e.g., site remediation, sediment 

resuspension, erosion, etc.). Residents living near National Priorities List (NPL) sites that contain DDT 

may be exposed to higher ambient levels of DDT and its metabolites than the general population through 

dermal (deposition onto plants, objects), ingestion (contaminated food, water), and/or inhalation of DDT 

vapor, although inhalation exposures are thought to be insignificant compared to dietary sources of DDT. 

These exposures can lead to greater body burdens of DDT and its metabolites; for example, residents 

living near a pesticide dump site in Aberdeen, North Carolina, which is known to contain DDT, have 

higher age-adjusted mean levels of DDE in their blood than residents of neighboring communities 

(4.05 vs. 2.85 ppb). 

Exposures of wildlife to DDT and its metabolites have been declining since the early 1970s, as evidenced 

by marked decreases in the levels of these compounds in fish, shellfish, aquatic mammals, birds, 
DDT, DDE, and DDD 17 

2. RELEVANCE TO PUBLIC HEALTH 

invertebrates and many other species. Yet, DDT and its metabolites were detected in 94% of whole fish 

samples in the 1990s even though total DDT concentration in fish continues to decline. This is due to the 

presence of DDT in stream beds and the continued input of DDT to streams through long-range aerial 

transport and deposition of DDT and erosion of soil. The DDT metabolite, p,p’-DDE, was detected with 

the highest frequency in fish, followed by p,p’-DDD and p,p’-DDT; accumulation of p,p’-DDE in fish is 

due to absorption of DDE from the diet rather than to recent exposures to DDT. Differences in exposures 

of benthic and pelagic feeders to DDT and its metabolites have been observed. In the marine 

environment, there is evidence to suggest that benthic organisms feeding at the bottom of the oceans 

accumulate higher levels of p,p’-DDT and its metabolites, p,p’-DDE and o,p’-DDE, than do surface 

feeders. This is attributed to the absorption of p,p’-DDT, p,p’-DDE, and o,p’-DDE onto organic 

particulates and detritus that then settle to the ocean floor. However, in a fresh water lake system, it was 

found that bioaccumulation of GDDT was greater in the pelagic food web than in the benthic food web 

due to the increased mobility of pelagic organisms and their ultimate reliance on benthos. The GDDT 

that has accumulated in these aquatic food webs can then be further accumulated in organisms in higher 

trophic levels of the food chain. Exposure of wildlife to DDT and its metabolites can also occur through 

DDT-contaminated soils and sediments. 

Since the ban on DDT was instituted in the United States and most of the world in 1972, the 

environmental concentrations of DDT and its metabolites have been decreasing. Average adult intakes of 

DDT were estimated to be 62 μg/person/day in 1965 and 240 μg/person/day in 1970, before the DDT ban 

was instituted. The FDA Total Diet Studies show that the daily intakes have fallen since the ban, with 

daily intakes (for a 16-year-old, 70 kg male) averaging 6.51, 2.38, 1.49, and 0.97 μg/person/day for 

1978–1979, 1979–1980, 1984–1986, and 1986–1991, respectively. As would be expected from the 

decline in the concentrations of DDT in the environment, the levels of DDT, DDE, and DDD measured in 

foodstuffs have also fallen over the last 30 years. Yet, there are still measurable quantities of DDT, DDE, 

and DDD in many commodities. The mean concentrations of p,p’-DDE within specific food groups 

ranged from 0.1 to 25.7 ppb in dairy products, 0.7–6.5 ppb in meat, 0.2–9.2 ppb in fish, and 0.3–0.7 ppb 

in poultry, as determined from the data collected in the FDA Total Diet Studies conducted between 1991 

and 1999. DDT was detected in 312 out of 2,464 plain milk samples at a maximum concentration of 

0.92 ppm and in 8 out of 180 vitamin D fortified milk samples at a maximum concentration of 0.10 ppm 

in 1985–1991. In a recent FDA Total Diet Study, the mean concentrations of p,p-DDT, o,p-DDT, and 

p,p-DDE ranged from 0.0001 to 0.0257 ppm. People who eat fish caught in the Great Lakes were found 

to consume larger amounts of DDT in their diets. However, this route of exposure to DDT is expected to 

decline as the levels of DDT in the environment continue to diminish. Even so, it is anticipated that low 
DDT, DDE, and DDD 18 

2. RELEVANCE TO PUBLIC HEALTH 

levels of DDT and its metabolites will be present in the human diet for several more decades, especially in 

food items, such as fish, that are expected to contain bioaccumulated DDT residues. 

Exposures of the general public to DDT and its metabolites result in the accumulation of these 

compounds in adipose tissue. Due to the persistence of DDT and its metabolites, the concentrations of 

these compounds in adipose tissue are determined by both past and current exposures. Estimated national 

means for DDT concentrations in adipose tissue on a lipid basis were 189, 123, and 177 ppb in 1982, 

1984, and 1986, respectively. For DDE, the mean concentrations were found to be 1,840, 1,150, and 

2,340 ppb, respectively. In a study of people from northern Texas, the concentrations of DDT and DDE 

in adipose tissues decreased from 7,950 ppb in 1970 to 5,150 ppb in 1974, and then to 1,670 ppb in 1983. 

More current measurements of the mean concentrations (lipid basis) of DDT, DDE, and DDD in breast 

adipose tissue collected during 1995–1996 in the United States were 267.3, 709.1, and 24.0 ppb, 

respectively, (DDT+DDE+DDD=1,000.4 ppb), which are consistent with the expectation of continuing 

declines in GDDT body burdens. The adrenocorticolytic agent, 3-methylsulfonyl-DDE, has been 

detected in the human liver (1.1 ppb), lung (0.3 ppb), and adipose tissue (6.8 ppb) (based on wet weight), 

and also in breast milk (0.1 ppb). In a study of Swedish women, concentrations of 3-methylsulfonyl-DDE 

in breast milk ranged from 0.4 to 5 ppb. 

At risk populations include indigenous peoples of the arctic who have diets that are high in fatty tissues 

from the consumption of traditional foods like seal, cariboo, narwhales, etc. Intake of GDDT averaged 

24.2–27.8 μg/day for eastern Arctic communities and 0.51–1.0 μg/day for western Arctic communities. It 

is likely that fish-eating populations in the higher latitudes of the Southern Hemisphere may also be at 

risk; however, these exposures have not been well documented. Children are also at risk to exposure to 

DDT, mainly through dietary sources. The mean intakes of DDT and its metabolites during 1986–1991 

have been estimated to be 0.0448 and 0.0438 μg/kg body weight/day for a 6–11-month-old infant and a 

2-year-old child, respectively, which is roughly 4 times the intake per body weight for an adult. Increased 

infant exposure to DDT has been attributed to breast feeding. The average levels of DDT in human breast 

milk fat were about 2,000–5,000 ppb in the early 1970s, but have steadily declined at a rate of 11–21% 

per year since 1975. For example, Norén reported concentrations of p,p’-DDT in breast milk fat of 0.71, 

0.36, 0.18, and 0.061 ppm for the years 1972, 1976, 1980, and 1984–1985, respectively. These 

investigators also reported concentrations of p,p’-DDE of 2.42, 1.53, 0.99, and 0.50 ppm for these same 

years, respectively. The mean concentrations of p,p’-DDE are roughly 10 times greater than those 

obtained for p,p’-DDT concentrations. Some more recent measurements taken in 1992 show a mean 

DDE concentration of 222.3 ppb in breast milk of Canadian women. Measurements of DDE and DDT in 
DDT, DDE, and DDD 19 

2. RELEVANCE TO PUBLIC HEALTH 

Swedish women in 1997 show levels of 129 and 14 ppb, respectively. Consumption of cow’s milk is 

another route of potential exposure of DDT to children; however, the concentrations of DDT and its 

metabolites are typically lower in cow’s milk than in human breast milk. The mean level of p,p’-DDE in 

cow’s milk was measured at 96 ppb in 1970 and 16 ppb in 1985–1986, showing a sharp decline over the 

years. 

2.2 SUMMARY OF HEALTH EFFECTS 

Numerous studies have been conducted on the effects of DDT and related compounds in a variety of 

animal species, but the human data are somewhat limited. Most of the information on health effects in 

humans comes from studies of workers in DDT-manufacturing plants or spray applicators who had 

occupational exposure to DDT over an extended period and also from some controlled exposure studies 

with volunteers. Epidemiological studies of the general population are also available. Because of 

limitations inherent to all epidemiological studies, disease causality cannot be determined from them; 

however, epidemiological studies have been conducted that allow the evaluation of the potential role of 

DDT and related compounds in specific health outcomes. 

DDT is an organochlorine pesticide whose best known effect is impairment of nerve impulse conduction. 

Effects of DDT on the nervous system have been observed in both humans and animals and can vary from 

mild altered sensations to tremors and convulsions. Humans have been reported to tolerate doses as high 

as 285 mg/kg without fatal result, although because vomiting occurred, the absorbed dose is not known. 

There are no documented unequivocal reports of a fatal human poisoning occurring exclusively from 

ingestion of pure DDT, but deaths have been reported following ingestion of commercial preparations 

containing also other substances. Death in animals following high exposure to DDT is usually caused by 

respiratory arrest. In addition to being a neurotoxicant, DDT is capable of inducing marked alterations on 

reproduction and development in animals. This is attributed to hormone-altering actions of DDT isomers 

and/or metabolites. Of all the DDT-related compounds, the o,p’-DDT isomer has the strongest estrogenlike properties, although it is still several orders of magnitude less potent than the natural hormone, 

17β-estradiol. p,p’-DDE, a metabolite of DDT and a persistent environmental pollutant, has 

antiandrogenic properties and has been shown to alter the development of reproductive organs when 

administered perinatally to rats. There have been studies in humans suggesting that high DDT/DDE 

burdens may be associated with alterations in end points that are controlled by hormonal function such as 

duration of lactation, maintenance of pregnancy, and fertility. High blood levels of DDE during 
 

DDT, DDE, and DDD 20 

2. RELEVANCE TO PUBLIC HEALTH 

pregnancy have also been associated with increased odds of having pre-term infants and small-forgestational-age infants and height abnormalities in children. 

Studies in animals have shown that DDT, DDE, and DDD can cause cancer, primarily in the liver. The 

possible association between exposure to DDT and various types of cancers in humans has been studied 

extensively, particularly breast cancer. Thus far, there is no conclusive evidence linking DDT and related 

compounds to cancer in humans. Possible genotoxic effects in humans have been reported in a few 

studies, but simultaneous exposure to other chemicals and lack of control for relevant confounders make 

the results inconclusive. For the most part, DDT and related compounds are not mutagenic in prokaryotic 

organisms. Studies in animals have shown that DDT can cause adverse liver effects, but studies of 

humans exposed occupationally or of volunteers given DDT have reported only mild liver alterations of 

no clinical significance. Very limited information exists on immunological effects of DDT in humans. A 

recent study of subjects residing near a waste site found that individuals with higher blood DDE levels 

had minor changes in some immune markers, which the authors considered of uncertain clinical 

importance. Studies in animals suggest that exposure to DDT may impair immunocompetence. There is 

sufficient information indicating that the adrenal gland is a target for o,p’-DDD toxicity in humans and 

animals and of the sulfonyl metabolite 3-MeSO2-DDE in animals. o,p’-DDD has been used in humans to 

treat adrenocortical carcinoma and benign Cushing’s disease. 

Discussion of some end points affected by exposure to DDT/DDE/DDD has been expanded below. These 

include end points that have generated considerable interest due to public health implications such as 

various types of cancer and reproductive/developmental effects that may be caused by alterations of the 

endocrine system. Also included are neurological effects, since the main effect of DDT is on the nervous 

system, and hepatic effects, since the liver is significantly affected in several animal species. 

Reproductive/Developmental Effects. In recent years, concern has been raised that many 

pesticides and industrial chemicals are endocrine-active compounds capable of having widespread effects 

on humans and wildlife. Hormones influence the growth, differentiation, and functioning of many target 

tissues, including male and female reproductive organs and ducts such as mammary gland, uterus, vagina, 

ovary, testes, epididymis, and prostate. Therefore, mimicking the effects of hormones or antagonizing 

hormonal effects can potentially affect a number of organs and systems, especially if this occurs at 

vulnerable times such as during development. Developing organisms respond to endocrine-disrupting 

chemicals very differently than adults. Fetuses lack feedback regulatory mechanisms and active 

metabolism of steroid hormones that regulate maintenance of secondary sex tissues, estrous cycling, and 
DDT, DDE, and DDD 21 

2. RELEVANCE TO PUBLIC HEALTH 

pregnancy. Consequently, low levels of exogenous hormones or xenobiotics may induce severe effects in 

the development of the reproductive organs. Thus far, there is no conclusive evidence that exposure to 

DDT/DDE/DDD at the levels found in the environment has affected reproduction and development in 

humans, but there is sufficient information from animal studies that these chemicals have the potential for 

doing so. 

Several studies in humans have examined possible associations between body burdens of DDT and 

analogues and the incidence of alterations in reproduction and development and in other end points in 

which hormones may play a pivotal role such as lactation, infertility, miscarriages, and pre-term labor. 

For example, duration of lactation was inversely related to the concentration of p,p’-DDE in a group of 

women from Mexico who had lactated previously and similar findings were reported in a study in the 

United States. A plausible explanation for this decrease in the longevity of lactation is the 

antiandrogenic/weak estrogenic effect of DDE, as estrogens inhibit milk secretion. In a study of 

240 women, there was no association between blood DDE levels (<6 ppb) and miscarriages, but there was 

suggestive evidence of an association in a study of 89 German women, 13 of whom had DDE blood 

levels >2.5 ppb; the highest DDE level was 8.6 ppb. A weak association between infertility and DDT 

blood levels was observed in an additional study of 489 German women. Studies in India and Israel have 

reported higher levels of DDT, DDE, and DDD in maternal blood and in placental tissue in mothers who 

gave birth to premature infants or who spontaneously aborted fetuses compared to mothers who gave 

birth to full-term infants, but the possible role of other potentially endocrine-disrupting chemicals, such as 

PCBs and other chlorinated pesticides, could not be ruled out. A smaller study in the United States of 

only 20 women did not find the same association between serum DDE levels and pre-term delivery. Yet 

another study found no differences in DDT levels in maternal blood at delivery between full-term babies 

and cases of pre-term delivery; however, they found that umbilical cord blood from pre-term newborns 

had higher levels of DDT than full-term newborns. A recent study of 361 pre-term infants and 221 smallfor-gestational-age cases found that the odds for both outcomes increased steadily with increasing 

maternal blood concentrations of DDE in samples taken in the third trimester of pregnancy. The 

association was evident at DDE concentrations $10 ppb, but there was essentially no relation at lower 

DDE concentrations. It is important to point out that while levels of DDE in the general population in the 

United States continue to decline and current serum levels average <10 ppb, it is not unusual to find 

higher levels in populations from countries where DDT is still used for controlling vector-transmitted 

diseases. High blood levels of DDE during pregnancy have also been associated with height 

abnormalities in children. 
DDT, DDE, and DDD 22 

2. RELEVANCE TO PUBLIC HEALTH 

Technical DDT is a mixture of various isomers of which p,p’-DDT is the most prevalent (65–80%), 

followed by o,p’-DDT (15–21%). Although p,p’-DDT is more acutely toxic than o,p’-DDT, it has been 

known for several decades that the latter had marked effects on reproduction in animals. In most 

experiments, o,p’-DDT mimicked the action of the natural ligand, 17β-estradiol. In general, results from 

in vivo studies indicate that DDT and analogues have much lower estrogenic potency 

(1,000–1,000,000-fold lower) than the endogenous hormone, 17β-estradiol. Experiments in rats have 

shown that o,p’-DDT can induce delayed vaginal opening and increased uterine and ovarian weights. In 

another study in rats, using the uterine glycogen response assay, it was observed that o,p’-DDT was the 

most potent among several isomers. p,p’-DDT and o,p’-DDE were approximately 16 times less potent 

than o,p’-DDT, and o,p’-DDD was inactive in this assay. In yet another study, the lowest dose of 

o,p’-DDT needed to produce estrogenic effects in the immature rat uterus was about 1x105

 times greater 

than that of DES. An in vivo assay with rats and mink showed that o,p’-DDT had uterotropic activity, 

whereas p,p’-DDT had only slight activity, and the activity of technical DDT was dependent on the level 

of o,p’-DDT that it contained. Many other in vivo studies have shown the estrogenic potential of DDT 

and related analogues. Most recent studies have been conducted using a wide variety of in vitro assays 

and the results vary greatly. Still, the estrogenicity of DDT-related compounds in these tests has been 

determined to be 104

–106 times lower than 17β-estradiol, depending on the testing conditions. 

Safe conducted a mass/potency balance exercise to assess the impact of environmental estrogens as 

causative agents of adverse health effects, primarily reproductive disturbances and breast cancer. Besides 

environmental estrogens (xenoestrogens), humans are exposed to several structural classes of naturally 

occurring estrogens including plant bioflavonoids and various mycotoxins. The estrogenic activities of 

these substances has been investigated in in vivo and in vitro cell systems and in ER binding assays; most 

elicit multiple estrogenic responses. Also, a number of foodstuffs contain 17β-estradiol and estrone. At 

the same time, several different structural classes of chemicals present in the human diet show 

antiestrogenic activity, including 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and related halogenated 

aromatics. Polynuclear aromatic hydrocarbons (PAH), found in cooked foods, and indole-3-carbinol 

(IC), a major component of cruciferous vegetables, have also shown antiestrogenic properties. From 

published data, Safe estimated a dietary estrogen equivalent level of 2.5x10-6 μg/day from estrogenic 

pesticides and of 102 μg/day from bioflavonoids. Studies from the literature also suggested that the 

estrogenic potencies of bioflavonoids relative to 17β-estradiol (potency=1) are 0.001–0.0001, whereas for 

pesticides, the estrogenic potency is about 0.000001. Therefore, the estrogen equivalent from dietary 

intake of flavonoids was estimated to be 4x107

 times higher than that from estrogenic pesticides. A 

similar exercise with the antiestrogenic compounds led Safe to the conclusion that the sum of antiestrogen 
DDT, DDE, and DDD 23 

2. RELEVANCE TO PUBLIC HEALTH 

toxic equivalents are orders of magnitude higher than the estimated dietary intakes of estrogenic pesticide 

equivalents. These conclusions are not necessarily contradictory to the recommendation that intake of 

pesticides should continue to be monitored and kept to a minimum for the protection of developing 

organisms, but rather point out the multiple sources of potentially endocrine-active substances. One 

additional fact to keep in mind is that DDT and analogues bioconcentrate in the food chain and 

accumulate in the body. 

Recent studies in vivo have also demonstrated that p,p’-DDE, the most environmentally relevant DDT 

derivative, has antiandrogenic properties when administered to rodents at relatively high doses. Studies 

in vitro have confirmed the findings in vivo. In particular, p,p’-DDE functions as an antagonist after it 

has bound to the androgen receptor. This androgen antagonism or antiandrogenic activity can explain a 

number of reproductive and developmental effects seen in male rats of various ages exposed to p,p’-DDE. 

These effects include reduced anogenital distance and retention of thoracic nipples in pups exposed 

during gestation and lactation; delayed puberty in rats exposed either during juvenile development or at 

very high doses during gestation and lactation; and reduced accessory sex organ weights in exposed adult 

males. In addition to being antiandrogenic by an action on the androgen receptor, p,p’-DDE was shown 

to increase liver CYP enzyme systems that hydroxylate testosterone in perinatally exposed rats. This 

effect can be also considered an antiandrogenic effect since hydroxylated testosterone is more polar than 

testosterone and can be conjugated and excreted. It is not known, however, whether the increase in 

activity of enzymes that hydroxylate testosterone resulted in a reduction of circulating androgens since 

the authors did not measure serum testosterone. In general, altering steroid metabolism is another 

possible mechanisms by which DDT and related compounds can alter reproduction and development. 

p,p’-DDE also was shown to increase liver aromatase activity in adult male rats; aromatase plays a critical 

role in steroidogenesis by catalyzing the conversion of C19 androgens into estrogens. Thus, like DDT, 

DDE can have an overall feminizing effect on animals by antagonizing the androgen receptor at the same 

time that it increases estrogens concentration. 

Administration of DDT in utero or to neonates during sensitive periods in nervous system development 

has caused behavioral and neurochemical changes in adult mice. In the series of experiments conducted 

by Eriksson and co-workers, exposure of 10-day-old mice to a single relatively low dose of 0.5 mg of 

technical DDT/kg resulted in altered motor behavior at the age of 4–5 months and in alterations in 

neurotransmitter receptors in the brain. The findings from the Eriksson’s studies served as the basis for 

derivation of an acute oral minimum risk level (MRL) for DDT. 
 

DDT, DDE, and DDD 24 

2. RELEVANCE TO PUBLIC HEALTH 

Cancer. The possibility that exposure to DDT and related compounds might increase the risk of 

certain cancers in humans has been explored in many studies. Special attention has been devoted, but not 

limited, to breast cancer. Moreover, the interest is not limited to DDT/DDE only, but to persistent 

organochlorine compounds in general (e.g., PCBs, dieldrin, hexachlorobenzene). In addition to factors 

inherent to epidemiological studies that play a role in trying to establish associations between exposure 

and health outcome, for breast cancer, one must consider additional factors that are thought to play 

important roles such as menopause, estrogen receptor status, and concurrent exposure to other estrogenic 

and antiestrogenic compounds. Several methods for assessing combined current and past exposures to 

DDT have been used. Most DDT (and the most prominent environmental residue, DDE) is stored in the 

fat, so an accurate method of assessing body burden is measuring levels in adipose tissue samples. 

However, procedures for obtaining adipose tissue samples are usually invasive and not easy to obtain 

(unless done during cancer surgery). Alternatively, serum DDT/DDD/DDE levels can be measured; 

however, since DDT/DDD/DDE are fat soluble, the serum levels may vary with the fat content of the 

blood. Variability in serum DDT/DDE/DDD that is attributable to variability in serum lipids can be 

accounted for by using lipid-adjusted serum DDT/DDE/DDD levels. Adjusting for lipid content provides 

a standardized body burden estimation and makes comparisons between studies easier. Studies have 

shown that on a fat-weight basis, DDT residue levels in serum correlate well with adipose tissue samples, 

making serum an adequate biological medium for estimating DDT body burden. 

The second issue related to appropriate exposure assessment in human breast cancer studies pertains to 

the timing of the exposure assessment relative to the etiology of cancer. Cancer is a chronic disease and 

can have a latency time of 15–20 years after initiation. Theoretically, DDT could contribute to breast 

cancer either by being a complete carcinogen, an initiator, or a promoter. Thus, for DDT to cause or 

contribute to breast cancer, exposure needs to occur at a time substantially before the time of diagnosis of 

the cancer. A common methodological approach in many of the cancer studies reported below has been 

to assess DDT body burdens at or shortly before the time of breast cancer diagnosis. This method is 

limited by the assumption that a “snapshot” of DDT body burden obtained near the time of diagnosis 

represents DDT exposures at the time of cancer initiation or early promotion, perhaps as much as 

15–20 years earlier. This assumption could be in error for several reasons. First, a large amount of the 

DDT body burden measured could represent more recent exposures, but it could also represent residues 

originating from past high exposures to the pesticide. In developed countries where DDT use is banned 

and average exposures in food have declined considerably with time, current DDT burdens most likely 

represent past exposure. Second, the presence of the cancer itself or other conditions associated with the 

cancer might influence DDT metabolism, excretion, or the use of fat stores and the mobilization of fat­
DDT, DDE, and DDD 25 

2. RELEVANCE TO PUBLIC HEALTH 

stored DDT/DDE/DDD to the blood, and hence, the concentration of DDT/DDE/DDD in blood, fat, and 

other tissues. Several studies have had access to blood samples collected several years before the 

diagnosis of cancer was made; however, there is no information available regarding possible exposures in 

utero or during adolescence that may have contributed to the development of cancer later in life. Third, 

recent reports have shown that obese and lean women metabolize organochlorines at a different rate. 

Therefore, there may be misclassification of exposure depending on time of measurement. 

Thus far, the majority of the epidemiological studies conducted in the United States and abroad, 

particularly the most recent ones and those that included larger sample sizes, have found no association 

between DDT/DDE and breast cancer. However, some studies did find positive associations between 

DDT/DDE body burdens and breast cancer. The study of Canadian women by Demers et al. provided a 

new and important finding that deserves special attention. Demers et al. found that while mean plasma 

concentrations of DDT and DDE in women with breast cancer were not significantly different from those 

of controls, there seemed to be an association between DDE and cancer aggressiveness, as judged by 

increased tumor size and axillary lymph node involvement. A similar finding was reported by Woolcott 

et al., who found that the concentrations of DDE and DDT were higher in adipose tissue from cases with 

more poorly differentiated tumors than cases with moderately or well differentiated tumors. It seems that 

replication of these results is necessary before further speculating on the clinical significance of these 

findings. Details of these studies can be found in Section 3.2.2.7. 

Possible associations with other cancers that have been investigated include pancreatic cancer, lymphoma, 

multiple myeloma, prostate and testicular cancer, endometrial cancer, and liver cancer. Overall, in spite 

of some positive correlations, there is no clear evidence that exposure to DDT/DDE causes cancer in 

humans. The case of pancreatic cancer deserves special mention because, as discussed by Hoppin et al., 

this cancer is characterized by cachexia, and the impact of this on serum organochlorine levels is difficult 

to predict. As the lipid pool decreases, the concentration of a fixed amount of lipid soluble chemical 

increases. While in the Hoppin et al. study, cases and controls had almost the same DDE serum levels 

before adjusting for lipid contents, adjusted values were much higher in cases than in controls. This 

points to the need to better understand changes in pharmacokinetics that occur as a result of a disease state 

and how these changes can influence the interpretation of the results. While there is no clear evidence of 

carcinogenicity of DDT/DDE/DDD in humans, there is sufficient evidence of carcinogenicity of these 

compounds in rodents, that developed tumors primarily in the liver following long-term oral treatment 

with these compounds. Two earlier studies in monkeys treated chronically with DDT reported no 

evidence of carcinogenicity, but a more recent study that involved Rhesus and Cynomolgus monkeys 
 

 

DDT, DDE, and DDD 26 

2. RELEVANCE TO PUBLIC HEALTH 

administered p,p’-DDT in the diet for up to 130 months reported that 2 out of 13 Cynomolgus monkeys 

developed malignant tumors, one hepatocellular carcinoma and one adenocarcinoma of the prostate. This 

corresponds to an incidence of 15%, which is significant compared to no neoplasms found in a group of 

nine contemporary untreated control monkeys. 

The EPA has assigned DDT, DDE, and DDD to Group B2, probable human carcinogens. The 

International Agency for Research on Cancer (IARC) has determined that DDT, DDE, and DDD are 

possibly carcinogenic to humans (Group 2B). The Department of Health and Human Services (DHHS) 

has determined that DDT, DDE, and DDD may reasonably be anticipated to be human carcinogens. 

Neurological Effects. The main neurophysiological effect of DDT is to prolong the sodium current 

in axons drastically, which accounts for the increase in depolarizing after-potential and the resultant 

repetitive afterdischarges in nerve fibers and synaptic junctions. As many basic functions such 

respiratory and cardiovascular functions are controlled by the nervous system, exposure to high amounts 

of DDT is expected to produce a wide array of symptoms and central and peripheral signs of toxicity. 

High oral doses of DDT in humans result in paresthesia of the tongue, lips, and face, as well as 

apprehension, hypersusceptibility to external stimuli, irritability, dizziness, vertigo, tremor, and tonic and 

clonic convulsions. In general, results from studies in animals are consistent with observations in 

humans. Earlier studies of workers exposed to DDT (approximately 42 mg/man/day) and also studies 

with volunteers who ingested up to 35 mg DDT/man/day up to 18 months did not report adverse 

neurological effects, as assessed by tests of motor and sensory functions. However, a recent study of 

27 workers in Costa Rica who used DDT for malaria control between 1955 and 1986 found that exposed 

subjects performed worst on tests of verbal attention and visuomotor speed and sequencing than a group 

of controls. However, because exposure levels were not available and current body burdens that could 

have been used to back estimate exposure were not determined, it is difficult to reconcile these findings 

with those from the earlier studies. 

Hepatic Effects. Although there is no conclusive evidence in the available studies that the human 

liver is a primary target for DDT toxicity or that exposure to DDT causes liver toxicity in humans, many 

studies have reported hepatic effects in animals following exposure to DDT and related compounds. 

Effects observed include induction of microsomal enzymes, increased serum transaminase activities of 

hepatic origin, liver hypertrophy, hyperplasia, and necrosis, and liver cancer. Induction of microsomal 

enzymes is important because it can lead to altered metabolism of exogenous and endogenous substrates, 

including steroid hormones. In fact, as previously mentioned, alteration of hormone metabolism has been 
DDT, DDE, and DDD 27 

2. RELEVANCE TO PUBLIC HEALTH 

suggested to be at least a contributing factor in the DDT-induced reproductive and developmental effects. 

Based on induction profiles obtained in rats, DDT and related compounds are considered primarily, 

although not exclusively, phenobarbital-type inducers. Liver effects were the most sensitive effects 

observed in animals treated with DDT by relevant routes of exposure. Thus, an intermediate-duration 

study in rats that provided dose-response information for liver effects was used to derive an intermediate 

oral MRL for DDT. 

2.3 MINIMAL RISK LEVELS 

Inhalation MRLs 

No MRLs were derived for inhalation exposure since adequate experimental data were not available by 

this route of exposure. 

Oral MRLs 

C An acute oral MRL of 0.0005 mg /kg/day was derived for DDT. 

This MRL is based on a LOAEL of 0.5 mg/kg/day for neurodevelopmental effects in mice reported in a 

series of studies from the same group of investigators (Eriksson and Nordberg 1986; Eriksson et al. 

1990a, 1990b, 1992, 1993; Johansson et al. 1995, 1996; Talts et al. 1998). An uncertainty factor of 

1,000 was used (10 for use of a LOAEL, 10 for animal to human extrapolation, and 10 to account for 

intrahuman variation). 

The most significant finding was the presence of altered motor behavior in adult mice treated with DDT 

perinatally. Groups of 10-day-old male NMRI mice were treated by gavage with a single dose of 

0 (vehicle control) or 0.5 mg DDT/kg in a fat emulsion vehicle by gavage (Eriksson et al. 1990a) (this 

same single dose level was used in all studies from the series). At the age of 4 months, the mice were 

subjected to behavioral tests of spontaneous activity (locomotion, rearing, and total activity). Tests were 

conducted for 1 hour, and scores were summed for three 20-minute periods. During the last 40 minutes of 

testing, the treated mice showed significantly more activity than untreated controls. This was interpreted 

as disruption of a simple, nonassociative learning process (i.e., habituation), or a retardation in adjustment 

to a new environment. These same results were reported in a later paper (Eriksson et al. 1990b) in which 

the authors also reported results of neurochemical evaluations conducted 2–3 weeks after behavioral 
DDT, DDE, and DDD 28 

2. RELEVANCE TO PUBLIC HEALTH 

testing. They measured muscarinic acetylcholine (MACh) receptors density and choline acetyltransferase 

(ChAT) activity in the cerebral cortex and hippocampus (MACh also in striatum) and also measured 

K+

-stimulated ACh release from cerebral cortex slices. In addition, five 10-day-old mice were 

administered 0.5 mg 14C-DDT and the radioactivity in the brain was assayed 24 hours, 7 days, or 1 month 

after dosing. The results showed that K+

-evoked ACh release in treated mice was significantly increased 

relative to controls, ChAT activity was not changed in the cerebral cortex or hippocampus, and the 

density of MACh was not significantly changed in the hippocampus or striatum, but a decreasing trend 

was seen in the cerebral cortex. DDT-derived radioactivity could be detected until day 7 after dosing, but 

none could be detected 1 month after dosing. 

Previous studies have shown a significant increase in density of MACh in the cerebral cortex of 10-dayold mice 7 days after dosing, but not at 1 day postexposure compared to controls (Eriksson and Nordberg 

1986). No increased binding was noted in the hippocampus either 1 or 7 days posttreatment. This was 

further investigated by evaluating the proportion of high- and low-affinity binding sites and affinity 

constants of the muscarinic receptors. A significant increase in the percentage of low-affinity binding 

sites accompanied by a significant decrease in high-affinity binding sites was measured in the cerebral 

cortex 7 days postexposure. No significant changes in affinity constants were noted. According to the 

authors, these low-affinity binding sites correspond to the M1 receptor in the cerebral cortex, which are 

thought to be associated with neuronal excitation. No changes were observed in the sodium-dependent 

choline uptake system in the cerebral cortex 7 days postexposure. 

In a follow-up study, Eriksson et al. (1992) treated 3-, 10-, and 19-day-old mice, and conducted 

behavioral testing and neurochemical evaluations at 4 months of age. As previously published, mice 

treated at 10 days old exhibited hyperactivity relative to controls and also had a significant decrease in the 

density of MACh in the cerebral cortex. No such changes were seen in mice treated at 3 or 19 days old. 

The authors suggested that the changes in MACh density and behavior might be the consequence of early 

interference with muscarinic cholinergic transmission specifically around the age of 10 days. In 

subsequent studies by the same group, mice were tested at 5 and 7 months old (Eriksson et al. 1993; 

Johansson et al. 1995). At both time points, mice treated with DDT perinatally showed increased 

spontaneous motor activity relative to controls, and decreased density of MACh in the cerebral cortex. 

No changes were seen regarding percentages of high- or low-affinity muscarinic binding sites in the 

cerebral cortex. Mice in these studies were also treated orally with the type I pyrethroid insecticide, 

bioallethrin, and tested for motor activity at 5 months old (Eriksson et al. 1993) and 7 months old 

(Johansson et al. 1995). In general, mice treated with DDT at the age of 10 days and later with 
DDT, DDE, and DDD 29 

2. RELEVANCE TO PUBLIC HEALTH 

bioallethrin showed increased motor behavior relative to those treated with bioallethrin alone, suggesting 

a DDT-induced increased susceptibility to bioallethrin. Mice treated first with DDT and later with 

bioallethrin showed increased difficulties in learning a skill, such as the swim maze test, compared with 

untreated mice, mice treated with DDT alone, or mice treated with bioallethrin alone (Johansson et al. 

1995). Also, treatment with DDT followed by bioallethrin significantly increased the density of 

muscarinic receptors in the cerebral cortex relative to DDT alone. This increase was later attributed to 

increased expression of muscarinic receptor m4 mRNA (Talts et al. 1998). In yet another study from this 

group, paraoxon replaced bioallethrin, and the mice were tested at 5 and 7 months old (Johansson et al. 

1996). In addition, acetylcholinesterase activity was measured in cerebral cortex of 5-month-old mice 

and MACh and nicotinic cholinergic receptors were measured in the cortex of 7-month-old mice. 

Relevant new findings include that: (1) DDT did not significantly alter acetylcholinesterase activity; 

(2) DDT did not alter the effects of paraoxon on acetylcholinesterase activity (decreased); (3) DDT 

altered (increased or decreased) some of motor responses due to paraoxon alone at 7 months, but not at 

5 months; (4) none of the treatments altered performance in the swim maze test; and (5) none of the 

treatments altered the density of nicotinic cholinergic receptors in the cortex. 

In this series of studies, two responses seem to be consistent from study to study in mice treated with 

DDT perinatally and tested as adults, a decrease in the density of muscarinic cholinergic receptors in the 

cerebral cortex and increased spontaneous motor activity. Whether or not there is a causal relationship is 

not clear. DDT also altered some motor responses induced by other pesticides, but a pattern was not 

always clear. The investigators interpreted the latter findings as DDT inducing changes early in the brain 

that translated into increased susceptibility to other pesticides later in life. The DDT-induced increase in 

spontaneous motor activity at the dose of 0.5 mg/kg is considered a less serious lowest-observed-adverseeffect level (LOAEL). 

C An intermediate oral MRL of 0.0005 mg/kg/day was derived for DDT. 

This MRL is based on a no-observed-adverse-effect level (NOAEL) of 0.05–0.09 mg/kg/day for liver 

effects in Osborne-Mendel rats administered technical DDT in the diet at dosage levels of 0, 1, 5, 10, or 

50 ppm for 15–27 weeks (Laug et al. 1950). In this dietary study, the amount of DDT added to the food 

was measured, but the actual food consumption and body weights of the rats were not measured. The 

calculated values for the NOAEL range from 0.05 to 0.09 mg/kg/day depending on the food consumption 

values used to calculate the actual dose of DDT consumed. Details about the dose calculation are 

discussed in Appendix A. The approximate doses provided in the diet in this study were 0.05–0.09, 
DDT, DDE, and DDD 30 

2. RELEVANCE TO PUBLIC HEALTH 

0.25–0.5, 0.5–0.9, and 2.5–4.6 mg/kg/day. An uncertainty factor of 100 was used (10 for interspecies 

extrapolation and 10 for human variability). The selected effects for the intermediate-duration oral MRL 

represent reliable effects that occurred at the lowest administered dose (including endocrine effects), as 

seen in Table 3-1 (see Chapter 3); therefore, the intermediate oral MRL is protective for endocrine effects 

and other non-cancer effects. 

This study was essentially designed to examine whether DDT accumulates in adipose tissue and to what 

extent, how age and dose level affect accumulation, and how rapidly it is eliminated. Seventy-seven rats 

were used for microscopic evaluation of only the liver and kidney. This was based on findings from a 

previous study from the same group (Fitzhugh and Nelson 1947, see below) in which higher dietary levels 

of DDT had been used. Based on the previous findings, only the liver was expected to show microscopic 

changes. Although not explicitly stated, it is assumed that morphologic evaluations were conducted at the 

times when DDT levels in fat were determined (after 15, 19, 23, and 27 weeks of treatment). 

There were no morphologic alterations in the kidneys. Liver alterations were noticed at the 5 ppm 

(0.25–0.5 mg/kg/day) dietary level of DDT and higher, but not at 1 ppm (0.05–0.09 mg/kg/day). Liver 

changes consisted of hepatic cell enlargement, especially in central lobules, increased cytoplasmic 

oxyphilia with sometimes a semihyaline appearance, and more peripheral location of the basophilic 

cytoplasmic granules. Necrosis was not observed. The severity of the effects was dose-related, and males 

tended to show more hepatic cell changes than females. Changes seen at the 5 ppm level 

(0.25–05 mg/kg/day) were considered by the authors as ”minimal”. 

In the Fitzhugh and Nelson (1947) study, 16 female Osborne-Mendel rats were fed a diet containing 

1,000 ppm technical DDT (approximately 96 mg/kg/day) for 12 weeks. Sacrifices were conducted at 

cessation of dosing and at various intervals after a DDT-free period. Liver changes were similar to those 

seen in the Laug et al. (1950) study, although of increased severity, and were still present in rats killed 

after 2 weeks in a DDT-free diet. Minimal liver changes were apparent after 4–6 weeks of recovery, and 

complete recovery was seen after 8 weeks. Hepatic effects ranging from increased liver weight to cellular 

necrosis have been reported in animals after chronic exposure to DDT in the diet. 

An oral MRL for chronic-duration exposure was not derived because of the inadequacy of the available 

data on liver effects in animals to describe the dose-response relationship at low dose levels. The liver 

appears to be the most sensitive target of DDT for chronic duration exposures. In a brief communication, 

Fitzhugh (1948) stated that histopathological lesions occurred in the liver of rats fed 10 ppm DDT in the 
DDT, DDE, and DDD 31 

2. RELEVANCE TO PUBLIC HEALTH 

diet for 2 years, but no experimental details were given, so the quality of the study cannot be evaluated. 

This dietary level was the lowest level tested in the study, but was still higher than the lowest level 

resulting in hepatic effects in the Laug et al. (1950) study used for derivation of the intermediate-duration 

MRL. 

DDT, DDE, and DDD 33 

3. HEALTH EFFECTS 

3.1 INTRODUCTION 

The primary purpose of this chapter is to provide public health officials, physicians, toxicologists, and 

other interested individuals and groups with an overall perspective on the toxicology of DDT, DDE, and 

DDD. It contains descriptions and evaluations of toxicological studies and epidemiological investigations 

and provides conclusions, where possible, on the relevance of toxicity and toxicokinetic data to public 

health. Information on the effects of DDT, DDE, and DDD in humans and in animal species traditionally 

used in laboratory experiments is presented in Section 3.2 Discussion of Health Effects by Route of 

Exposure, whereas information on the effects of these compounds in wildlife is presented in Appendix D, 

Health Effects in Wildlife Potentially Relevant to Human Health. 

A glossary and list of acronyms, abbreviations, and symbols can be found at the end of this profile. 

While this document is specifically focused on the primary forms or isomers of DDT, DDE, and DDD 

(namely p,p’-DDT, p,p’-DDE, and p,p’-DDD), other isomers of these compounds will be discussed when 

appropriate. In some cases, the term DDT will be used to refer to the collective forms of DDT, DDE, and 

DDD. Should this not be clear from the context, the term ΣDDT (Σ is used to mean sum of) will be used. 

Typically, people are not exposed to DDT, DDE, or DDD individually, but rather to a mixture of all three 

compounds since DDE and DDD are degradation and metabolic products of DDT. In addition, DDT, 

DDE, and DDD each can exist in three isomeric forms based on the relative position of the chlorine 

substitutions on the two chlorophenyl rings (Chapter 4). The most prevalent isomer of DDT, DDE, or 

DDD in the environment is the p,p’-isomer. Technical-grade DDT contains 65–80% p,p’-DDT, 15–21% 

o,p’-DDT, and up to 4% of p,p’-DDD (Metcalf 1995). When the toxicity of the isomers of DDT, DDE, 

or DDD reported in the experimental data differ in an organ system, such as the reproductive or 

developmental systems, isomer-specific results are presented, when available. Therefore, the data 

presented in this document include some relevant toxicity information on DDE and DDD analogues and 

on the o,p’- and p,p’-isomers of DDT and technical-grade DDT. Levels of Significant Exposure (LSE) 

for the o,p’-isomers are presented in separate tables to emphasize the point that these isomers have a 

somewhat different spectrum of effects than the p,p’-isomers. 
DDT, DDE, and DDD 34 

3. HEALTH EFFECTS 

3.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE 

To help public health professionals and others address the needs of persons living or working near 

hazardous waste sites, the information in this section is organized first by route of exposure (inhalation, 

oral, and dermal) and then by health effect (death, systemic, immunological, neurological, reproductive, 

developmental, genotoxic, and carcinogenic effects). These data are discussed in terms of three exposure 

periods: acute (14 days or less), intermediate (15–364 days), and chronic (365 days or more). 

Levels of significant exposure for each route and duration are presented in tables and illustrated in 

figures. The points in the figures showing no-observed-adverse-effect levels (NOAELs) or 

lowest-observed-adverse-effect levels (LOAELs) reflect the actual doses (levels of exposure) used in the 

studies. LOAELS have been classified into "less serious" or "serious" effects. "Serious" effects are those 

that evoke failure in a biological system and can lead to morbidity or mortality (e.g., acute respiratory 

distress or death). "Less serious" effects are those that are not expected to cause significant dysfunction 

or death, or those whose significance to the organism is not entirely clear. ATSDR acknowledges that a 

considerable amount of judgment may be required in establishing whether an end point should be 

classified as a NOAEL, "less serious" LOAEL, or "serious" LOAEL, and that in some cases, there will be 

insufficient data to decide whether the effect is indicative of significant dysfunction. However, the 

Agency has established guidelines and policies that are used to classify these end points. ATSDR 

believes that there is sufficient merit in this approach to warrant an attempt at distinguishing between 

"less serious" and "serious" effects. The distinction between "less serious" effects and "serious" effects is 

considered to be important because it helps the users of the profiles to identify levels of exposure at which 

major health effects start to appear. LOAELs or NOAELs should also help in determining whether or not 

the effects vary with dose and/or duration, and place into perspective the possible significance of these 

effects to human health. 

The significance of the exposure levels shown in the Levels of Significant Exposure (LSE) tables and 

figures may differ depending on the user's perspective. Public health officials and others concerned with 

appropriate actions to take at hazardous waste sites may want information on levels of exposure 

associated with more subtle effects in humans or animals (LOAELs) or exposure levels below which no 

adverse effects (NOAELs) have been observed. Estimates of levels posing minimal risk to humans 

(Minimal Risk Levels or MRLs) may be of interest to health professionals and citizens alike. 
DDT, DDE, and DDD 35 

3. HEALTH EFFECTS 

Levels of exposure associated with carcinogenic effects (Cancer Effect Levels, CELs) of DDT, DDE, and 

DDD are indicated in Table 3-1 and Figure 3-1. Because cancer effects could occur at lower exposure 

levels, Figure 3-1 also shows a range for the upper bound of estimated excess risks, ranging from a risk of 

1 in 10,000 to 1 in 10,000,000 (10-4 to 10-7), as developed by EPA. 

Estimates of exposure levels posing minimal risk to humans (Minimal Risk Levels or MRLs) have been 

made for DDT, DDE, and DDD. An MRL is defined as an estimate of daily human exposure to a 

substance that is likely to be without an appreciable risk of adverse effects (noncarcinogenic) over a 

specified duration of exposure. MRLs are derived when reliable and sufficient data exist to identify the 

target organ(s) of effect or the most sensitive health effect(s) for a specific duration within a given route 

of exposure. MRLs are based on noncancerous health effects only and do not consider carcinogenic 

effects. MRLs can be derived for acute, intermediate, and chronic duration exposures for inhalation and 

oral routes. Appropriate methodology does not exist to develop MRLs for dermal exposure. 

Although methods have been established to derive these levels (Barnes and Dourson 1988; EPA 1990a), 

uncertainties are associated with these techniques. Furthermore, ATSDR acknowledges additional 

uncertainties inherent in the application of the procedures to derive less than lifetime MRLs. As an 

example, acute inhalation MRLs may not be protective for health effects that are delayed in development 

or are acquired following repeated acute insults, such as hypersensitivity reactions, asthma, or chronic 

bronchitis. As these kinds of health effects data become available and methods to assess levels of 

significant human exposure improve, these MRLs will be revised. 

A User's Guide has been provided at the end of this profile (see Appendix B). This guide should aid in 

the interpretation of the tables and figures for Levels of Significant Exposure and the MRLs. 

3.2.1 Inhalation Exposure 

Occupational exposure to DDT involves multiple routes of exposure. The primary routes of exposure 

were probably inhalation and dermal; however, absorption of DDT from the lungs may not have been 

significant, and ingestion due to the mucociliary apparatus of the respiratory tract is more likely. 

Therefore, with the exception of a report on lung cancer, epidemiological studies of occupational 

exposure will be discussed under oral exposure and the following section refers to nonoccupational 

inhalation exposure. 
 

 

DDT, DDE, and DDD 36 

3. HEALTH EFFECTS 

3.2.1.1 Death 

No studies were located regarding death in humans or animals after inhalation exposure to DDT, DDE, or 

DDD. 

3.2.1.2 Systemic Effects 

No studies were located regarding cardiovascular, gastrointestinal, hematological, musculoskeletal, 

hepatic, renal, or dermal effects in humans or animals after inhalation exposure to DDT, DDE, or DDD. 

Respiratory Effects. Volunteers were exposed by inhalation of aerosols containing DDT at 

concentrations that left a white deposit on the nasal hair (Neal et al. 1944). Except for moderate irritation 

of the nose, throat, and eyes, which may have been related to the vehicle to disperse DDT in an aerosol, 

no significant changes were reported. This study had several limitations. The study did not provide 

information concerning conditions of exposure, dose, or information on persons exposed. 

No studies were located regarding the respiratory effects in animals after inhalation exposure to DDT, 

DDE, or DDD. 

Ocular Effects. Reports of ocular effects in humans exposed to DDT in the air are limited to the study 

by Neal et al. (1944). In this study, moderate, nonspecific eye irritation was reported by volunteers 

exposed to an aerosol containing DDT. This effect is assumed to have been caused by direct contact of 

the aerosol with the eye and not by inhalation of the aerosol. No information on the length of exposure or 

the concentration of DDT in air was provided. Irritation may have been related to the vehicle in which 

DDT was dissolved. 

No studies were located regarding the ocular effects in animals after inhalation exposure to DDT. 
DDT, DDE, and DDD 37 

3. HEALTH EFFECTS 

No studies were located regarding the following effects in humans or animals after inhalation exposure to 

DDT, DDE, or DDD: 

3.2.1.3 Immunological and Lymphoreticular Effects 

3.2.1.4 Neurological Effects 

3.2.1.5 Reproductive Effects 

3.2.1.6 Developmental Effects 

3.2.1.7 Cancer 

Occupational exposure to DDT was associated with increased lung cancer in a case control study of the 

Uruguayan work force (De Stefani et al. 1996). Elevated, but not statistically significant, odds ratios 

(OR) for any type of lung cancer were observed in 34 workers who had been exposed for 1–20 years 

(OR=1.6; 95% confidence interval [CI]=0.9–4.6), in 16 workers who had been exposed for greater than 

20 years (OR=2.0; 95% CI=0.9–4.7), and in 50 workers who had ever been exposed to DDT (OR=1.7; 

95% CI=1.0–2.8). Significantly elevated odds ratios were reported in a subset of 33 DDT-exposed lung 

cancer patients with small cell cancer (OR=3.6; 95% CI=1.5–8.9) or in 57 with adenocarcinoma (OR=2.3; 

95% CI=1.2–4.7). Analyses were adjusted for age, residence, education, tobacco smoking, and alcohol 

consumption. 

Several additional studies of workers occupationally exposed to DDT are discussed in Section 3.2.2.8 as 

exposure most probably occurred by ingestion of trapped particles rather than by inhalation. 

No studies were located regarding cancer in animals after inhalation exposure to DDT, DDE, or DDD. 

EPA (IRIS 2001a, 2001b, 2001c) calculated an inhalation unit risk of 9.7x10-5 per μg/m3

 for DDT from 

oral data in animals (Section 3.2.2.8). The air concentrations corresponding to excess cancer risk levels 

of 1x10-4, 1x10-5, 1x10-6, and 1x10-7 are 1x10-3, 1x10-4, 1x10-5, and 1x10-6 mg/m3

, respectively. 

3.2.2 Oral Exposure 

Tables 3-1 and 3-2 and Figures 3-1 and 3-2 present the health effects observed in humans and animals 

associated with levels of significant oral exposure for each designated exposure duration. The levels of 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) 

K~Ytoa Species frequency NOAEL Less serious 

Exposure/ 

duration/ 

figure (Strain) (Specific route) System (mglkglday) (~sncs/daY) 

- Oral 

LOAEL 

Serious 

(mglkglday) 

Reference 

Chemical Form 

0 

-2 

a 

0 

0 

rn 

(u 

0 

0 

ACUTE EXPOSURE 

300 (LD50) Ben-Dykeet 

al. 1970 

DDT, NS 

400 (LD50) Ben-Dykeet 

al. 1970 

DDD, NS 

4 

5 

Rat 

(Sherman) 

Rat 

(Sherman) 

Once 

Once 

(G) 

800 (LD50) 

113 (LD50) 

4000 (LD50) 

Cameronand 

Burgess 1945 

DDT, NS 

Gaines 1969 

DDT, p,p' 

Gaines 1969 

" DDD, Tech 

0 

I 

rn 

$

-I

I 

rn 

n 

n 

rn 

0 

-I 

(I, 

6 Rat 

(Sherman) 

Once 

(G) 

880 (LD50) Gaines 1969 

DDE, Tech 

7 

8 

9 

10 

Rat 

(NS) 

(NS) 

(NS) 

(NS) 

Rat 

Rat 

Rat 

4d 

(G) 

(G) 

(GI 

(GI 

4d 

Once 

Once 

279.2 

285.6 

355.2 

194.5 

(4-day LD50, preweanling; 

cumulative dose) 

(4-day adult LD50; 

cumulative dose) 

(LD50, weanling rats) 

(LD50, adult rats) 

Lu et al. 1965 

DDT, Tech 

Lu et al. 1965 

DDT, Tech 

Lu et al. 1965 

DDT, Tech 

Lu et al. 1965 

DDT, Tech W 

00 
Table 3-1. Levels of Significant Exposureto DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0 

-2 

0 

Exposure1 LOAEL 0rn 

duration/ a, 

K~Ytoa Species frequency NOAEL Less serious Serious Reference a3 

figure (Strain) (Specific route) System (mglkglday) (mglkglday) (mglkg/day) Chemical Form 0 

4000 (LD50, newborn rats) Lu et al. 1965 U 

11 Rat 

DDT, Tech (NS) 

12 Rat 437.8 (LD50, preweanlingrats) Lu et a!. 1965 

DDT, Tech (NS) 

300 M (LD50) Kashyap et al. 13 Mouse 

1977 (Inbred 

Swiss) DDT, Tech

14 Mouse 85.7 F (50% of mice killedafter a Okey and 

6-day feeding period) Page 1974 w

(C3W DDT, P,P' 

15 Mouse 1466 (LD50) Tomatis et al. 

1972 

(CF1) m-n DDD, P,P' -n 

m 

16 Mouse 237 (LD50) Tomatis et al. 

* 1972 

(CF1) DDT, Tech 

17 Gn Pig 400 (LD50) Cameron and Burgess 1945 

(NS) DDT, NS 

18 Rabbit 300 (LD50) Cameron and 

Burgess 1945 

(NS) DDT, NS 

Systemic 

19 Monkey Once Hepatic 150 (increased serum LDH, Agarwal et al. 

AP, 1978 and transaminases) (Rhesus) (GI DDT, NS 

w

W 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0

0 

-I 

Exposure1 LOAEL 0 

0

duration/ rn 

K~Y frequency Reference toa Species NOAEL Less nl 3 serious Serious 

figure (Strain) (Specific route) 

20 Rat 12 d 

(GI (Wistar) 

System 

Hepatic 

(,,,glkglday) (m@?May) 

40 (18% increase in relative 

liver weight; increased 

enzyme activities) 

liver GSH and AHH 

(mglkgldw) Chemical Form 

DeWaziers 

and Azais 

1987 

DDT, NS 

21 Rat 

(Sprague 

Dawley) 

Once 

(GO) 

Endocr 25 50 (reduced capacity to 

concentrate iodine in 

thyroid) 

Goldman 

1981 

DDT. Tech 

22 

23 

Rat 

(Wistar) 

Mouse 

(CFI1 

14 d 

1 x/d 

(GO) 

1 wk 

(F) 

Hepatic 

Hepatic 

12 M (increased relative liver 

weight; necrotic 

changes) 

42.9 (29% increase absolute 

liver weight; increase 

cytochrome-c reductase 

and P-450) 

Kostka et al. 

2000 

DDT, Tech 

Pasha 1981 

DDE, NS 

24 Dog Once 

(C) (Mongrels 

and Beagles) 

Endocr 200 (reversible degenerative 

adrenal changes) 

Powers et al. 

DDD, Tech 

" 1974 

25 Dog 6 d or 

(C) 

(Mongrels 15d 

and Beagles) 

Endocr 100 (necrosis, adrenal) Powers et al. 

1974 

DDT, Tech 

26 

Immunological/Lymphoreticular 

Rabbit 10 d 

(GI (New 

Zealand) 

4.3 Shiplov et al. 

1972 

DDT, NS 

27 

Neurological 

Human Once 

(F) 

10.3 16 (convulsions) Hsieh 1954 

DDT, NS 

Q

0 

0

0 

w 

I 

5

rn 

I 

rn 

-n 

-n 

m 

0 

-I 

(n 

P 

0 
Table 3-1.Levels ofSignificant Exposureto DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 

Exposure1 LOAEL 

duration/

Key toa Species frequency 

figure (Strain) (Specific route) System 

NOAEL 

(mflglday) 

Less serious 

(rnglkglday) 

Serious 

(mglkglday) 

28 Monkey 

(Rhesus) 

Once 

(G) 

150 (decreaseCNStotal lipids, phospholipids and 

cholesterol) 

29 Rat 160 (tremors) 

(Wistar) 

30 Rat 25 50 (tremors) 

(Fischer) 

31 Rat 100 (myoclonus) 

(Albino 

Sprague 

Dawley) 

32 Mouse 160 M (tremors) 

(albino) 

33 Mouse 200 (convulsions) 

(Albino) 

34 Gn Pig 160 (paralysisof hind legs) 

(NS) 

35 Hamster 160 

(NS) 

Reproductive 

36 

(Long-Evans)4d 

Rat 1x/d 200 M (reduced seminal vesicle 

and ventral prostate 

weight) 

(GO) 

Reference 

Chemical Form 

Sanyal et al. 

1986 

DDT, Tech 

Hietanen and 

Vainio 1976 

DDT, NS 

Hong et al. 

1986 

DDT, p,p' 

Hwang and 

Van Woert 

1978 

DDT, p,P' 

Hietanen and 

Vainio 1976 

DDT, p,p' 

Matin et at. 

1981 

DDT, P,P' 

Hietanen and 

Vainio 1976 

DDT, NS 

Hietanen and 

Vainio 1976 

DDT, NS 

Kelce et al. 

1995 

DDE, P,P' 

w 

I 

rn

F 

--I

I 

rn 

n 

n 

rn 

0 

-I 

v) 

f 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0 

0 

-I 

K~Yto" Species 

figure (Strain) 

37 Rat 

(SpragueDawley) 

Exposure1

duration/ 

frequency 

(Specific route) 

1 x/d 

(GO) 

5d 

System 

NOAEL 

(mglkglday) 

LOAEL 

Less serious 

(mgmwy) 

200 M (reduced seminal vesicle 

and ventral prostate 

weight) 

Serious 

(mglkglday) 

Reference 

Chemical Form 

Kelce et al. 

1997 

DDE, P,P' 

0 

0 

&I 

3 

a 

0 

0 

0 

rn 

38 Rabbit 

(New 

Zealand) 

Gd7-9 

(GO) 

50 (increased resorptions, 1.8% Hart et al. 

in controls, 25% in treated) l971 

DDT, P,P' 

39 Rabbit 

(New 

Zealand) 

Gd7-9 

(GO) 

10 (increased resorptions, 1.3% Hart et al. 

in controls, 9.5% in treated) l97* 

DDT, P,P' 

0 

40 

41 

Developmental 

Rat Gd15-19 

(G) (SpragueDawley) 

Rat 5d 

(Long- Gd 14-1 8 

(GO) Evans) 

28 

00 1 (significant decrease in 

ventral prostate weight) 

Gellert and Heinrichs 

1975 

DDT, P,P' 

.,. Gray et al. 

1999 

DDE, P,P' 

I 

rn

5

I 

rn 

n 

n 

rn 

0 

-I 

0, 

42 Rat 

(SpragueDawley) 

5d 

Gd 14-18 

(GO) 

100 M (decreased weight of glans Gray et a\. 

penis; epididymis and ventral l999 

prostate; reduced anogenital DDE, p,p' 

distance) 

43 Rat 1 x/d 

(GO) 

(Long-Evans)Gd 14-1 8 

100 M (reduced anogenital distance Kelce et al. 

at birth: ppd 13 retained 

thoracic nipples) DDE, P,P' 

1995 

44 Rat 

(Holtzman) 

1 x/d 

(GO) 

Gd 14-18 

10 M 50 M (reduced anogenital distance Loeffler and 

on ppdl and relative ventral Peterson l999 

prostate weight on ppd21) DDE, p,p' 

P 

h) 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 

Exposure1 LOAEL I duration/ 

Key tofigure Species frequency

(Strain) (Specific route) System 

NOAEL 

(mglkglday) 

Less serious 

(msn(s/day) 

Serious 

(mglkglday) 

Reference 

Chemical Form 

45 Rat 1xld 10 100 M (reduced anogenital distance You et al. 

(GO) 

(Long-Evans) Gd 14-18 on ppd2; retained thoracic 

nipples on ppdl3) 

1998 

DDE, P,P' 

10 M (ppdl3 males retained 

thoracic nipples) 

You et al. 

1998 

DDE, P,P' 

O.!jb M (increased motor 

activity at 4 months) 

Eriksson and 

Nordberg 

1986; 

Eriksson et al. 

1990a, 1990b 

DDT, Tech 

48 Mouse Once 0.5 M (decrease in cerebral Eriksson et al. 

(NMRI) 

(GO) 

PPd10 cortex muscarinic 

acetylcholine receptors 

at 4 months) 

1992 

DDT, Tech 

49 Mouse 

(NMRI) 

once 

(GO) 

PPd10 

0.5 M (decreased muscarinic 

receptors in cerebral 

cortex; increased 

spontaneous activity at 5 

Johansson et 

al. 1995 

DDT, Tech 

months) 

50 Mouse 

(NMRI) 

once 

(GO) 

PPd10 

0.5 M (decreased muscarinic 

receptors in cerebral 

cortex; increased 

spontaneous activity at 5 

Johansson et 

al. 1996 

DDT, Tech 

and 7 months) 

51 Rabbit 

(New 

Zealand) 

Gd4-7 

(GI 

1 (decreased fetal weight) Fabro et al. 

1984 

DDT, NS 

-5

0 

0 

0 

rn 

@I

3 

Q 

w 

I 

rn

5

I 

rn 

n 

n

rn 

2 

v, 

P 

W 
Table 3-1, Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0 

0 

-4 

Exposure1 LOAEL 

duration1 

K~Y toaSpecies frequency NOAEL Less serious Serious

figure (Strain) (Specific route) System (mglkgl&,y) (mglkdday) (mglkdday) 

52 Rabbit Gd7-9 50 (22% decreased offspring 

weight ) (New 

Zealand) (GO) 

0 

0 

rn 

Reference ry

3 

Q

Chemical Form 0 

0

0 Hart et al. 

1971 

DDT, P,P' 

53 Rabbit Gd7-9 10 (11% decreased fetal 50 (19% decreased fetal weight Hart et at. 

(New weight on day 28) on day 28) 1972 

(GO) Zealand) DDT,P,P' 

w 

I 

rn

9 

I

5 

P 

P 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0 

0 

-4

- Exposure1 LOAEL 0 

rn 

U 

duration/ a,

Key to" Species frequency NOAEL Less serious Serious Reference 3 

Q. 

figure (Strain) (Specific route) System (mgkdday) (~gncway) (mg/kglday) Chemical Form 0 

0 

0 

INTERMEDIATE EXPOSURE 

Death 

54 Monkey 2,4or 50 (death of 616 in 14weeks) Cranmer et al. 

1972a 

(Squirrel) 6 mo 

DDT, P,P' 

(G) 

55 Mouse 6wk 35 F (4 out 5 died) NCI 1978 

DDT, Tech (B6c3F1) (F) 

66 M (4 out of 5 died) NCI 1978 56 Mouse 6wk 

DDE, p,p' I (B6C3F1) (F) rn

5 Systemic I 

rn 

57 Rat 120 d Endocr 0.2 M (degeneration of adrenal Chowdhury et n

n cortex and medulla) al. 1990 rn 

(albino) 0 (GO) DDT, Tech 4

v, 

Bd Wt 0.2 M (30% reduced body weight ' 

gain) 

58 Rat 3wk Hepatic 15 (significant increase in Gupta et al. 

1989 liver weight and in (Wistar) (GO) cytochrome P-450 DDT, P,P' 

enzymes) 

6.6 (focal Jonsson et al. 59 Rat 36 wk Hepatic 1981 necrosislregeneration) (Sprague 7dlwk 

Dawley) DDT, NS 

(F) 

60 Rat 15-27wk Hepatic 0.05' 0.25 (cellular hypertrophy, Laug et al. 

1950 cytoplasmic (OsborneMendel) (F) eosinophilia) DDT, Tech 

P

ul 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) 

Exposure/

duration/ 

K~Ytoa Species frequency 

figure (Strain) (Specific route) System (,,,@@day) 

NOAEL Less serious 

(mgWCfay) 

61 Rat 6wk Bd Wt 28 M 50 M (16% reduced body 

weight gain) (F) (OsborneMendel) 

- Oral (continued) 

LOAEL 

Serious 

(mg/kg/day) 

97 F (45% reduced body weight 

gain) 

Reference 

Chemical Form 

NCI 1978 

DDT, Tech 

0 

0 

-I 

0 

0 

rn 

L1, 

a 

0 

3 

0

0 

62 Rat 

(OsbomeMendel) 

6wk 

(F) 

Bd Wt 88 M (1 1 % decrease in body 

weight gain) 

157 M (22% decrease in body 

weight gain) 

NCI 1978 

DDE, p,p' 

63 Rat 

(Sherman) 

2-18 mo 

(F) 

Hepatic 5 M 

20 F 

(minor liver vacuolation, 

hypertrophy and cell 

margination) 

Ortega 1956 

DDT. Tech 

64 

65 

Mouse 

(B6C3F1) 

Mouse 

(B6C3F1) 

6wk 

(F) 

6wk 

(F) 

Bd Wt 

Bd Wt 

310 F 

101 F 

NCI 1978 

DDD, Tech 

NCI 1978 

DDE, p,p' 

w 

I 

rn

5

I 

rn 

rn 

-n 

-n 

66 Mouse 6wk Bd Wt 35 NCI 1978 

(B6C3F1) (F) '- DDT, Tech 

67 Mouse 

(NMRI) 

28d 

(G) 

Hepatic 6.25 (increased liver weight 

and P-450 induction) 

Orberg and 

Lundberg 

1974 

DDT, P,P' 

Immunological/Lymphoreticular 

68 Rat 

(Wistar) 

4wk 

(F) 

2.3 M 5.7 M (decreased IgG and IgM, 

increased 

albumin/globulin ratio) 

Banerjee et al. 

1995 

DDT, P,P' 

69 Rat 

(albino) 

31 d 

(F) 

24hr/d 

1.9 (decreased mast cells) Gabliks et al. 

1975 

DDT, NS 

P 

m 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0 

0 

--I 

Exposure1 0 

0 LOAEL rn duration/ Ill K~Ytoa Species frequency NOAEL Serious Reference =I Less serious figure (Strain) (Specific route) System (mg/kg/day) Q 

Chemical Form 0 (mglkglday) (mgk3lday) 0 

0

70 Mouse 3-12wk 10.5 21 M (decreased IgM antibody Banerjee 

1987a titer) (Hissar) (F) DDT, NS 

Neurological 

5 50 (staggering, weakness, loss Cranmer et al. 71 Monkey 2,4, or 

equilibrium) 1972a 

(Squirrel) 6 mo 

DDT, P,P'

(GI 

72 Monkey 100d 10 (15-20% decrease in Sanyal et al. 

1986 brain lipids, CNS (Rhesus) (G) w phospholipids, and DDT, Tech 

cholesterol) 

73 Rat 26 wk 16 F (body tremors) NCI 1978 

m 

n (Osborne- DDT, Tech -n

Mendel) (F) rn

0 

-I 

74 Rat 9wk 34.1 F (tremors in 80% of females Rossi et al. v) 

after 9 weeks of treatment). 1, 1977 (Wistar) (F) DDT, Tech 

Reproductive 

75 Rat 36 wk 6 12 (sterility) Jonsson et al. 

1976 (Sprague 

Dawley) (F) DDT, Tech 

76 Rat 1 x/d 100 M (delayed onset of puberty by Kelce et al. 

5 days) 1995 

(Long-Evans)ppd 2l-57 

DDE, P,P'

(GO) 

Kornburst et 77 Rat 9 wk 10 

al. 1986 (Sprague 

Dawley) (G) DDE, P,P' 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0

U 

--I

- Exposure1 LOAEL U 

0 

rn duration/ 

K~Ytoa Species frequency NOAEL Less serious Serious Reference al

3 

a 

figure (Strain) (Specific route) System (mglkglday) (mglkglday) (msntdday) Chemical Form 0 

0 Krause 1977 0 78 Rat 3 wk 100 (marginal, but significant 

decrease in testosterone DDT, NS (Wistar) 3 dwk 

in the testis) (GO) 

Krause et al. 79 Rat 200 (decreased absolute 1975 testis weight) (NS) DDT, NS 

80 Mouse 60-90 d 51.4 (78% decreased fertility) Bernard and 

Gaertner 1964 

(C-57) (F) DDT, Tech 

Ledoux et al. 81 Mouse 86 d 3.4 I 1977 

5 (B6C3F1) rn 

(F) DDT, Tech 

I 

rn 82 Mouse 1.67 (decreased implanted ova, Lundberg n increased-persistent- estrus) 1974 n (NMRI) rn DDT, P,P' 2 

(I) 

83 Mouse 1.67 (decreased corpora lutea Lundberg

1974 and implants) (NMRI) DDT, P,P' 

Orberg and 84 Mouse 28 d 6.25 Lundberg (NMRI) 1974 (G) 

DDT, P,P' 

Ware and 85 Mouse 120 d 1.3 Good 1967 

(BALBlc) (F) DDT. Tech 

Lindenau et 86 Rabbit 3 F (reduced ovulation rate al. 1994 and slight decrease (New 

Zealand) circulating progesterone DDT, Tech 

post-insemination) 

P 

W 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - oral (continued) 

Exposurd LOAEL I duration/ 

Key tofigure Species frequency

(Strain) (Specific route) System 

NOAEL 

(mglkdday) 

Less serious 

(mgkglday) 

Serious 

(mglkglday) 

Developmental 

87 Rat 42 d 1.7 16.8 (decreased growth of 42.1 (pup death by 10 days) 

(Wistar) Gd 1-21 

Ld 1-21 

nursing pups) 

(F) 

88 Mouse Gd 1-21 34.3 (decreased maze 

(CFI) Ld 1-21 

(F) 

performance learning at 

1 and 2 months in 

survivors) 

Cancer 

89 Mouse 15-30wk 42.8 (liver hepatomas) 

(CF1) (F) 

Reference 

Chemical Form 

Clement Okey 1974 and 

DDT, P,P' 

Craig and 

Ogilvie 1974 

DDT, Tech 

Tomatis et al. 

1974b 

DDT, P,P' 

0 

0 

-i 

0 

0 

rn 

ffl 3 

a 

0 

0 

0 

w 

I 

rn

5

I 

rn 

-n 

-n 

rn

9 cn 

P

W 

Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p’-Isomers) - Oral (continued) 

Exposure/

duration/ 

LOAEL 

K~Ytoa Species figure (Strain) frequency (Specific route) System NOAEL 

(mdkglday) 

Less serious 

(mgWday) 

Serious 

(mglkglday) 

Reference 

Chemical Form 

97 Rat 27 mo Resp 20 al. Deichmann1967 et 

(Osborne Mendel) (F) Hemato 20 (hemolysis in spleen) 

DDT, NS 

Hepatic 20 (focal hepato-cellular 

necrosis) 

Renal 20 (some tubular epithelial 

necrosis and polycystic 

degeneration; small 

hemorrhages) 

98 Rat 

(OsborneMendel) 

2 Yr 

(F) 

Hepatic 7 (focal hepatocellular 

necrosis) 

Fitzhugh and 

Nelson 1947 

DDT, Tech 

u 

I 

rn

2

I 

99 Rat 

(OsborneMendel) 

78 wk 

(F) 

Resp 

Cardio 

45 M 

45 M 

NCI 1978 

DDT, Tech 

rn 

n 

n 

rn

2 u) 

Gastro 45 M 

Musc/skel 45 M 

Hepatic 23 M (fatty metamorphosis) 

Renal 45 M 

Endocr 45 M 

Dermal 45 M 

Bd Wt 32 F (20% decrease in body 

weight gain) 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Orat (continued) 0 

0 

-4 

Exposure/ LOAEL 

duration/ 

K~Ytoa Species frequency NOAEL Less serious Serious 

figure (Strain) (Specific route) System (mfldday) (mWdday) (mWdday) 

100 Rat 78wk Resp 231 M 

(OsborneMendel) (F)

Reference 

Chemical Form 

NCI 1978

DDD, Tech

0

0 

rn 

(u

a 

3-

0 

Cardio 231 M 

Gastro 231 M 

Musc/skel 231 M 

Hepatic 

Renal 

231 M 

66 F (chronic inflammation of 

the kidney) 

Endocr 

Dermal 

Bd Wt 

231 M 

231 M 

66 F (26% decrease in body 

weight gain) 

w

-F 

L 

rn

F

-I

I 

101 Rat 

(OsborneMendel) 

78wk Resp 

Cardio 

59 M 

59 M 

NCI 1978

DDE, p,p' 

rn

-n 

n 

rn

2

vj 

Gastro 59 M 

Musc/skel 59 M 

Hepatic 31 M (fatty metamorphosis) 

Renal 59 M 

Endocr 59 M 

Dermal 59 M 

Bd Wt 19 F (21% decrease in body 

weight gain) 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p’-Isomers) - Oral (continued) 0

0 

Exposure1

duration/ 

K~Ytoa Species frequency figure (Strain) (Specific route) System 

102 Mouse 78wk Resp 

(B6c3F1) (F) 

Cardio 

Gastro 

Musclskel 

Hepatic 

Renal 

Endocr 

Dermal 

Bd Wt 

103 Mouse 78wk Resp 

(B6c3F1) (F) 

Cardio 

Gastro 

Musclskel 

Hepatic 

Renal 

Endocr 

Dermal 

Bd Wt 

-I 

LOAEL 0 

0 

rn 

NOAEL 

(mg/kg/day) 

30.2 F 

Less serious 

(mg/kglday) 

Serious 

(m9Wday) 

Reference 

Chemical Form 

NCI 1978 

a, 

Q 

0 

0 

0 

DDT, Tech 

30.2 F 

30.2 F 

30.2 F 

3.7 M (amyloidosis) 

30.2 F 

30.2 F 

30.2 F 

30.2 F w 

142 F 

142 F 

142 F 

142 F 

NCI 1978 

DDD, Tech 

I 

rn

F 

-I

I 

rn 

n 

n 

rn 

0 

-I 

142 F 0) 

142 F 

142 F 

142 F 

71 F (17% decrease in body 142 F (28% decrease in body 

weight gain) weight gain) 
0 Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) W 

--I 

Exposure1

duration/ 

LOAEL 

Q

0 

rn 

K~Ytoa Species figure (Strain) frequency (Specific route) System NOAEL 

(mfldday) 

Less serious 

(mWdday) 

Serious 

(m9ncslday) 

Reference 

Chemical Form 

a, 

Q 

0 

0 

3 

104 Mouse 78wk Resp 49 F NCI 1978 

(B6C3F1) (F) 

Cardio 49 F 

DDE, p,p' 

Gastro 49 F 

Musc/skel 49 F 

Hepatic 49 F 

Renal 27 M (chronic inflammation of 

the kidney) 

Endocr 49 F 

Dermal 49 F w 

105 Hamster 

(Syrian) 

life 

(F) 

Bd Wt 

Hepatic 

Bd Wt 

20 

40 

40 (focal necrosis) 

28 F (29% decrease in body 

weight gain) 

Cabral et al. 

1982a 

DDT, Tech 

I 

rn

5

I 

rn 

n 

n 

rn 

(n 

7 

106 Hamster 

(NS) 

life 

(F) 

Hepatic 67 (50% increase in relative 

liver weight) 

Graillot et al. 

1975 

DDT. Tech 

107Hamster 

(Syrian) 

128 wk 

(F) 

Hepatic 47.5 (liver necrosis) Rossi et al. 

1983 

DDE, p,p' 

Bd Wt 47.5 (unquantified reduction 

in body weight gain) 

39-40 mo Hepatic 16 80 (focal or diffuse liver 160 (severe liver damage) Lehman 1965 

(F) 

alterations) DDT, Tech 

cn 

P 

Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0 

0 

-i 

118 Mouse 

(SwissWebster) 

K~Ytoa Species 

figure (Strain) 

3 gen 

(F) 

Exposure/

duration/ 

frequency 

(Specific route) System 

NOAEL 

(mdkg/day) 

Less serious 

(mglkglday) 

LOAEL 

Serious 

(mg/kg/day) 

20 (decreased fertility) 

Reference 

Chemical Form 

Keplinger et 

al. 1970 

DDT, NS 

0

0 

a, 

Q 

0 

rn 

3 

0

U 

2.4 Wolfe et al. 

1979 

DDT, Tech 

120 Dog 7dlwk-

(Beagle) F2 gen 

(F) 

Developmental 

121 Rat 2 gen 

(F) (Sprague 

Dawley) 

10 

1.9 18.6 (tail abnormalities, 

constriction rings) 

16.5 (decreased neonatal 

survival) 

Ottoboni et al. 

1977 

DDT. Tech 

Ottoboni 1969 

DDT, Tech 

Del Pup et al. 

1978 

DDT, Tech 

w 

I 

rn

5

I 

rn 

rn 

(I) 

n 

n 

2 

123 Mouse 

(B6C3F1) 

7dlwk2gen 

(F) 

10 50 (increased preweanling 

death) 

Tomatis et al. 

1972 

DDT, P,P' 

124 Mouse 

(CFV 

life 

(F) 

8.3 41.3 (increased in preweanling 

death) 

Turusov et al. 

1973 

DDT, Tech 

Cancer 

125 Rat life 

(MRC Porton) (F) 

19.7 F (liver tumors, NS) Cabral et al. 

1982b 

DDT, Tech 

126 Rat 

(OsborneMendel) 

78wk 

(F) 

116 M (CEL: thyroid follicular cell 

adenoma and carcinoma) 

NCI 1978 

DDD, Tech 
Table 3-1. Levels of Significant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - oral (continued) 0 

0 

-i 

Exposure/

duration/ 

LOAEL 0

0 

rn 

K~Ytoa Species 

figure (Strain) 

127 Rat 

frequency 

(Specific route) 

120 wk 

System NOAEL (,@@Jay) Less serious 

(mglkglday) 

Serious 

(m&Way) 

34.1 M (liver cell tumors; 33.3% 

Reference 

Chemical Form 

Rossi et al. 

Q 

0 

m 

3 

U 

0 

(Wistar) (F) 

incidence, 0% in controls) 

DDT, Tech 

1977 

128Mouse 

(C57BU6N) 

81 wk 

(F) 

28 (liver tumors, NS) lnnes et al. 

1969 

DDT, Tech 

129 Mouse 

(Swiss) 

80 wk 

(F) 

16.5 (lymphomas; lung and liver 

tumors NS) Kashyap et al. 

1977 

DDT, Tech 

130 Mouse 78wk 27 M (CEL: hepatocellular NCI 1978 w 

(B6C3F1) 

131 Mouse 

(A strain) 

(F) 

5 gen 

(GI 

1.3 

carcinomas; 17/47) 0/19, 7/41, 

(lung tumors, NS,lung

adenomas) 

DDE, p,p' 

Shabad et al. 

1973 

DDT, Tech 

I 

rn

5

I 

rn 

-n 

n 

rn 

7 

132 Mouse 

(BALB/c) 

life 

6 gen 

(F) 

0.4 (lung adeno-carcinomas in ',.Wan and 

F2, leukemia in F3) 

DDT, P,P' 

Kemeny 1969 

0) 

133 Mouse 

(BALB/c) 

life 

2-gen 

50 F (liver tumors in FO and F1) Terracini et al. 

1973 

DDT, Tech 

(F) 

134 Mouse 2 Yr 15.8 (liver tumors, NS) Thorpe and 

Walker 1973 

(CF1) (F) DDT, P,P' 

135 Mouse 

W1) 

life 

multi-gen 

(F) 

0.38 M (liver tumors in FO and F1) 

45.5 F (liver tumors in FO and F1) 

Tomatis et al. 

1972 

DDT, Tech 
Table 3-1. Levels ofSignificant Exposure to DDT, DDE, DDD (Except o,p'-Isomers) - Oral (continued) 0

0 

-I 

Exposurd LOAEL 0 

0 duration/ rn 

a, K~Y toa Species frequency NOAEL Less serious Serious Reference 3

figure (Strain) (Specific route) System (,,,@@day) a (mg/kg/day) (F.dWdaY) Chemical Form 8- 136 Mouse 42.6 M (significant increase in liver Tomatis et al. 0 

tumors) 1974a (CF1) 

DDE, P,P' 

137 Mouse 42.6 M (significant increase in lung Tomatis et al. 

and liver tumors) 1974a (CF1) 

DDD, P.P' 

138 Mouse 0.33 (liver tumors, NS) Turusov et al. 

1973 (CF1) DDT, Tech 

139 Hamster 95 (CEL: adrenal neoplasms; et al. I*' 

(Syrian) 14% in controls, 34% in 1983 I 

treated) DDT, Tech rn 

I

5 

140 Hamster 47.5 (CEL: hepatocellular tumors; Rossi et al. rn 

n 0/73, 11/69, 14/78) 1983 (Syrian) n

DDE, P,P' rn 

2 

vj 

The number corresponds to entries in Figure 3-1. 

'Used to derive an acute oral minimal risk level (MRL) of 0.0005 mag-day for DDT. The MRL was derived by dividing the LOAEL by an uncertainty factor of 1000(10 to account 

for intraspecies variability, 10 for interspecies variability, and 10 for the use of a LOAEL) . 

'Used to derive an intermediate minimal risk level (MRL) of 0.0005 mgkglday for DDT. The MRL was derived by dividing the NOAEL by an uncertainty factor of 100 (10 to account 

for interspecies variability and 10 for intraspecies variability). See Appendix A, ATSDR Minimal Risk Level Worksheets, for explanation of how dietary concentrations were converted 

to doses. 

Bd Wt =body weight; (C) =capsule; Cardio =cardiovascular; CEL = cancer effect level; CNS = central nervous system; d =day(s); (F) = food; Endocr - endocrine; F = female; 

(G) =gavage; gastro =gastrointestinal; Gd = gestation day; gen = generation(s); (GO) = gavage in oil; Hemato = hematological; hr = hour(s); kg = killigram; Id = lactation day; 

LD, = lethal dose, 50%kill; LOAEL = lowest-observable-adverse-effect level; M = male; mg = milligram; mo = month(s): Musc/skel = musculoskeletal; 

NOAEL = no-observable-adverse-effect level; NS = not specified; PND = postnatal day; ppd = postpartum day; Resp = respiratory; wk = week(s); x = times 
10000 

1000 

100 

10 

1 

0.I 

0.01 

0.001 

c-Cat 

d-Dog

r-Rat 

p-Pig

q-Cow 

Figure 3-1. Levels of Significant Exposure to DDT, DDE, DDD - Oral (Except o,p' isomers) 

Acute (114 days) 

A27 

A27 

-Humans f-Ferret n-Mink +Cancer Effect Level-Animals Cancer Effect Level-Humans 

k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans 

m-Mouse e-Gerbil CD LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans 

h-Rabbit s-Hamster 0NOAEL -Animals A NOAEL- Humans 

a-Sheep g-Guinea Pig _ _ _ ~ ______ 

030r 

030r 

LD501LC50 

; Minimal Risk Level 

; for effects 

other than 

Cancer 

0 

0 

-I 

W 

a 

w 

I 

F'

rn 

-I

I 

rn 

n 

n 

9

rn 

v) 

Ln 

(D 

0.0001 
mglkglday 

10000 

Figure 3-1. Levels of Significant Exposure to DDT, DDE, DDD - Oral (Except 0.p' isomers) (continued) 

Acute (114 days) 0 

0 

-I 

0

O 

rn 

0) 

a 

0 

0 

0 

1000 

100 

036r 037r 

038h u 

10 039h 

1 

947m 048m 049m 050m 

0.1 

0.01 

0.001 w 

c-Cat -Humans f-Ferret n-Mink +Cancer Effect Level-Animals V Cancer Effect Level-Humans LD50/LC50 a d-Dog k-Monkey j-Pigeon o-Other 0LOAEL, More Serious-Animals A LOAEL, More Serious-Humans : Minimal Risk Level 

r-Rat m-Mouse e-Gerbil 0 LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans : for effects 

p-Pig h-Rabbit s-Hamster 0NOAEL -Animals NOAEL - Humans A other than 

q-Cow a-Sheep g-Guinea Pig Cancer 

0.0001 

0 
Figure 3-1. Levels of Significant Exposure to DDT, DDE, DDD - Oral (Except o,p' isomers) (confinued) 

Intermediate (15-364days) 

mglkglday 

1000 

100 

10 

1 

0.1 

0.01 

0.001 

c-Cat 

d-Dog

r-Rat 

p-Pig 

q-cow 

-Humans 

k-Monkey

m-Mouse 

h-Rabbit 

a-sheep 

064m 

Q79r .62r 

.76r Q78r 

.80m .87r +89m @88m 

@87r 

075r 077r 

084m 07% Q68r 071k 

081m @86h 

068r Q69r 087r .82m 88% 

*Doses represent the lowest dose tested per study that produced a tumorigenic 

response and do not imply the existence of a threshold for the cancer endpoint. 

n-Mink +Cancer Effect Level-Animals 'ICancer Effect Level-Humans LD501LC50 f-Ferret 

j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans I Minimal Risk Level 

e-Gerbil LOAEL, Less Serious-Anlmals A LOAEL, Less Serious-Humans I for effects 

s-Hamster 0NOAEL -Animals A NOAEL - Humans A other than Cancer g-(;uinea Pig -~ - _~~_____~ ----p_____-pp

w 

0.0001 

2 
Figure 3-1. Levels of Significant Exposure to DDT, DDE, DDD - Oral (Except for o,p' isomers) (continued) 

Chronic (2365 days) 

0 

U 

rn 

rJl 

3 

Q 

U 

0 

0 

1000 

100 

I0 090k 

1 

A94 A94 A94 I 

rn 

0.1 

0.01 

0.001 

0.0001 

1E-5 

1E-6 

1E-7 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals 'ICancer Effect Level-Humans LD50/LC50

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans : Minimal Risk Level 

r-Rat m-Mouse e-Gerbil aLOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans ; foreffects 

p-Pig h-Rabbit s-Hamster 0NOAEL - Animals A NOAEL - Humans other than 

__q-Cow a-Sheep- &Guinea Pig Cancer- . ____ - - .. . - -
Figure 3-1. Levels of Significant Exposureto DDT, DDE, DDD - Oral (Except o,p' isomers) (continued) 

Chronic (2365days) 0 

U 

-I 

mglkglday 

1000 

0 1OOr 

100 

0 103m 0103m 

.118m 

10 .110k 0120d 0115 

0112m 

1 

A94 A109 

w 

-P 

0.1 i

I 

rn 

n

n 

0.01 

rn

2 0) 

0.001 

0.0001 

1E-5 

1E-6 

1E-7 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals 'ICancer Effect Level-Humans LD50/LC50 

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans Minimal Risk Level I

I for effects W

ul 

r-Rat m-Mouse e-Gerbil (D LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans 

p-Pig h-Rabbit s-Hamster 0NOAEL-Animals A NOAEL - Humans A other than 

q-cow a-SheefL __gIGumeaPig-. .- -~- -- _- ~ - ~- - __ __ - __ - Cancer _- -____ 
mglkglday 

1000 

100 

10 

1 

0.1 

0.01 

0.001 

0.0001 

1E-5 

1 E-6 

1E-7 

Figure 3-1. Levels of Significant Exposure to DDT, DDE, DDD - Oral (Except o,p' isomers) (continued) 

Chronic (2365 days) 

+I39s e126r 

0121r +I31m 

+I32m +I35m +I 38m 

*Doses represent the lowest dose tested per study that produced a tumorigenic 

response and do not imply the existence of a threshold for the cancer endpoint. 

104 

10-5 

10-c 

10-z 

Estimated 

Upper-Bound 

Human Cancer 

Risk Levels 

c-Cat -Humans f-Ferret n-Mink +Cancer Effect Level-Animals Cancer Effect Level-Humans LD50/LC50

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans : Minimal Risk Level 

0LOAEL, Less Serious-Animals aLOAEL, Less Serious-Humans : for effects r-Rat m-Mouse e-Gerbil p-Pig h-Rabbit s-Hamster 0NOAEL -Animals A NOAEL- Humans A other than 

Cancer 

0 

0 

rn 

u 

rn

5

I 

rn 

n 

n 

rn 

0 

m 

P 
U 

Table 3-2. Levels of Significant Exposure to o,p'-DDT, -DDE, -DDD - Oral -40 

0 

Exposure1 LOAEL 0rn 

duration/ Iu 

~eytoa Species frequency NOAEL Less serious Serious Reference a figure (Strain) (Specific route) System (rngkg/day) (mdkCdday) (mWCdday) Chemical Form 0 0

U 

ACUTE EXPOSURE 

Death 

810 (LD50) Tomatis et al. 1 Mouse Once 1972 

(CF1) (GI DDE, o,p' 

Systemic 

2 Dog 14 d Cardio 50 (decreasein contractile Cueto 1970 

force) DDD, o,p' (NS) 

Endocr 50 (decreasedplasma w 

I glucocorticoids) rn 

3 Dog 10 d Endocr 138.5 (adrenalhemorrhage) Kirk et al. 5

I 1974 rn 

n (NS) (C) DDD, o,p' n 

rn 

0 

Reproductive 0) 

-I 

4 Rat 7d 100 (increased uterus weight; Clement and 

Okey 1972 premature vaginal (Wistar) opening) DDT, o,p' 

(F) 

5 Rat 3d 10 F 100 F (significantincrease in Die1et al. 

wet uterine weight) 2000 

(DNHan) 1 x/d 

DDT, o,p'

03 

Developmental 

Gellert and 6 Rat Gdl5-19 28 

Heinrichs 

(Sprague 1975 (GI Dawley) DDE, o,p' 

m 

u1 
Table 3-2. Levelsof Significant Exposure to o,p'-DDT, -DDE, -DDD 

- Oral (continued) 

00

-+ 

Exposure/

duration/ 

LOAEL 0 

0 

rn 

K~Ytoa Species frequency figure (Strain) (Specific route) 

7 Rat Gdl5-19 

System 

NOAEL 

(mmgday) 

28 

Less serious (mglkglday) 

Serious 

(mglkglday) 

Reference 

Chemical Form 

Gelled and 

a,

3 

0. 0E0 

(Sprague 

Dawley) 6) 

Heinrichs 

1975 

DDT, o,p' 

8 Rat 

(SpragueDawley) 

Gdl5-19 

(G) 

28 (delayed vaginal 

opening) 

Gelled and 

Heinrichs 

1975 

DDD, o,p' 

9 Mouse 

(CD-1) 

7d 

Gd 11-71 

(GO) 

0.018 M (decreased testes 

weight) 

Palanza et al. 

1999 

DDT, o,p' w 

I 

10 Mouse 

(CF1) 

1 x/d 

(GO) 

Gd 11-17 

0.018 M 1.82 M (altered behavior; 

increase urine marking in 

a novel territory) 

vom Saal et 

al. 1995 

DDT, o,p' 

rn5

I 

rn 

n 

n 

rn 

0 

-I 

cn 
Table 3-2. Levels ofSignificant Exposure to o,p'-DDT, -DDE, -DDD 

figure (Strain) (Specific route) System (mglkglday) (mg/kglday) 

Exposure/ 

duration1 

K~Ytoa Species frequency NOAEL Less serious 

INTERMEDIATE EXPOSURE 

- Oral (continued) 

LOAEL 

(mg/kg/day) 

Serious 

Chemical Form 

Reference 

0 

Q 

0 

-4 

-% 

0 

a, 

a 3

0 

0 

Systemic 

11 Dog 

(NS) 

36-150 d 

(C) 

Endocr 50 (adrenocortical necrosis) Kirk and 

Jenson 1975 

DDD, o,p' 

IrnmunologicallLymp horeticular 

12 Rat 

(SpragueDawley) 

16 d 

(F) 

121 (atrophy thymus) Harnid et al. 

1974 

DDD, o,p' w 

13 

Reproductive 

Rat 20 wk 

(F) (Wistar) 

Developmental 

4.0 Wrenn et al. 

1971 

DDT, o,p' 

rn 

I 

rn 

n 

n 

rn 

5 

n

4 

(I, 

14 Rat 

(Wistar) 

me-conceDtion 

Gd 1-21 

Ld 1-21 

16.8 84 (17% less weight than 

controls at age 21 days) 

Clement and 

Okey 1974 

DDT, o,p' 

Figure 3-2. Levels of Significant Exposure to o,p'-DDT, -DDE, -DDD - Oral 

Acute (114 days) 

Svstemic 

1000 

100 

10 

.2d @Zd 

06r 07, @8r w 

1 

rn 

0 

-I 

(I) 

0.1 

0.01 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals Cancer Effect Level-Humans LD50/LC50

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans : Minimal Risk Level 

r-Rat m-Mouse e-Gerbil @ LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans : foreffects 

p-Pig h-Rabbit s-Hamster 0NOAEL -Animals A NOAEL- Humans A other than 

q-Cow a-Sheep g-Guinea Pig Cancer 
Figure 3-2. Levels of Significant Exposure to o,p’-DDT, -DDE, -DDD - Oral (Continued) 

Intermediate (15-364 days) 

1000 

1oc u 

OIld 

@14r I 

rn

5

I 

rn 

n 

n 

rn

2 

v, 

014r 

I( 

0 13r 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals Cancer Effect Level-Humans LD50/LC50

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals A LOAEL, More Serious-Humans : Minimal Risk Level 

r-Rat m-Mouse e-Gerbil @ LOAEL, Less Serious-Animals A LOAEL, Less Serious-Humans ; for effects 

p-Pig h-Rabbit s-Hamster 0NOAEL -Animals NOAEL - Humans other than 

q-Cow a-Sheep g-Guinea Pig Cancer 
Figure 3-2. Levels of Significant Exposure to o,p’-DDT, -DDE, -DDD - Oral (Continued) 

Chronic (2365days) 

rnglkglday 

1 

015r 

0.1 

c-Cat -Humans f-Ferret n-Mink Cancer Effect Level-Animals 

d-Dog k-Monkey j-Pigeon o-Other LOAEL, More Serious-Animals 

r-Rat m-Mouse e-Gerbil (3LOAEL, Less Serious-Animals 

p-Pig h-Rabbit s-Hamster 0NOAEL -Animals 

q-Cow a-S heep g-Guinea Pig 

Cancer Effect Level-Humans 

A LOAEL, More Serious-Humans 

A LOAEL, Less Serious-Humans 

NOAEL- Humans 

0 

0 

rn 

0 

0 

0 

u 

rn

F

4

I 

rn 

0 

-I 

cn 

LD50/LC50

; Minimal Risk Level 

: for effects 

other than 

Cancer 

2 
DDT, DDE, and DDD 72 

3. HEALTH EFFECTS 

exposure to humans below which the risk of adverse effects (other than cancer) is presumed to be minimal 

are also presented. 

Occupational exposure to DDT has involved multiple routes of exposure, the most significant of which 

involved ingestion of DDT particles via the mucociliary apparatus of the respiratory tract. Therefore, 

epidemiological studies of occupational exposure to DDT are discussed in this section. 

3.2.2.1 Death 

Only one case of fatal poisoning in humans after accidental oral exposure to DDT has been documented 

(Hill and Robinson 1945). One ounce of 5% DDT in kerosene was ingested by a 1-year-old child. 

Clinical signs included coughing and vomiting followed by tremors and convulsions. The child then 

became comatose and died; however, the contribution of the kerosene solvent to DDT toxicity was not 

clear. Doses as high as 285 mg DDT/kg body weight have been accidentally ingested by humans with no 

fatal results (Garrett 1947). 

In a follow-up study to Morgan and Lin (1978) (see Section 3.2.2.2), Morgan et al. (1980) analyzed 

morbidity and mortality in an extensive cohort of male and female workers exposed (formulators, 

applicators, farmers) to organochlorine pesticides. The follow-up included 73% of an original cohort of 

2,600 workers. The median duration of work in a pesticide-related occupation was 13 years. Disease 

incidence rates were studied in relation to job exposure categories and to serum levels of organochlorine 

pesticides measured in the original study. There were no significant differences in mortality patterns 

between pesticide-exposed workers and controls except for an excess of deaths by accidental trauma in 

workers engaged in structural pesticide application. 

A historical prospective mortality study was conducted on 3,600 white male workers employed between 

1935 and 1976 in occupations that involved exposures to various brominated compounds, organic and 

inorganic bromides, and DDT (Wong et al. 1984). Among individuals exposed to DDT, overall 

mortality, expressed as the standardized mortality ratio (SMR), was not elevated over expected values. 

However, there was an excess in deaths from respiratory cancer. Several factors confound these results: 

those individuals exposed to DDT also were potentially exposed to other chemicals, and smoking history 

was not included in the analysis. Brown (1992) conducted an update of an historical prospective 

mortality study of workers in five pesticide manufacturing plants. In the plant where primary exposure 
DDT, DDE, and DDD 73 

3. HEALTH EFFECTS 

was DDT (320 persons and 90 deaths since 1964), there was a significant excess of deaths (11) from 

stroke. The SMR was 2.38. The study is limited by insufficient exposure data, possible confounding 

exposures, and by relatively small numbers of deaths from stroke. 

The LD50 values reported in rats exposed to single oral doses of DDT ranged from 113 to 800 mg/kg 

(Ben-Dyke et al. 1970; Cameron and Burgess 1945; Gaines 1969). Results from a study by Lu et al. 

(1965) revealed age-dependent LD50 values in rats. The LD50 values in newborn, preweanling, weanling, 

and adult rats were >4,000, 438, 355, and 195 mg/kg, respectively. However, when preweanling and 

adult rats were administered one-quarter of the LD50 daily for 4 days, there was no significant difference 

in the 4-day LD50 between the two age groups. Lu et al. (1965) suggested that the elimination mechanism 

in the preweanling rats is less well developed, thus making them more susceptible to repeated doses than 

adults. The age-dependent susceptibility to single high oral doses of DDT in rats was confirmed by 

others who suggested that seizures and hyperthermia, observed in the adults but not in young rats, as well 

as less resistance to hypoxia, contribute to the apparent higher sensitivity of the adult rat (Henderson and 

Wooley 1969b, 1970). 

In male and female mice, a single oral dose of >237 mg technical-grade DDT/kg caused death (Kashyap 

et al. 1977; Tomatis et al. 1972), and a daily dietary dose of about 85.7 mg p,p’-DDT/kg killed 50% of a 

group of mice after a 6-day feeding period (Okey and Page 1974). Exposure of pregnant mice to 34.3 mg 

technical DDT/kg on gestation days 1–21 followed by cross-fostering of the pups resulted in preweaning 

death in 39% of the neonates exposed in utero and through lactation and 10% of the pups exposed only 

through lactation (Craig and Ogilvie 1974). No deaths occurred in pups exposed in utero only. The LD50 

values for guinea pigs and rabbits after oral exposure to DDT were 400 and 300 mg/kg, respectively 

(Cameron and Burgess 1945). Four out of five female B6C3F1 mice fed a diet that provided 

approximately 35 mg technical DDT/kg/day for 6 weeks died; the cause of death was not discussed (NCI 

1978). Four out of six monkeys treated by gavage with 50 mg p,p’-DDT/kg/day died after 4 weeks of 

treatment; two additional monkeys died during the weeks nine and fourteen of treatment, the cause of 

death was not specified (Cranmer et al. 1972a). In a 130-month study that administered approximately 

6.4–15.5 mg of p,p’-DDT/kg/day to 11 Rhesus and 14 Cynomolgus monkeys, there were 6 early deaths; 

the lowest dose associated with death was approximately 6.9 mg/kg/day (Takayama et al. 1999). A 10% 

mortality rate (relative to 0% in controls) was observed in the 78-week NCI (1978) bioassay for female 

B6C3F1 mice treated with approximately 15 mg technical DDT/kg/day; in the high-dose group 

(30.2 mg/kg/day), the mortality rate was 28%. There was no positive association between dose of DDT 

and mortality in male mice (NCI 1978). 
 

 

DDT, DDE, and DDD 74 

3. HEALTH EFFECTS 

DDE mortality studies also have been conducted. An LD50 of 880 mg/kg was reported for male Sherman 

rats (Gaines 1969). Death occurred in mice after single oral doses of o,p’-DDE ranging from 810 to 

880 mg DDE/kg (Tomatis et al. 1974a). 

An LD50 was reported for rats as a range of single oral doses (400–4,000 mg/kg) in which mortality was 

observed in 50% of rats exposed to DDD (Ben-Dyke et al. 1970). An LD50 for DDD in rats of 

>4,000 mg/kg was reported by Gaines (1969). Tomatis et al. (1974a) reported an LD50 in mice after a 

single oral dose ranging from 1,466 to 1,507 mg DDD/kg. 

The LD50 values for the various isomers and technical-grades of DDT, DDE, and DDD are recorded in 

Tables 3-1 and 3-2 and plotted in Figures 3-1 and 3-2. 

3.2.2.2 Systemic Effects 

The highest NOAEL values and all LOAEL values from each reliable study for each systemic effect in 

each species and duration category are recorded in Tables 3-1 and 3-2 and plotted in Figures 3-1 and 3-2. 

Respiratory Effects. No studies were located regarding respiratory effects in humans after oral 

exposure to DDT, DDE, or DDD. Rats fed a diet containing 20 mg DDT/kg/day for 27 months did not 

develop adverse respiratory effects with the exception of squamous bronchial metaplasia in one rat 

(Deichmann et al. 1967). In the 78-week chronic bioassay conducted by NCI (1978), no adverse effects 

on the respiratory system were observed in Osborne-Mendel rats treated in the diet with up to 45 mg 

technical DDT/kg/day, 59 mg p,p’-DDE/kg/day, or 231 mg technical DDD/kg/day. The same findings 

were reported for B6C3F1 mice treated with up to 30.2 mg technical DDT/kg/day, 49 mg 

p,p’-DDE/kg/day, or 142 mg technical DDD/kg/day (NCI 1978). 

Cardiovascular Effects. Cardiovascular performance was one of the parameters evaluated in male 

volunteers orally administered 3.5 or 35 mg DDT/day by capsule for 12–18 months either as 

recrystallized or technical-grade DDT (Hayes et al. 1956). This dosing regimen resulted in administered 

doses of 0.05–0.063 or 0.36–0.5 mg/kg/day for 12–18 months. The background concentrations measured 

in the food of both controls and test subjects were 0.0021–0.0038 mg DDT/kg/day. Although some 

variations among individuals in heart rate (resting and with exercise), systolic blood pressure, and pulse 

pressure were noted, these variations did not correlate with increasing dosage of DDT or with duration of 

exposure. The authors concluded that DDT at these doses did not result in adverse cardiac effects. 
 

 

DDT, DDE, and DDD 75 

3. HEALTH EFFECTS 

Tachycardia (increased heart rate) was reported in 1 of 11 cases in humans after accidental ingestion of 

approximately 286–1,716 mg DDT in food, which when adjusted for individual body weights, was 

5.1–120.5 mg/kg (Hsieh 1954). It is not known if the tachycardia was a direct result of damage to the 

heart or a result of a neurologically-mediated mechanism. A suggestive relationship between high serum 

levels of DDT and DDE and subsequent development of vascular disease, especially hypertension, was 

noted in pesticide-exposed individuals (Morgan et al. 1980). However, exposure to multiple pesticides is 

a significant confounding factor in this study. 

In the 78-week chronic bioassay conducted by NCI (1978), no significant chemical-related adverse effects 

on the cardiovascular system were observed in Osborne-Mendel rats treated in the diet with up to 45 mg 

technical DDT/kg/day, 59 mg p,p’-DDE/kg/day, or 231 mg technical DDD/kg/day. The same findings 

were reported for B6C3F1 mice treated with up to 30.2 mg technical DDT/kg/day, 49 mg 

p,p’-DDE/kg/day, or 142 mg technical DDD/kg/day (NCI 1978). 

Gastrointestinal Effects. No studies were located regarding the gastrointestinal effects in humans 

after oral exposure to DDT, DDE, or DDD. In the 78-week chronic bioassay conducted by NCI (1978), 

no significant chemical-related adverse effects on the gastrointestinal system were observed in OsborneMendel rats treated in the diet with up to 45 mg technical DDT/kg/day, 59 mg p,p’-DDE/kg/day, or 

231 mg technical DDD/kg/day. The same findings were reported for B6C3F1 mice treated with up to 

30.2 mg technical DDT/kg/day, 49 mg p,p’-DDE/kg/day, or 142 mg technical DDD/kg/day (NCI 1978). 

Hematological Effects. Hematological effects of DDT (Hayes et al. 1956; Laws et al. 1967; Morgan 

and Lin 1978; Ortelee 1958) and DDE (Dunstan et al. 1996; Morgan and Lin 1978) have been assessed, 

but the statistical power in some of these studies is practically nil because of the small number of subjects. 

An extensive study measuring organochlorine pesticide concentrations in the blood of 2,600 pesticideexposed individuals was conducted (Morgan and Lin 1978). One thousand controls with minimal 

exposure to pesticides were recruited and monitored. However, controls were not matched to pesticideexposed individuals for age, sex, or race. Various clinical tests, including extensive hematologic analyses 

were performed during 1967–1973, and p,p’-DDT and p,p’-DDE levels in blood were determined (not 

adjusted for lipid content). None of the many hematological parameters measured had a correlation 

coefficient with DDT and/or DDE blood levels >0.17 (meaning that at best, 2.89% of the variance could 

be attributed to the pesticide levels) although a number of coefficients showed statistical significance. 

Morgan and Lin (1978) acknowledged that because the analysis was based on such large numbers of 

measurements, some statistically significant differences were likely to be found. 
DDT, DDE, and DDD 76 

3. HEALTH EFFECTS 

Clinical and laboratory examinations were performed on 40 workers exposed to DDT, some of whom had 

also been exposed to other pesticides (Ortelee 1958). Exposure was reported to have occurred primarily 

via dermal and inhalation routes, but absorption by these routes may not have been significant. No 

protective equipment was used and the workers were often coated with concentrated DDT dust. 

Examinations included a complete medical history, physical and neurological examination, and 

hematological and blood chemistry analyses. In addition, plasma and erythrocyte cholinesterase levels 

were determined as well as urinary excretion of bis(p-chlorophenyl) acetic acid (DDA), which is the 

predominant final metabolite of DDT. On the basis of DDA excretion, it was estimated that these 

workers received absorbed doses equivalent to oral doses of 14–42 mg/person/day (approximately 

0.2–0.6 mg/kg/day). Despite this relatively high level of exposure, no correlation was found between 

DDT exposure and the frequency and distribution of abnormalities, including neurological effects, except 

for a few cases of minor skin and eye irritation. 

A study was conducted on 35 workers who had been involved in the manufacture and formulation of 

DDT for an average of 15 years (Laws et al. 1967). Extensive medical examinations and blood, urine, 

and fat analyses were performed in an attempt to find any correlation between any health problems and 

exposure to DDT. The authors concluded that the clinical findings for this group were not significantly 

different from those expected in an appropriate control group with no occupational exposure to DDT. 

However, this study did not include a control group and comparisons of data were made by using 

information derived from "the general population". 

Fifty-one male volunteers were exposed to 3.5 or 35 mg DDT/day, resulting in administered doses of 

0.05–0.063 or 0.36–0.5 mg DDT/kg/day, respectively, for 12–18 months (Hayes et al. 1956). The 

background concentration measured in food of both controls and test subjects was 0.0021–0.0038 mg 

DDT/kg/day. Although some variation among individuals in hemoglobin levels, red and white blood cell 

count, and percentage of polymorphonuclear leukocytes was noted, these variations did not correlate with 

increased dosage of DDT or with duration of exposure. 

In a case-control study of patients with chronic, debilitating fatigue lasting at least 6 months, the mean 

concentration of p,p’-DDE in blood serum was significantly higher in case subjects (11.9 ppb; n=14) than 

in controls (5.2 ppb; n=23) (Dunstan et al. 1996). When the 37 subjects were pooled and then redivided 

according to high serum DDE (>6 ppb) and low serum DDE (<6 ppb), the red blood cell distribution 

width (variation in erythrocyte cell width change; often a sign of anemia) was significantly greater in the 
 

DDT, DDE, and DDD 77 

3. HEALTH EFFECTS 

high DDE group than in the low DDE group; no differences were seen in other hematological parameters. 

In a chronic study, exposure of rats to DDT at 20 mg/kg/day for 27 months resulted in alterations in the 

spleen, which consisted of congestion and hemolysis exceeding that observed in untreated control rats 

(Deichmann et al. 1967). In addition, squirrel monkeys exposed orally to doses of 0.05–50 mg 

DDT/kg/day for up to 6 months exhibited no hematological changes; however, all monkeys in the highest 

dose group (six animals) died by week 14 (Cranmer et al. 1972a); the cause of death was not determined, 

but before death, the monkeys had apparently recovered from severe neurotoxic symptoms. 

Overall, the existing information does not suggest that hematological parameters are sensitive targets for 

DDT toxicity. The extent to which this is due to limitations in the available studies is uncertain. 

Musculoskeletal Effects. The relationship between serum levels of DDE and bone mineral density 

was examined in a group of 103 (50 black, 53 white) women who had participated in the Mount Sinai 

Medical Center Longitudinal Normative Bone Density Study (1984–1987) (Bohannon et al. 2000). 

Because bone metabolism is both estrogen- and androgen-dependent, and DDT and related compounds 

have shown hormone-like properties, the authors hypothesized that high levels of DDE (the main stored 

metabolite of DDT) would be associated with lower bone density in peri- and postmenopausal women 

than in premenopausal women. Bone mineral density was measured at the lumbar spine and radius at 

6 months intervals during a 2-year period. The results showed that black women had serum DDE levels 

significantly higher than white women (mean 13.9 vs. 8.4 ppb), but there was no correlation between 

either DDE and bone density or between DDE and the rate of bone loss. In a study of 68 sedentary 

Australian women (45–64 years of age), the authors state that an association was found between bone 

mineral density (lumbar spine) and serum DDE levels >2 ppb (Beard et al. 2000). However, the 

correlation coefficient was 0.269, meaning that only about 7.24% of the variance could be attributed to 

DDE; a stronger correlation was found for the relationship between bone mineral density and age. 

An additional study examined associations between bone mineral density and serum levels of 

organochlorines (five DDT-related compounds among them) in 115 men (40–75 years of age) from the 

general Swedish population (Glynn et al. 2000). The most concentrated DDT-related compound was 

p,p’-DDE, with a mean value of 784 ppb (lipid basis). After adjusting for confounding factors in linear 

regression analyses, there was no significant association between serum concentrations of single 

organochlorines and bone mineral density. 
 

DDT, DDE, and DDD 78 

3. HEALTH EFFECTS 

Limited information exists from studies in animals. In the 78-week chronic bioassay conducted by NCI 

(1978), no significant chemical-related adverse musculoskeletal effects were observed in Osborne-Mendel 

rats treated in the diet with up to 45 mg technical DDT/kg/day, 59 mg p,p’-DDE/kg/day, or 231 mg 

technical DDD/kg/day. The same findings were reported for B6C3F1 mice treated with up to 30.2 mg 

technical DDT/kg/day, 49 mg p,p’-DDE/kg/day, or 142 mg technical DDD/kg/day (NCI 1978). 

The existing information suggests that neither skeletal muscle nor bone are targets for DDT/DDE toxicity. 

Hepatic Effects. Based on the data available, there is no conclusive evidence that DDT, DDD, and 

DDE cause adverse liver effects in humans, but in animals, the liver appears to be a sensitive target for 

DDT. Morgan and Lin (1978) measured organochlorine pesticide concentrations in the blood of 

2,600 pesticide-exposed workers. Controls were not matched for age, sex, or race to pesticide-exposed 

individuals; data on alcohol consumption were not available. Weak correlations were reported between 

increasing serum DDT and DDE and elevated serum levels of alkaline phosphatase and lactic 

dehydrogenase. No significant correlation was found for serum glutamic oxaloacetic transaminase 

(SGOT; aspartate aminotransferase [AST]) and serum glutamic pyruvic transaminase (SGPT; alanine 

aminotransferase [ALT]). Morgan and Lin (1978) pointed out that the effects on serum enzyme activities 

were of such small magnitude that they were only detectable by exhaustive statistical analysis of data 

from thousands of subjects. In a follow-up study of Laws et al. (1967), Laws et al. (1973) reexamined 

31 of the 35 workers who had been involved in the manufacture and formulation of DDT for an average 

of 15 years. These men completed detailed questionnaires concerning their daily contact with DDT. In 

addition, liver function tests and DDT serum sample tests were performed on 21 subjects. Despite the 

fact that these workers had an average exposure to DDT for 21 years at levels estimated to correspond to 

3.6–18 mg daily (0.05–0.26 mg/kg/day), no evidence of hepatotoxicity, hepatic enlargement, or 

dysfunction (as measured by the Bromsulfthalein test, also known as sulfobromophthelein sodium) was 

found. Only one subject had an elevated Bromsulfthalein retention and slightly elevated serum levels of 

alkaline phosphatase and AST. Hayes et al. (1956) exposed 51 male volunteers to 3.5 or 35 mg 

DDT/day, resulting in administered doses of 0.05–0.063 or 0.36–0.5 mg DDT/kg/day for 12–18 months. 

The background concentration measured in food of both controls and test subjects was 0.0021–0.0038 mg 

DDT/kg/day. No signs of illness or adverse hepatic effects (as measured by liver function tests) were 

reported that were considered to be related to DDT exposure to humans. 

Epidemiological studies of DDT-exposed workers often show increased activity of hepatic metabolic 

enzymes. Kolmodin et al. (1969) studied drug metabolism in 26 workers with occupational exposure to 
DDT, DDE, and DDD 79 

3. HEALTH EFFECTS 

several pesticides, primarily DDT, chlordane, and lindane. An appropriate control group of 33 subjects 

was included. Xenobiotic metabolism was assessed by administering antipyrine in a single dose of 

10–15 mg/kg after which blood levels were monitored. Antipyrine was chosen because it is metabolized 

in the liver and not appreciably bound to plasma proteins. The half-life of antipyrine (both mean and 

median) was shorter in workers exposed to insecticides than in control subjects. In addition, the range of 

the half-life was much smaller in the exposed group. These results indicate that exposure to these 

insecticides induces hepatic enzyme activities, but the relative contribution of DDT to the observed 

enzyme induction is impossible to ascertain in this study. 

The effect of high DDT exposure on phenylbutazone and endogenous cortisol metabolism was 

investigated in exposed workers by Poland et al. (1970). Both of these compounds are metabolized via 

hepatic microsomal enzymes. Nineteen workers with an average of 14 years of employment in a plant 

producing only DDT were selected for study, and matched controls were obtained. Both groups were 

screened for medical history and given a physical examination, and venous blood samples were drawn. 

After a single dose of 400 mg phenylbutazone, serum was obtained at regular intervals beginning at 

24 and ending at 120 hours after drug administration. The serum half-life of phenylbutazone was 

significantly shorter in DDT-exposed workers than in the control population. However, the serum halflife of phenylbutazone did not correlate with the total concentration of DDT and related compounds in the 

serum of either group. The urinary excretion of the cortisol metabolite, 6β-hydroxycortisol, was 

increased by 57% in DDT-exposed individuals when compared to controls. However, there was no 

significant correlation found between the absolute serum concentration of total DDT-related compounds 

and the urinary excretion of 6β-hydroxycortisol. These results suggest that exposure to DDT can alter 

normal hepatic metabolic enzyme activity. 

In animals, the liver appears to be one of the primary targets of DDT toxicity. Acute, subchronic, and 

chronic oral administration of DDT has been shown to cause dose-related mild-to-severe hepatic effects 

in numerous animal studies. In addition, DDT has been demonstrated to be an inducer of microsomal 

mixed function oxidases of the liver by its ability to promote the biotransformation of various chemicals 

(Pasha 1981; Street and Chadwick 1967). 

Acute oral exposure to DDT is associated with a number of effects in animals including increased liver 

weights, increased serum levels of liver enzymes (suggestive of liver injury), and changes in the 

appearance of the liver. Both liver plasma membrane and serum gamma-glutamyl transpeptidase (GGTP) 

were increased 2-fold in rats treated with a single dose of 200 mg DDT/kg (Garcia and Mourelle 1984). 
DDT, DDE, and DDD 80 

3. HEALTH EFFECTS 

Plasma membrane enzyme activity returned to normal 48 hours after dosing, but enzyme activity in serum 

remained elevated for 48 hours. Relative liver weight increased 18% after 12 days exposure to 

40 mg/kg/day DDT in rats previously deprived of vitamin A for several generations; DDT treatment 

began 1 week after vitamin A supplementation was started to increase the rats’ vitamin A intake back to 

“normal” (de Waziers and Azais 1987). In both “normal” and vitamin A deficient rats, 12 days of DDT 

treatment increased liver CYP (P450) enzyme content per gram of liver by 2-fold, benzo(a)pyrene 

hydroxylase (a CYP enzyme) activity by 3-fold, and glutathione-S-transferase activity by 2-fold; all of 

these changes were statistically significant. A recent study showed that hepatic O-dealkylation activities 

associated with CYP2B were induced in a dose-related manner in rats treated orally with DDT for 

14 days (Nims et al. 1998). The dose range tested was approximately 0.17–36 mg/kg/day. Maximal 

induction was more than 21 times greater than control values. There was no effect on in vivo CYP2B 

induction at <0.17 mg/kg/day. Limited induction of ethoxyresorufin O-dealkylation was observed, and 

no induction of immunoreactive hepatic microsomal CYP1A protein was seen. DDT also induced 

CYP3A1 and CYP3A2 proteins. Induction activity was also accompanied by an increase in relative liver 

weight with doses of approximately 4.6 mg/kg/day and higher. Mice administered 42.9 mg DDT/kg/day 

for 1 week by oral gavage had increased liver weights and increased microsomal enzyme activities (Pasha 

1981) and Rhesus monkeys exposed to one oral dose of 150 mg DDT/kg had increased alkaline 

phosphatase (AP), LDH, AST, and ALT activities in serum (Agarwal et al. 1978). An increase in the 

levels of liver metabolic enzymes by itself is not considered adverse; however, continued microsomal 

enzyme induction may lead to hypertrophy and contribute to morphological changes in the liver. It 

should be noted also that microsomal enzyme induction may lead to the activation of some chemicals to 

toxic metabolites and the detoxification of other chemicals. In addition, increased serum levels of liver 

enzymes such as transaminases, LDH and AP can be predictive of more serious liver effects on prolonged 

exposure since this may indicate cell death and consequent lysis of membranes and leakage of enzymes. 

Hepatic cell hypertrophy, histopathologic alterations (proliferation of the smooth endoplasmic reticulum 

and concentric membrane arrays), and increased microsomal enzyme activity have also been seen 

following intermediate exposure of rats to DDT. These effects were observed in rats following 

3–27 weeks of dietary exposure at doses ranging from 0.25 to 20 mg DDT/kg/day (Gupta et al.1989; 

Laug et al. 1950; Ortega 1956). In general, as doses increased, hepatic effects in rats became more 

severe. Phase I metabolic enzymes were significantly increased in rats treated for 3 weeks with 15 mg 

p,p’-DDT/kg/day (Gupta et al. 1989). Minor hepatocyte vacuolation was seen in male rats receiving 

2.5 mg technical DDT/kg/day in the diet for 3 months; females exhibited liver hypertrophy at 

10 mg/kg/day (Ortega 1956). Rats administered approximately $(0.25–0.5) mg technical DDT/kg/day 
DDT, DDE, and DDD 81 

3. HEALTH EFFECTS 

(see Appendix A for dose calculation) in the diet for 12–27 weeks had liver changes consisting of hepatic 

cell enlargement, especially in the cental lobules, increased cytoplasmic oxyphilia with sometimes a 

semihyaline appearance, and more peripheral locations of the basophilic cytoplasmic granules (Laug et al. 

1950). The severity of the effects was dose-related; a dose of approximately 0.05–0.09 mg/kg/day was a 

NOAEL. This study (Laug et al. 1950) was used as the basis for the derivation of an intermediateduration oral MRL for DDT. Fitzhugh and Nelson (1947) fed female Osborne-Mendel rats a diet 

containing 1,000 ppm (96 mg/kg/day) technical-grade DDT for 12 weeks and observed reversible 

changes in the liver, including enlargement of centrolobular hepatocytes with accompanying histological 

changes. Rats exposed continuously to 6.6 mg DDT/kg/day in the diet for 36 weeks exhibited spotty 

cellular necrosis with moderate hepatocyte regeneration (Jonsson et al. 1981). Squirrel monkeys and 

mice had increased hepatic cytochrome P-450 enzyme activities and/or increased liver weights following 

short-term exposure by oral gavage to DDT. These effects were observed following doses of 1.67 or 

6.25 mg DDT/kg/day for 28 days in mice (Lundberg 1974; Orberg and Lundberg 1974) and 5 mg 

DDT/kg/day for up to 6 months in squirrel monkeys (Cranmer et al. 1972a). 

Hepatic effects ranging from increased liver weights to cellular necrosis have been reported in animals 

after chronic exposure to DDT in the diet. Necrosis, centrilobular hypertrophy, and hyperplasia have also 

been reported in rats exposed to 7–56 mg DDT/kg/day for 24–27 months (Deichmann et al. 1967; 

Fitzhugh and Nelson 1947). Increased incidence of fatty metamorphosis was seen in the liver of male rats 

treated with approximately 23 mg technical DDT/kg/day or more and of amyloidosis in the liver of male 

mice treated with about 3.7 mg DDT/kg/day for 78 weeks (NCI 1978). Increased relative liver weights, 

but no increase in serum ALT, LDH, AP, cholinesterase, or in liver ALT, AST, LDH, AP, or 

cholinesterase activities were reported in hamsters after exposure to 67–133 mg DDT/kg/day for life 

(Graillot et al. 1975). Cabral et al. (1982a) reported a significant increase in liver necrosis in hamsters 

exposed to approximately 71 mg DDT/kg/day in the diet for life, but not at lower doses. Both focal and 

diffuse liver alterations were observed in dogs exposed by diet to 80 mg DDT/kg/day for 39–40 months, 

but not at 16 mg/kg/day (Lehman 1965). In Rhesus monkeys given up to 3.9 mg DDT/kg/day for 

3.5–7.5 years, periodic liver biopsies showed no “significant” observable alterations in liver histology and 

the Bromsulfthalein clearance test was normal, indicating no functional liver changes (Durham et al. 

1963). However, in a 130-month dietary study in Rhesus and Cynomolgus monkeys, a dose of 

approximately 6.4 mg of p,p’-DDT/kg/day, the lowest dose tested, produced significant liver alterations 

including hepatocyte vacuolation, proliferation of bile duct cells, clear hepatocyte foci, and liver cell 

necrosis (Takayama et al. 1999). 
 

DDT, DDE, and DDD 82 

3. HEALTH EFFECTS 

Limited information exists on hepatic effects after oral exposure of animals to DDD or DDE. Rats 

exposed to two gavage doses of 350 mg DDE/kg/day exhibited an increase in enzyme levels (ornithine 

decarboxylase and cytochrome P-450) (Kitchin and Brown 1988), and mice showed increased liver 

weight, microsomal P450, cytochrome-C reductase, and serum total protein after daily gavage dosing 

with 42.9 mg/kg DDE/day for 7 days (Pasha 1981). A dose of 100 mg of p,p’-DDE/kg/day for 7 days 

caused elevated hepatic aromatase (enzyme involved in steroid metabolism) activity in adult rats, a 

potential mechanism by which DDE can affect reproduction (You et al. 2001). Chronic exposure of rats 

(78 weeks) to p,p,’-DDE resulted in increased incidence of fatty metamorphosis in the liver from males at 

a dose of approximately 31 mg/kg/day (NCI 1978). DDE produced significant induction of hepatic 

O-dealkylation activities associated with CYP2B in rats administered DDE in the diet at dose levels 

between approximately 0.15 and 36 mg/kg/day for 14 days (Nims et al. 1998). The NOEL for in vivo 

CYP2B induction was 0.17 mg/kg/day. No induction of immunoreactive hepatic microsomal CYP1A 

protein was observed, and induction of CYP1A associated activities was limited. DDE also induced 

CYP3A1 and CYP3A2 proteins. Induction activity was also accompanied by a significant increase in 

relative liver weight at a dose level of approximately 4.1 mg/kg/day and higher. No adverse liver effects 

were observed in mice chronically exposed to p,p’-DDE in the diet at up to 49 mg/kg/day for 78 weeks 

(NCI 1978). Chronic exposure to approximately 48 mg p,p’-DDE/kg/day resulted in focal necrosis of the 

liver in hamsters in a 128-week study (Rossi et al. 1983). 

DDD induced CYP2B associated activities in the livers of rats treated with the test material in the diet at 

dose levels of approximately 1.4 mg/kg/day and higher for 14 days (Nims et al. 1998). The NOEL was 

between 0.5 and 1.4 mg/kg/day. No induction of immunoreactive hepatic microsomal CYP1A protein 

was observed, and limited induction of CYP1A associated enzyme activity was seen. DDD also slightly 

induced CYP3A2, but not CYP3A1. No significant treatment-related nonneoplastic alterations were seen 

in the livers from rats or mice exposed orally to up to 231 or 142 mg technical DDD/kg/day, respectively, 

for 78 weeks (NCI 1978). Mice exposed orally to 42.9 mg DDD/kg/day in the diet for 1 week had no 

change in levels of P-450 microsomal enzymes, but showed a decrease in cytosolic-enzyme-mediated 

hydroxylation of 2- and 4-biphenyls with no change in liver weight (Pasha 1981). 

Renal Effects. No studies were located regarding renal effects in humans after oral exposure to DDT, 

DDE, or DDD. Male and female Osborne-Mendel rats exposed orally to 20 mg DDT/kg/day in the diet 

for 27 months had tubular polycystic degeneration, tubular epithelial necrosis, and hemorrhage of the 

kidney (Deichmann et al. 1967). In the 78-week chronic bioassay conducted by NCI (1978), no 

significant chemical-related adverse renal effects were observed in Osborne-Mendel rats treated in the 
 

DDT, DDE, and DDD 83 

3. HEALTH EFFECTS 

diet with up to 45 mg technical DDT/kg/day, or in mice treated with up to 30.2 mg DDT/kg/day. 

Similarly, no adverse renal effects were seen in rats treated with up to approximately 59 mg 

p,p’-DDE/kg/day for 78 weeks (NCI 1978). However, chronic inflammation of the kidneys was seen in 

male mice treated with approximately 27 mg p,p’-DDE/kg/day (NCI 1978); no such effect was seen in 

female mice treated with up to 49 mg/kg/day. Technical DDE, at a dose level of approximately 

66 mg/kg/day, increased the incidence of chronic inflammation of the kidney in female rats in the 

78-week NCI (1978) bioassay. No significant chemical-related increase in kidney lesions was seen in 

mice treated for 78 weeks treated with up to 231 mg technical DDD/kg/day (NCI 1978). 

Endocrine Effects. Exposure to DDT and DDT-related compounds, particularly during 

development, can adversely affect the development and function of the reproductive system of both 

female and male animals. This is due primarily to the ability of some of these compounds to disrupt the 

action of natural steroids and bind to receptors for estrogens and androgens. Hormonal effects of DDT 

and residues that lead to altered reproduction and/or development are discussed in Sections 3.2.2.5 

Reproductive Effects, 3.2.2.6 Developmental Effects, and 3.6.2 Mechanisms of Toxicity. Reported 

endocrine effects unrelated to reproduction or development are briefly mentioned below. 

No studies were located regarding endocrine effects in humans after oral exposure to DDT or DDE. 

However, o,p’-DDD has been used to treat adrenocortical carcinoma for almost four decades (Bergenstal 

et al. 1960; Wooten and King 1993). The therapeutic action is based on the activation of the compound 

by local cytochrome P-450 (CYP11β) to a reactive metabolite that binds to macromolecules in the adrenal 

cortex (see Section 3.6.2, Mechanisms of Toxicity, for further details). 

In rats administered 0.2 mg technical-grade DDT/kg/day for 120 days, atrophy of all zones of the adrenal 

gland, except the zona glomerulosa, was observed (Chowdhury et al. 1990). Hyalin degeneration of the 

medulla and cortex and a decrease in adrenal gland weight were also reported. Dogs given 1–15 oral 

doses of 100 mg/kg/day technical-grade DDD showed degenerative changes in mitochondria in both the 

zona fasciculata and zona reticularis of the adrenal cortex, which were reversible 56 weeks after dosing 

ceased (Powers et al. 1974). Similar degenerative adrenal changes were reported by Kirk et al. (1974) 

after dosing dogs (by capsule) with 138.5 mg/kg/day of o,p'-DDD for 10 days. Plasma levels of cortisol 

were decreased and a decreased response to ACTH stimulation was observed. In one animal, there was 

hemorrhage, invasion by lymphocytes, and necrosis of the adrenal cortex (Kirk et al. 1974). 

Adrenocortical necrosis was also reported in dogs treated for 36–150 days with 50 mg o,p’-DDD/kg/day 

in a capsule (Kirk and Jensen 1975). Studies in animals have also shown that the adrenal gland is affected 
 

 

DDT, DDE, and DDD 84 

3. HEALTH EFFECTS 

by sulfonyl metabolites of DDT. The metabolic pathways leading to these metabolites are discussed in 

Section 3.5.3, Metabolism, whereas information on their mechanism of toxic action can be found in 

Section 3.6.2, Mechanisms of Toxicity. 

Reduced iodine concentrating capacity in the thyroid was reported in adult male Sprague-Dawley rats 

given a single oral gavage dose of 50 mg or higher of technical DDT/kg; the-no-observed-effect dose 

level was between 25 and 50 mg/kg (Goldman 1981). 

In the 78-week chronic bioassay conducted by NCI (1978), no treatment-related adverse effects on the 

endocrine system (pituitary, adrenals, thyroid, parathyroid) were observed in Osborne-Mendel rats treated 

in the diet with up to 45 mg technical DDT/kg/day, 59 mg p,p’-DDE/kg/day, or 231 mg technical 

DDD/kg/day. The same findings were reported for B6C3F1 mice treated with up to 30.2 mg technical 

DDT/kg/day, 49 mg p,p’-DDE/kg/day, or 142 mg technical DDD/kg/day (NCI 1978). 

Dermal Effects. Clinical examinations with laboratory workups were performed on 40 workers 

exposed to DDT, some of whom had also been exposed to other pesticides (Ortelee 1958). Exposure was 

reported to occur primarily via dermal and inhalation routes; however, no protective equipment was used, 

and the workers were often coated with concentrated DDT dust. Information collected on each worker 

included a complete medical history, physical and neurological examination results, and hematological 

and blood chemistry analyses results. Plasma and erythrocyte cholinesterase levels were determined as 

well as urinary excretion of DDA. On the basis of DDA excretion, it was estimated that these workers 

received absorbed doses equivalent to oral doses of 14–42 mg/day (approximately 0.2–0.6 mg/kg/day). 

Despite the relatively high estimated exposure, no correlation was found between DDT exposure and the 

frequency and distribution of skin abnormalities, except for a few cases of minor skin irritation. 

No studies were located indicating adverse dermal effects in animals after oral exposure to DDT, DDE, or 

DDD. 

Ocular Effects. The only information available is that from an earlier report by Ortelee (1958) 

regarding 40 workers exposed to DDT, some of whom had also been exposed to other pesticides. Based 

on information on DDA excretion, it was estimated that these workers received absorbed doses equivalent 

to oral doses of 14–42 mg/day (approximately 0.2–0.6 mg/kg/day). No correlation was found between 

DDT exposure and the frequency and distribution of abnormalities, except for a few cases of minor eye 

irritation, which were probably due to direct contact of the pesticides with the eye. 
 

DDT, DDE, and DDD 85 

3. HEALTH EFFECTS 

Body Weight Effects. No treatment-related effects on body weight were observed in a group of 

51 male volunteers given daily doses of up to 0.5 mg technical DDT/kg for up to 18 months (Hayes et al. 

1956). 

Body weight gain was reduced by 30%, relative to controls, in male albino rats treated with 0.2 mg 

technical DDT/kg by gavage for 120 days; food consumption data were not provided (Chowdhury et al. 

1990). Female and male Osborne-Mendel rats fed a diet that provided approximately 97 or 50 mg 

technical DDT/kg/day, respectively, for 6 weeks had a 45 and 16% reduction in body weight, 

respectively, at the end of the treatment period relative to controls (NCI 1978). Male and female B6C3F1 

mice treated similarly with up to 35 mg technical DDT/kg/day showed no significant treatment-related 

alterations in body weight gain (NCI 1978). In a 78-week chronic study, female Osborne-Mendel rats 

showed a 20% decrease in body weight gain due to dietary administration of approximately 32 mg 

technical DDT/kg/day compared to controls; males exhibited a 16% decrease in weight gain with a 

dietary level of approximately 45 mg/kg/day (NCI 1978). B6C3F1 mice treated in the diet with up to 

approximately 30.2 mg technical DDT/kg/day for 78 weeks did not show treatment-related alterations in 

body weight (NCI 1978). Hamsters treated for life with up to 71 mg technical DDT/kg/day in the diet 

showed no significant treatment-related effects on body weight (Cabral et al. 1982a), but a dietary level of 

approximately 95 mg/kg/day of technical DDT for 128 weeks caused an unspecified reduction in body 

weight gain relative to controls (Rossi et al. 1983). 

A 6-week treatment period with up to approximately 157 mg p,p’-DDE/kg/day in the diet induced a 22% 

decrease in body weight gain in male Osborne-Mendel rats, and a dose level of 88 mg/kg/day caused an 

11% decrease in weight gain relative to controls (NCI 1978); a dose of 50 mg/kg/day was a NOAEL. No 

significant treatment-related effects on body weight were seen in female rats treated with up to 305 mg 

p,p’-DDE/kg/day or in B6C3F1 mice treated with up to 101 mg/kg/day (NCI 1978). Chronic 

administration of approximately 19 mg p,p’-DDE/kg/day in the diet to female Osborne-Mendel rats 

caused a 21% decrease in body weight gain; this was the lowest dose tested (NCI 1978). Female B6C3F1 

mice administered approximately 28 mg p,p’-DDE/kg/day for 78 weeks had a 29% reduction in body 

weight gain (this was also the lowest dose tested); body weight gain in male mice was unaffected by 

doses of up to 47 mg p,p’-DDE/kg/day (NCI 1978). Hamsters fed a diet that provided approximately 

47.5 mg p,p’-DDE/kg/day for 128 weeks showed an unspecified reduction in body weight gain compared 

to controls (Rossi et al. 1983). 
DDT, DDE, and DDD 86 

3. HEALTH EFFECTS 

Treatment of female Osborne-Mendel rats with approximately 97 mg technical DDD/kg/day for 6 weeks 

resulted in a 39% decrease in body weight gain relative to controls, but treatment with 172 mg/kg/day 

resulted in only a 4% reduction in weight gain (NCI 1978). In males, the highest dose tested, 

279 mg/kg/day, caused a 10% reduction in weight gain at the end of the treatment period (NCI 1978). In 

the 78-week NCI (1978) bioassay, female rats treated with approximately 66 mg technical DDD/kg/day 

had a 26% decrease in weight gain compared to controls at the end of the study; this was the lowest dose 

level tested. In males, there was a 28% decrease in weight gain in a group dosed with approximately 

116 mg/kg/day and a 39% decrease in a group receiving about 231 mg/kg/day of technical DDD. 

3.2.2.3 Immunological and Lymphoreticular Effects 

Limited information was located regarding immunological effects in humans after oral exposure to DDT, 

DDE, or DDD. In a study in which humans were challenged with an injection of Salmonella typhimurium 

vaccine, serum agglutinin titers were significantly higher in three volunteers given capsules containing 

5 mg DDT/day (0.07 mg/kg) for 20 days when compared to volunteers who received only the bacterial 

antigen; immunoglobulin levels were unaffected by treatment with DDT (Shiplov et al. 1972). The 

volunteers exhibited no apparent symptoms of DDT exposure. 

A different type of study assessed parameters of immunocompetence in a group of 23 men chronically 

exposed to DDT through consumption of fish from the Baltic Sea (Svensson et al. 1994). The levels of 

DDT in the fish were not provided. None of the subjects had symptoms of any infectious disease at the 

time of the study. Twenty men with almost no fish consumption served as controls. The parameters 

examined included white cell counts, lymphocyte levels, serum immunoglubulin levels, and lymphocyte 

subsets. Of all the parameters examined, only the level of natural killer (NK) cells was reduced in the fish 

eaters, but the difference between groups was not statistically significant. Weekly intake of fatty fish 

correlated significantly (r=0.32, p<0.04) with the reduction in NK cells. A correlation (r=0.72, p=0.02) 

was observed between NK cell level and plasma level of p,p’-DDT in 12 subjects. The toxicological 

significance of these findings is unknown. 

Further information on the potential immunological effects of DDT is available in a study of 

302 individuals residing near a waste site in North Carolina (Vine et al. 2001). Of 20 organochlorines 

tested, only DDE was detected in the blood of the participants; the median concentration in plasma was 

2 ppb and the highest concentration was 32 ppb. End points evaluated included white blood cell counts, 

lymphocyte phenotypes, immunoglobulins, and mitogen-induced lymphoproliferative activity; a skin test 
DDT, DDE, and DDD 87 

3. HEALTH EFFECTS 

was also conducted to evaluate delayed-type hypersensitivity. When DDE levels were divided into five 

categories from #1 ppb or $7.6 ppb, subjects with higher blood DDE levels had lowered mitogen-induced 

lymphoproliferative activity (concanavalin A) and slightly increased total lymphocytes and 

immunoglobulin A levels. Results from the skin test showed no differences in response by blood DDE 

levels. Vine et al. (2001) concluded that relatively low DDE levels were associated with changes in 

immune markers, although the magnitude of the effects are of uncertain clinical importance. 

Evidence of DDT-induced compromises in immune function has been obtained from studies conducted in 

animals. The effects of DDT on the humoral immune response were studied in mice (Banerjee 1987a; 

Banerjee et al. 1986, 1997b), in rats (Banerjee 1987b; Banerjee et al. 1995, 1996; Gabliks et al. 1975), 

and in rabbits (Shiplov et al. 1972). There was no evidence that DDT adversely affected the humoral 

response in rabbits; however, dosing was for only 10 days by oral gavage at a single dose level 

(4.3 mg/kg/day) (Shiplov et al. 1972). Following dietary dosing at levels to provide 13 mg DDT/kg/day 

for 3–12 weeks, mice showed an immunosuppression particularly of the secondary humoral immune 

response to immunization with sheep red blood cells (SRBC) or Escherichia coli lipopolysaccharide 

(LPS), which are T-cell dependent and T-cell-independent antigens, respectively. There was a significant 

reduction in splenic plaque-forming cells as well as decreased IgM titers to lipopolysaccharide compared 

to controls (Banerjee 1987a; Banerjee et al. 1986). Rats immunized with diphtheria toxoid were fed DDT 

in the diet for 31 days at an intake of 1.9 or 19 mg/kg/day. When challenged with diptheria toxoid, the 

severity of the anaphylactic response was decreased and the number of mast cells (producing histamine) 

in mesenteries was significantly decreased compared to controls; antibody titers to the toxoid were not 

decreased (Gabliks et al. 1975). Exposure of adult mice to DDT in the diet for 16 weeks at doses of 

0.006 mg/kg/day had no effect on the humoral immune response but doses of 0.06 or 0.63 mg/kg/day for 

16 weeks significantly stimulated the primary IgM response to SRBC and the lymphoproliferative 

response to LPS-coated SRBC (Rehana and Rao 1992). Dosing for longer periods (20 or 24 weeks) 

caused a sharp reduction in both responses. Dosing of dams during pregnancy and lactation caused 

suppression of the immune system in offspring as discussed in Section 3.2.2.6 (Rehana and Rao 1992); 

however, because appropriate controls were not used, the significance of this finding is not clear. 

Effects in lymphoid organs were noted in rabbits (Street and Sharma 1975) and rats (Deichmann et al. 

1967; Hamid et al. 1974). Rabbits administered an oral dose of 0.18 mg DDT/kg/day in the diet for 

8 weeks exhibited a significant increase in gamma globulin levels and atrophy of the thymus whereas a 

much higher dose, 6.54 mg/kg/day, was found to significantly decrease the skin sensitivity to tuberculin 

(cell mediated immunity) (Street and Sharma 1975). Rats exposed to doses of 121 mg o,p’-DDD/kg/day 
DDT, DDE, and DDD 88 

3. HEALTH EFFECTS 

for 16 days had decreased (no statistical analysis) numbers of plaque-forming cells and rosette-forming 

cells in the spleen and thymus compared to controls and displayed atrophy of the thymus and adrenal 

gland (Hamid et al. 1974). 

Dietary DDT increased the growth of the leprosy bacterium, Mycobacterium leprae, in mouse foot pads 

(Banerjee et al. 1997a). DDT was fed to male albino Rockfeller strain mice in the diet at 0, 20, 50, or 

100 ppm (approximately 0, 4.3, 10.7, and 21.4 mg/kg/day) for 24 weeks following inoculation of the 

footpads of mice with the leprosy bacterium. At $50 ppm, bacillary growth was significantly increased. 

Both humoral and cell-mediated responses were adversely affected in groups of 8–12 male Wistar rats fed 

diets containing 200 ppm DDT, DDE, or DDD (approximately 22.2 mg/kg/day) for 6 weeks (Banerjee et 

al. 1996). Compared to controls, effects on the humoral immune system were seen after dietary exposure 

to each of the compounds, including significantly increased serum albumin/globulin ratio, suppressed 

serum IgM, IgG after ovalbumin immunization, and decreased antibody titer. Likewise, cell-mediated 

effects were seen in rats fed DDT, DDE, and DDD, including increased inhibition of leucocyte and 

macrophage migration, and decreased footpad thickness. Mean relative liver weight was increased 

compared to controls in rats fed DDT and DDE, while relative spleen weight was decreased in DDD-fed 

rats; no other signs of toxicity were observed. 

The influence of dietary protein on the humoral and cell-mediated immunotoxicity of DDT was evaluated 

in groups of male Wistar rats fed diets containing 0, 20, 50, or 100 ppm DDT (approximately 0, 2.3, 5.7, 

and 11.4 mg/kg/day, respectively) for 4 weeks (Banerjee et al. 1995). Dose groups were divided into 

subgroups that received either 3, 12, or 20% protein in the diet; half of each dietary subgroup, in turn, 

received a tetanus toxoid immunostimulant, resulting in allocation of 10–12 rats to each treatment group. 

The serum albumin/globulin ratio was significantly increased in mid- and high-DDT dose groups 

compared to treatment controls, but only in rats fed the low-protein diet and receiving the 

immunostimulant injection. Serum IgM and IgG were significantly reduced in mid- and high-DDT 

groups that were fed the 3% protein diet, regardless of whether they had been immunostimulated with 

tetanus toxoid injections. No effects were seen in any group fed diets containing 12 or 20% protein. 

DDT enhanced stress-induced humoral immune response suppression in mice (Banerjee et al. 1997b). 

Groups of 80–90 albino male Hissar mice were fed 0, 20, 50, or 100 ppm p,p’-DDT in the diet 

(approximately 0, 4.1, 10.1, and 20.3 mg/kg/day, respectively). Equal subgroups were selected for SRBC 

or plaque-forming cell (PFC) assays. Assay subgroups were in turn divided into five equal groups, 
DDT, DDE, and DDD 89 

3. HEALTH EFFECTS 

resulting in treatment groups of 8–10 mice per group; one group received no stressor, and the other four 

groups received one of four different combinations of temperature (3 hours at 4 EC) and restraint stresses. 

Relative to controls, DDT alone caused no significant alterations in primary antibody titer to SRBC or in 

PFC response. DDT in combination with stress, and in a dose-related manner, significantly reduced both 

responses to a greater extent than did each stressor alone. The results of this study indicate the following: 

(1) certain types of stressors are sufficient to suppress the humoral immune response in mice, in the 

absence of known chemical stressors, and (2) dietary DDT at sufficient levels can enhance the 

immunosuppressant effect of other stressors that otherwise have no statistically discernable effect. 

Sheep red blood cell (SRBC) antibody titers were reduced in groups of 16 or 20 male Wistar rats fed 

100 or 200 ppm in the diet for 8 weeks (approximately 10.3 and 20.6 mg/kg/day, respectively); DDT 

dietary exposure groups were divided into two subgroups, one received ascorbic acid by gavage at 

100 mg/kg/day and the other received no ascorbic acid (Koner et al. 1998). Dose-related decreases in 

SRBC antibody titers were seen in the DDT-treated rats compared to controls, with or without concurrent 

treatment with ascorbic acid; the decrease was statistically significant in the 200-ppm group. In the DDTtreated rats, the absolute decrease appeared to be attenuated in rats co-treated with ascorbic acid compared 

to those not receiving ascorbic acid, although statistical tests were not reported. 

In the 78-week chronic bioassay conducted by NCI (1978), no treatment-related adverse effects on the 

thymus, spleen, or lymph nodes were observed in Osborne-Mendel rats treated in the diet with up to 

45 mg technical DDT/kg/day, 59 mg p,p’-DDE/kg/day, or 231 mg technical DDD/kg/day. The same 

negative findings were reported for B6C3F1 mice treated with up to 30.2 mg technical DDT/kg/day, 

49 mg p,p’-DDE/kg/day, or 142 mg technical DDD/kg/day (NCI 1978). Immunocompetence was not 

evaluated in this study. 

The overall evidence suggests that DDT and related compounds can affect immunocompetence in 

animals, but there is no conclusive evidence regarding effects in humans. 

The highest NOAEL values and all LOAEL values from each reliable study for immunological effects in 

each species and duration category are recorded in Tables 3-1 and 3-2 and plotted in Figures 3-1 and 3-2. 
DDT, DDE, and DDD 90 

3. HEALTH EFFECTS 

3.2.2.4 Neurological Effects 

The nervous system appears to be one of the primary target systems for DDT toxicity in humans after 

acute, high exposures. A number of investigators conducted experimental studies on humans in the 1940s 

and 1950s at controlled doses that produced effects. Other data come from accidental poisonings where 

dose levels were crudely estimated. Persons exposed to 6 mg DDT/kg administered orally by capsule 

generally exhibited no illness, but perspiration, headache, and nausea have been reported (Hayes 1982). 

Convulsions in humans have been reported at doses of 16 mg DDT/kg or higher (Hsieh 1954). Velbinger 

(1947a, 1947b) exposed volunteers to oral doses of 250, 500, 750, 1,000, or 1,500 mg DDT 

(approximately up to 22 mg/kg) suspended in an oil solution. Variable sensitivity of the mouth (defined 

by the author as a prickle at the tip of the tongue, lower lip, and chin area) was reported in volunteers 

exposed to 250 and 500 mg DDT/person. Six hours after exposure to 750 or 1,000 mg DDT, disturbance 

of sensitivity of the lower part of the face, uncertain gait, malaise, cold moist skin, and hypersensitivity to 

contact were observed. Prickling of the tongue and around the mouth and nose, disturbance of 

equilibrium, dizziness, confusion, tremors, malaise, headache, fatigue, and severe vomiting were all 

observed in volunteers within 10 hours after oral exposure to 1,500 mg DDT. All volunteers exposed to 

DDT orally had achieved almost complete recovery from these acute effects within 24 hours after 

exposure. Similar symptoms were reported in persons after accidental or intentional ingestion of DDT 

(Francone et al. 1952; Garrett 1947; Hsieh 1954; Mulhens 1946). 

Few studies have explicitly evaluated neurotoxicity in humans following chronic exposure. There was no 

correlation between DDT exposure and neurological effects in workers whose estimated exposure, based 

on DDA excretion data, was approximately 14–42 mg/person/day (Ortelee 1958). Occupational exposure 

involved all possible routes of exposure, but most of the intake is considered to be from the oral route. 

Inhaled dusts are poorly absorbed because of size, but they are cleared by the mucociliary mechanisms 

and a fair portion is then ingested. None of the subjects had any evidence of hyperexcitability, and the 

results of the neurological examinations were normal. No neurological effects related to DDT were noted 

in volunteers who ingested 3.5 or 35 mg DDT/day (0.05–0.05 or 0.36–0.5 mg/kg/day) for 12–18 months 

(Hayes et al. 1956). The subjects displayed no loss of coordination and there was no indication of 

tremors. Other tests (over 20) were negative and showed no peripheral neuropathy or central nervous 

system functional deficits. Background DDT levels in food of both controls and test subjects were 

0.0021–0.0038 mg DDT/kg/day. Somewhat different findings were described in a recent study of retired 

malaria-control workers in Costa Rica (van Wendel de Joode et al. 2001). Neurological tests assessing 

cognitive, motor, and sensory functions were conducted on 27 former workers and 27 matched controls. 
DDT, DDE, and DDD 91 

3. HEALTH EFFECTS 

The men had been involved in DDT use between 1955 and 1986; the mean number of years of DDT 

application was 4.6 years. The exposed group had overall poorer performance; verbal attention and 

visuomotor speed and sequencing differed most between the groups. When years of DDT application was 

entered in the analysis as the explanatory variable, significant exposure-effect relations were seen for five 

tests in cognitive, motor, and sensory domains. Although the population sample was small, the study 

seems to have been well controlled. However, neither exposure levels nor past or current levels of DDT 

(or DDE) in blood or in adipose tissue were available. 

As in humans, the nervous system appears to be one of the primary targets in animals after acute, 

subchronic, and chronic oral exposure to DDT. Acute oral exposure to high doses of DDT has been 

associated with DDT-induced tremors or myoclonus (abrupt, repeated involuntary contractions of skeletal 

muscles), hyperexcitability, and convulsions in several species. These effects have been observed in rats 

after single oral gavage doses of 50–600 mg DDT/kg/day (Henderson and Wooley 1969b; Herr and 

Tilson 1987; Herr et al. 1985; Hietanen and Vainio 1976; Hong et al. 1986; Hrdina and Singhal 1972; 

Hrdina et al. 1973; Hwang and Van Woert 1978; Philips and Gilman 1946; Pranzatelli and Tkach 1992; 

Pratt et al. 1986; Tilson et al. 1987). Mice receiving a single oral gavage dose of 160 mg DDT/kg had 

tremors (Hietanen and Vainio 1976), and single doses of 200–600 mg p,p’-DDT/kg/day induced 

convulsions (Matin et al. 1981). In guinea pigs and hamsters similarly dosed, no tremors were observed 

at 160 mg DDT/kg, but hind leg paralysis occurred in guinea pigs (Hietanen and Vainio 1976). 

Acute oral exposure of animals to DDT has also been associated with increases in brain biogenic amine 

and neurotransmitter levels. Alterations in the metabolite 5-HIAA (5-hydroxy-indoleacetic acid), the 

degradation product of serotonin, have been reported to correlate with DDT-induced tremors; doses at 

50 mg/kg/day or greater resulted in increases in the levels of 5-HIAA in the brain (Hong et al. 1986; 

Hrdina et al. 1973; Hudson et al. 1985; Hwang and Van Woert 1978; Tilson et al. 1986). Alterations in 

the levels of other neurotransmitters have been found. The neurotransmitter changes observed are 

consistent with one of the putative mechanisms for DDT toxicity; DDT is thought to influence membrane 

ion fluxes and consequently potentiate neurotransmitter release (see Section 3.6.2). Acetylcholine and 

norepinephrine decreased in rats after acute exposure to 400 mg/kg DDT (Hrdina et al. 1973). Also, 

aspartate and glutamine were increased in brain tissue of rats (Hong et al. 1986; Hudson et al. 1985; 

Tilson et al. 1986). 
DDT, DDE, and DDD 92 

3. HEALTH EFFECTS 

The effects of acute DDT exposures on adult learning and motor activity are discussed in Section 3.2.2.6 

Developmental Effects, where they are compared to similar, but more pronounced, behavioral effects 

resulting from perinatal exposure. 

Body tremors and hunched appearance were observed in female Osborne-Mendel rats after 26 weeks of 

treatment in the diet with approximately 16 mg technical DDT/kg/day in a 78-week study (NCI 1978). 

Similar findings were reported by Rossi et al. (1977) in female Wistar rats after 9 weeks of treatment with 

approximately 34 mg technical DDT/kg/day in the diet. Intermediate and acute exposure to DDT resulted 

in changes in brain lipid metabolism that affected total brain lipids and the relative brain lipid ratios in 

nonhuman primates. Rhesus monkeys exhibited a decrease in total brain lipids and the relative amount of 

cholesterol to phospholipid after oral exposure to 10 mg technical DDT/kg/day for 100 days or a single 

oral dose of 150 mg/kg/day (Sanyal et al. 1986). The results were more pronounced in the intermediate 

group than the acute group. Decreases in the levels of brain phospholipids and cholesterol may result in 

altered neuronal transmission (Sanyal et al. 1986). Lipids associated with the myelin sheath were not 

affected by DDT. Staggering, weakness, and loss of equilibrium were observed in monkeys treated for up 

to 14 weeks with 50 mg p,p’-DDT/kg/day, a dose level that was also lethal (Cranmer et al. 1972a). No 

such manifestations of toxicity were seen with a 5 mg/kg/day dose. Hunched appearance was reported in 

male rats after 8 weeks of treatment with approximately 59 mg of p,p’-DDE/kg/day and in male mice 

after 22 weeks of treatment with approximately 27 mg of p,p’-DDE/kg/day in 78-week duration studies 

(NCI 1978). 

Some Rhesus and Cynomolgus monkeys exposed to approximately 6.9 mg of p,p’-DDT/kg/day or more 

in a 130-month dietary study experienced severe tremors (Takayama et al. 1999). Hyperactivity and 

tremors were also reported in chronic studies in mice at doses up to 8.3 mg DDT/kg/day (Kashyap et al. 

1977; Turusov et al. 1973). Even after a change of diet or a decrease in dose, the tremors persisted for 

several weeks. No clinical signs of neurotoxicity were observed in hamsters fed diets to provide doses up 

to 95 mg technical DDT or p,p’-DDE/kg/day for life (Rossi et al. 1983). 

The highest NOAEL values and all LOAEL values from reliable studies for neurological effects in each 

species and duration category are recorded in Table 3-1 and plotted in Figure 3-1. 
DDT, DDE, and DDD 93 

3. HEALTH EFFECTS 

3.2.2.5 Reproductive Effects 

Effects discussed in this section are those that result from alterations affecting reproduction from the time 

of gametogenesis through implantation of the conceptus in the endometrium. In addition, this section 

includes descriptions of effects on male and female reproductive organs; these effects may be direct 

effects or triggered by hormonal changes. Developmental effects are distinguished from reproductive 

effects by evaluation of the conceptus after it is implanted and can be categorized as structural 

abnormalities, altered growth, functional deficiencies, and death of the developing organisms. 

Developmental effects are discussed in Section 3.2.2.6. 

The potential association between DDT (and DDT metabolites) and reproductive end points has been 

examined in numerous studies. For example, a study of 240 women showed no difference in maternal 

DDT blood levels between 120 individuals who had a miscarriage and 120 controls; in both groups, the 

mean blood DDT levels were similar and below 6 ppb (Leoni et al. 1989). A more recent study of 

89 women with a history of at least two miscarriages found that 15% of them had DDE blood levels 

above a reference value of 2.5 ppb; the mean was 1.2 ppb and the range was between 0.01 ppb (detection 

limit ) and 8.6 ppb (Gerhard et al. 1998). Hexachlorobenzene, pentachlorophenol, and PCBs were also 

above reference values in 15–22% of the women. A study by Ron et al. (1988) reported no significant 

differences in DDT blood levels between 20 cases of premature rupture of fetal membranes and 

15 controls; the mean concentrations of DDT in blood were comparable, approximately 24 ppb. 

A difference in maternal DDE blood levels for women delivering full-term and premature infants was 

reported in a brief communication by O'Leary et al. (1970a). Twenty-three women who delivered early 

had infants with mean DDE blood levels of 19–22.1 ppb, while 44 women who delivered at full-term had 

infants with mean DDE blood levels of 4.9–6.1 ppb; DDE was the only organochlorine monitored. 

Studies of women from India reported that total DDT residues (DDT plus DDE and DDD) in maternal 

blood from 25 full-term cases ranged from 7 to 73 ppb (mean 26 ppb) compared to 48–481 ppb (mean 

66 ppb) in 15 pre-term and 90–1,280 ppb (mean 394 ppb) in 10 cases of spontaneous abortion (Saxena et 

al. 1980, 1981). Mean DDT residues in the respective placentas were 24, 66, and 162 ppb. Other 

organochlorine pesticides were also elevated in the pre-term and abortion cases relative to full-term cases. 

A study of 54 women from Brazil found no differences in maternal blood (at delivery) DDT residue 

levels between 30 cases of full-term delivery (31 ppb) and 24 cases of pre-term delivery (30 ppb) 

(Procianoy and Schvartsman 1981). However, analyses of umbilical cord blood showed significantly 
DDT, DDE, and DDD 94 

3. HEALTH EFFECTS 

elevated levels of DDT in pre-term newborns compared to full-term. No other chemical was monitored in 

this study. A similar study of Israeli women reported that maternal blood of 10 full-term cases had a 

concentration of 27 ppb of total DDT residues vs. 120 ppb in blood from 17 pre-term cases (Wassermann 

et al. 1982); other organochlorines such as PCBs, lindane, heptachlor epoxide, and dieldrin were also 

elevated in pre-term delivery cases. A more recent study of New York City women found no association 

between blood DDE levels and pre-term birth after comparing 20 cases with 20 matched controls 

(Berkowitz et al. 1996). The median DDE blood levels were 1.4 ppb in cases and 1.3 ppb in controls. 

Recently, Longnecker et al. (2001) presented the results of an analysis of 2,380 children, of whom 

361 were born pre-term and 221 were small-for-gestational-age. The children were born between 1959 

and 1966 and were part of the U.S. Collaborative Perinatal Project. Serum samples were taken from the 

mothers in the third trimester of pregnancy and were stored at -20 EC without recorded thaws. Analyses 

of DDE in serum were conducted between 1997 and 1999. DDE concentrations were evaluated in 

relation to odds of pre-term birth and small-for-gestational-age birth with logistic regression models. The 

median concentration of DDE in serum was 25 ppb and the range was 3–178 ppb. The authors found that 

the adjusted odds ratios for pre-term birth increased steadily with increasing DDE concentrations (trend 

p<0.0001). Quadratic spline models showed that the odds of pre-term birth began to increase at a DDE 

concentration of 10 ppb. Adjusted odds for small-for-gestational-age also increased, but less consistently 

(trend p=0.04). 

A case-control study of women with endometriosis (n=86) and a matched control group of women 

without the condition (n=70) found no association between plasma concentration of DDT compounds and 

the occurrence of endometriosis (Lebel et al. 1998). No differences were observed between cases and 

controls in crude geometric mean total plasma DDT (DDT plus DDE) or in geometric mean total plasma 

DDT adjusted for age, body mass index, and symptoms indicating a need for laparoscopy. No differences 

were observed in crude geometric mean p,p’-DDE or p,p’-DDT concentrations in plasma between cases 

and controls. 

An ecological study evaluated the relationship between several factors, including the concentration of 

p,p’-DDE in tree bark as a measure of DDT contamination in the environment, and birth rate among 

countries or pregnancy rate among states in the United States (Cocco 1997). In both multivariate 

regression analyses, no relationship between DDE exposure (as measured by the environmental 

concentration) and response was observed. A possible association between chlorinated hydrocarbons in 

blood and infertility was investigated in a group of 489 infertile women from Germany (Gerhard et al. 

1999). Only eight of the women had total DDT blood levels higher than a reference value of 2.5 ppb; the 
DDT, DDE, and DDD 95 

3. HEALTH EFFECTS 

mean concentration was 1.0 ppb and the range span was 0–9.6 ppb. The authors reported that elevated 

o,p’-DDT and p,p’-DDT were found less often in women who conceived during the study than in women 

who did not. In addition, the conception rate was reduced with increasing total DDT levels from 18% for 

DDT levels below 0.5 ppb to 10% for DDT levels above 1.0 ppb. 

Concentration of p,p’-DDE in human milk was inversely related to duration of lactation in women of 

Tlahualilo, Mexico, who had lactated previously, but not among women having their first lactation 

(Gladen and Rogan 1995). Median lactation duration declined from 7.5 months among women with 

lipid-adjusted DDE concentration of 0–2.5 ppm, to 3 months among women with $12.5 ppm in their 

milk; the difference was statistically significant. Significantly elevated crude hazard ratios (defined by 

the authors as estimated ratios of the hazard of weaning relative to the 0–2.5 ppm group; hazard was 

defined as the instantaneous probability of weaning) were observed in groups with milk DDE levels 

$7.5 ppm; hazard ratios adjusted for various determinant factors were elevated above unity in the group 

with milk DDE $12.5 ppm. A similar inverse relationship between milk DDE and lactation duration was 

seen in women in the United States (Rogan et al. 1987). In contrast, no correlation was seen between 

DDE concentration in maternal milk fat and birth weights, head circumference, or neonatal jaundice, but 

the authors indicated that higher levels of DDE (>4 ppm) in maternal milk fat were associated with 

hyporeflexia in infants (Rogan et al. 1986). The effects of DDE on lactation duration are likely due to the 

ability of DDE to disrupt the normal endocrine regulation of lactation, as estrogen is a potent inhibitor of 

milk secretion (Guyton and Hall 2000) 

It has long been suspected that technical-grade DDT has estrogen-like properties based on findings in 

wildlife exposed to the pesticide (Bishop et al. 1991; Fry and Toone 1981; Fry et al. 1987; Guillette et al. 

1994); results from studies in laboratory animals have left little doubt (Bitman and Cecil 1970; Clement 

and Okey 1972; Singhal et al. 1970; Welch et al. 1969). The estrogenic activity is largely due to the 

o,p’-isomer of DDT, which is present as a 15–21% contaminant of technical DDT (Metcalf 1995). For 

example, estrogenic effects on the uterus (uterotrophic: increased weight and glycogen content) and 

premature vaginal opening were observed in immature rats given diets that provided doses of 100 mg/kg 

o,p’-DDT from days 23–30 of life suggesting an agonistic action with estrogen; p,p’-DDT did not have 

such estrogenic activity (Clement and Okey 1972). Also, Singhal et al. (1970) showed that injection of a 

single dose of 100 mg o,p’-DDT/kg to ovariectomized rats mimicked estrogens in increasing uterine 

weight, glycogen content, and the activities of several enzymes involved in glycolysis and the hexose 

monophosphate shunt pathway. Recently, Diel et al. (2000) showed that oral administration of o,p’-DDT 

to juvenile ovariectomized rats produced the typical increase in uterine weight, although it induced a 
DDT, DDE, and DDD 96 

3. HEALTH EFFECTS 

pattern of gene expression different from ethinylestradiol. Much of the research in this area has been 

conducted in experimental animals administered the test substance parenterally or, more recently, in in 

vitro systems. Representative studies are discussed in Section 3.6.2, Mechanisms of Toxicity. 

No spermatotoxic effects were observed in an acute screening test in rats given a single oral gavage dose 

of 100 mg p,p’-DDT/kg/day or five daily doses of 50 mg/kg/day (Linder et al. 1992). In juvenile male 

rats dosed by oral gavage on days 4 and 5 of life with 500 mg/kg/day or from day 4 to day 23 with 

200 mg DDT/kg/day, a decrease in testis weight was observed (Krause et al. 1975). After two doses, 

significant decreases were seen at 34 days, and after repeated lower doses, decreases were significant at 

days 18, 26, and 34. Treated males were mated with healthy females on days 60 and 90. The number of 

fetuses and implantations was decreased 30% at the 60-day mating but not at the 90-day mating of rats 

dosed on days 4 and 5. For rats receiving multiple doses, the decreases were 95 and 35% after mating at 

days 60 and 90, respectively. In adult male rats dosed with 200 mg DDT/kg every other day (oral 

gavage) for 2 weeks, serum testosterone levels were decreased (statistically significant) compared to 

controls, but serum levels of luteinizing hormone (LH) and follicle stimulating hormone (FSH) were not 

significantly altered (Krause 1977). In the same study, testicular levels of testosterone were decreased, 

although not significantly, in adult male rats treated with 100 mg DDT/kg 3 times/week for 3 weeks, and 

no histologic effects on spermatogenesis were seen (Krause 1977). Oral gavage dosing of male mice at 

6.25 mg/kg/day of p,p’-DDT for 28 days had no effect on testis weight (Orberg and Lundburg 1974). In a 

dominant lethal test in male rats with a single oral gavage dose of 100 mg/kg of p,p’-DDT, a clear 

genotoxic response was not reported, but the proportion of mated females with dead implants was 

somewhat increased. 

The effect of DDT on fertility was examined in male and female rats fed diets providing an intake of 

0.56 mg DDT/kg/day for 60 days before mating and continuing throughout gestation (Green 1969). A 

75% depression of fertility was found, but there was no effect on litter size. The F1 pups from these dams 

were completely infertile when mated. Jonsson et al. (1976) reported that dietary dosing of female rats 

with technical-grade DDT for 36 weeks caused sterility at 12 mg/kg/day, and this was accompanied by a 

decrease in serum progesterone; 6 mg/kg/day had no effect on fertility. When 21-day-old female rats 

were fed diets to provide o,p’-DDT levels of 0.1–4.0 mg/kg/day for up to 14 weeks, no effects on age of 

vaginal opening were seen, and after mating, there were no effects on litter size or pup weight at birth or 

weaning (Wrenn et al. 1971). No adverse reproductive effects were reported in rats treated with 10 mg 

p,p’-DDE/kg/day for 5 weeks before mating and during gestation and lactation (Kornbrust et al. 1986). 

Reproductive parameters examined included percent sperm positive, percent pregnant, gestation length, 
DDT, DDE, and DDD 97 

3. HEALTH EFFECTS 

litter size, sex ratio of pups, and milk production and composition. Significantly decreased fertility was 

observed in female mice fed technical-grade DDT for 60 days at a dose level of 51.4 mg/kg/day (Bernard 

and Gaertner 1964). Female mice exposed to 2.0 mg p,p’-DDT/kg/day for periods that included 

premating and gestation had decreased number of implanted ova, lengthening of the estrus cycle, 

decreased corpora lutea, and decreased implants (Lundberg 1973, 1974). Administration of up to 

approximately 3.4 mg technical DDT/kg/day in the diet to mice for 86 days, including mating, resulted in 

no adverse reproductive effects, but treated mice had larger litter sizes relative to controls (Ledoux et al. 

1977). Treatment of male and female mice for 120 days with approximately 1.3 mg technical 

DDT/kg/day (30 days before mating plus 90 additional days) had no effect on fertility, fecundity, or litter 

size (Ware and Good 1967). Treatment of pregnant rabbits with $10 mg p,p’-DDT by gavage on 

days 7–9 of gestation resulted in premature deliveries and a significant increase in the number of 

resorptions (Hart et al. 1971, 1972); however, no such effects were seen after treatment on gestation 

days 21–23 (Hart et al. 1972). Neither fertility nor pre- or postimplantation embryonic losses were 

significantly affected in female rabbits administered 3 mg technical DDT/kg (only dose level tested) 

3 days/week for 12–15 weeks before artificial insemination and throughout gestation (Seiler et al. 1994). 

However, this exposure regime significantly reduced ovulation rate, but did not cause treatment-related 

histopathological alterations in the ovarian cortex, corpora lutea, or uterus (Lindenau et al. 1994). 

In recent years, evidence has been presented that the persistent DDT metabolite, p,p’-DDE, is an 

androgen receptor antagonist; this evidence is discussed in Section 3.6.2, Mechanisms of Toxicity. A 

number of experiments have investigated in vivo antiandrogenic effects. Kelce et al. (1995) conducted a 

series of experiments in male rats administered p,p’-DDE at different ages. Effects from prenatal plus 

juvenile exposure are discussed in Section 3.2.2.6, Developmental Effects. Treatment of adult male rats 

(120 days old) with 200 mg p,p’-DDE/kg/day for 4 days significantly reduced androgen-dependent 

seminal vesicle and ventral prostate weight relative to controls (Kelce et al. 1995). These rats had been 

castrated and implanted with testosterone-containing Silastic capsules to maintain a constant serum 

testosterone level. Prostates from treated rats had a 13-fold increase in androgen-repressed testosteronerepressed prostatic message 2 (TRPM-2) messenger RNA levels and a 35% decline in androgen-induced 

prostate binding subunit 3 (C3) mRNA levels relative to control rats (Kelce et al. 1995). These findings, 

coupled with results from in vitro studies (Section 3.6.2) on receptor (androgen and estrogen) binding, 

suggested that the antagonistic effects of DDT on the male reproductive system are mediated by 

p,p’-DDE through competitive inhibition of binding of androgens to the androgen receptor (AR) and 

subsequent inhibition of transcriptional activity. In a subsequent study, Kelce et al. (1997) observed a 

significant reduction in immunohistochemical staining of androgen receptor in epididymal nuclei of adult 
DDT, DDE, and DDD 98 

3. HEALTH EFFECTS 

rats given 200 mg p,p’-DDE/kg/day for 5 days as well as a significant increase in TRPM-2 and a decrease 

in testosterone-induced C3 (third subunit of prostatein or prostate specific binding protein mRNA). Kelce 

et al. (1997) also reported a significantly reduced seminal vesicle and ventral prostate weight in the rats. 

Reproductive studies of chronic dietary DDT exposure or multigeneration studies have not generally 

indicated reproductive toxicity. However, the doses used were sufficiently low so that tremors, 

convulsions, and death would not be a confounding factor. Keplinger et al. (1970) conducted a sixgeneration dietary study in mice with two matings per generation; mice in all generations were mated at 

4 months of age. No effects were observed at 5 mg/kg/day; 20 mg/kg/day caused decreased fertility 

evidenced by decreased viability and lactation indices, and the 50 mg technical-grade DDT/kg/day dose 

caused frank toxic effects and was discontinued after three generations. No adverse effects on litter size 

were reported after 15 months of exposure of field mice to technical-grade DDT at dietary levels 

providing up to 2.4 mg/kg/day (Wolfe et al. 1979). No adverse effects on reproduction (end points 

assessed limited mostly to fecundity and fertility) were observed in rats fed up to 18.6 mg technical-grade 

DDT/kg/day in the diet for 2 generations (Ottoboni 1969), 1.25 mg/kg/day for 3 generations (Treon et al. 

1954), or 1.7 mg/kg/day for 11 breedings (Ottoboni 1972). Duby et al. (1971) found no reproductive 

effects in two successive generations of rats fed technical-grade DDT (0.5 mg/kg/day), p,p’-DDT 

(1.5 mg/kg/day) or o,p’-DDT (0.3 mg/kg/day). A three-generation study was conducted in Beagle dogs 

given daily oral doses of 0, 5, or 10 mg technical-grade DDT from weaning to termination (Ottoboni et al. 

1977); all dogs were sacrificed by 28 months of age and all females were mated at the first estrus only. 

The parental generation consisted of 4 males and 7–10 females/group. In the F1 and F2 generations, 

8–19 females/dose group were mated with males of the same generation. The only significant effect 

noted was a 2–3-month reduction in the age at first estrus in females of the F2 generation, but the 

relationship of this effect to DDT administration is questionable. A total of 650 pups were produced. No 

effects were observed on length of gestation, fertility, litter size, viability, gestation, or lactation indices. 

In another study in dogs, Deichman et al. (1971) reported that daily administration of p,p’-DDT by 

capsule for 14 months at a level of 12 mg/kg/day caused subnormal reproduction in dogs. However, this 

was a one-generation study with several chlorinated pesticides, only one dose of DDT (without aldrin) 

was administered to four females, and these females were mated to males that had been fed DDT plus 

aldrin. In addition, the age of the dogs at initiation of study was not provided, and mating in some dogs 

took place up to 19 months after dosing was discontinued. No correlation between fertility and levels of 

DDT in adipose tissue at time of mating could be made, and no clear conclusions on the effect of DDT on 

reproduction can be determined. 
DDT, DDE, and DDD 99 

3. HEALTH EFFECTS 

In the 78-week chronic bioassay conducted by NCI (1978), no treatment-related adverse effects on the 

ovaries, uterus, mammary gland, or prostate were observed in Osborne-Mendel rats treated in the diet 

with up to 45 mg technical DDT/kg/day, 59 mg p,p’-DDE/kg/day, or 231 mg technical DDD/kg/day. The 

same findings were reported for B6C3F1 mice treated with up to 30.2 mg technical DDT/kg/day, 49 mg 

p,p’-DDE/kg/day, or 142 mg technical DDD/kg/day (NCI 1978). Reproductive function was not 

evaluated in this study. 

In summary, studies in humans suggest that high DDT/DDE burdens may be associated with alterations in 

end points that are controlled by hormonal function such as duration of lactation, maintenance of 

pregnancy, and fertility. High blood levels of DDE during pregnancy have also been associated with 

increased odds of having pre-term infants and small-for-gestational-age infants. Perinatal exposure of 

animals to DDT/DDE has caused alterations in the reproductive organs and infertility. 

The highest NOAEL values and all LOAEL values from reliable studies for reproductive effects in each 

species and duration category are recorded in Tables 3-1 and 3-2 and plotted in Figures 3-1 and 3-2. 

3.2.2.6 Developmental Effects 

The proper development of many systems and functions depends on the timely action of hormones, 

particularly sex steroids; therefore, interfering with such actions can lead to a wide array of effects that 

may include altered metabolic, sexual, immune, and neurobehavioral functions. Effects of this type, that 

occur following exposure during fetal life via the placenta or early in life caused by either direct exposure 

to chemicals or exposure via maternal milk, are discussed in this section. 

A prospective longitudinal study of children from North Carolina examined whether prenatal or 

lactational exposure to background levels of DDE (or PCB) were associated with altered pubertal growth 

and development (Gladen et al. 2000). Exposure was estimated by measuring p,p’-DDE in maternal 

breast milk, maternal blood, cord blood, and placenta. The original cohort consisted of 856 children, and 

follow-up analysis at puberty was conducted on 594 of these individuals (316 girls and 278 boys). Most 

of the children were between 12 and 14 years old when they were first contacted and the duration of 

participation ranged from 1 to 5 years. The transplacental DDE index ranged from 0.3 to 23.8 ppm 

(median 2.4 ppm) and the lactational DDE index from 0.2 to 96.3 mg (median 6.2 mg). Height and 

weight adjusted for height of boys at puberty increased with transplacental exposure to DDE; those with 

the highest exposures were 6.3 cm taller and 6.9 kg larger than those with the lowest exposures; no such 
DDT, DDE, and DDD 100 

3. HEALTH EFFECTS 

effect was seen among girls. There was no effect on the age at which pubertal stages were attained. 

Lactational exposures to DDE had no apparent effects. Children from this cohort had normal birth weight 

and normal growth in the first year of life. 

A study of German children examined the potential association between height and serum levels of DDE 

(Karmaus et al. 2002). The children were from three different regions from Germany; two regions were 

industrial and agricultural regions, but no information was given regarding the third region, or the 

distribution of children by region. Data on weight and height from birth to the age of 4 years (seven 

measurements) were obtained from medical records; measurements at the ages of 8, 9, and 10 years old 

were conducted by the investigators. Blood was collected only once, when the children were 8 years old, 

and was analyzed for DDE and PCBs. The number of children from whom weight and height were 

measured at each time point varied between 202 (at 43–48 months) and 323 (at birth). At all time points, 

boys outnumbered girls. Whole blood DDE concentrations were stratified in quartiles: 0.08–0.2, 

0.21–0.29, 0.3–0.43, and 0.44–40.4 ppb. Analysis of the results controlling for relevant confounders 

showed that DDE was a significant predictor of height in girls starting at 1.3 months and remained 

significant through 8 years of age. On the average, girls in the highest DDE quartile were 1.8 cm shorter 

than girls in the lowest quartile. At 10 years old, no association was evident between height and DDE. 

The fact that only one DDE measurement was made, at 8 years of age, casts some uncertainty on the 

significance of the results. 

Evaluation of Inuit infants exposed to p,p’-DDE (and other organochlorines) in utero revealed no 

association between immunological parameters (white blood cell counts, lymphocyte subsets, 

immunoglobulins) and prenatal exposure to organochlorines (Dewailly et al. 2000). Organochlorine 

concentration in breast milk fat was used as an index of exposure and evaluations were conducted at 3, 7, 

and 12 months of age. The authors also examined whether organochlorine exposure was associated with 

the incidence of infectious diseases and found that during the second evaluation period, that the risk of 

otitis media increased with prenatal exposure to p,p’-DDE, hexachlorobenzene, and dieldrin. They also 

found that the relative risk for 4- to 7-month-old infants in the highest tertile of p,p’-DDE exposure was 

1.87 (95% CI, 1.07–3.26) compared to infants in the lowest tertile. In addition, the relative risk of otitis 

media over the entire first year of life also increased with prenatal exposure to hexachlorobenzene and 

p,p’-DDE. Dewailly et al. (2000) pointed out that “because all organochlorines originate from the same 

few food items and share a number of properties (persistence, liposolubility), their concentrations in 

breast milk are correlated with each other. Therefore, associations of health outcomes with p,p’-DDE or 

hexachlorobenzene may be due to other organochlorines found in the organochlorine mixture.” 
DDT, DDE, and DDD 101 

3. HEALTH EFFECTS 

In animals, DDT can produce embryotoxicity, fetotoxicity, and abnormal development of the sex organs. 

Estrogen-like effects on the developing reproductive system have been reported. The o,p’-isomers of 

DDT and DDE have greater estrogen-like effects than other isomers and when administered to rats in the 

first few days of life can seriously affect the development and maturation of the reproductive system. 

Developmental effects have been observed in animals after acute oral exposure to DDT during gestation 

or in the early perinatal development period; the seriousness of these effects is dependent on the isomeric 

form, the dose, and the timing of exposure. Exposure during early gestation resulted in a 25% decrease in 

fetal body weights and in significant decreases in fetal brain and kidney weights in the offspring of 

pregnant rabbits given oral gavage doses of 1 mg DDT/kg/day on gestation days 4–7 (Fabro et al. 1984). 

Offspring from rabbit dams orally exposed to a dose level of $10 mg p,p’-DDT/kg/day by gavage on 

days 7–9 of gestation showed a significant reduction in weight on day 28 (Hart et al. 1971, 1972). 

However, treatment late in gestation (days 21, 22, and 23) did not induce such an effect (Hart et al. 1972). 

Gellert and Heinrichs (1975) exposed pregnant rats orally to 28 mg o,p’-DDT, p,p’-DDT, o,p’-DDE, or 

o,p’-DDD/kg/day on days 15–19 of gestation. No significant effects on body weight, weight of the 

ovaries and pituitary, estrous cycle, or vaginal opening in the offspring were noted with the exception of a 

small but significant delay (2 days) in vaginal opening with the o,p’-DDD isomer. The p,p’-isomer is 

most prevalent in the environment, accounting for approximately 85% of the total amount of DDT, DDE, 

or DDD found; also, technical-grade DDT contains between 65–80% p,p’-DDT, between 15–21% 

o,p’-isomer, and up to 4% p,p’-DDD (Metcalf 1995). 

As previously mentioned in Section 3.2.2.5 Reproductive Effects, the DDT metabolite, p,p’-DDE, has 

been found to have antiandrogenic activity (Kelce et al. 1995, 1997). In vivo antiandrogenic effects on 

adult rats are discussed in that section. Kelce and co-workers showed that male pups from Long-Evans 

dams exposed during gestation days 14–18 to 100 mg p,p’-DDE/kg/day and then exposed indirectly to 

maternally stored p,p’-DDE via breast milk had significantly reduced anogenital distance at birth and 

retained thoracic nipples on postnatal day 13. Female rats normally have a shorter anogenital distance 

than males. Treatment of weanling male rats from either day 21 or 25 (specific day unclear in text) until 

day 57 of age with 100 mg p,p’-DDE/kg/day resulted in a statistically significant delayed onset of puberty 

(measured by the age of preputial separation) by 5 days. The fact that serum levels of testosterone were 

not reduced suggested that the antiandrogenic effects were not confounded by the reported ability of 

DDT-related chemicals to increase steroid metabolism. Other gestational exposure studies have 

confirmed these findings. For example, anogenital distance was not affected in male Sprague-Dawley 

rats on postnatal day 2 after treating the dams with up to 100 mg p,p’-DDE/kg on gestation days 14–18, 

but was significantly reduced in similarly exposed Long-Evans pups (You et al. 1998). A 10 mg/kg dose 
DDT, DDE, and DDD 102 

3. HEALTH EFFECTS 

to the dams was without effect in the Long-Evans pups. Anogenital distance was not affected in female 

pups from either strain. Treatment of the dams with 10 mg p,p’-DDE/kg resulted in retention of thoracic 

nipples in Sprague-Dawley pups, but only the higher dose (100 mg/kg) had this effect in Long-Evans 

pups. Treatment with p,p’-DDE also resulted in an apparent reduction of androgen receptor expression in 

male sex organs from mainly high-dose Sprague-Dawley pups, as shown by immunochemical staining; 

however, there were no changes in androgen receptor steady state mRNA levels in the high-dose SpragueDawley rats, but mRNA were increased 2-fold in the high-dose Long-Evans rats. Exposure of the pups to 

p,p’-DDE during gestation and lactation had no significant effect on the onset of puberty in either strain. 

Further studies by the same group showed that prenatal exposure to 100 mg/kg p,p’-DDE reduced 

prostate ventral weight in 85-day-old males, but did not alter the weight of the testes, epididymis, and 

seminal vesicles (You et al. 1999a); serum levels of testosterone and LH were not significantly altered by 

DDE, but treatment with DDE was associated with expression of TRPM-2, an androgen-repressed gene. 

A similar study in Holtzman rats exposed during gestation days 14–18 to doses between 1 and 200 mg 

p,p’-DDE/kg (offspring were exposed to p,p’-DDE in utero and via breast milk) found reduced anogenital 

distance in males on postnatal day 1 and reduced relative ventral prostate weight on postnatal day 21 at 

50 mg p,p’-DDE/kg and higher, but not at 10 mg p,p’-DDE/kg (Loeffler and Peterson 1999). On 

postnatal day 4, anogenital distance was reduced only at 200 mg/kg/day. Nipple retention on postnatal 

day 13 was significantly increased at 100 and 200 mg/kg/day. Doses up to 100 mg/kg/day to the dams 

had no effect on the onset of puberty, but 200 mg/kg/day did significantly delay puberty in males by less 

than 2 days. Androgen receptor staining in the ventral prostate was also reduced on postnatal day 21 at 

100 mg/kg/day (the only dose tested). Serum levels of testosterone or 3α-diol androgens were not 

significantly altered at any time. mRNA levels of androgen regulated genes from both ventral and 

dorsolateral prostate were not significantly changed in treated animals on postnatal day 21. This study 

also reported that at the 100 mg/kg dose level, cauda epididymal sperm number was reduced by 17% on 

postnatal day 63 relative to controls. No measurements of DDE body burden were made in the 

200 mg/kg/day offspring postnatally, so it is difficult to determine whether effects on puberty were due to 

the previous gestational plus lactational exposures or directly due to the effects of DDE present near the 

time of puberty. Gray et al. (1999) conducted similar studies and reported that gestational exposure 

(gestation day [Gd] 14–18) to p,p’-DDE (100 mg/kg/day) resulted in a significant decrease in ventral 

prostate weight at 15 months of age in Sprague-Dawley rats and also in decreases in weights of glans 

penis, ventral prostate, and epididymis in 10-month-old Long-Evans rats. Newborn Long-Evans rats also 

showed reduced anogenital distance and increased mean number of retained nipples. Exposure of 

pregnant mice to 0.18 mg/kg of the estrogenic o,p’-DDT during gestation days 11–17 had no significant 
DDT, DDE, and DDD 103 

3. HEALTH EFFECTS 

effect on preputial gland weight or testes weight at 6 months of age, but a lower dose of 0.018 mg kg/day 

reduced absolute testes weight by 13% (Palanza et al. 1999b). This was associated with a decreased 

intensity of aggressive behavior in the mice. 

There are limited data suggesting that if mice exposed to DDT in utero and during lactation are further 

exposed to DDT postnatally, responses in both immunological plaque forming assays and lymphoproliferative assays are reduced (Rehana and Rao 1992). The results of this study are difficult to interpret 

because of experimental design issues, including the lack of a comparison unexposed group and 

ambiguity about whether statistical testing was done. 

Administration of DDT in utero or to neonates during sensitive periods in nervous system development 

has caused behavioral and neurochemical changes in adult mice (Craig and Ogilvie 1974; Eriksson and 

Nordberg 1986; Eriksson et al. 1990a, 1990b, 1992, 1993; Johansson et al. 1995, 1996; Palanza et al. 

1999b; vom Saal et al. 1995). Offspring of dams given 34.3 mg/kg/day DDT by oral gavage during 

gestation and lactation displayed impaired learning and decreased memory function in a maze when tested 

1 and 2 months after weaning (Craig and Ogilvie 1974). However, the dose was sufficiently high in this 

study to cause 39% mortality in the pups before weaning. 

Ten-day-old mice treated once with 0.5 mg technical DDT/kg/day demonstrated behavioral changes when 

tested at 4 months of age (Eriksson et al. 1990a, 1990b). The test entailed placing animals in a new cage 

with infrared motion detecting beams for 1 hour. Locomotion (horizontal movement), rearing (vertical 

movement), and total activity (vibration, tremors, grooming, movement) were scored and summed for 

each 20-minute period. Initially, motor activity was fairly comparable between DDT-treated mice and 

controls; locomotion and activity were similar, although rearing was increased. During the last 

40 minutes in the box, the DDT-treated animals had significantly more locomotion, rearing, and activity 

than the controls. When control mice were placed in a new environment, such as the test cage, they 

initially increased their activity as they explored. Eventually, this exploratory activity decreased as they 

become familiar with their new environment; this phenomenon is known as habituation. Thus, the 

authors concluded that perinatal DDT exposure interferes with habituation, which is considered a simple, 

nonassociative learning process. The neurobehavioral effects in the 4-month-old mice only appeared after 

dosing at 10 days and not after similar dosing at day 3 or 19 of age. The induction period was limited to a 

peak in the rate of brain development and development of muscarinic cholinergic receptors (Ericksson et 

al. 1992). Since this behavioral assay used motor exploration as the end point, an alternate interpretation 

of the observed results is that the DDT directly affected neural regulation of activity rather than learning; 
DDT, DDE, and DDD 104 

3. HEALTH EFFECTS 

however, a problem with appropriate control of motor behavior would also be a functional deficit of 

concern. Hyperactivity has been observed in DDT-treated adult mice (Kashyap et al. 1977; Rossi 1977; 

Turusov et al. 1973). 

One month after behavioral testing, the potassium-evoked release of acetylcholine from cortical slices 

was significantly increased in treated mice compared to controls. This is consistent with the wellcharacterized effect of DDT in slowing the closure of sodium channels and thus causing general central 

nervous system stimulation and neurotransmitter release. Apparently, behavioral effects seen in adult 

mice represented a persistent neurological change since no residual DDT from their neonatal exposure 

remained in the brain (Eriksson et al. 1990b). Radiolabeled DDT was found in the brain at 1 and 7 days 

but not at 30 days following a single oral dose of 0.5 mg DDT/kg to 10-day-old mice. The classification 

of this effect as developmental, rather than an immediate effect of current DDT body burden, hinges on 

the apparent clearance of DDT before the behavioral testing. 

In parallel experiments, neurochemical changes were also observed shortly after 10-day-old mice were 

treated with DDT. However, all of these specific neurochemical changes did not persist into adulthood. 

A single low dose of DDT (0.5 mg/kg) administered orally to 10-day-old (preweaning) mice affected the 

muscarinic cholinergic receptors in the brain. A significant increase in the density of specific muscarinic 

receptors was observed in the cerebral cortex, but not the hippocampus, at 7 days postexposure, but was 

not evident as early as 24 hours postexposure (Eriksson and Nordberg 1986). Furthermore, a significant 

decrease in the percentage of muscarinic high-affinity binding sites and a corresponding increase in the 

percentage of muscarinic low-affinity binding sites were measured, indicating that most of the increased 

binding could be attributed to low affinity sites. According to the study authors, the low-affinity 

muscarinic binding sites are thought to correspond to the M1 receptors in the cerebral cortex, which are 

postulated to be associated with neuronal excitation (McKinney and Richelson 1984). The persistence of 

these observed neurochemical changes in muscarinic receptors in the cerebral cortex was also evaluated 

(Eriksson et al. 1990b). The density of the muscarinic receptors in the cerebral cortex, hippocampus, and 

striatum of the 3-month-old NMRI mice that had been given a single dose of 0.5 mg DDT/kg at 10 days 

of age was determined. The study authors reported a tendency towards a decrease in the amount of 

specific binding in the cerebral cortex, but no significant changes in the hippocampus and striatum. No 

significant changes were noted in choline acetyltransferase activity in the cerebral cortex, hippocampus, 

or striatum. However, significant (p<0.01) decreases in cerebral cortex muscarinic receptors at 4, 5, and 

7 months of age were seen in similar later studies from the same group of investigators (Eriksson et al. 

1992, 1993; Johansson et al. 1995, 1996). Tests of spontaneous motor activity conducted at the ages of 
DDT, DDE, and DDD 105 

3. HEALTH EFFECTS 

5 and 7 months old revealed that mice treated perinatally at the age of 10 days with DDT still exhibited 

hyperactive behavior relative to controls (Johansson et al. 1995, 1996). Nicotinic cholinergic receptor 

density was not significantly altered in 5- or 7-month-old mice treated with DDT at the age of 10 days. In 

these more recent studies, Johansson et al. (1995, 1996) also compared the motor responses to bioallethrin 

(a type I pyrethroid insecticide) and paraoxon in mice treated with and without DDT perinatally. From 

the graphic presentation of the results, it appears that pretreatment with DDT altered (increase or 

decrease) some of the responses attributed to paraoxon alone, but the results with bioallethrin were much 

less clear. However, treatment with DDT followed by bioallethrin significantly increased the density of 

muscarinic receptors in the cerebral cortex relative to DDT alone. This increase was later attributed to 

increased expression of muscarinic receptor m4 mRNA (Talts et al. 1998). 

Behavioral effects have also been observed in animals exposed as adults. Inhibition of nonassociative 

learning processes (habituation response) occurred in adult rats acutely exposed to DDT (Sobotka 1971). 

Open field activity was assessed in rats, and activity (crossing squares on a checkered board) was 

measured over a 5-minute period; the normal response is increased activity for the first 2 minutes and 

decreased activity in the last 2 minutes (e.g., habituation). This "habituation" response was significantly 

affected (continued activity) in adult rats when tested 24 hours after administration of a single dose of 

25 mg/DDT/kg but not at the lower doses tested (1 and 10 mg DDT/kg). The open field test used by 

Sobotka was not as sensitive a test of habituation as the tests on the Eriksson studies reviewed above. No 

differences in problem solving, locomotion speed, or reaction to stress were found between untreated rats 

and rats given oral doses of DDT up to 30 mg/kg/day (Khairy 1959). However, the scores for the pattern 

of locomotion significantly increased with dose (Khairy 1959). These doses are approximately 50 times 

that administered to neonates in the Eriksson studies. 

A different behavioral response, the rate of urine marking in a novel territory, was examined in a study by 

vom Saal et al. (1995). Pregnant mice were administered o,p’-DDT by gavage in doses ranging from 

approximately 0.018 to 91 mg/kg/day on gestation days 11–17. Behavioral testing of male mice was 

conducted at the age of 70 days. Exposure to $1.82 mg/kg/day resulted in a significant increase in the 

number of urine marks deposited by the mice during a 1-hour test in a novel environment. The authors 

(vom Saal et al. 1995) stated that since marking a territory is a central feature of the reproductive strategy 

of male mice, alteration of this behavior could markedly impact the social structure of the species. In a 

different study, the same group of investigators reported that exposure of mice to o,p’-DDT (0.018 or 

0.18 mg/kg/day) during gestation days 11–17 had no significant effect on the behavior of the male 

progeny at 30 days of age towards mouse pups (Palanza et al. 1999b). However, the lowest dose tended 
DDT, DDE, and DDD 106 

3. HEALTH EFFECTS 

to increase the proportion of males exposed in utero that attacked a male intruder, although the intensity 

of the attacks was lower than observed in controls. 

Intermediate-duration oral exposure to DDT in animals has been shown to produce developmental effects 

such as infertility, mortality, and slow development in offspring of exposed dams (Clement and Okey 

1974; Craig and Ogilvie 1974; Deichmann and Keplinger 1966). Exposure throughout gestation and 

lactation is more fetotoxic in rats than exposure only in gestation. Clement and Okey (1974) exposed 

pregnant rats to 1.7, 16.8, or 42.1 mg p,p’-DDT/kg/day for an unspecified period before mating and 

during gestation and/or lactation. An increase in mortality of pups exposed perinatally via dams receiving 

42.1 mg p,p’-DDT/kg/day was reported, along with a decrease in growth of the pups after exposure via 

nursing from dams receiving 16.8 or 42.1 mg p,p’-DDT/kg/day or 84 mg o,p’-DDT/kg/day (Clement and 

Okey 1974). In mice exposed from mothers receiving 34.3 mg/kg/day technical-grade DDT, preweaning 

mortality was observed in 10% of the neonates exposed to DDT prenatally and nursed on unexposed or 

"foster" mothers, while 39% of the neonates exposed perinatally (during gestation and through lactation) 

died before weaning (Craig and Ogilvie 1974). In addition, mice exposed perinatally subsequently 

showed learning impairment and decreased memory function in the maze test (Craig and Ogilvie 1974). 

Developmental effects, including preweaning mortality and premature puberty, were reported in animals 

in multigeneration studies. An increase in preweaning mortality was observed in the offspring of mice 

chronically exposed to 41.3 mg technical or p,p’-DDT/kg/day (Tomatis et al. 1972; Turusov et al. 1973). 

Del Pup et al. (1978) found a decrease in 30-day survival of neonatal mice after exposing successive 

generations of dams to 16.5 mg DDT/kg/day in the diet for a total of 70 weeks. Increases in abortions, 

stillbirths, and pup mortality were reported in mice exposed to 1.3–6.5 mg DDT/kg/day in a 

multigeneration study; however, most of the females in the 6.5-mg/kg group died before delivery (Shabad 

et al. 1973). Green (1969) reported that there was an increase in the number of resorptions in rats 

exposed to 0.56 mg DDT/kg/day and that no litters were produced by the second-generation animals in a 

multigeneration study. However, it is unclear from the study whether females were mated to exposed 

males or to untreated males. Treon et al. (1954) conducted a multigeneration study in which dams were 

fed 0.125–1.25 mg DDT/kg/day throughout gestation and lactation. No reductions in litter size were 

noted. Changes in preweaning mortality were reported but were not considered to be dose related. Other 

developmental aspects evaluated were negative. No conclusions concerning the developmental effects of 

DDT could be drawn from the Treon et al. study. 
DDT, DDE, and DDD 107 

3. HEALTH EFFECTS 

In a multigeneration study, Ottoboni et al. (1977) reported an increase in the incidence of premature 

puberty that increased with dose and with each consecutive generation among female dogs (4–19 dogs per 

group) dosed with 1, 5, or 10 mg p,p’-DDT/kg/day. However, the increase was significant only in the 

two high-dose groups when all generations were combined. A significant increase in constricting rings of 

the tail was seen in the offspring of rats fed 18.6 mg DDT/kg/day through two generations (Ottoboni 

1969). 

In summary, there is no conclusive evidence that DDT/DDE at the levels found in the environment cause 

developmental effects in humans, although two studies suggested that there may be an association 

between prenatal exposure and/or exposure early in life to DDE and alterations in height in children. In 

animals, DDT/DDE can produce embryotoxicity, fetotoxicity, and abnormal development of the sex 

organs. o,p’-DDT has been shown to have estrogenic properties, whereas p,p’-DDE has shown antiandrogenic properties. In addition, adult mice administered DDT early in life showed neurobehavioral 

alterations when tested later in life. 

The highest NOAEL values and all LOAEL values from reliable studies for developmental effects in each 

species and duration category are recorded in Tables 3-1 and 3-2 and plotted in Figures 3-1 and 3-2. 

Based on the studies by Eriksson and Nordberg (1986) and Eriksson et al. (1990a, 1990b), an acute oral 

MRL of 0.0005 mg/kg/day was calculated as described in the footnote in Table 3-1 and in Section 2.3. 

3.2.2.7 Cancer 

Numerous studies in the United States and abroad have examined the possible association between 

exposure to DDT and related compounds and cancer in humans. Studies have been conducted of 

members of the general population as well as of occupationally exposed individuals. Exposure has been 

traditionally assessed by measuring DDT residues (most often DDE because of its persistence in the body 

and in the environment) in blood or in adipose tissue. Current data indicate that DDT in either media is a 

valid biomarker of exposure to DDT residues and each media correlate well with one another providing 

that the concentration of residues is adjusted for lipid content. The drawback of trying to associate 

current residue levels determined at or near the time of diagnosis with the occurrence of cancer is that 

levels at diagnosis may be very different than those at the time when cancer began to develop, and this is 

particularly relevant for cancers that exhibit a long latency. These and other issues inherent to 

epidemiological studies make it difficult to draw definite conclusions about exposure to DDT/DDE/DDD 

and cancer. However, taking all factors into consideration, the existing information does not support the 
 

DDT, DDE, and DDD 108 

3. HEALTH EFFECTS 

hypothesis that exposure to DDT/DDE/DDD increase the risk of cancer in humans. The human studies 

below are presented in the following order: breast cancer, pancreatic cancer, Hodgkin’s disease and 

non-Hodgkin’s lymphoma, multiple myeloma, prostate and testicular cancer, endometrial cancer, and the 

occurrence of any cancer. 

Human Studies 

Breast Cancer. Many epidemiological studies have investigated the association between breast cancer 

and levels of DDT and DDT-derived compounds in blood or adipose tissue from humans. Some studies 

have suggested a positive association (Aronson et al. 2000; Dewailly et al. 1994; Falck et al. 1992; Güttes 

et al. 1998; Romieu et al. 2000; Wasserman et al. 1976; Wolff et al. 1993), while others do not support 

such an association (Demers et al. 2000; Dorgan et al. 1999; Helzlsouer et al. 1999; Laden et al. 2001a; 

Krieger et al. 1994; Liljegren et al. 1998; Lopez-Carrillo et al. 1997; Mendonca et al. 1999; Moysich et al. 

1998; Schecter et al. 1997; Unger et al. 1984; van’t Veer et al. 1997; Ward et al. 2000; Wolff et al. 2000a, 

2000b; Zheng et al. 1999, 2000). Below, studies that found an association between DDT/DDE burdens 

and breast cancer are summarized first, followed by those that found no significant association. In each 

case, serum DDT/DDE as a biomarker of exposure is discussed before adipose tissue. 

Evidence of a positive association between breast cancer and exposure to DDT and DDT-related 

compounds was found in a case-control analysis nested within a prospective study in which blood 

samples of New York City women attending a mammography clinic were collected between 1985 and 

1991 (Wolff et al. 1993). Serum DDE (not lipid-corrected) was determined on the archived blood 

samples of women who were diagnosed with breast cancer within 6 months of entering the study (n=58) 

and in matched, cancer-free control women from the same cohort (n=171). Controls were matched to 

case patients with respect to menopausal status, age at entry into the study, number and dates of blood 

donations, and day of menstrual cycle (if premenopausal) at time of first blood drawing. Mean serum 

DDE was significantly higher in case patients (11.0 ppb) than in control subjects (7.7 ppb) (p=0.031). 

The adjusted odds ratio for breast cancer in the highest quintile of serum DDE was significantly elevated 

(OR=3.68; 95% CI=1.01–13.50), using the lowest quintile as the referent group. A significant positive 

trend was identified between the odds ratio for breast cancer and increasing quintile (p=0.035). A 

significant positive trend was also obtained for the relation between the adjusted odds ratio and serum 

DDE when serum DDE was evaluated as a continuous variable using conditional multiple logistic 

regression (p=0.0037). The odds ratios were adjusted for first-degree family history of breast cancer, 

lifetime months of lactation, and age at first full-term pregnancy; other potential confounders, including 
DDT, DDE, and DDD 109 

3. HEALTH EFFECTS 

age at menarche, history of benign breast disease, history of tobacco and alcohol use, and race were found 

not to affect the outcome in preliminary evaluations. This study supported a positive relation between 

serum DDE and breast cancer, in spite of a short follow-up period and a relatively small number of 

subjects. See below for follow-up data from this study (Wolff et al. 2000b). 

The association between elevated serum DDE levels and risk of breast cancer was examined among a 

subsample of women from a larger breast cancer case-control study of women living in Mexico City 

(Romieu et al. 2000). The final analysis was restricted to 120 cases and 126 controls who had given birth 

to at least one child and had complete information on all key variables. Major predictors of DDE levels 

included DDT levels, age, duration of lactation, parity, and socioeconomic level. Mean lipid-adjusted 

serum DDE levels were higher among cases (3.84 ppm) than in controls (2.51 ppm) (p=0.02). This 

difference was more apparent for postmenopausal women; among premenopausal women, the difference 

was only marginally significant. Age-adjusted serum DDE levels were significantly related to breast 

cancer risk, but serum DDT levels were not. When DDE levels were examined as quartiles, after 

adjusting for age at menarche, duration of breast feeding, quetelet index, and menopausal status, there 

was a positive trend in the risk of breast cancer with increasing levels of serum DDE that was marginally 

significant (OR Q1-Q4=2.16; CI 95%, 0.85–5.50; p for the trend=0.06). The association was found to be 

stronger among those women who had experienced menopause; no association was found among 

premenopausal women. 

Several studies have reported positive associations between DDT/DDE/DDD body burden, measured in 

adipose tissue, and breast cancer. p,p’-DDE and p,p’-DDT levels in breast adipose tissue were 

significantly elevated in Hartford, Connecticut patients with breast cancer (n=20) in relation to agematched controls (n=20) who had benign breast disease (Falck et al. 1992); samples were obtained from 

biopsy or mastectomy tissue, near the time of diagnosis in 1987. Women from Quebec City with estrogen 

receptor (ER)-positive breast cancer cells (n=9) had significantly higher p,p’-DDE levels in breast 

adipose tissue (p=0.01) and plasma (p=0.052) than in controls with benign breast disease (n=17) near the 

time of diagnosis during 1991–1992 (Dewailly et al. 1994); the difference was not statistically significant 

for women with ER-negative breast cancer cells. Wasserman et al. (1976) found that the concentrations 

of p,p’-DDT and o,p’-DDD in malignant breast tissue of hospitalized Brazilian women (n=9) collected 

after diagnosis (dates not reported) were significantly greater than those found in adjacent “normal” breast 

tissue in the same women (p<0.01 and p<0.1, respectively). Concentrations of p,p’-DDE, p,p’-DDD, 

o,p’-DDT, and o,p’-DDE were also elevated in malignant breast tissue over adjacent normal tissue, but 

the differences were not statistically significant. A similar comparison was performed in a 1993–1994 
DDT, DDE, and DDD 110 

3. HEALTH EFFECTS 

case-control study of German women, in which the age-adjusted geometric mean concentration of 

p,p’-DDE in malignant breast tissue of recently mastectomized women (n=45) was found to be 62% 

higher (p=0.017) than the age-adjusted geometric mean concentration of p,p’-DDE in benign breast tissue 

in the control group (n=20); however, no statistically significant difference was found with respect to 

p,p’-DDT concentrations (Güttes et al. 1998). A hospital-base case-control study of Canadian women 

found a weak association between p,p’-DDE in breast adipose tissue and breast cancer (Aronson et al. 

2000). Biopsy tissues from 217 cases and 213 benign controls were analyzed for p,p’-DDT, p,p’-DDE, 

and other organochlorines. The geometric mean concentrations of p,p’-DDE in cases and controls were 

693 and 596 ppb, respectively; those of p,p’-DDT were 22.0 and 19.3 ppb, respectively. The only 

increased risk for p,p’-DDE was among a subgroup excluding current hormone replacement therapy 

users. Further evaluation of this cohort showed that the odds ratio of DDE was higher with risk of 

estrogen receptor-negative breast cancer than estrogen receptor-positive breast cancer (Woolcott et al. 

2001). A possible explanation suggested by the investigators was that tumors that are receptor-negative 

when diagnosed may have a faster rate of progression than tumors that are receptor-positive when 

diagnosed. 

Several studies found no association between serum DDT/DDE/DDD and breast cancer incidence in 

developed countries using blood samples collected prior to the diagnosis of breast cancer (Dorgan et al. 

1999; Helzlsouer et al. 1999; Hunter et al. 1997; Krieger et al. 1994; Ward et al. 2000). Cholesteroladjusted plasma DDE was determined in blood samples collected during 1989–1990 in a prospective 

study of the health of 121,700 married nurses in the United States (Hunter et al. 1997). Historical plasma 

DDE was no different between women who developed breast cancer before June 1992 (n=236) and pairwise matched control women who did not subsequently develop breast cancer. The lack of an association 

was also observed within strata of menopausal status, age, age at menarche, age at birth of first child, 

number of children, and history of lactation. A follow-up report to the Hunter et al. (1997) study included 

the results of adding to the analysis 143 postmenopausal cases and controls and of adjusting plasma 

organochlorines for triglycerides in addition to cholesterol (Laden et al. 2001a). Median (lipid-adjusted) 

concentrations of DDE in cases and controls were 0.768 and 0.817 ppm, respectively. The study found 

no evidence of an association between DDE and breast cancer within strata of age, age at menarche, age 

at birth of first child, number of children, history of benign breast disease, or family history of breast 

cancer. For DDE, the multivariate relative risk of breast cancer for women in the highest quintile 

(1.5–6.0 ppm) of exposure as compared with women in the lowest quintile (0.007–0.43 ppm) was 

0.82 (95% CI=0.49–1.37). 
DDT, DDE, and DDD 111 

3. HEALTH EFFECTS 

A prospective nested case-control study of blood samples collected between 1964 and 1971 compared 

DDE concentration in blood serum of northern California women who were later diagnosed with breast 

cancer (n=150; 50 each of white, black, and Asian women) with serum DDE of paired control women 

who did not develop breast cancer in the interval at least 6 months after the blood was drawn through the 

end of 1990 (Krieger et al. 1994). Serum DDE was not adjusted for serum lipid content. These blood 

serum samples were collected prior to the U.S. ban on DDT in 1972, and the serum DDE levels were 

much higher than those measured in the Wolff et al. (1993) study. Serum samples were collected an 

average of 14 years prior to cancer diagnosis. No significant difference in DDE serum level was observed 

in the combined analysis of 150 pairs of cases and controls (43.1 vs. 43.3 ppb). When the odds ratios 

were calculated for data stratified according to the amount of time between the serum sample collection 

and the breast cancer diagnosis, the results did not change. However, when ethnic groups were compared 

separately, high serum DDE concentrations were correlated with breast cancer incidence in Caucasian and 

African American women, but not Asian American women. For black women (50 pairs), serum DDE 

levels were higher in case patients than paired controls by an average of 5.7 ppb, although the difference 

was not statistically significantly higher (95% CI on the difference ranged from -3.3 to 14.8). The study 

also found that serum DDE was significantly higher in black and Asian women compared to white 

women. 

Serum DDE (both lipid-adjusted and unadjusted) was measured in blood samples collected in 1974 or 

1989 in a breast cancer case-control study nested within a prospective cohort study, consisting primarily 

of residents of Washington County, Maryland (Helzlsouer et al. 1999). Blood samples were collected 

from 20,305 residents of Washington County in 1974, and from 25,080 residents in 1989. A group of 

346 women who were diagnosed with breast cancer by June 1994 after having donated blood (for a total 

follow-up period of up to 20 years), who were residents of Washington County at the time of donating 

blood, and who had no other invasive cancers comprised the case group; 346 cancer-free control women 

were matched for age, race, menopausal status, and date of blood donation. Mean and median lipidadjusted DDE concentrations were nonsignificantly higher in controls than in women who subsequently 

developed breast cancer. Separate risk analyses conducted for quintiles of DDE concentration in samples 

from 1974 and for tertile samples from 1989 of lipid-adjusted serum DDE showed no association between 

DDE and breast cancer. Women who donated in both 1974 and 1989 and who were diagnosed with 

breast cancer after 1989 were included in both program-specific analyses. 

A hospital-based case-control study investigated risk for breast cancer associated with organochlorine 

exposure among 175 cancer patients, 181 control patients with benign breast disease, and 175 women in a 
DDT, DDE, and DDD 112 

3. HEALTH EFFECTS 

second hospital control group (Wolff et al. 2000a). Fifty-seven percent of the subjects were Caucasian, 

21% Hispanic, and 22% African-American. Chemicals evaluated in blood serum included p,p’-DDT, 

p,p’-DDE, trans-nonachlor, and higher and lower chlorinated biphenyls. Also, four tumor markers 

(estrogen, progesterone, p53, and erbB-2) were examined among cases. Most blood samples were 

collected prior to surgery, and none were taken more than 2 months after surgery. African-American 

women had significantly higher lipid-adjusted DDE levels (1.0 ppm, cases geometric mean) than 

Hispanic (0.71 ppm) and Caucasian (0.48 ppm) women, whereas DDT was highest among Hispanic 

women. However, organochlorine levels were not associated with risk of breast cancer, nor did 

organochlorines differ with respect to tumor stage or tumor markers. Higher DDE levels were associated 

with increasing body mass index (BMI), but with decreasing level of education, frequency of nulliparity, 

and frequency of family history of breast cancer. Wolff et al. (2000a) attributed their findings to 

historical patterns of exposure and to metabolic differences in organochlorines related to body mass 

index. 

An additional study by Wolff et al. (2000b) provides prospective information on the cohort studied earlier 

(Wolff et al. 1993) by these investigators. Because the authors were interested in assessing the half-lives 

of DDE and PCBs, the study was restricted to only cases with at least three yearly blood donations, after 

the first year of serum analysis. Lipid measurements were available on 110 cases and 213 controls. Halflives for DDE and PCBs were calculated for 84 cases and 196 controls who had contributed more than 

one blood donation. Odds ratios were calculated using conditional logistic regression analysis with DDE 

and PCB levels as both quartiles and continuous variables; some analyses, however, were conducted 

tertiles rather than quartiles. Adjustments were made for the following potential confounders: age at 

menarche, number of full-term pregnancies, age at first full-time pregnancy, first degree family history of 

breast cancer, months of lactation, height, BMI, and an interaction term for BMI and menopausal status. 

Breast cancer was significantly associated with nulliparity and family history of breast cancer. DDE 

levels were similar in cases and controls, whether lipid adjusted or not (geometric mean 977 ppb adjusted 

in cases vs. 1,097 ppb in controls). The lack of significance of these differences was maintained when 

estrogen receptor status of tumors was considered. Half-lives of DDE and PCBs were similar in cases 

and controls, and there was a strong correlation between the two chemicals. Half-life of DDE, but not 

serum level, was correlated with BMI (shorter among leaner women). The reason for the discrepancy 

between the earlier results (Wolff et al. 1993) and those from the present study was not apparent, and the 

authors could not rule out their earlier finding was a chance observation. 
DDT, DDE, and DDD 113 

3. HEALTH EFFECTS 

Demers et al. (2000) conducted a case-control study among women from the Quebec City area to assess 

not only cancer risk, but also disease aggressiveness in relation to plasma levels of several 

organochlorines, including p,p’-DDT and p,p’-DDE. The cohort consisted of 315 women newly 

diagnosed with breast cancer, 219 hospital-based controls, and 307 population controls. Mean 

concentrations of p,p’-DDT and p,p’-DDE in women with breast cancer were not significantly different 

from those of hospital or population controls. Moreover, high concentrations of p,p’-DDT and p,p’-DDE 

(concentrations divided in quintiles) were not related to breast cancer risk. Additional analyses restricted 

to cases showed that after adjusting for confounders, the relative risk of axillary lymph node involvement 

increased significantly with p,p’-DDE tertiles (and with other organochlorines as well). When tumor size 

and lymph node involvement were considered in the same model, odd ratios of having involved nodes 

after adjustment for tumor size and confounding factors increased with tertiles of p,p’-DDE. However, 

odd ratios of having a large tumor after adjustment for lymph node involvement and confounding factors 

did not increase with p,p,’-DDE plasma concentrations. There was no interaction between exposure to 

DDE/DDT and the hormonal status of the tumor (estrogen positive or negative) with regard to either 

tumor size or axillary lymph node involvement. 

A case-control study of 240 breast cancer patients and 477 controls was nested within a prospective study 

initiated in 1976 to investigate the fate of 7,712 women in the Copenhagen City Heart Study (Høyer et al. 

1998). Odds ratios for breast cancer were not increased in upper quartiles of lipid-adjusted serum total 

DDT, p,p’-DDT, or p,p’-DDE, compared to the lowest quartile of lipid-adjusted serum concentrations. 

The analysis was adjusted to account for the following potential confounders: weight, height, number of 

full-term pregnancies, alcohol consumption, smoking, physical activity, menopausal status, household 

income, marital status, and education. Blood samples were collected in 1976, and breast cancer diagnosis 

occurred up to 17 years following the sample. Exclusion of women who developed breast cancer within 

5 years of serum sampling did not alter the results. Median serum levels of p,p’-DDE in samples taken 

5 years after the first examination were not significantly changed (1,197 vs. 1,169 ppb), but p,p’-DDT 

levels were about 3 times lower in the more recent analysis (144 vs. 46 ppb) (Høyer et al. 2000). When 

the values for p,p’-DDT from the two sampling periods were averaged and the means were stratified in 

quartiles, a positive association between the average serum concentration of p,p’-DDT and breast cancer 

risk was observed (p for the trend was 0.02). 

A study of 150 breast cancer cases and 150 controls from Norway found no association between 

p,p’-DDT or p,p’-DDE in serum levels and breast cancer (Ward et al. 2000). In this study, the interval 

between blood collection and diagnosis ranged from 2 to 18.2 years with a mean of 8.8 years. The mean 
DDT, DDE, and DDD 114 

3. HEALTH EFFECTS 

lipid-adjusted DDE concentrations in cases and controls were 1,230 and 1,260 ppb, respectively; the 

corresponding values for DDT were 120 and 138 ppb. Analyses of the data stratified by age at diagnosis, 

interval between blood collection and diagnosis, and estrogen and progesterone receptor status showed 

that DDE was higher in women cases age 50 years or older at diagnosis and in cancer cases with 

$10 years between blood sample and diagnosis. Neither one of these differences were statistically 

significant. 

The association between breast cancer and DDT exposure was evaluated in 105 breast cancer cases 

relative to 208 matched controls in a case-control study nested within a prospective breast cancer study 

(Dorgan et al. 1999). Breast cancer was diagnosed up to 9.5 years after blood samples were obtained 

(between 1977 and 1987) from a cohort of women who participated in the Columbia, Missouri Breast 

Cancer Serum Bank study. DDT exposure was estimated using lipid-adjusted serum levels of o,p’-DDT, 

p,p’-DDT, o,p’-DDE, p,p’-DDE, p,p’-DDD, and total DDT. The percent of participants with serum DDT 

compound levels above the assay detection limit was not elevated in cases compared to controls. Risk 

ratios calculated by quartile of DDT compound level were evaluated for total DDT, p,p’-DDT, and 

p,p’-DDE, and found not to be elevated. 

A case-control study of only postmenopausal women in western New York State (154 cases; 

192 controls) found no association between risk of breast cancer and current age- and lipid-corrected 

serum p,p’-DDE concentrations, based on an odds ratio analysis (Moysich et al. 1998). Samples were 

collected between 1986 and 1991, within several months after diagnosis. The analysis controlled for a 

variety of breast cancer determinant factors, including age, education, familial breast cancer, parity, 

quetelet index (body mass index expressed as kg/m2

), age at first birth, duration of lactation, years since 

last pregnancy, fruit and vegetable intake, and serum lipids. A recent study by Zheng et al. (2000) 

examined the relationship between serum DDE (lipid-adjusted) and breast cancer risk by menopausal 

status, parity, and lactation, and by cases’ estrogen receptor status among 475 cases and 502 control 

women in Connecticut. Age-adjusted mean serum DDE levels were comparable in cases and controls, 

460 and 456 ppb, respectively. Stratification by disease stage showed that DDE levels in 20 patients with 

later stage disease were insignificantly lower (402 vs. 456 ppb) than in 389 patients with early stage 

disease. The authors also found no significant differences for mean serum levels of DDE between 

controls and the various types of treatment groups, estrogen receptor status, parity and lactation status, 

and menopausal status. 
DDT, DDE, and DDD 115 

3. HEALTH EFFECTS 

Two studies evaluated hospital populations of breast cancer patients in countries where DDT has been 

used in the recent past for malaria control programs, and found no relationship between risk of breast 

cancer and serum concentrations of p,p’-DDT or metabolites. Lopez-Carrillo et al. (1997) investigated 

the association between current serum levels of DDE and p,p’-DDT and the occurrence of breast cancer in 

hospital patients in Mexico City from 1994 to 1996. Neither arithmetic mean nor geometric mean serum 

DDE (either wet weight or lipid weight basis) were significantly different between Mexican women with 

breast cancer (n=141) and age-matched control women with no breast cancer. Arithmetic mean serum 

p,p’-DDT also was not different between the groups, either on a wet weight or lipid weight basis. Control 

women were recruited from various diagnostic areas of the hospitals, excluding oncology and 

gynecology; mean serum DDE was not statistically significantly different across diagnostic areas among 

control women. No significant effect of serum DDE on breast cancer risk was found for all subjects, or 

for subsets of subjects based on menopausal status, using an odds ratio analysis adjusted for age, quetelet 

index, breast feeding with first birth, parity, familial history of breast cancer, and time elapsed since first 

birth. Among northern Vietnamese women, no difference was found in 1994 plasma p,p’-DDE and 

p,p’-DDT concentrations between newly diagnosed breast cancer patients (n=21) and age- and residencematched control patients who had fibrocystic breast disease (Schecter et al. 1997). Plasma levels of DDE 

and DDT were not adjusted for plasma lipid content. Relative risk of breast cancer was not significantly 

elevated for subjects in the higher tertiles of plasma DDE, DDT, and total DDT (DDT plus molaradjusted DDE), compared to the lowest tertile. No differences were observed between controls and cases 

with respect to age, age at menarche, age at first pregnancy, parity, history of lactation, and maximum 

attained body weight. The authors concluded that exposure to p,p’-DDT is not an important factor in the 

etiology of breast cancer among northern Vietnamese women, although they acknowledge that the study 

was limited by small sample size. 

Two studies of women from Brazil also found no association between blood DDT/DDE levels and breast 

cancer incidence. The first study investigated 177 cases of invasive breast cancer and 350 controls; cases 

and controls were residents of a large metropolitan region (Mendonca et al. 1999). The mean and median 

DDE concentrations in serum from patients were 5.1 and 3.1 ppb, respectively; the corresponding values 

in controls were 4.8 and 3.1 ppb. The age-adjusted odds ratio of breast cancer for women in the upper 

quintile compared with those in the lowest quintile was 0.90 (95% CI 0.47–1.73). The second study of 

46 women with breast cancer and 152 controls found that mean total DDT serum levels was 13.6 ppb in 

cases and 15.9 in controls (Matuo et al. 2000). However, cases who never breast fed had significantly 

higher serum DDT than controls who never breast fed. 
DDT, DDE, and DDD 116 

3. HEALTH EFFECTS 

Studies that found a lack of association between DDT/DDE in adipose tissue and breast cancer include 

one by Liljegren et al. (1998) who found that the DDE concentration in breast tissue fat of women with 

malignant breast cancer (n=43) was no different from the DDE concentration in breast adipose tissue of 

women with benign breast disease (n=35), whether or not the study group was divided into pre- and 

postmenopause subgroups; samples were collected during surgical procedures in 1993–1995, after 

diagnosis. Odds ratio analysis of the entire study group, the postmenopausal subgroup, and the 

ER-positive subgroup revealed no association between breast cancer malignancy and tissue DDE 

concentration. Similarly, Unger et al. (1984) found that there was no difference in mean DDE levels in 

biopsied extractable breast fat tissue between newly diagnosed breast cancer patients and noncancer 

patients; the dates of biopsies were not provided. Zheng et al. (1999) also found no significant 

association between breast adipose tissue levels of DDT and DDE and breast cancer. The study included 

304 cases of breast cancer and 186 benign breast disease controls in Connecticut. The age-adjusted 

geometric mean concentrations of DDE in breast adipose in cases and controls were 737 and 784 ppb, 

respectively; the corresponding values for DDT were 52 and 56 ppb. There was no increased risk of 

breast cancer associated with increasing breast adipose tissue levels of either DDT or DDE. An 

additional study compared breast adipose tissue levels of DDT/DDE/DDD between 73 breast cancer 

cases and 73 women undergoing breast reduction surgery in California (Bagga et al. 2000). Analysis of 

unadjusted mean concentrations revealed no significant differences for DDT and DDD between cases and 

controls, but DDE was significantly elevated in cases relative to controls. However, when adjusted for 

age (cases were significantly older than controls), there was no statistically significant relationship 

between either DDT, DDE, or DDD, or the sum of the three chemicals and breast cancer. Another casecontrol study evaluated postmenopausal women from several European countries, showing no significant 

difference in current p,p’-DDE concentration in buttocks adipose tissue between women with breast 

cancer (n=265, 1.35 ppm) and cancer-free women matched for age and study center (n=341, 1.51 ppm) 

(van’t Veer et al. 1997). The dates of tissue sampling were not reported. 

A combined analysis of the data from five studies (Helzlsouer et al. 1999; Laden et al. 2001a; Moysich et 

al. 1998; Wolff et al. 2000a; Zheng et al. 2000) including 1,400 case patients and 1,642 control subjects 

was recently published (Laden et al. 2001a). The authors calculated pooled odds ratios and 95% 

confidence intervals by use of the random-effects model. When women in the fifth quintile of lipid 

adjusted values for DDE were compared with those in the first quintile, the multivariate pooled odds ratio 

for breast cancer was 0.99 (95% CI, 0.77–1.27), suggesting that there was no association between breast 

cancer risk and circulating DDE concentrations. 
DDT, DDE, and DDD 117 

3. HEALTH EFFECTS 

Aschengrau et al. (1998) explored the relationship between occupational exposure to estrogenic chemicals 

and the occurrence of breast cancer. Investigators compared probable exposure to DDT (inferred from 

subjectively reported employment history and a job exposure analysis that related job type to probable 

xenobiotic exposure) in women from Cape Cod, Massachusetts who were diagnosed with breast cancer 

during 1983–1986 (n=261) with age- and race-matched controls (n=753) from Cape Cod. No relationship 

was found; the incidence of women who were occupationally exposed to DDT was regarded as “rare” and 

was not numerically reported in the study. No information was provided regarding the estimated timing 

of DDT exposure versus diagnosis of breast cancer. 

Pancreatic Cancer. Pancreatic cancer was weakly associated with exposure to DDT in a nested casecontrol mortality study following up a cohort of 5,886 chemical manufacturing workers who were 

potentially exposed between 1948 and 1971 (Garabrant et al. 1992). Deaths due to pancreatic cancer 

occurred between 1953 and 1988. The mortality registry had been updated and periodically analyzed. 

An elevated mortality from pancreatic cancer was first seen in the cohort in 1987. The association 

between pancreatic cancer and exposure to 429 chemicals or groups of related chemicals was examined 

for 28 cases of pancreatic cancer and 76 matched controls. Only 16 of the 28 cases were medically 

verified; medical records were not available for 12 subjects whose death certificates indicated pancreatic 

cancer as the underlying cause of death. Only 11 of the cases had been exposed to DDT or related 

materials. The relative risk (RR) for exposure to DDT alone was 4.8 (6 cases, 7 controls; 95% 

CI=1.2–17.6) and for DDD was 4.3 (9 cases, 12 controls; 95% CI=1.5–12.4). The RR for DDT increased 

with greater-than-median exposure duration and more than 20 years of latency. Multivariate analyses to 

correct for confounding factors (e.g., cigarettes, decaffeinated coffee, antacid use) did not change the RR 

markedly. When exposures to other chemicals (e.g., ethylan, nitrofen, dinocap, carbon tetrachloride, 

dispersing agents) found associated with pancreatic cancer (i.e., RR>1) were added to the multivariate 

model the RR for DDT remained relatively stable, ranging from 3.1 to 5.4 (statistical significance was not 

reported). It was concluded that DDT was an independent risk factor for pancreatic cancer. The evidence 

was not strong enough to conclude that DDD was an independent risk factor. The study was limited 

because of the small number of pancreatic cancer cases in DDT-exposed persons and the large number of 

exposures. Although confounding factors and biases have been minimized, the evidence for association is 

weak. 

The association between pancreatic cancer and self-reported exposure to organochlorine pesticides (DDT 

among them) was examined in group of 66 pancreas cancer cases and 131 controls in southeastern 

Michigan (Fryzek et al. 1997). Analysis of the results showed a nonsignificant increased odds ratio of 
DDT, DDE, and DDD 118 

3. HEALTH EFFECTS 

1.6 for pancreatic cancer (95% CI=0.8–3.1; cases=17, controls=23) for patients who were ever exposed to 

DDT, as compared to patients who had not been exposed. 

A more recent study examined serum levels of p,p’-DDE and other organochlorine compounds in 

108 pancreatic cancer cases and 82 controls (Hoppin et al. 2000). The study was part of an ongoing 

population-based case-control study of pancreatic cancer in the San Francisco Bay area. Controls were 

matched to cases by sex and age. After lipid adjustment, cases had higher concentrations of 

organochlorine than controls; the median p,p’-DDE concentration in cases was 1.3 ppm vs. 1.0 ppm in 

controls (p=0.05). When the p,p’-DDE concentration was measured by tertiles, the odds ratio for the 

highest tertile (2.1, 95% CI=0.9–4.7) was elevated compared to the lowest tertile, although it did not 

achieve statistical significance. The odds ratio of 2.1 decreased to 1.1 after controlling for PCB level. As 

part of the same ongoing study, Slebos et al. (2000) investigated the association between two molecular 

markers, K-ras oncogene and p53 tumor suppressor gene, and risk factors for pancreatic cancer in a group 

of 61 newly diagnosed patients. Activating point mutations of K-ras are common oncogene alterations in 

pancreatic carcinoma. Conversely, inactivation of the p53 tumor suppressor gene is common in almost all 

human cancers. The study found that patients with K-ras positive tumors tended to have lower serum 

levels of DDE than those with K-ras negative tumors and also that serum levels of DDE were not 

significantly different between the p53 positive and negative groups. In contrast to the findings of Slebos 

et al. (2000), Porta et al. (2000) found in a study of 51 subjects with pancreatic cancer that serum levels of 

p,p’-DDT and p,p’-DDE were significantly higher than among 26 controls. Among the 51 cases, 34 had 

mutated K-ras tumors and 17 had wild-type K-ras. A comparison of nonlipid adjusted values revealed 

that serum p,p’-DDE levels from wild-type cases did not differ from controls, but levels from mutated 

K-ras cases were about double those in controls; a similar comparison, but adjusted for lipid content, was 

not conducted. 

Lymphoma. Hodgkin’s disease and non-Hodgkin’s lymphoma were investigated in a retrospective study 

of 31 cases of Hodgkin’s disease, 93 cases of non-Hodgkin’s lymphoma, and 204 referents. A crude ratio 

of 2.2 for subjectively reported DDT exposure was calculated for Hodgkin’s disease, but no association 

was found when confounding factors were included into the logistic regression analysis (Persson et al. 

1993). For non-Hodgkin’s lymphoma, the logistic odds ratio was 2.0 (90% CI=0.3–13). Data were 

obtained by questionnaire from Swedish workers diagnosed between 1975 and 1984. The numbers of 

exposed cases and controls were too small to provide evidence of an association between DDT exposure 

and the diseases; DDT-exposed subjects included only one case of Hodgkin’s disease, four cases of 

non-Hodgkin’s lymphoma, and three referents. Exposure to phenoxy herbicides and fresh wood among 
DDT, DDE, and DDD 119 

3. HEALTH EFFECTS 

sawmill workers, lumberjacks, and paper pulp workers were found to be significant risk factors for 

Hodgkin’s disease, whereas welding, working as a lumberjack, nursing, and past-smoking were 

significant increased risks for non-Hodgkin’s lymphoma. 

No significant difference in mean concentration of p,p’-DDE in abdominal adipose tissue was found 

between a group of 28 newly diagnosed patients with non-Hodgkin’s lymphoma and 17 surgical controls 

in Sweden (Hardell et al. 1996). The mean p,p’-DDE concentrations in patients and controls were 

1.4 ppm (lipid basis) and 1.1 ppm, respectively. However, almost all of 34 PCB congeners were 

significantly higher in 27 of the cases relative to the 17 controls. 

Analysis of pooled data from three case-control studies in the United States revealed tenuous evidence of 

an association between DDT exposures and the occurrence of non-Hodgkin’s lymphoma among male 

farmers (Baris et al. 1998). Exposure was reported subjectively, and was categorized into the following 

three groups: DDT use on crops and animals, DDT use on animals only, and DDT use on crops only. 

Odds ratios and confidence intervals were obtained by logistic regression. Using a reference group of 

nonfarmers and pooled data for farmers from four Midwestern states, the age- and state-adjusted odds 

ratio for the occurrence of non-Hodgkin’s lymphoma were 1.5 (95% CI=1.1–2.1) for farmers who used 

DDT on crops (n=74) and 1.6 (95% CI=1.1–2.3) for farmers who personally handled DDT that was 

applied to crops (n=63). The odds ratios were lower and not statistically significantly elevated above 

unity for using or handling DDT applied to animals or applied to animals and crops combined. When 

adjusted for use of other individual pesticides or pesticide groups, when evaluated by type of 

non-Hodgkin’s disease, or when stratified by co-exposure to 2,4-D and organophosphate pesticides, no 

significant odds ratios were observed. No association was observed between estimated duration of DDT 

use and occurrence of non-Hodgkin’s lymphoma, adjusted for use of other pesticides. 

Lipid-corrected total serum DDT (sum of o,p’-DDT, p,p’-DDT, o,p’-DDE, and p,p’-DDE) was not 

associated with risk for non-Hodgkin’s lymphoma in patients diagnosed with the disease between 1975 

and 1989 (n=74) when compared to matched, cancer-free controls (n=147) in a case-control study nested 

within a prospective cohort study of Maryland residents that was begun in 1974 (Rothman et al. 1997). A 

separate evaluation for each DDT compound was not reported. The median lipid-adjusted concentration 

of total DDT in archived blood of patients who subsequently developed non-Hodgkin’s lymphoma 

(3.2 ppm) was not significantly higher than in patients who did not develop the disease (2.8 ppm) 

(Wilcoxon signed rank test; p=0.2). Odds ratios for non-Hodgkin’s lymphoma in serum DDT 
DDT, DDE, and DDD 120 

3. HEALTH EFFECTS 

concentration quartiles, compared to the lowest quartile, were not statistically significant and ranged 

from 1.2 in the second quartile to 1.9 in the highest quartile. 

A population case-control study of the association between subjectively reported DDT use and 

non-Hodgkin’s lymphoma was conducted on farmers in Iowa and Minnesota with newly diagnosed or 

confirmed disease (n=622) and population-based controls (n=1,245) (Cantor et al. 1992). Significantly 

elevated maximum likelihood estimates of odds ratios were obtained for ever having handled DDT 

applied to crops (OR=1.7; CI=1.2–2.6; cases=57 and controls=75) and for having handled DDT applied 

to crops only prior to 1965 (OR=1.8; CI=1.1–2.7; cases=45 and controls=57). Odds ratios were not 

significant for exposure to DDT applied as an animal insecticide. In spite of the fact that the odds ratios 

were adjusted using logistic analysis for vital status, age, state, smoking, family history of lymphopoeitic 

cancer, high-risk occupation, and high-risk exposure, causality could not be established because of 

exposure to multiple pesticides. 

Multiple Myeloma. Several studies evaluated a possible association between DDT exposure and risk of 

multiple myeloma. Eriksson and Karlson (1992) investigated the several environmental factors that are 

possibly related to the occurrence of multiple myeloma in Sweden. Using univariate statistical analysis, a 

significantly increased relative risk of multiple myeloma of 1.75 (90% CI=1.19–2.64) was obtained for 

any DDT exposure, but not for DDT exposures stratified by exposure duration. Relative risk decreased 

and was nonsignificant when multivariate analysis was used to correct for confounding risk factors. For a 

group of 20 Iowa farmers who reported that they mixed, handled, or applied DDT as a crop insecticide, a 

nonsignificant odds ratio for multiple myeloma of 1.7 (95% CI=0.9–3.1) was obtained using a control 

group of 52 nonfarmers (Brown et al. 1993). An odds ratio of 1.8 was obtained for farmers who failed to 

use protective equipment, although statistical significance cannot be determined since a confidence 

interval was not reported. A nonsignificant odds ratio of 1.1 (95% CI=0.6–1.9) for multiple myeloma 

was estimated for 20 farmers who mixed, handled, or applied DDT as an animal insecticide, using a 

control group of 84 unexposed individuals. The combined evaluation of crop DDT exposure plus animal 

DDT exposure was not provided. 

Proportional mortality analysis was conducted on death certificates of a cohort of 590 persons who 

applied DDT or inspected DDT application areas, and 453 unexposed workers, in a malaria eradication 

campaign in Sardinia, Italy conducted during 1946–1950 (Cocco et al. 1997a, 1997b). Based on 

information on annual use of DDT, its concentration in the pesticide mix, and the concentration applied to 

surfaces, the investigators estimated that DDT exposure concentrations ranged from 170 to 600 mg/m3

 in 
DDT, DDE, and DDD 121 

3. HEALTH EFFECTS 

outdoor operations. They also estimated that on average, for a man working 6 hours/day in pesticide 

application, the minimum indoor exposure would have been 254 g/day (exactly how this was estimated is 

unclear). Proportional mortality ratios (PMR) were provided for numerous types of cancer, for both 

DDT-exposed and unexposed groups, and the general Italian male population was used as the reference 

group for estimating expected mortality. For most types of cancer, the PMR was either elevated in both 

exposed and unexposed groups, or was not elevated in either group. For deaths due to multiple myeloma, 

however, the PMR was elevated (statistically significant) in exposed workers (PMRx100=341) [where a 

value of 100 indicates no difference between observed and expected mortality from multiple myeloma, 

and values above 100 indicate that the observed mortality was greater than the expected]; 95% 

CI=110–795; n=5), but not in unexposed workers (PMRx100=94; 95% CI=1–522; n=1). These results 

provided only marginal support for a positive association between occupational DDT exposure and 

multiple myeloma, because of the small numbers of cases. No consistent pattern of association was 

observed when data were categorized by estimated duration of exposure in days. 

In a case-control study of workers in Italy who worked in an agricultural profession at some time, the 

odds ratio for multiple myeloma was significantly elevated above unity for self-reported exposure to 

chlorinated pesticides, including DDT (OR=1.6; 95% CI=1.1–2.4) (Nanni et al. 1998). Very similar, 

albeit marginally significant results were obtained for exposure specifically to DDT (OR=1.6; 95% 

CI=1.0–2.5). The referent group was comprised of age- and gender-matched individuals who had never 

worked in agriculture. A higher, but not statistically significant, odds ratio was obtained for exposure to 

DDT among workers whose primary occupation was agricultural (OR=2.6; 95% CI=0.9–7.8). 

Prostate and Testicular Cancer. An ecologic study evaluated the relationships between p,p’-DDE 

concentration in subcutaneous fat, or p,p’-DDE in tree bark, and mortality from prostate and testicular 

cancers using multivariate statistical techniques (Cocco and Benichou 1998). Adipose DDE was obtained 

from samples collected in the EPA Human Monitoring Program in 1968 for people in 22 states, tree bark 

DDE data were available for 18 states representing the years 1992–1995, and age-adjusted mortality rates 

from prostate and testicular cancers during 1971–1994 were available by state from the National Center 

for Health Statistics. Numerous demographic factors were considered as possible confounders, and data 

were obtained from public sources. Separate analyses were conducted for whites and African Americans 

since mean adipose DDE in African Americans was 74% higher than in whites (p<0.001). The authors 

concluded that study results do not support an association between prostate and testicular cancer mortality 

and DDE exposure. 
DDT, DDE, and DDD 122 

3. HEALTH EFFECTS 

Endometrial Cancer. No association was found between newly-diagnosed endometrial cancer and lipidcorrected blood serum concentrations of p,p’-DDE in a multicenter case-control study of 180 women in 

the United States (Sturgeon et al. 1998). The median blood lipid-adjusted concentrations of p,p’-DDE 

were very similar between cases and controls (about 1.4 ppm). The adjusted relative risk of endometrial 

cancer in the highest quartile of exposure compared with women in the lowest quartile was 0.7 (95% 

CI=0.2–2.0). The blood concentration of o,p’-DDT was slightly higher among controls relative to cancer 

cases, but there was no significant association between increased cancer risk and this isomer. In contrast, 

cancer cases had higher blood levels of p,p’-DDT and a slightly higher risk of endometrial cancer. 

Similar overall results were obtained when the analysis was restricted to women over 50 years of age, as 

the authors sought to evaluate the influence of menopausal status. 

The mean concentration of 'DDT (sum of p,p’-DDT, o,p’-DDT, p,p’-DDE, and p,p’-DDD) was 

significantly higher (p<0.001) in uterine leiomyomatous tissue (0.845 ppm) obtained from 25 recent 

hysterectomies of 36- to 55-year-old women than in normal uterine tissue (0.103 ppm) obtained from 

25 recent autopsies of women from the same age range (Saxena et al. 1987). The mean concentration of 

each of the individual forms was nonsignificantly elevated above control levels in the leiomyomatous 

tissue. Dates that tissue samples were obtained were not reported. 

The occurrence of endometriosis, a benign proliferation of endometrial tissue outside the endometrial 

cavity, was not associated with DDT exposure as measured by lipid-adjusted plasma p,p’-DDT, 

p,p’-DDE, and 'DDT in a study of 86 cases and 70 controls confirmed by laparoscopy in 1994 (Lebel et 

al. 1998). Geometric mean plasma concentrations of DDT compounds were not elevated in all cases 

compared to all controls, nor in subgroups of cases compared to controls matched by symptomatology 

(e.g., pelvic pain, infertility, and tubal fulguration). 

All Cancer. In a prospective cancer mortality study of 919 adults from Charleston, South Carolina 

(Austin et al. 1989), serum levels of total DDT (DDT plus DDT molar equivalents of DDE) were 

estimated in 1974–1975, and 10 years later, the individuals were followed up and the cause of death was 

determined for the deceased. The cohort was divided into exposure tertiles based on total serum DDT; 

the lowest tertile was used as the reference group for calculating relative mortality rates. Adjustments 

were made for age, race, gender, years of education, and smoking habits. Relative risk of death, and 

specifically of death due to any cancer, was not significantly elevated in the high serum DDT tertile 

groups. No consistent positive trend in risk of cancer mortality relative to serum DDT was observed. 
DDT, DDE, and DDD 123 

3. HEALTH EFFECTS 

However, there was some evidence of a dose-response relation between serum DDT and respiratory 

cancer, but the point estimates were unstable and the trend was not statistically significant. 

A historical prospective mortality study was conducted on 740 white male workers employed between 

1935–1976 in occupations that involved exposures to DDT (Wong et al. 1984). Follow-up was 

conducted in 1976, and expected age- and cause-specific mortality rates were calculated from U.S. rates 

for white males for 5-year periods during 1935–1975. Among individuals exposed to DDT, overall 

mortality (all cancers), expressed as the SMR, was not elevated over expected values. Standardized 

mortality ratios were also not significantly elevated for individual cancers of the digestive, respiratory, 

urogenital, and lymphohaematopoietic systems. Several factors confound these results: individuals 

exposed to DDT also were potentially exposed to other chemicals, and smoking history was not included 

in the analysis. 

Mortality and health impairments in a cohort of 2,620 pesticide workers and 1,049 controls were tracked 

during 1971–1977 (Morgan et al. 1980), and cancer rates were found to be significantly elevated in at 

least one category of exposed workers for internal cancers, leukemias, and lymphomas (evaluated 

together), as well as for skin cancers and all cancers. However, geometric mean total serum DDT (DDT 

plus DDE) measured in 1971–1973 was not significantly different between subjects who developed 

cancer of leukemia by the end of 1977 (n=43) and a group of subjects without the disease (n=45). 

Exposure to multiple pesticides was a significant confounding factor in this study. 

A recent study examined the association of the 1968 adipose DDE levels of population samples from 

22 U.S. states with age-adjusted mortality rates between 1975 and 1994 for multiple myeloma, 

non-Hodgkin’s lymphoma, and breast, uterus, liver, and pancreas cancer (Cocco et al. 2000). Past body 

burdens were estimated by using half-lives for DDT and DDE of 7 years and several decades, 

respectively. Separate analyses were conducted for gender and race. The authors found no association 

for pancreatic cancer and multiple myeloma. Breast cancer mortality was inversely correlated with 

adipose DDE levels among both white and African American women. Significant inverse correlations 

were also observed for uterine cancer among white women, while no association was observed for 

African Americans and for non-Hodgkin’s lymphoma among white (men and women) and African 

American women. Liver cancer mortality significantly increased with adipose DDE levels in white males 

and females, but not among African Americans. A previous study of Italian workers exposed to DDT in a 

malaria eradication campaign in Sardinia, Italy also found a significant increase in the PMR for liver and 
 

DDT, DDE, and DDD 124 

3. HEALTH EFFECTS 

biliary tract cancer; however, liver cancer was also elevated among unexposed workers and it did not 

show a trend with duration of exposure (Cocco et al. 1997a). 

Overall, in spite of some positive associations for some cancers within certain subgroups of people, there 

is no clear evidence that exposure to DDT/DDE causes cancer in humans. 

Animal Studies 

DDT is one of the most widely studied pesticides in animals, and data are available on a number of 

carcinogenicity studies in several species. Intermediate exposures, in which animals were exposed to 

DDT in food, caused cancer increases in mice but not in rats or hamsters. Mice that were observed for 

50–105 weeks after cessation of treatment developed liver hepatomas following dietary exposure to 

42.8 mg p,p’-DDT/kg/day for 15–30 weeks (Tomatis et al. 1974b). DDT did not produce increases in the 

tumor incidence in rats exposed to 10–20 mg/kg/day in the food for up to 45 weeks (Kimbrough et al. 

1964; Laug et al. 1950; Numoto et al. 1985) or in hamsters fed 50 mg DDT/kg/day for 30 weeks (Tanaka 

et al. 1987). 

Chronic exposure (>1 year) to DDT caused cancer in multiple strains of mice but not in dogs; most 

studies in nonhuman primates have also been negative. Chronic exposure to DDT produced liver 

neoplasms in mice strains ([C57BL/6 x C3H/Anf]F1, [C57BL/6 x AKR]F1, BALB/c, and CF1) fed DDT 

at doses as low as 0.38 mg DDT/kg/day for a minimum of 78 weeks (Innes et al. 1969; Kashyap et al. 

1977; Terracini et al. 1973; Thorpe and Walker 1973; Tomatis et al. 1972, 1974a; Turusov et al. 1973). 

An increased incidence of pulmonary adenomas was observed in mice after chronic gavage administration 

(Shabad et al. 1973). Malignant lymphomas and lung and liver tumors were also observed in mice treated 

with DDT in the food (Kashyap et al. 1977). 

Some evidence exists to indicate that DDT may be carcinogenic in the rat. Rats maintained on diets 

containing DDT for more than 2 years or at doses higher than 25 mg DDT/kg/day developed liver tumors, 

primarily in female rats (Cabral et al. 1982b; Fitzhugh and Nelson 1947; Rossi et al. 1977). Liver tumors 

occurred in rats at doses of 19.7 mg DDT/kg/day for 2 years (Cabral 1982b). In contrast, no evidence of 

carcinogenicity was seen in rats receiving up to 45 mg technical DDT/kg/day for 78 weeks in the NCI 

(1978) bioassay. No significant increases in tumor incidence were observed in mice administered DDT at 

doses of 3–30 mg/kg/day (Del Pup et al. 1978; NCI 1978). Long-term exposure failed to induce 

significant increases in tumors in monkeys at doses of 3.9–20 mg/kg/day for up to 5 years (Adamson and 
DDT, DDE, and DDD 125 

3. HEALTH EFFECTS 

Sieber 1979, 1983; Durham et al. 1963); or in dogs at 80 mg/kg/day for 49 months (Lehman 1965). A 

more recent study that involved 11 Rhesus and 13 Cynomolgus monkeys administered approximately 

6.4–15.5 p,p’-DDT/kg/day in the diet for up to 130 months reported that 2 out of 13 Cynomolgus 

monkeys (15%) developed malignant tumors, one hepatocellular carcinoma and one adenocarcinoma of 

the prostate (Takayama et al. 1999). No neoplasms were found in a group of nine Cynomolgus and eight 

Rhesus untreated control monkeys. 

Evidence of carcinogenicity of DDT in hamsters is equivocal. Rossi et al. (1983) reported an increased 

incidence (14% in controls, 34% in treated hamsters) of adrenal neoplasms in hamsters administered 

approximately 95 mg DDT/kg/day via the diet for 30 months. At lower doses, Cabral et al. (1982a) did 

not observe a statistically significant increase in adrenal gland tumors; however, the incidence in males 

was increased compared to controls in animals receiving 71 mg DDT/kg/day via the diet for 28 months. 

Other studies in hamsters did not indicate any carcinogenic effects of DDT; however, early deaths 

occurred in one study (Agthe et al. 1970) and the duration of exposure was shorter in another (Graillot et 

al. 1975). 

Several multigeneration studies have been conducted in mice. In these studies, exposure of the F1 and 

subsequent generations to DDT was initially perinatal (i.e., in utero and through lactation) and was 

followed postweaning by oral exposure to DDT in the diet. In a study by Tarjan and Kemeny (1969), 

exposure to 0.4 mg p,p’-DDT/kg/day resulted in significant increases in leukemia and pulmonary 

carcinomas in the F2 generation and occurred with increasing frequency with each subsequent generation 

of mice. Liver tumors (0.3–0.4 mg/kg/day) (Tomatis et al. 1972; Turusov et al. 1973) and pulmonary 

tumors (1.3 mg/kg/day) (Shabad et al. 1973) in the F1 generation had a shorter latency period than in the 

parental generation, but the tumor incidence was comparable and did not increase with consecutive 

generations. 

There are several studies of the potential carcinogenicity of DDE and DDD in rats, mice, and hamsters. 

DDE administered chronically in the diet produced liver tumors in male and female mice at doses of 

27–43 mg/kg/day for 30–78 weeks (NCI 1978; Tomatis et al. 1974a) and in hamsters dosed with 

approximately 48 mg p,p’-DDE/kg/day for 128 weeks (Rossi et al. 1983). DDE did not induce 

significant increases in tumor incidence in rats exposed to DDE in the diet at doses ranging from 12 to 

42 mg/kg/day for 78 weeks (NCI 1978), but doses of approximately 43 mg/kg/day for 130 weeks 

significantly increased the incidence of liver tumors in mice (Tomatis et al. 1974a). DDD induced liver 

tumors and lung adenomas in CF-1 mice at doses of approximately 43 mg/kg/day (Tomatis et al. 1974a), 
DDT, DDE, and DDD 126 

3. HEALTH EFFECTS 

but it was not tumorigenic in B6C3F1 mice in a 78-week study at doses of approximately 142 mg/kg/day 

(NCI 1978). In the NCI (1978) bioassay, the combined incidences of thyroid follicular cell adenoma and 

follicular cell carcinomas were 1/19, 16/49, and 11/49 in controls, low-dose (116 mg/kg/day), and highdose (231 mg/kg/day) male rats, respectively. The difference between the control and low-dose group 

was significant according to the Fisher Exact test. However, NCI (1978) pointed out that the variation of 

these tumors in control male rats in the study did not permit a more conclusive interpretation of the lesion. 

EPA has estimated an oral cancer potency factor (q1

*

) for DDT of 3.4x10-1 (mg/kg/day)-1, which was 

derived using the linearized multistage model (IRIS 2001a). This potency factor is derived from the 

geometric mean of potency factors based on the incidence of liver tumors in mice and rats as reported by 

Cabral et al. (1982b), Rossi et al. (1977), Terracini et al. (1973), Thorpe and Walker (1973), Tomatis and 

Turusov (1975), and Turusov et al. (1973). At this potency, 3.4x10-1 (mg/kg/day)-1, the lifetime average 

daily doses that would result in risks of 1x10-4, 1x10-5, 1x10-6, and 1x10-7 are 2.9x10-4, 2.9x10-5, 2.9x10-6, 

and 2.9x10-7 mg/kg/day, respectively. 

The oral cancer potency factor (q1*) for DDD is 2.4x10-1 (mg/kg/day)-1 based on the incidence of liver 

tumors in CF-1 mice (Tomatis et al. 1974a) and the oral q1* for DDE is 3.4x10-1 based on the geometric 

mean of six slope factors for liver tumors in both sexes of B6C3F1 mice (NCI 1978), CF-1 mice (Tomatis 

et al. 1974a), and Syrian hamsters (Rossi et al. 1983). At these potencies, the lifetime average doses that 

would result in risk of 1x10-4, 1x10-5, 1x10-6, and 1x10-7 are 4.2x10-4, 4.2x10-5, 4.2x10-6, and 

4.2x10-7 mg/kg/day, respectively, for DDD and 2.9x10-4, 2.9x10-5, 2.9x10-6, and 2.9x10-7 mg/kg/day, 

respectively, for DDE. 

Cancer Effect Levels (CELS) are recorded in Table 3-1 and plotted in Figure 3-1. 

3.2.3 Dermal Exposure 

Occupational exposure to DDT involved multiple routes of exposure. The primary contact was probably 

inhalation and dermal; however, absorption of DDT from the lungs may not have been significant, and 

ingestion via the mucociliary apparatus of the respiratory tract is more likely. Therefore, epidemiological 

studies of occupational exposure are discussed under oral exposure. 
 

DDT, DDE, and DDD 127 

3. HEALTH EFFECTS 

3.2.3.1 Death 

The dermal LD50 of DDT in rats was reported by Ben-Dyke et al. (1970), Cameron and Burgess (1945), 

and Gaines (1969) to range from 2,500 to 3,000 mg DDT/kg. In guinea pigs, a single dose of 1,000 mg 

DDT/kg resulted in death of 50% of the animals (Cameron and Burgess 1945). The LD50 in rabbits was 

300 mg DDT/kg (Cameron and Burgess 1945) and 4,000–5,000 mg DDD/kg (Ben-Dyke et al. 1970). In 

the study by Cameron and Burgess (1945), the animals were dermally exposed to various doses of DDT 

in solvents including kerosene (1 or 10%), ethyl alcohol, acetone, or ether. It is uncertain what 

contribution these solvents made to the toxic effects observed; the authors stated that kerosene itself may 

have caused some deaths. 

3.2.3.2 Systemic Effects 

No studies were located regarding gastrointestinal, musculoskeletal, or ocular effects in humans or 

animals after dermal exposure to DDT, DDE, or DDD. 

All of the information on the systemic and neurological effects of dermal exposure to DDT in animals is 

derived from a study by Cameron and Burgess (1945). In this study, rabbits, guinea pigs, and rats were 

dermally exposed to various doses of DDT in solvents including kerosene (1 or 10%), ethyl alcohol, 

acetone, or ether. It is uncertain what contribution these solvents made to the toxic effects observed; the 

authors stated that kerosene itself may have caused some deaths. The only information reported on the 

method of application stated that the skin area was shaved 24 hours before application of DDT 

impregnated in cloth and that precautions were taken to prevent animals from licking contaminated skin. 

The duration of exposure was not clearly reported. In addition, the species and the number of animals 

exhibiting specific toxic symptoms were not clearly reported and no statistical analysis was conducted. 

Respiratory Effects. No studies were located regarding respiratory effects in humans after dermal 

exposure to DDT, DDE, or DDD. In rats, guinea pigs, and rabbits exposed to acute dermal doses ranging 

from 50 to 200 mg DDT/kg, pulmonary edema and respiratory failure were reported (Cameron and 

Burgess 1945). 
 

 

 

 

 

DDT, DDE, and DDD 128 

3. HEALTH EFFECTS 

Cardiovascular Effects. No studies were located regarding cardiovascular effects in humans after 

dermal exposure to DDT, DDE, or DDD. Cameron and Burgess (1945) exposed rats, guinea pigs, and 

rabbits to acute dermal doses ranging from 50 to 200 mg DDT/kg and reported fat in the fibers of the 

heart. 

Hematological Effects. No studies were located regarding hematological effects in humans after 

dermal exposure to DDT, DDE, or DDD. Cameron and Burgess (1945) exposed rats, guinea pigs, and 

rabbits to acute dermal doses ranging from 50 to 200 mg DDT/kg. A decrease in hemoglobin and 

leukocytosis was reported. 

Hepatic Effects. No studies were located regarding hepatic effects in humans after dermal exposure 

to DDT, DDE, or DDD. Cameron and Burgess (1945) exposed rats, rabbits, and guinea pigs to acute 

dermal doses of 10, 50, or 100 mg DDT/kg and reported fatty degeneration, calcification, and necrosis in 

the liver. 

Renal Effects. No studies were located regarding renal effects in humans after dermal exposure to 

DDT, DDE, or DDD. Cameron and Burgess (1945) exposed rats, rabbits, and guinea pigs to acute dermal 

doses ranging from 50 to 100 mg DDT/kg and reported fat deposits, tubular changes, calcification, and 

necrosis of the kidneys. 

Dermal Effects. Cameron (1945) conducted a series of experiments on volunteers wearing 

undergarments impregnated with 1% DDT in order to determine whether this treatment would protect 

soldiers against body lice. Several individuals had transient dermatitis, but no other symptoms were 

observed. The length of exposure via this route was not specified. Cameron and Burgess (1945) exposed 

rats, rabbits, and guinea pigs to 10, 50, or 100 mg DDT/kg and reported inflammation, edema, and 

destruction of the epidermis. Guinea pigs were dosed 5 days a week for 3 weeks with 322–400 mg 

DDT/kg (Kar and Dikshith 1970). A decrease in skin amino acids, disruption and degeneration of the 

basal cell layer, and morphologic changes in the cells were reported. 

3.2.3.3 Immunological and Lymphoreticular Effects 

No studies were located regarding immunological effects in humans or animals after dermal exposure to 

DDT, DDE, or DDD. 
DDT, DDE, and DDD 129 

3. HEALTH EFFECTS 

3.2.3.4 Neurological Effects 

No studies were located regarding neurological effects in humans after dermal exposure to DDT, DDE, or 

DDD. Cameron and Burgess (1945) exposed rats, rabbits, and guinea pigs to acute, dermal doses ranging 

from 50 to 200 mg/kg DDT and reported tremors and nervousness. 

No studies were located regarding the following effects in humans or animals after dermal exposure to 

DDT, DDE, or DDD: 

3.2.3.5 Reproductive Effects 

3.2.3.6 Developmental Effects 

3.2.3.7 Cancer 

No case studies or epidemiological investigations concerning the carcinogenic effects in humans after 

dermal exposure exclusively to DDT, DDE, or DDD were located. Occupational studies that probably 

involved both dermal and inhalation routes of exposure are discussed in Section 3.2.2.8. 

Dermal exposure (skin painting) of mice to DDT did not result in a significant increase in tumor 

incidence when applied in a 5% solution in kerosene once weekly for 52 weeks (Bennison and Mostofi 

1950) or at 8 mg/kg twice weekly for 80 weeks (Kashyap et al. 1977). No information on dermal 

exposure of rats or hamsters to DDT or dermal exposure to DDE or DDD was located. 

3.3 GENOTOXICITY 

DDT, DDE, and DDD have been tested in several genotoxicity studies in animals and in bacterial 

systems, but studies in humans are limited. Tables 3-3 and 3-4 report the results of in vivo and in vitro 

studies, respectively. Chromosomal aberrations have been reported in human studies, but simultaneous 

exposure to other chemicals and/or lack of control of potential covariates (such as smoking) make the 

findings inconclusive. For example, blood cultures of men occupationally exposed to several pesticides, 

including DDT, exhibited an increase in chromatid lesions (Yoder et al. 1973). Rabello et al. (1975) 

reported chromosomal aberrations in workers occupationally exposed to DDT. When all workers were 

considered, regardless of direct or indirect exposure, a significant increase in the incidence of 
Table 3-3. Genotoxicity of DDT, DDE, and DDD In Vivo 

Species (test system) End point Results Reference Mammalian cells: 

Human (plasma) Chromosomal aberrations + Rabello et al. 1975 

Human (plasma) Chromosomal aberrations + Yoder et al. 1973 

Human (lymphocytes) Chromosomal aberrations + Rupa et al. 1991 

Human (lymphocytes) Micronuclei – Vine et al. 2001 

Mouse Chromosomal aberrations + Clark 1974 

Rat Chromosomal aberrations – Legator et al. 1973; 

Rabbit (fetus’ liver) Chromosomal aberrations – Hart et al. 1972 

Mouse (bone marrow) Chromosomal aberrations + Johnson and Jalal 1

1974 

Mouse Dominant lethal + Clark 1974 

Rat Dominant lethal (+) Palmer et al. 1973 

Mouse (inhibition of testicular synthesis) DNA synthesis – (DDE) Seiler 1977 Rat (liver) DNA lesions + Hilpert et al. 1983 

Host-mediated assays: 

Serratia marcescens Gene mutation – (DDT, DDE) Buselmaier et al. 197

(Mouse hosted-mediated) + (DDD) Neurospora crassa Gene mutation – Clark 1974 

Invertebrate systems: 

Drosophila melanogaster Dominant lethal + Clark 1974 

– = negative result; + = positive result; (+) = weakly positive result; DNA = deoxyribonucleic acid 
Table 3-4. Genotoxicity of DDT, DDE, and DDD In Vitro 

Results 

With Without 

Species (test system) End point activation activation Re

Prokaryotic organisms: 

Salmonella typhimurium (TA1535, TA1537, TA98, TA100) Escherichia coli (Pol-A) E. coli (Back mutation)Escherichia marcescens (glucose prototrophy)Bacillus subtilis (rec-assay)

E. coli (col E1 plasmid DNA) 

E. coli (DNA cell binding assay) 

Fungal and plant cells: 

Neurospora crassaSaccharomyces cerevisiae

Mammalian cells: 

Chinese hamster V79 cells Chinese hamster (B14F28 cells [chromosomal damage])Kangaroo rat (cells)Human (hepatocyte-mediated cell)Chinese hamster (V79 cells [6-thioguanine resistent mutation])Mouse (L51784 lymphoma cells)Rat (liver epithelial cell)Human SV-40 (unscheduled DNA synthesis)Mouse (hepatocytes-UDS)Rat (hepatocytes-UDS)Hamster (hepatocytes-UDS)

Gene mutation Gene mutation Gene mutation Gene mutation DNA damage DNA damage DNA damageRecessive lethal Mitotic gene conversion Chromosomal aberrations Chromosomal aberrations Chromosomal aberrationsGene mutation Gene mutation Gene mutation Gene mutationDNA damage DNA damage DNA damage DNA damage

 – – – – – – –– – + (DDE) – (DDT) + + – – + –– – No data No data 

– – No data No data No data No data No data No data No data No data No data No data No data – No data No data No data – – – – 

McFluFahFahShiGriKubClaFahKel197MaPalTonTsuAmTelAhmProProMa198

– = negative result; + = positive result; (+) = weakly positive; DNA = deoxyribonucleic acid 
DDT, DDE, and DDD 132 

3. HEALTH EFFECTS 

chromosomal damage was reported; no information was provided regarding control for relevant 

covariates. In pesticide sprayers who were exposed to DDT as well as seven other pesticides, increased 

frequencies of sister chromatid exchanges and chromosomal aberrations in peripheral lymphocytes were 

reported, compared to controls (Rupa et al. 1988). Increases in sister chromatid exchanges, the 

proliferation rate index, and the mitotic index were also reported in pesticide sprayers exposed to DDT 

along with several other pesticides (Rupa et al. 1989). A recent study of 302 individuals residing near a 

waste site in North Carolina found that plasma DDE levels were not associated with a higher frequency of 

micronuclei (Vine et al. 2001). 

Studies in animals have provided mixed results (Table 3-3). In a dominant lethal assay, treatment of male 

rats with a single dose of 100 mg p,p’-DDT/kg resulted in a statistically significant increase in the 

proportion of females with one or more dead implantations only in animals mated during the postmeiotic 

stage of spermatogenesis (Palmer et al. 1973). In a dominant lethal assay in mice, DDT was administered 

orally to male mice at 150 mg/kg/day for 2 days (acute), or 100 mg DDT/kg twice weekly for 10 weeks 

(intermediate); the final dose was given 24 hours before sequential mating began (Clark 1974). 

Significant increases occurred in the number of dead implants per female. Acute doses resulted in 

maximum sensitivity in the induction of dominant lethal effects in week 5 and chronic doses in week 2, 

with continued increases above control through week 6. Repeated dosing caused significant reductions in 

testes weight, sperm viability, and a reduction of cell numbers in all stages of spermatogenesis. With 

acute treatment, the meiotic stage of spermatogenesis appeared to be the most sensitive. Acute treatment 

produced a significantly increased frequency of chromosome breakage, univalents, and stickiness in 

spermatocytes. BALB/C mice exposed in vivo to DDT exhibited chromosomal aberrations of the bone 

marrow (Johnson and Jalal 1973; Larsen and Jalal 1974). 

Rats treated orally (by gavage) with DDT in single doses of 50–100 mg/kg or daily doses of 

20–80 mg/kg/day for 5 days did not show a dose-related increase in percent of chromosomal aberrations 

over the solvent control (Legator et al. 1973). DDE, when administered in a single oral dose to male mice 

at the rate of 50 mg/kg, did not inhibit testicular DNA synthesis (Seiler 1977). 

Administration of up to 50 mg p,p’-DDT/kg by gavage to rabbits on gestation days 7–9 did not affect 

chromosomal number distribution or the percentage of aberrations compared with controls (Hart et al. 

1972). In addition, the distribution of chromosomes in liver samples from fetuses of DDT-treated rabbits 

and the percentage of chromosomal aberrations in these fetuses did not differ from controls (Hart et al. 

1972). 
DDT, DDE, and DDD 133 

3. HEALTH EFFECTS 

As shown in Table 3-4, DDT and related compounds were, for the most part, nonmutagenic in 

prokaryotic organisms and did not induce DNA damage under the conditions tested. However, a few 

studies with mammalian cells in vitro found positive results. For example, Mahr and Miltenburger (1976) 

reported chromosomal damage in the B14F28 Chinese hamster cell line after exposure to DDT, DDE, or 

DDD. Palmer et al. (1972) also observed these same results in kangaroo rat cells (Potorus tridactylis) in 

vitro after exposure to DDT, DDE, or DDD. Kelly-Garvert and Legator (1973) reported a significant 

increase in chromosomal aberrations in Chinese hamster V79 cells after exposure to DDE, but not DDT. 

Collectively, the data do not suggest that DDT and related compounds present a genotoxic hazard at 

environmentally relevant concentrations. 

3.4 HEALTH EFFECTS IN WILDLIFE POTENTIALLY RELEVANT TO HUMAN HEALTH 

The 1972 EPA decision to ban DDT for most uses in the United States was significantly influenced by a 

large body of scientific information documenting adverse effects to wildlife (EPA 1975). These observed 

effects were severe, including the lethality of DDT to birds and fish and the DDE-induced reproductive 

effects in birds, particularly eggshell thinning (EPA 1975). Although it is difficult to draw firm 

conclusions about adverse effects to human health based on those observed in wildlife, it is impossible to 

ignore that the documented effects to wildlife have motivated the investigation of human health effects. It 

is reasonable to assume that the adverse effects observed in wildlife may also be a concern to humans and 

that wildlife are possible "sentinels" for human health (NRC 1991). In order to completely address 

potential concerns for human health, it is necessary to review and evaluate the observed effects of 

DDT/DDE/DDD to terrestrial wildlife. Exposures for wildlife to DDT and its metabolites in the natural 

environment are primarily associated with the accumulation and persistence of these contaminants in both 

aquatic and terrestrial food chains. Ingestion of contaminated food results in the deposition of 

DDT/DDE/DDD in tissues with subsequent reproductive, developmental, and neurological effects. The 

most important reproductive effect observed in wildlife concerns eggshell thinning in birds. These and 

other effects on terrestrial wildlife are discussed in greater detail in Appendix D with the purpose of 

providing a qualitative synopsis of effects in terrestrial wildlife to address potential concerns that these 

effects from DDT/DDE/DDD exposure may also occur in humans. 
 

DDT, DDE, and DDD 134 

3. HEALTH EFFECTS 

Eggshell Thinning in Birds. Eggshell thinning in birds reached widespread public awareness in the 

1960s and 1970s largely because of field observations in wild raptor populations including the bald eagle, 

peregrine falcon, and osprey, and the association of these effects with abrupt population declines. 

Experimental studies established a scientific link between DDT/DDE/DDD exposure, particularly DDE, 

and avian eggshell thinning, which weighed significantly in the decision to ban most domestic crop uses 

of DDT in the 1970s (EPA 1975). In general, raptors, waterfowl, passerines, and nonpasserine ground 

birds were more susceptible to eggshell thinning than domestic fowl and other gallinaceous birds, and 

DDE appears to have been a more potent inducer of eggshell thinning than DDT (Cooke 1973b; EPA 

1975; Lundholm 1997; WHO 1989). Further, reproductive disturbances associated with DDT/DDE/DDD 

exposure continue to be reported in North American populations of predatory birds and/or birds that 

migrate to regions such as South America where DDT is still used (Lundholm 1997). 

Numerous experimental studies have shown that dietary exposures to DDT/DDE/DDD are associated 

with eggshell thinning and breakage in wild birds including the barn owl (Tyto alba) (Mendenhall et al. 

1983), the American kestrel (Porter and Wiemeyer 1969), the mallard duck (Anas platyrhynchus) (Heath 

et al. 1969; Risebrough and Anderson 1975; Vangilder and Peterle 1980), the black duck (Anas rubripes) 

(Longcore et al. 1971), the Japanese quail (Coturnix coturnix japonica) (Kenney et al. 1972), the 

bobwhite quail (Colinus virginianus) (Wilson et al. 1973) and the the Ringed turtle doves (Streptopelia 

risoria) (Haegele and Hudson 1973; Peakall 1970; Peakall et al. 1975). These experimental results have 

verified that the field observations of eggshell thinning and reductions in wild raptor populations are 

associated with releases of DDT. Possible mechanisms of eggshell thinning in birds have been 

extensively studied and reviewed (Cooke 1973b; EPA 1975; Lundholm 1997; Peakall et al. 1975; WHO 

1989). The leading hypothesis for DDE-induced thinning involves an inhibition by p,p’-DDE (but not by 

o,p’-DDE or DDT or DDD isomers) of prostaglandin synthesis in the shell gland mucosa (Lundholm 

1997). Overall, there is still some question as to the primary mechanism and reviewers have suggested 

that these may differ between bird species or differ with environmental conditions or physiological state 

for a given species. There is some question, however, as to the relevance of avian eggshell thinning to 

human health. There is no anatomical or physiological counterpart of the shell gland, a specialized 

segment of the oviduct, in humans. The shell gland lays down calcite (CaCO3, calcium carbonate) onto 

the developing avian egg to form the eggshell (EPA 1975). Mechanisms of action that involve a direct 

action of DDT/DDE/DDD on the shell gland itself probably have no human relevance, but mechanisms of 

action that involve intermediate effects, such as reduced blood calcium, may have relevance to human 

health. 
 

DDT, DDE, and DDD 135 

3. HEALTH EFFECTS 

Reproductive Effects. Exposure to DDT/DDD/DDE is associated with reproductive toxicity in 

avian wildlife including embryolethality (Heath et al. 1969; Longcore et al. 1971; Porter and Wiemeyer 

1969), decreased egg size and weight (Jefferies 1969; Peakall 1970; Wilson et al. 1973), delayed 

oviposition after mating (Cecil et al. 1971; Jefferies 1967, 1969; Peakall 1970; Richie and Peterle 1979; 

Vangilder and Peterle 1980), ovarian effects (Bitman et al. 1968; Gish and Chura 1970; Keith and 

Mitchell 1993) and testicular effects (Burlington and Lindeman 1950; George and Sunararaj 1995; Gish 

and Chura 1970; Locke et al. 1966). Several authors have speculated that these effects are associated 

with DDT-induced hormonal imbalances (Jefferies 1967) such as DDT induced estrogen-like inhibition of 

FSH and LH secretion by the pituitary inhibiting ovary development. 

In most studies, egg production is not affected by DDT/DDD/DDE exposure (Azevedo et al. 1965; Chen 

et al. 1994; Davison et al. 1976; Davison and Sell 1972; Heath et al. 1969; Jefferies 1969; Longcore et al. 

1971; Mendenhall et al. 1983; Porter and Wiemeyer 1969; Risebrough and Anderson 1975; Scott et al. 

1975; Shellenberger 1978; Vangilder and Peterle 1980; Wilson et al. 1973). There are, however, a few 

reported cases of decreased egg production especially in birds with restricted diets (Cecil et al. 1971; Gish 

and Chura 1970; Haegele and Hudson 1973; Kenney et al. 1972). Egg fertility and hatchability are not 

consistently affected by DDT/DDD/DDE exposure. Some studies report significantly decreased fertility 

and hatchability (Porter and Wiemeyer 1969; Vangilder and Peterle 1980; Wilson et al. 1973), while 

others do not document significant effects (Azevedo et al. 1965; Haegele and Hudson 1973; Jones and 

Summers 1968; Scott et al. 1975; Shellenberger 1978). When considered collectively, egg production, 

fertility, and hatchibility were largely unaffected in numerous studies in a variety of bird species. This 

may be inconsequential to the overall reproductive success of birds since DDT/DDD/DDE exposure is 

clearly associated with decreased embryonic survival or fledgling success (Keith and Mitchell 1993). 

DDT exposure has been shown to be associated with reduced post-hatch survival in avian wildlife. This 

effect has been observed in laboratory testing with mallards, pheasant, black duck, chicks; Japanese quail 

and ringed turtle doves (Azevedo et al.1965; Genelly and Rudd 1956; Haegele and Hudson 1973; Heath 

et al. 1969; Jones and Summers 1968; Keith and Mitchell 1993; Longcore et al. 1971; Porter and 

Wiemeyer 1969; Shellenberger 1978). The mechanism of DDT-induced reduced survival after oral 

exposures to DDT or DDE in maternal birds is hypothesized to be associated with increased body burdens 

of DDT/DDD/DDE in chicks as either a result of direct toxicity to the chick, or a reduction in parental 

care-giving among treated birds resulting in chick malnutrition and poor survival. 
 

DDT, DDE, and DDD 136 

3. HEALTH EFFECTS 

The implications of these observed effects in wildlife to human health is uncertain as the mechanisms of 

toxicity are not thoroughly understood. The consistency of the observed reproductive effects to avian 

wildlife and the field observations of effects to birds and reptiles have stimulated the investigation of 

reproductive effects in mammalian models that are more directly relevant to humans. In vitro mechanism 

of action studies have resulted in the identification of some DDT isomers and metabolites as androgen 

antagonists and estrogen agonists. There have been a number of intriguing mechanistic studies of DDT 

isomers and metabolites in fish that relate to reproductive and developmental effects (Das and Thomas 

1999; Faulk et al. 1999; Khan and Thomas 1998; Loomis and Thomas 1999; Sperry and Thomas 1999; 

Thomas 1999). There are some interesting parallels between mammalian wildlife and human health 

studies. Similar to the associations made between DDT and preterm deliveries in humans (Saxena et al. 

1980, 1981; Wassermann et al. 1982), premature births in California sea lions (Zalophus californianus 

californianus) are associated with elevated DDT concentrations (DeLong et al. 1973). However, the 

effect could not be solely, causally isolated to DDT due to the presence of PCBs. 

Developmental Effects. The developmental effects of DDT/DDD/DDE on reptiles and avian 

wildlife have received considerable attention. Studies of alligator populations at Lake Apopka in Florida, 

where a pesticide spill occurred in 1980, have reported various reproductive effects including reduced 

clutch viability (Woodward et al. 1993), altered steroidogenesis (Crain et al. 1997; Guillette et al. 1995), 

abnormal ovarian morphology and plasma 17β-estradiol levels (Guillette et al. 1994), and reductions of 

phallus size and serum testosterone (Guillette et al. 1994, 1995, 1996, 1997, 1999). The authors 

hypothesized that the estrogenicity of DDT and other contaminants induced hormonal imbalance in the 

alligators, causing the observed effects (Guillette and Crain 1996). The contribution of DDT/DDE/DDD 

(only one component of the mixture of pesticides present) to the observed effects is uncertain. However, 

other experimental findings support the hypothesis that certain DDT-related compounds induce estrogenic 

effects in reptiles which could potentially adversely affect reproduction in a population (in ovo DDE 

exposures in alligators by Matter et al. 1998). In general, reptiles may be particularly susceptible to the 

endocrine-altering effects of DDT/DDE/DDD, as sex in many species are determined by environmental 

factors (temperature, etc.,) compared to the genetic sex determining mechanisms in birds and mammals 

(Crain and Guillette 1998). Organochlorine contaminants in general and p,p’-DDE, specifically, are 

thought to influence sexual dimorphism in the common snapping turtle (Chelydra serpentina)(de Solla et 

al. 1998). Snapping turtles in Ontario, Canada, lacked the normal sexual dimorphism in the distance 

between cloaca and plastron which was attributed to the antiandrogenic effects of p,p’-DDE. 
 

DDT, DDE, and DDD 137 

3. HEALTH EFFECTS 

DDT exposure is also associated with developmental abnormalities in amphibians and avian wildlife. 

DDT exposures are associated with delayed tadpole metamorphosis in the frog (Rana temporaria) (Cooke 

1972, 1973a) and altered facial features (Cooke 1970a). Developmental effects in avian wildlife 

associated with exposure to DDT/DDD/DDE include a reduced growth (Seidensticker 1968), decreased 

ability to thermoregulate (Vangilder and Peterle 1980), behavioral alterations (Heinz 1976), reduced 

testicular development (Burlington and Lindeman 1952), and development of ovarian tissue and oviducts 

in genetic males (Fry and Toone 1981). Wildlife species may be appropriate sentinels of developmental 

effects in humans because certain effects, particularly reduced early survival in young, occurred 

consistently across several species under various exposure conditions. 

Neurological and Behavioral Effects. Neurological effects (tremors, convulsions, hyperactivity, 

and behavioral changes) were observed in mammalian wildlife, amphibians, and avian wildlife 

experimentally exposed to DDT or DDE, particularly after administration of lethal doses or after 

administration of lower doses when food intake was restricted. Tremors were the most commonly 

reported neurological effect and have been reported in the brown bat (the short-tailed shrew (Blarina 

brevicauda) (Blus 1978), the free-tailed bat (Tadarida brasiliensis) (Clark and Kroll 1977) the big brown 

bat (Eptesicus fuscus) (Luckens and Davis 1964) and Pipistrelle bats (Jefferies 1972). Studies generally 

did not offer explanations as to the possible mechanisms that caused tremors, although it is reasonable to 

assume a mechanism similar to that seen in laboratory animals. Diets were experimentally restricted in 

several studies to simulate the health effects of DDT/DDE/DDD mobilized from fat during periods of 

energetic stress in the wild such as may occur, for example, during periods of nesting, migration, or 

thermal or other stress. Reviews (EPA 1975; WHO 1989) have postulated that during periods of energy 

stress, DDT mobilized from fat is redistributed to the brain (presumably because of the high lipid content 

in brain tissue) where it induces neurological effects and death. A study in bats (Clark and Kroll 1977) 

demonstrated that DDT residues in the brain increase substantially when the diet was restricted. Although 

a direct action on the central nervous system in wildlife has not been confirmed by observations of brain 

lesions, one study documented significant decreases in brain neurotransmitter levels associated with 

increased brain DDE residue levels after sublethal dietary exposures (Heinz et al. 1980). Alterations in 

neurotransmitter levels may explain changes in bird behavior that were observed in several species. 

Neurological effects observed in amphibians exposed to DDT/DDE/DDD in water include uncoordinated 

movement (Cooke 1970b) and hyperactivity (Cooke 1972, 1973a), tremors, lack of muscular coordination 

and weakness (Harri et al. 1979). Most available data suggest that wildlife species exhibit neurological 

effects similar to those observed in humans. These neurological effects, however were observed in 

wildlife at lethal exposure levels or in energy-stressed animals at lower exposure levels. 
 

DDT, DDE, and DDD 138 

3. HEALTH EFFECTS 

In avian wildlife DDT/DDD/DDE exposures are associated with decreased brain dopamine levels with a 

significant negative correlation observed between neurotransmitter levels and DDE residues in the brain 

(Heinz et al. 1980). Tremors have also been observed in bald eagles (Haliaetus leucocephalus) (Chura 

and Stewart 1967; Locke et al. 1966), kestrels (Falco sparverius) (Porter and Wiemeyer 1972), 

double-crested cormorants (Phalacrocorax auritus) (Greichus and Hannon 1973), pheasants (Azevedo et 

al. 1965), bobwhite quail (Colinus virginianus) (Hill et al. 1971), Japanese quail (Davison et al. 1976; 

Gish and Chura 1970), house sparrows (Passer domesticus) (Boykins 1967), cardinals (Richmondena 

cardinalis), and blue jays (Cyanocitta cristata) (Hill et al. 1971), homing pigeons (Columba livia) 

(Jefferies and French 1971) (Jefferies and French 1972), and domestic chickens (Glick 1974). Balance 

disturbances have also been observed (in some cases prior to death) in pheasants that (Azevedo et al. 

1965), Bobwhite quail (Colinus virginianus) (Hill et al. 1971), cardinals (Richmondena cardinalis), and 

blue jays (Cyanocitta cristata) (Hill et al. 1971). Neurological effects in avian wildlife are also 

manifested as behavioral effects. These include the delayed onset of nocturnal restlessness indicative of 

normal migratory behavior (Mahoney 1975), significantly decreased courting behavior (Haegele and 

Hudson 1977), and decreased nest attendance by parental birds (Keith and Mitchell 1993). 

Other adverse effects observed in wildlife species are described in detail in Appendix D. This section 

only provides a summary of reproductive, developmental and neurological effects that are the primary 

adverse effects to terrestrial wildlife associated with DDT/DDD/DDE exposure. 

3.5 TOXICOKINETICS 

Overview. Although oral exposure is considered the most significant route of entry in humans, DDT, 

DDE, and DDD are absorbed following inhalation, oral, and dermal exposures. Oral exposure to DDT, 

DDE, and DDD results in preferential absorption by the intestinal lymphatic system. Some absorption 

into the portal blood also occurs. Limited data exist regarding the rate and extent of DDT absorption in 

humans. DDT, DDE, and DDD are readily distributed in the lymph and blood to all body tissues and 

ultimately stored in proportion to the lipid content of the tissue, regardless of the route of exposure. 

Metabolism of DDT in humans appears similar to that seen in rats, mice, and hamsters, except that not all 

intermediate metabolites detected in animals have been identified in humans. Excretion of DDT in the 

form of its metabolites (e.g., DDA and its conjugates) is largely via the urine, regardless of route of 

exposure, but DDT excretion may occur via feces, semen, and breast milk. Some experiments have 
DDT, DDE, and DDD 139 

3. HEALTH EFFECTS 

suggested that fecal excretion may be the major route of elimination at high doses; however, this has not 

been confirmed by later investigations. 

3.5.1 Absorption 

3.5.1.1 Inhalation Exposure 

Absorption of DDT by the lung is considered to be a minor route of entry. It is assumed that the large 

particle size of DDT (crystalline) prevents it from entering the deeper, smaller spaces of the lung, and that 

it is deposited on the upper respiratory tract mucosa and then eventually swallowed because of the action 

of the mucociliary apparatus (Hayes 1982). Some crystalline DDT, however, could be small enough to 

pass through the tracheal-bronchial passages. In occupational settings, human exposure has occurred by a 

mixture of routes, including inhalation with subsequent oral ingestion, and dermal absorption. Evidence 

of DDT absorption was indicated by the appearance of DDA (a DDT metabolite) in the urine (Laws et al. 

1967; Ortelee 1958) and the presence of DDT in adipose tissue (Laws et al. 1967) and plasma or serum 

(Morgan et al. 1980; Rabello et al. 1975). However, no studies were located that quantify the rate or 

extent of absorption of DDT, DDE, or DDD in humans after inhalation exposure. No studies were 

located regarding the absorption of DDT, DDE, or DDD after inhalation exposure in animals. 

3.5.1.2 Oral Exposure 

Absorption following ingestion of DDT, DDE, and DDD is evident in humans both from measurements 

of serum and adipose tissue concentrations of these chemicals and from measurements of DDA in the 

urine (Hayes et al. 1971; Morgan and Roan 1971, 1974). Indirect evidence of absorption is provided in 

the development of toxicity following accidental or intentional (suicidal) ingestion of DDT (Hsieh 1954). 

In subjects chronically exposed to oral doses of DDT up to 20 mg/day (approximately 0.3 mg/kg/day), 

DDT appeared in the serum and reached peak serum concentrations 3 hours after ingestion (Morgan and 

Roan 1971). Serum levels remained elevated but returned to near pre-dose values 24 hours after each 

dose. 

Gastrointestinal absorption can be inferred in animals. The presence of urinary metabolites in mice, rats, 

and hamsters (Fawcett et al. 1987; Gold and Brunk 1982, 1983, 1984), the presence of DDT and its 

metabolites in bile collections (Jensen et al. 1957), and the induction of tumors and other toxic effects in 

animals after oral administration of DDT, DDE, or DDD provide evidence of gastrointestinal absorption. 
DDT, DDE, and DDD 140 

3. HEALTH EFFECTS 

In animals, absorption of orally administered DDT is enhanced when it is dissolved in digestible oils 

(Keller and Yeary 1980). Approximately 70–90% of the administered dose is absorbed by rats after oral 

exposure to DDT in vegetable oils (Keller and Yeary 1980; Rothe et al. 1957). DDT is absorbed 

1.5–10 times more effectively in laboratory animals when given in digestible oils than when dissolved in 

nonabsorbable solvents (Hayes 1982). 

Gastrointestinal absorption by way of the intestinal lymphatic system plays a major role in the uptake of 

DDT in animals (Noguchi et al. 1985; Pocock and Vost 1974; Sieber 1976; Turner and Shanks 1980). 

For example, Sieber (1976) showed that 12–24% of the administered dose was recovered in the 24-hour 

lymph after intraduodenal administration of 14C-isomers to thoracic duct-cannulated rats, and most of the 

radioactivity was attributed to parent compounds. Other studies indicate that little DDT is absorbed from 

the gastrointestinal tract directly into the blood (Palin et al. 1982; Rothe et al. 1957). 

3.5.1.3 Dermal Exposure 

Dermal absorption of DDT in humans and animals is considered to be limited, but can be inferred by 

observation of toxicity after dermal application of DDT. Acute toxicity studies in several species 

demonstrate that toxicity, expressed as an LD50, is less when DDT is applied dermally than when given by 

gavage or by injection, which reflects the difference in the amount of DDT absorbed by the dermal route. 

The data indicate that DDT is 4 times more toxic when given by intraperitoneal injection than when 

administered orally and 40 times more potent when given by intraperitoneal injection than when 

administered by the dermal route (Hayes 1982). Absorption of DDT from soil applied to the abdomen of 

monkeys, as extrapolated from urinary excretion data, was 3.3% of the applied dose in 24 hours (Wester 

et al. 1990). 

3.5.2 Distribution 

The distribution and storage of DDT in humans and animals has been extensively studied. DDT and its 

metabolites, DDE and DDD, are lipid-soluble compounds. Once absorbed, they are readily distributed 

via the lymph and blood to all body tissues and are stored in these tissues generally in proportion to organ 

tissue lipid content (Morgan and Roan 1971). 

Hayes et al. (1971) and Morgan and Roan (1971, 1974) evaluated the distribution of orally administered 

DDT, DDE, or DDD in volunteers. Morgan and Roan (1971, 1974) and Roan et al. (1971) measured the 
DDT, DDE, and DDD 141 

3. HEALTH EFFECTS 

concentration of DDT, DDE, DDD, and DDA in blood, fat, and urine after oral dosing. The administered 

doses ranged from 5 to 20 mg DDT/kg/day for up to 6 months; the ratio of concentration of DDT stored 

in adipose tissue to that present in blood was estimated to be 280:1. 

DDT uptake into tissues is a function of the blood flow, lipid content of that tissue, and the partition 

coefficient for DDT between the blood and lipids in specific organs. The ratio of DDT concentrations in 

adipose tissue to blood may remain relatively constant; however, the amount of DDT from past exposure 

cannot be determined from present blood levels only. DDT, DDE, and DDD have been reported to be 

distributed to and retained in the adipose tissue of humans (Morgan and Roan 1971). The affinity for 

storage in adipose tissue is related to each chemical's lipophilicity and increases in the order p,p’-DDD #

o,p,’-DDT < p,p’-DDT < p,p’-DDE (Morgan and Roan 1971). 

DDT and DDE selectively partition into fatty tissue and into human breast milk, which has a higher fat 

content than cow’s milk. Takei et al. (1983) reported concentrations from the 1969–1970 U.S. national 

human milk study. The p,p’-isomer of DDT and DDE was found in 100% of the samples tested, with 

mean concentrations of 0.19 and 1.9 ppm (lipid-basis), respectively. However, variance in levels of DDT 

and its metabolites may be influenced by such factors as number of parity, children nursed, diet, and 

cigarette smoking (Bouwman et al. 1990; Bradt and Herrenkohl 1976; Mes et al. 1993; Rogan et al. 

1986). A steady decrease in the levels of DDT and its metabolites in human milk has been reported as a 

result of decreased intake of DDT. In Finland, samples taken between 1973 and 1982 indicate a reduction 

of more than 50% in total DDT concentration in human milk (Wickstrom et al. 1983). Using data from 

the United States and Canada, Smith (1999) estimated that since 1975, there has been an 11–21% decline 

in average DDT in breast milk. Table 6-4 shows levels of DDT and related compounds in breast milk 

from some recent studies. 

LaKind et al. (2000) formulated a model to estimate doses of DDE to nursing infants. Included in the 

model were nursing variables (duration of breast-feeding, volume of breast milk consumed, percent lipid 

in breast milk, and milk density), infant variables (percent absorption of chemical, half-life of chemical, 

and body weight), and maternal depuration rate of the chemical via breast-feeding (time-dependent 

concentration of the chemical in the mother’s milk). Several exposure scenarios were evaluated: 3, 6, 9, 

and 12 months, and a randomly simulated duration. The concentration of DDE in milk at time zero used 

in the model was 222.3 ppb (lipid basis), which was the mean value obtained in a Canadian breast milk 

evaluation of 497 donors from across Canada conducted in 1992 (Newsome et al. 1995). The results 

showed that, regardless of the exposure scenario simulated, infant body burdens of DDE increase rapidly 
 

DDT, DDE, and DDD 142 

3. HEALTH EFFECTS 

at the start of lactation, but decrease after approximately 5–6 months, even if nursing continues. The 

maximum mean body burden of DDE was about 70 μg/kg, and by 24 months postpartum, it was 

<10 μg/kg, regardless of the duration of breast-feeding. 

DDT and DDT metabolites have been shown to cross the placenta. A study of 90 mother/infant pairs 

from Mexico found that all 90 cord blood samples had detectable levels of p,p’-DDE, 9 had detectable 

levels of o,p’-DDT, and 44 had measurable levels of p,p’-DDT (Waliszewki et al. 2000). The mothers 

were volunteers from the general population with no known occupational exposure to DDT. The 

concentrations of the pesticide residues in maternal blood were very similar to those in cord blood. 

p,p’-DDE was the most concentrated (lipid-adjusted means, 4.4 ppm maternal; 4.7 ppm cord blood), 

followed by p,p’-DDT (1.8 ppm maternal, 2.8 ppm cord blood), and o,p’-DDT (0.30 ppm maternal, 

0.35 ppm cord blood). A study of Spanish women reported a geometric mean of 2.2 ppb (not lipidadjusted) for p,p’-DDE in maternal blood at delivery (n=72) and 0.83 ppb in cord blood (n=69) (Sala et 

al. 2001). The women in the study lived in the vicinity of an organochlorine-compound factory. 

The distribution of p,p’-DDT in newborn rats from dams administered p,p’-DDT in the diet before mating 

and throughout gestation was evaluated before and after suckling had occurred (Woolley and Talens 

1971). In newborn rats sacrificed 0–1 hours after birth, before suckling had occurred, levels of p,p’-DDT 

were noted in the brain, liver, kidneys, and stomach. These results demonstrate that DDT readily passes 

through the placental barrier to enter tissues of the developing fetus. In newborn rats sacrificed after 

suckling, the tissue levels were relatively higher than levels in newborns before suckling. This could be 

attributed in part to the higher levels of DDT in the maternal milk, compared to DDT levels in maternal 

plasma (Woolley and Talens 1971). Recently, You et al. (1999b) studied the transplacental and 

lactational transfer of p,p’-DDE in rats. Pregnant rats were administered 10 or 100 mg p,p’-DDE/kg/day 

on gestation days (Gd) 14–18. In dams killed on day 15, the concentration of p,p’-DDE in the placenta 

was about 3-fold higher than in fetal tissues; similar observations were made in rats killed on Gd 17. 

Using a cross-fostering scheme, the authors found that lactational transfer provided the pups with a 

p,p’-DDE tissue burden far greater than that provided through placental transfer. For example, 10-dayold pups exposed only in utero to the low-dose had no detectable p,p’-DDE in blood, liver, or brain (only 

samples collected); p,p’-DDE could be detected only in the liver of high-dose pups. In contrast, in pups 

exposed only via dam’s milk or through combined gestational and lactational exposure, p,p’-DDE was 

detected at much higher concentrations in the samples in the order of liver > brain > blood, and no 

significant differences were seen between the two groups. In 78-day-old rats, p,p’-DDE was detected 

only in fat in the three exposure groups, and rats exposed only in utero had at least two orders of 
DDT, DDE, and DDD 143 

3. HEALTH EFFECTS 

magnitude less p,p’-DDE in fat than the other two groups. In dams, at the end of nursing, p,p’-DDE 

levels in tissues and plasma were approximately 1/3 those at the end of gestation, suggesting that a large 

portion of the p,p’-DDE in storage sites was mobilized during lactation. See Section 3.5.5 for a 

description of physiologically based pharmacokinetic (PBPK) models developed by You et al. (1999b) to 

describe transplacental and lactational transfer of p,p’-DDE. 

The time course of distribution of DDT (evaluated as o,p’- and p,p’-isomers) in reproductive tissues of 

female rabbits was examined by Seiler et al. (1994). The rabbits were administered the test material by 

gavage 3 days/week for 12–15 weeks before artificial insemination and throughout gestation. The authors 

examined DDT residues in oviductal and uterine luminal fluid, cleavage stage embryos (day 1 

postcoitum), blastocytes (day 6 postcoitum), and in fetuses, exocoelic fluid and placenta (day 11 

postcoitum). No demonstrable residues were detected in cleavage stage embryos and tubal flushings. 

Relative to controls, DDT residues were significantly increased in blastocytes (14-fold), uterine fluid 

(7-fold), fetuses (40-fold), and exocoelic fluid (700-fold). The higher concentration of DDT residues 

found in fetuses relative to blastocytes suggested that transplacental passage may be more easily 

accomplished than passage into blastocytes via uterine secretions. 

Mühlebach et al. (1991) examined the kinetics of distribution of p,p’-DDE in rats after a single 

intravenous dose of 5 mg/kg of the radiolabeled material. Peak concentrations of DDE were observed 

before 1 hour in the liver and muscle, at 3 hours in the skin, and between 1 and 4 days in adipose tissue. 

Between 4 and 14 days after exposure, the tissue/blood concentration ratio was about 6 for liver and 

muscle, 35 for skin, and 400 for adipose tissue. The results also showed that the distribution kinetics of 

DDE was characterized by a redistribution from blood to liver and muscle, to skin, and ultimately to 

adipose tissue, and this process appeared to take about 1 day. 

3.5.3 Metabolism 

The metabolism of DDT, DDE, and DDD has been studied in humans and a variety of other mammalian 

species. The metabolism in rats, mice, and hamsters is similar to that in humans; however, not all of the 

intermediary metabolites identified in animals have been identified in humans. It has been proposed by a 

number of investigators that in mammals, the major urinary metabolite of DDT, 2,2-bis(p-chlorophenyl) 

acetic acid (DDA), is produced by a sequence involving reductive dechlorination, dehydrochlorination, 

reduction, hydroxylation, and oxidation of the aliphatic portion of the molecule (Gingell 1976; Peterson 

and Robison 1964). In this proposed pathway (Model I: see Figure 3-3a), DDT is initially metabolized in 
DDT, DDE, and DDD 144 

3. HEALTH EFFECTS 

the liver to two intermediary metabolites, DDE (Mattson et al. 1953; Pearce et al. 1952) and DDD (Klein 

et al. 1964). In rats, DDE is slowly converted in the liver to 1-chloro-2,2-bis(p-chlorophenyl)ethene 

(DDMU), and then to DDA in the kidney by way of 1,1-bis(p-chlorophenyl)ethene (DDNU) (Datta 1970; 

Datta and Nelson 1970). DDD is rapidly detoxified by way of DDMU to 1-chloro-2,2-bis(p-chlorophenyl)ethane (DDMS) and then to DDNU (Datta 1970). Metabolism of DDMS to DDNU occurs in both 

the liver and kidney, but the kidney is the primary site (Datta 1970). DDNU is then further metabolized, 

primarily in the kidney, to 2,2-bis(p-chlorophenyl)ethanol (DDOH) then to 

2,2-bis(p-chlorophenyl)ethanal (DDCHO) (Suggs et al. 1970), which is further oxidized to DDA 

(Peterson and Robison 1964). 

Other evidence in mice and hamsters suggests an alternative metabolic scheme (Gold and Brunk 1982, 

1983) (Model II: see Figure 3-3b). In these studies, one dose of radioactively-labeled DDT or DDD was 

administered by gavage, and urine was collected for 72 hours for analysis. The principal urinary 
R2CHCCl3 R= Cl

DDT 

LIVER LIVER 

R2C=CCl2 R2CHCHCl2 

DDE DDD

SLOW 

FASTER LIVER LIVER 

R2-C CHCl R2C=CHCl 

DDMU O 

LIVER 

DDMU epoxide 

R2CHCH2Cl

LIVER DDMS 

-HCl KIDNEY & 

LIVER (some) electrophile R2C=CH2 

DDNU 

KIDNEY 

R2CH-CH2OH 

DDOH 

-2H KIDNEY 

R2CH-CHO 

DDCHO 

R2CH-COOH CONJUGATION

DDA 

*Adapted from Peterson and Robinson 1964 

DDT, DDE, and DDD 145 

3. HEALTH EFFECTS 

Figure 3-3a. Model I Metabolic Scheme for DDT* 
Figure 3-3b. Model II Metabolic Scheme for DDT* 

Cl 

R2-CH-CCl R = 3

DDT

H Deuteration here

affects rate to DDA

R2CHCHCl2 R2C=CHCl DDD minor path DDMU 

major path 

O OH 

R2CH-C-Cl2 O electrophile 

-HCl 

R2C CHCl 

DDMU-epoxide O 

H2O 

electrophile 

R2CHCCl 

DDA-Cl 

R2C(OH)CHO 

R [H] 

R2C(Cl)

2C(Cl)CHO

αOH-DDCHO 

αCl-DDCHO 

O 

R2C=

R

O 2CH-COOH 

R2C(Cl)CHO DDA 

αCl-DDA 

R2CH

H2O DD

H CONJUGATION 

R2C(OH)COOH 

R2CHCH2OH 

OH-DDA *Derived from Gold and Brunk 1982 DDOH 
DDT, DDE, and DDD 147 

3. HEALTH EFFECTS 

metabolites were DDA and its conjugates. Few metabolite intermediates on the Model I pathway were 

isolated in the urine of DDT- or DDD-treated animals; however, if DDMU, a Model I pathway 

intermediate, was administered, downstream Model I metabolites were found in greater quantities. Based 

on this and the fact that deuterium labeling studies (Gold and Brunk 1984) indicate that removal of the 

alpha hydrogen from DDD is a rate limiting step in DDA formation, Gold et al. (1981) hypothesized that 

DDD is hydroxylated at the chlorinated 1-ethane side-chain carbon to yield 2,2-bis(p-chlorophenyl)acetyl 

chloride (DDA-Cl), which in turn can be hydrolyzed to the major urinary metabolite, DDA. Epoxidation 

of DDMU, a minor metabolite of DDD in Model II (note that in Model I it is not minor), yields DDMUepoxide (Gold et al. 1981). The major (Model II) urinary metabolite DDA-Cl, which is an electrophile 

capable of acylating nucleophilic cellular molecules, and the minor DDMU-epoxide metabolite may 

contribute to the known tumorigenicity of DDT and DDD via the formation of covalent DDA adducts in 

the mouse. Urinary metabolites were similar in both female Swiss mice and female Golden Syrian 

hamsters (Gold and Brunk 1982, 1983, 1984). The metabolic disposition of DDT, DDD, and DDMU in 

the hamster is similar to that of the mouse and proceeds by the same oxidative metabolic pathways in both 

species (Gold and Brunk 1983). Therefore, it is unlikely that the observed differences in species 

sensitivity to the DDT-induced tumorigenicity in the mouse and the resistance to tumor production from 

DDT exposure in the hamster are due to differences in the production of the DDMU-epoxide or DDA-Cl. 

However, there was a species difference in the metabolic conversion of DDT to DDE. DDE was detected 

at much higher levels in the urine of mice after both acute and chronic studies than in hamsters (Gingell 

1976; Gold and Brunk 1983). The data indicate that the hamster was less efficient than the mouse in the 

conversion of DDT to DDE. Fawcett et al. (1987) evaluated the metabolism of radiolabeled DDT, DDE, 

DDD, and DDMU in male Wistar rats. These results suggested that the metabolism of DDT to DDA 

proceeding via the acid chloride (DDA-Cl), as had also been found in the mouse and hamster (Gold and 

Brunk 1982, 1983, 1984). 

Several investigators have isolated DDT metabolites from human urine, serum, and adipose tissue. The 

DDT metabolites in humans are the same as some of those produced in animals and it can be inferred that 

the metabolic pathways in humans and animals are similar. In humans, ingested DDT undergoes 

reductive dechlorination to DDD, which is further degraded and readily excreted as DDA (Roan et al. 

1971). DDT is also converted by dehydrodechlorination to DDE, although at a much slower rate than the 

DDT-to-DDD pathway (Morgan and Roan 1971). Morgan and Roan (1971) concluded that the 

conversion of DDT to DDE occurs with considerable latency and that the extent of the conversion was 

estimated to be less than 20% over the course of the 3-year study. Further metabolism of DDE is 

apparently slow, and DDE is retained in adipose tissue (Hayes et al. 1971; Morgan and Roan 1971). 
DDT, DDE, and DDD 148 

3. HEALTH EFFECTS 

According to Roan et al. (1971) and Morgan and Roan (1971), oral administration of DDT or DDD to 

volunteers resulted in an increased urinary excretion of DDA, but no increase in excretion of DDA above 

predose values was noted after oral ingestion of DDE. According to the study authors, the data indicate 

that DDD, not DDE, is the precursor for DDA in humans and that little if any DDE is further converted to 

DDA. 

After Phase I metabolism (reactions involving oxidation, reduction, and hydrolysis), many of the DDT 

metabolites ultimately are excreted in the conjugated form. Conjugates have been reported to include 

glycine, bile acid conjugates, serine, aspartic acid, and glucuronic acid (Gingell 1975; Pinto et al. 1965; 

Reif and Sinsheimer 1975). 

DDT induces microsomal mixed function oxidases that are involved in the catabolism of both xenobiotics 

and many endogenous hormones. DDT has also been shown to induce its own metabolism in rats 

(Morello 1965) and hamsters (Gingell and Wallcave 1974). DDE and DDD have also caused the 

induction of hepatic cytochrome P-450 microsomal enzymes (Pasha 1981). DDT, DDE, and DDD were 

each found to be phenobarbital-type cytochrome P-450 inducers in male F344 rats, causing induction of 

hepatic CYP2B and less of CYP3A, but not CYP1A proteins (Nims et al. 1998). Limited induction of 

CYP1A1-associated enzyme activity was observed. DDT and DDE appeared to have similar inducing 

potencies, whereas the potency of DDD was within one order of magnitude lower than the other two 

compounds. Nims et al. (1998) further demonstrated that CYP2B induction resulting from DDT 

administration was due to the combined inductive effects of DDT, DDE, and DDD, and not exclusively to 

the DDE generated metabolically. 

The metabolism of DDT can also produce methylsulfonyl metabolites, which are potent toxicants, 

particularly in the adrenal gland, after metabolic activation. Methylsulfonyl metabolites of DDT 

(specifically 3- and 2-methylsulfonyl-DDE) were first identified in seal blubber from the Baltic Sea 

(Jensen and Jansson 1976); they have later been found in several species of animals (Bergman et al. 1994) 

and in humans (Weistrand and Norén 1997). Methylsulfonyl-DDE is formed as follows: products of the 

reaction between arene oxides, formed in phase I metabolism, and glutathione are degraded and excreted 

in the bile into the large intestine where they undergo cleavage by a microbial C-S lyase (Bakke et al. 

1982; Preston et al. 1984). The thiols formed are methylated, reabsorbed, and the sulfur is further 

oxidized to the corresponding methylsulfones, which are distributed by the blood (Haraguchi et al. 1989). 

Figure 3-4 shows the proposed pathway for sulfonyl metabolites. Further information regarding the 
DDT, DDE, and DDD 149 

3. HEALTH EFFECTS 

toxicity of this metabolite is presented in Section 3.7, Toxicities Mediated Through the Neuroendocrine 

Axis. 

3.5.4 Elimination and Excretion 

Excretion of DDT has been studied in humans and a variety of animals. The major route of excretion of 

absorbed DDT in humans appears to be in the urine, but some excretion also occurs by way of feces (via 

biliary excretion) (Jensen et al. 1957) and breast milk (Takei et al. 1983). Results of studies with mice, 

rats, and hamsters indicate that the metabolites of DDT and small amounts of unmetabolized DDT are 

excreted primarily in the urine and feces (Gold and Brunk 1982, 1983, 1984). 

The biological half-lives for the elimination of these compounds are ranked as follows: 

DDE>DDT>DDD. This relationship is based on the chemical stability of each compound in the body, 

efficiency of excretory mechanisms, and possibly transport in and out of fat depots (Morgan and Roan 

1971). The excretion of DDT was investigated in volunteers who ingested DDT. Hayes et al. (1971) 

reported that, in subjects receiving 35 mg/day (approximately 0.5 mg/kg) for up to 18 months, urinary 

excretion of DDA increased rapidly for the first few days and a steady state level of approximately 

13–16% of the daily dose was reached and remained stable for 56 weeks. However, although the rate of 

excretion of DDA was relatively constant in each individual, there were marked differences observed 

between men receiving the same dose. Urinary excretion of DDA fell rapidly after cessation of dosing. It 

appears that a steady state for storage was reached within 12–18 months of daily dosing, after which 

humans were apparently able to eliminate the entire daily dose of 35 mg/day. Although no excretion of 

DDT metabolites was detected in the feces, the authors stated that there were probably DDT metabolites 

in the feces which were organic unextractable polar conjugates (Hayes et al. 1971). Since only 

5.7 mg/day of all DDT isomers were found in the urine at steady state, it was postulated that other routes 

of excretion, such as biliary transport, may be involved. This was shown by Paschal et al. (1974). Roan 

et al. (1971) reported that increased urinary excretion of DDA is detectable within 24 hours of ingestion 

of DDT (5, 10, or 20 mg/day), DDD (5 mg/day), or DDA (5 mg/day). DDA excretion returns to predose 

levels within 2–3 days of dose termination for DDA, but continues significantly above predose levels for 

over 4 months after termination of DDD or DDT doses. Ingestion of DDE (5 mg/day) failed to produce 

any increase in DDA excretion. 
Cl 

C 

Cl Cl phase I Cl 

C enzymes
 

Cl
 C Cl 

P450 Cyp-2B 

C Cl Cl

O 

phase II 

enzymes, 

glutathione 

Cl Cl Cl 

C 

Cl C bile 

Cl C Cl Cl C Cl

flora 

(C-S lyase) SH S-glutathione methylation of
 

thiol,
 

oxidation of thiol
 

Cl Cl 

C 

Cl C Cl

SO2CH3 

Source: Bergman et al. 1994; Letcher et al. 1998; Weistrand and Norén 1997 

DDT, DDE, and DDD 150 

3. HEALTH EFFECTS 

Figure 3-4. Proposed Metabolic Pathway for the Conversion of p,p’-DDE 

to its Methylsulfone Derivative 
DDT, DDE, and DDD 151 

3. HEALTH EFFECTS 

DDT excretion in the feces may be a major route of excretion at high doses of DDT. DDT and DDTrelated metabolites have been identified in the feces of humans receiving 35 mg/day (Hayes et al. 1956); 

however, this result has not been confirmed by later investigations (Hayes et al. 1971). According to 

Jensen et al. (1957), biliary excretion is the major source of DDT metabolites found in the feces of rats, as 

demonstrated in bile-cannulated rats given an intravenous injection of DDT. Following intravenous 

administration of DDT in rats, bile cannulation results indicated that enterohepatic circulation was 

occurring for conjugated DDA, a DDT metabolite (Gingell 1975; Pinto 1965). 

Analysis of urine from humans occupationally exposed to DDT showed the presence of DDA (Laws et al. 

1967; Ortelee 1958; Ramachandran et al. 1984). By comparing the urinary excretion of DDA with that of 

volunteers given known doses of DDT, the average occupational exposure can be estimated (WHO 1979). 

The observations by Laws et al. (1967) and Ortelee (1958) indicate that the urinary excretion of DDA is 

correlated with the level of exposure to DDT. The concentration of DDA in the urine in occupationally 

exposed workers was reported to be greater than that observed in the general population, while DDE 

excretion was reported to be only slightly higher than in the general population. It is of interest that 

monkeys fed DDT stored DDE in their fat (not DDT) and when feeding ceased, it was rapidly lost, 

probably by urinary excretion (Durham et al. 1963). 

Parameters of elimination of DDE and DDE-derived radioactivity in rats were estimated by Mühlebach et 

al. (1991) after administration of a single intravenous dose. Over a 14-day period following dosing, 34% 

of the administered dose was excreted in the feces and 1% in the urine. In the feces, 10% of the excreted 

radioactivity represented unchanged DDE, whereas no unchanged DDE could be detected in the urine. 

No hexane-extractable lipophilic metabolites were found in the feces. The average total DDE recovered 

in tissues and excreta was 90%. The total body burden half-life was 120 days. 

3.5.5	 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) 

Models 

Physiologically based pharmacokinetic (PBPK) models use mathematical descriptions of the uptake and 

disposition of chemical substances to quantitatively describe the relationships among critical biological 

processes (Krishnan et al. 1994). PBPK models are also called biologically based tissue dosimetry 

models. PBPK models are increasingly used in risk assessments, primarily to predict the concentration of 

potentially toxic moieties of a chemical that will be delivered to any given target tissue following various 

combinations of route, dose level, and test species (Clewell and Andersen 1985). Physiologically based 
DDT, DDE, and DDD 152 

3. HEALTH EFFECTS 

pharmacodynamic (PBPD) models use mathematical descriptions of the dose-response function to 

quantitatively describe the relationship between target tissue dose and toxic end points. 

PBPK/PD models refine our understanding of complex quantitative dose behaviors by helping to 

delineate and characterize the relationships between: (1) the external/exposure concentration and target 

tissue dose of the toxic moiety, and (2) the target tissue dose and observed responses (Andersen et al. 

1987; Andersen and Krishnan 1994). These models are biologically and mechanistically based and can 

be used to extrapolate the pharmacokinetic behavior of chemical substances from high to low dose, from 

route to route, between species, and between subpopulations within a species. The biological basis of 

PBPK models results in more meaningful extrapolations than those generated with the more conventional 

use of uncertainty factors. 

The PBPK model for a chemical substance is developed in four interconnected steps: (1) model 

representation, (2) model parametrization, (3) model simulation, and (4) model validation (Krishnan and 

Andersen 1994). In the early 1990s, validated PBPK models were developed for a number of 

toxicologically important chemical substances, both volatile and nonvolatile (Krishnan and Andersen 

1994; Leung 1993). PBPK models for a particular substance require estimates of the chemical substancespecific physicochemical parameters, and species-specific physiological and biological parameters. The 

numerical estimates of these model parameters are incorporated within a set of differential and algebraic 

equations that describe the pharmacokinetic processes. Solving these differential and algebraic equations 

provides the predictions of tissue dose. Computers then provide process simulations based on these 

solutions. 

The structure and mathematical expressions used in PBPK models significantly simplify the true 

complexities of biological systems. If the uptake and disposition of the chemical substance(s) is 

adequately described, however, this simplification is desirable because data are often unavailable for 

many biological processes. A simplified scheme reduces the magnitude of cumulative uncertainty. The 

adequacy of the model is, therefore, of great importance, and model validation is essential to the use of 

PBPK models in risk assessment. 

PBPK models improve the pharmacokinetic extrapolations used in risk assessments that identify the 

maximal (i.e., the safe) levels for human exposure to chemical substances (Andersen and Krishnan 1994). 

PBPK models provide a scientifically sound means to predict the target tissue dose of chemicals in 

humans who are exposed to environmental levels (for example, levels that might occur at hazardous waste 
 

DDT, DDE, and DDD 153 

3. HEALTH EFFECTS 

sites) based on the results of studies where doses were higher or were administered in different species. 

Figure 3-5 shows a conceptualized representation of a PBPK model. 

Models of the pharmacokinetics of p,p'-DDE, a principle metabolite of DDT, have been proposed by You 

et al. (1999b). The You et al. (1999b) models are PBPK models of p,p'-DDE uptake and disposition in 

pregnant and lactating/nursing rats. The models are based on experimental studies in which pregnant 

Sprague-Dawley rats were administered gavage doses of p,p'-DDE, and the kinetics of p,p'-DDE tissue 

levels in the dams, fetuses, and pups were measured. The models provide an approach to estimating 

tissue doses in fetuses and pups associated with maternal exposures to p,p'-DDE and can be used to 

explore dose-response relationships for the developmental effects of p,p'-DDE in the Sprague-Dawley rat. 

Description of the model. Figures 3-6 and 3-7 show conceptualized representations of the gestation 

model and the lactation/nursing models, respectively. Parameters used in the models are shown in 

Tables 3-5 and 3-6. The gestation model simulates the absorption of an oral dose of p,p’-DDE from the 

of p,p’-DDE from the dam to the nursing pup from mammary milk, followed by exchanges with pup fat, 

kidney, and other richly-perfused and poorly-perfused tissues in the pups. 

All exchanges with blood plasma, in both models, are simulated as flow-limited processes, with the 

exception of the following. Exchanges between maternal fat and a deep fat compartment are assumed to 

be diffusion-limited and are represented with first order rate constants. Exchanges between the 

embryo/fetus and placenta are modeled as diffusion-limited processes and are represented with diffusion 

coefficients (L/day). Parameters used in the model were either taken from the literature, estimated by 

using the SIMUSOLV simulation program, or set by visually inspecting the fit of the data. 

Elimination pathways in the maternal model include metabolism, transfer from mammary tissue to 

maternal milk (in the lactation model), and fecal excretion from the gastrointestinal tract, including 

transfer from bile to the gastrointestinal tract. A fecal pathway from liver (presumably through bile and 

the gastrointestinal tract) is included in the pup model. 
DDT, DDE, and DDD 154 

3. HEALTH EFFECTS 

Table 3-5. Tissue:Blood Partition Coefficients and Pharmacokinetic
 

Constants for Modeling DDE Disposition in the Pregnant Rata
 

Tissue:blood partition coefficients 

Liver 7 

Fat 450 

Poorly-perfused tissues 12 

Well-perfused tissues 6 

Kidney 6 

Uterus 6 

Placenta 2 

Mammary gland 12 

Pharmacokinetic constants 

KAS (L/day) Portal absorption rate constant 24 

KLY (L/day) Lymphatic absorption rate constant 74 

KFX (L/day) Fecal excretion rate constant 230 

KB (L/day) Biliary excretion rate constant 1.2 

PAF (L/day) Fat diffusion coefficient 5 

PA1 (L/day) Placenta-to-embryo/fetus diffusion coefficient 1.6 

PA2 (L/day) Embryo/fetus-to-placenta diffusion coefficient 1.9 

K12/K21 Diffusion to deep fat 1.0/0.1 

Tdel (day) Delay in time 0.1 

a

Adapted from You et al. 1999b 
DDT, DDE, and DDD 155 

3. HEALTH EFFECTS 

Table 3-6. Physiological Constants Used in the PBPK Model for the
 

Lactating Dam and the Nursing Pupa
 

Dam Pup 

Body weight (kg) (BW) 0.290–0.340 0.0061–0.58 

Tissue volumes (% of body weight) 

Liver, VL 4 4 

Well-perfused tissues, VWP 8 8 

Poorly-perfused tissues, VPP 76-VMT 76 

Fat, VF 7 0.0199*pBW+1.664 

Mammary tissue, VMT 4.4–9.6 

Milk, Vmilk 0.002L 

Cardiac output (L/h) 14*pBW0.75 18*pBW0.74 

Blood flows (% of cardiac output) 

Liver, QL 25 25 

Well-perfused tissues, QWP 41-QMT 49 

Poorly perfused tissues, QPP 25 25 

Fat, QF 7 1 

Mammary tissue, QMT 9–15 

a

Adapted from You et al. 1999b 
DDT, DDE, and DDD 156 

3. HEALTH EFFECTS 

Figure 3-5. Conceptual Representation of a Physiologically
 

Based Pharmacokinetic (PBPK) Model for a 

Hypothetical Chemical Substance
 

Source: adapted from Krishnan et al. 1994 

Note: This is a conceptual representation of a physiologically based pharmacokinetic (PBPK) model for a hypothetical 

chemical substance. The chemical substance is shown to be absorbed via the skin, by inhalation, or by ingestion, 

metabolized in the liver, and excreted in the urine or by exhalation. 
DDT, DDE, and DDD 157 

3. HEALTH EFFECTS 

Figure 3-6. Diagrammatic Representation of the PBPK Model for Gestation*. 

LYMPH 

DEEP 

FAT 

FAT 

POORLYPERFUSED 

WELLPERFUSED 

KIDNEY 

LIVER 

UTERUS 

MAMMARY 

TISSUE 

K12 K21 

QF 

QPP 

QU 

QMT 

QL 

QK 

QWP 

GI TRACT 

bile KB 

KAS 

KLY 

KFX 

feces 

Metabolism 

x N 

QDEC 

YOLK SAC 

PLACENTA 

x N 

CHORIOALLANTOIC 

PLACENTA 

QCAP 

EMBRYO/ 

FETUS 

PA2 PA1

PA2 PA1

p,p'-DDE 

oral dose 

N = number of concepti
 

Terms are Defined in Table 3-6 and Table 3-7.
 

*Adapted from You et al. 1999b
 
DAM PUP 

DEEP 

FAT 

FAT 

POORLYPERFUSED 

WELLPERFUSED 

KIDNEY 

LIVER 

MAMMARY 

TISSUE 

MILK 

K12 K21 

QF 

QPP 

QMT 

KMILK 

QL 

QK 

QWP 

LIVER 

bile feces 

x N 

PAM 

FAT 

POORLYPERFUSED 

WELLPERFUSED 

KIDNEY 

pQK 

pQWP 

pQPP 

pQF 

pQL 

DDT, DDE, and DDD 158 

3. HEALTH EFFECTS 

Figure 3-7. Diagrammatic Representation of the PBPK Model for the Lactating 

Dam and Nursing Pup*. 

N = number of pups
 

Portal and lymphatic absorption routes for dams are not shown (see Figure 3-6).
 

Terms are Defined in Table 3-5 and Table 3-6.
 

*Adapted from You et al. 1999b
 
 

 

 

 

 

DDT, DDE, and DDD 159 

3. HEALTH EFFECTS 

Validation of the model. The models have been calibrated with data from experimental studies. 

Pregnant Sprague-Dawley rats were administered gavage doses of 0, 10, or 100 mg p,p'-DDE on 

Gd 14–18. A subset of the dams were killed 4 hours after each dosing, and tissue levels of p,p'-DDE 

were measured in the dams, placenta, and fetuses. A subset of pups in each dose group was cross-fostered 

to assess p,p'-DDE transfer to tissues from maternal milk. The gestational model was calibrated by 

adjusting variables to achieve agreement between estimated and observed concentrations of p,p'-DDE in 

maternal and pup plasma, liver, and fat, and in placenta and fetal tissue. The lactation/nursing model was 

calibrated with data on p,p'-DDE concentrations in pup blood, liver, and fat. Reasonable agreement was 

achieved between model output and observations. 

Risk assessment. The models provide an approach to estimating tissue doses in fetuses and pups 

associated with maternal gavage doses of p,p'-DDE to pregnant Sprague-Dawley rats. This information 

can be used to interpret the results of developmental toxicity bioassays of p,p'-DDE in terms of maternal, 

fetal, and pup tissue dose-response relationships. 

Target tissues. Output from the You models that are described in You et al. (1999b) are estimates of 

the p,p'-DDE concentrations in maternal and pup blood and plasma, liver, and fat, and in placenta and 

embryo/fetus. Estimates of p,p'-DDE concentrations in other tissues may also be feasible, although they 

are not described in model evaluations thus far reported (You et al. 1999b) 

Species extrapolation. The models have been calibrated to predict p,p'-DDE pharmacokinetics in 

pregnant and lactating/nursing Sprague-Dawley rats. Extrapolation to other physiological states (e.g., 

immature rats, nonpregnant adults, senescent rats), other rat strains, or other species would require 

modification to the models to account for different tissue masses, blood flows, and possibly other kinetic 

variables. 

Interroute extrapolation. The models are calibrated to simulate the pharmacokinetics of p,p'-DDE 

when it is administered by gavage doses in corn oil or in a similar lipid vehicle. The gestation model 

includes a lymphatic absorption pathway from the gastrointestinal tract that is intended to simulate the 

absorption kinetics of p,p'-DDE dissolved in a lipophilic vehicle. The kinetics would be expected to be 

different for aqueous vehicles. Therefore, the output of the models cannot be extrapolated to other types 

of vehicles or to other exposure pathways (e.g., dietary, drinking water) or routes (e.g., dermal, 

inhalation) without re-evaluation and possibly recalibration and modification of the models. 
DDT, DDE, and DDD 160 

3. HEALTH EFFECTS 

3.6 MECHANISMS OF ACTION 

3.6.1 Pharmacokinetic Mechanisms 

DDT, DDE, and DDD are efficiently absorbed, distributed, and stored because of their high lipophilicity. 

There are not thought to be specific transport or distribution mechanisms for DDT, DDE, or DDD other 

than their affinity for other hydrophobic fats. These chemicals are absorbed in the stomach and, to a 

larger extent, in the intestine. Sieber (1976) studied lymphatic absorption of DDT and related compounds 

in rats and found these compounds to be preferentially absorbed via the intestinal lymphatic system with 

some absorption into blood. Pocock and Vost (1974) and Sieber et al. (1974) reported that most DDT 

absorbed into lymph is carried in the lipid core of chylomicrons. Once absorbed into the lymphatic 

system, DDT is carried throughout the body and incorporated into fatty tissues. In addition to facilitating 

the absorption of these compounds from the gastrointestinal tract, the lipophilicity of DDT, DDE, and 

DDD enables them to cross the blood-brain barrier readily without a specific transporter. Uptake into 

tissues is a function of the blood flow, lipid content of the specific tissue, and the partition coefficient for 

DDT between the blood and lipids in specific tissues. 

Studies in humans suggest that DDT and related compounds are primarily transported in the blood bound 

to protein. In occupationally exposed workers, Morgan et al. (1972) found that less than 18% of 

p,p’-DDT and p,p’-DDE in human blood is carried in erythrocytes. Less than 1% of all DDT-related 

compounds is carried by the chylomicrons in plasma of normal fat content. Instead, DDT-related 

compounds are carried by proteins and are undetectable in plasma from which protein has been 

precipitated. Following ultracentrifugation, p,p’-DDT and p,p’-DDE are found in relation to lipoproteins 

of various densities, but mainly in the triglyceride-rich, low density (LDL), and very low density 

lipoproteins (VLDL). Results from electrophoresis experiments showed that plasma albumin, and 

secondarily the smaller globulins are the principal plasma constituents associated with blood-borne 

p,p’-DDT and p,p’-DDE. In agreement with the findings of Morgan et al. (1972), Norén et al. (1999) 

found that in blood from humans, p,p’-DDE and its methylsulfonyl metabolite, 3-MeSO2-DDE, were 

associated to a great extent (almost 80%) with a fraction containing primarily albumin. In this case, the 

subjects (three men and two women) had no known occupational exposure to organochlorine compounds. 

Of the remaining DDE, 12% was recovered associated with LDL, 9% with VLDL, and 6% with high 

density lipoprotein (HDL). The corresponding figures for 3-MeSO2-DDE were 8, 4, and 11%. Results 

from studies in vitro are consistent with findings in vivo. Gómez-Catalán et al. (1991) showed that 82.3% 

of p,p’-DDE in human blood was associated with plasma and only 17.7% with cells. In plasma, 
 

DDT, DDE, and DDD 161 

3. HEALTH EFFECTS 

approximately 60% was associated with protein and 9–15% with the various lipid fractions. Similar 

results were reported by Mohammed et al. (1990) in an in vitro study with radiolabeled DDT. 

Extensive information regarding the metabolism of DDT and DDT-related compounds was presented in 

Section 3.5.3. DDT and DDT-related compounds are metabolized primarily in the liver and kidney, but 

important activation reactions also occur in other tissues such as the lung and adrenal glands. The 

metabolism of DDT can be considered both a detoxification reaction as well as an activation reaction. 

For effects associated with the parent compound, such as neurotoxicity, metabolism means detoxification. 

For other toxicities, such as antiandrogenicity and adrenal effects, biotransformation of DDT may be an 

activation process. In general, the metabolism of DDT in animals is similar to that in humans; however, 

there are also differences between species, and within species, there are differences between tissues. 

DDT, DDE, and DDD induce CYP2B-associated activities in rat liver, which makes them phenobarbitaltype cytochrome P-450 inducers. Studies in humans have shown that the major route of excretion of 

absorbed DDT is the urine and that DDA is the main urinary metabolite. 

3.6.2 Mechanisms of Toxicity 

Mechanisms for the major effects of DDT are discussed in this section. In animals, these adverse effects 

include neurotoxicity, hepatotoxicity, metabolic effects, reproductive effects, and cancer. 

Neurological Effects. DDT acts on the central nervous system by interfering with the movement of 

ions through neuronal membranes. There appear to be at least four mechanisms by which DDT affects 

ion movement, all possibly functioning simultaneously. DDT both delays the closing of the sodium ion 

channel and prevents the full opening of the potassium gates (Ecobichon 1995; Narahashi and Haas 

1967). DDT has been shown to target a specific neuronal adenosine triphosphatase (ATPase) thought to 

be involved in the control of the rate of sodium, potassium, and calcium fluxes through the nerve 

membrane (Matsumura and Patil 1969). This ATPase plays a vital role in neuronal repolarization. In 

addition, it has been suggested that DDT inhibits the ability to transport calcium ions in nerves possibly 

by binding with a hydrophobic site on calmodulin and secondarily effecting a Ca/Mg ATPase; however, 

this mechanism has not been confirmed (Matsumura 1985). Calcium ions are essential to the release of 

neurotransmitters. These actions combine to effectively maintain the depolarization of the nerve 

membrane, potentiating the release of transmitters and leading to central nervous system excitation 

manifested as hyperexcitability, tremors, and convulsions, along with secondary effects from convulsions, 

such as tachycardia, metabolic acidosis, and hyperthermia. 
 

DDT, DDE, and DDD 162 

3. HEALTH EFFECTS 

The action of DDT as a central nervous system stimulant may also be related to a deficiency of brain 

serotonin (Hwang and Van Woert 1978). Much research has focused on changes in levels of biogenic 

amines and amino acids resulting from DDT exposure. It is possible that some of these changes could be 

a result of the effects of DDT on membrane ion transport and consequent neurotransmitter release. 

Several researchers have reported increased levels of 5-HIAA, 3-methoxy-4-hydroxyphenylglycol 

(MHPG), aspartate, and glutamate in the brains of rats exposed to DDT (Hong et al. 1986; Hudson et al. 

1985; Pratt et al. 1985). 5-HIAA and MHPG are metabolites of the biogenic amines serotonin and 

norepinephrine, respectively, which act as inhibitors in the central nervous system. The increases in 

levels of 5-HIAA and MHPG indicate increased breakdown of these inhibitory neurotransmitters. Both 

aspartate and glutamate function as excitatory amino acids. The overall effect of increased levels of 

aspartate and glutamate is the induction of a state of increased excitability in the neurons of the central 

nervous system. This leads to hyperexcitability in the exposed animal and contributes to tremors and 

convulsions, which have been observed following DDT exposure. Serotonin, in addition to being active 

in the inhibition of neurons, plays an important role in the regulation of body temperature. It has been 

suggested that the increased turnover of serotonin in DDT-exposed organisms may be responsible for 

DDT-induced hyperthermia (Hudson et al. 1985). Other data suggest that DDT-induced hyperthermia 

may be due to heat generated by muscular movement during tremors or convulsions (Herr et al. 1986). 

Reproductive and Developmental Effects. DDT intake, particularly during sexual 

differentiation, can adversely affect the reproductive system of male animals. Such effects have been 

attributed to DDT and related compounds acting in any of the following manners or in any combination 

of them: (1) mimicking endogenous hormones, (2) antagonizing endogenous androgenic hormones, 

(3) altering the pattern of synthesis or metabolism of hormones, and (4) modifying hormone receptor 

levels. DDT is primarily suspected of influencing reproduction and development through its interaction 

with steroid hormones receptors for estrogens and androgens. 

Androgens and estrogens are very lipid soluble and thus, diffuse easily through the cell membrane into 

the cytosol and nucleus. Unoccupied steroid receptors bound to a complex of molecular chaperones are 

in dynamic equilibrium between the cytosol and the nucleus (DeFranco 1999). Once a steroid binds to its 

receptor and the hormone-receptor complex reaches the nucleus, it binds to hormone response elements in 

the enhancers, silencers, or promoters upstream of the genes controlled by the steroid in question. The 

hormone-receptor complex acts as a transcription factor to either stimulate or repress transcription of 

RNA from the steroid responsive gene. This RNA is spliced to form mRNA, which directs the synthesis 

of proteins that cause the characteristic responses to the steroid hormone. Some ancillary models for 
DDT, DDE, and DDD 163 

3. HEALTH EFFECTS 

immediate action of steroid hormones have recently been proposed and involve either cross-talk 

interactions with, or possibly direct binding to, other hormone and growth factor cell surface receptors, 

and cell surface receptor mediated uptake of serum carrier proteins bound to steroid hormones (Chen and 

Farese 1999). It was long thought that estrogen mediated its effects by binding to a single receptor, 

estrogen receptor α (ERα). Recently, however, a second estrogen receptor, ERβ, has been discovered 

(Kuiper et al. 1996; Mosselman et al. 1996). None of the studies examining the difference in estrogen 

receptor binding affinity for DDT isomers and metabolites differentiated between ERα and ERβ. It has 

not been fully established what effects are mediated by each estrogen receptor. However, there appears to 

be tissue-specific distribution of the two receptors (Kuiper et al. 1997), which may allow for tissuespecific effects by estrogens. In tissues where both receptors are expressed, ligand binding to the 

receptors results in heterodimer formation (an ERα receptor pairs up with an ERβ receptor) (Cowley et al. 

1997; Pace et al. 1997; Pettersson et al. 1997), which may result in different patterns of gene regulation 

than seen with homodimeric pairing (an ERα with an ERα or an ERβ with an ERβ). Additionally, each 

different estrogenic compound might act as an estrogen agonist at one receptor type and an estrogen 

antagonist at the other receptor type. 

Results from numerous studies using a wide range of experimental approaches suggest that binding to the 

estrogen receptor and subsequent events are the predominant mechanism by which estrogenic effects are 

expressed. Tests for estrogenicity fall into two general categories, in vivo and in vitro (Zacharewski 

1998). Examples of the former include the uterotrophic (increase in uterine wet weight) and vaginal cell 

cornification assays. In vitro assays include (i) measuring the activities of enzymes involved in steroid 

synthesis, (ii) competitive ligand binding assays using binding globulins and receptors, (iii) cell 

proliferation assays, and (iv) gene expression assays in mammalian cells and yeast. If possible, both in 

vivo and in vitro tests should be conducted since, for instance, in vitro assays can give false positives if 

the chemical is not absorbed or distributed to the target tissue or is rapidly metabolized. Conversely, false 

negatives may arise from the lack of an activation system in the in vitro system; other factors may play a 

role too. Also, both in vivo and in vitro tests are available for assessing androgenic effects such as 

measurement of anogenital distance in new born male rodents and androgen receptor binding assays in 

prostate cytosol, respectively. Some representative studies that examined mechanistic aspects of 

estrogenicity and antiandrogenicity of DDT and related compounds are briefly summarized below. 

Bulger and Kupfer (1983) reviewed the literature and references to most of the earlier studies on 

estrogenic effects can be found therein. 
 

DDT, DDE, and DDD 164 

3. HEALTH EFFECTS 

p,p’-DDT (200 mg/kg twice) and o,p’-DDD (200 mg/kg up to 5 times) administered intraperitoneally to 

rats had almost no estrogenic activity for initiating implantation, in contrast with responses observed after 

administration of o,p’-DDE (200 mg/kg once) and o,p’-DDT (200 mg/kg once) (Johnson et al. 1992). 

The latter two chemicals not only initiated implantation but also maintained pregnancy when 

administered repeatedly. p,p’-DDT was as efficient as o,p’-DDT in a uterine response assay in rats in 

vitro, but exhibited low estrogenic activity in vivo (Galand et al. 1987). Welch et al. (1969) reported an 

estrogenic activity ranking of o,p’-DDT > technical DDT > p,p’-DDT in immature female rats treated 

intraperitoneally. Singhal et al. (1970) showed that o,p’-DDT (100 mg/kg) was much more effective in 

increasing uterine weight in young rats than p,p’-DDT after a single intramuscular injection. Shelby et al. 

(1996) examined the estrogenicity of o,p’-DDT in three different assays; (1) competitive binding with the 

mouse uterine receptor, (2) transcriptional activation in HeLa cells transfected with plasmids containing 

an estrogen receptor and an estrogen response element linked to a reporter gene, and (3) the uterotropic 

assay in mouse. The first two assays are in vitro assays, whereas the third one was carried out by 

injecting immature mice subcutaneously with o,p’-DDT on three consecutive days and determining 

uterine weight on the fourth day. o,p’-DDT gave positive estrogenic responses in the three assays, but 

with a potency that was several orders of magnitude weaker than 17β-estradiol and diethylstilbestrol 

(DES). o,p’-DDT, o,p’-DDD, and p,p’-DDT were full estrogenic agonists in the in vitro E-screen test, 

p,p’-DDE and p,p’-DDD were partial agonists and technical DDT was a full agonist (Soto et al. 1997). 

The E-screen test uses breast cancer estrogen-sensitive MCF-7 cells, and the rationale of the assay is that 

a human serum-borne molecule specifically inhibits the proliferation of human estrogen-sensitive cells, 

and estrogens induce proliferation by neutralizing this inhibitory effect. 107

 times more o,p’-DDT, 

o,p’-DDD, p,p’-DDT, p,p’-DDE, p,p’-DDD, and technical DDT was needed to produce maximal cell 

yields than 17β-estradiol. p,p’-DDE also showed partial antagonistic effects on MCF-7 cell proliferation 

(Soto et al. 1998). 

Gaido et al. (1997) used a yeast system to express the estrogen receptor, a reporter gene regulated by two 

estrogen receptor response elements, and a yeast gene that is supposed to enhance steroid receptor 

mediated gene transcription. o,p’-DDT and o,p’-DDD appeared to stimulate estrogen receptor responsive 

genes at very high doses, but a complete curve of expression versus dose was not done for these two 

chemicals. The authors extrapolated EC50 (effective concentration for 50% maximal response) values for 

o,p’-DDT and o,p’-DDD assuming that these chemicals would have been as effective as estradiol in 

inducing reporter gene expression if a complete dose response experiment had been done. The calculated 

EC50 values were 1.8 mM for o,p’-DDT (8x106

 less potent than estradiol) and 3.32 mM for o,p’-DDD 

(15x107

 less potent than estradiol) compared to an actual estradiol EC50 of 2.25 nM. o,p’-DDE produced 
DDT, DDE, and DDD 165 

3. HEALTH EFFECTS 

little apparent response. Sohoni and Sumpter (1998) investigated the ability of o,p’-DDT and p,p’-DDE 

to induce estrogen receptor regulated gene expression in yeast expressing the human estrogen receptor 

and a secreted reporter gene controlled by an estrogen response element. A complete dose-response curve 

was generated for o,p’-DDT and it was found to be about 105

 less potent than 17β-estradiol in inducing 

estrogen regulated gene transcription. There was little response to p,p’-DDE even at concentrations as 

high as 0.5 mM. In a similar study using recombinant receptor-reporter gene assays in stably transfected 

MCF-7 and HeLa cells, both o,p’-DDE and p,p’-DDE showed weak estrogenic activity (Balaguer et al. 

1999). o,p’-DDE was slightly more potent than p,p’-DDE , but it was still approximately 106

 times less 

potent than 17β-estradiol on a molar basis. In an additional study using HeLa cells cotransfected with an 

expression vector containing ER and an RE-responsive chloramphenicol acetyltransferase reporter, 

p,p’-DDT, p,p’-DDE, and p,p’-DDD showed no appreciable transcriptional induction (Tully et al. 2000). 

The highest concentration tested was 10 μM; for comparison, 0.01 nM 17β-estradiol produced 

approximately a 3-fold induction over the vehicle. Moreover, when the chemicals were tested in binary 

combinations, the mixtures showed no additional estrogenicity (Tully et al. 2000). 

Several investigators have demonstrated that o,p’-DDT can compete with estradiol for binding to the 

estrogen receptor. An IC50 is the concentration at which 50% of the maximal inhibition of binding to the 

labeled standard (in this case, estradiol) is achieved. Danzo (1997) showed that o,p’-DDE (IC50=40 μM) 

competed with 17β-estradiol (IC50=2.7 nM) for binding to the estrogen receptor in rabbit uterine extracts 

and p,p’-DDT and p,p’-DDE were much less effective competitors. Kelce et al. (1995) showed in uterine 

extracts from immature rats that o,p’-DDE (IC50=5 μM) competed with 17β-estradiol (IC50=0.8 nM) for 

estrogen receptor binding and that p,p’-DDT, p,p’-DDE, and p,p’-DDD were relatively ineffective 

competitors (IC50>1,000 μM). The observations of the lack estrogenicity of the p,p’-isomers were 

consistent with results from earlier in vivo studies (Bitman and Cecil 1970; Gellert et al. 1972; Nelson 

1974). 

Kelce et al. (1995) characterized the binding of p,p’-DDT, p,p’-DDE, o,p’- DDT, and p,p’-DDD to the 

androgen and estrogen receptors in vitro in uterine cytosolic extracts from immature rats and in rat ventral 

prostate cytosol, respectively. In competitive androgen receptor binding assays using a radiolabeled 

synthetic androgen (R1881), the four chemicals showed dose-dependent competitive inhibition. 

p,p’-DDE was the greatest competitor with an inhibition constant (Ki

) of 3.5 μM, which was similar to 

that of DES and about 30 times weaker than 17β-estradiol. The other three isomers were 12–20-fold less 

effective than p,p’-DDE. Based on IC50 values for displacement of the synthetic androgen from the 

androgen receptor or 17β-estradiol from the estrogen receptor, o,p’-DDT was 20-fold more effective as a 
 

DDT, DDE, and DDD 166 

3. HEALTH EFFECTS 

competitor of estrogen binding to the estrogen receptor than androgen binding to the androgen receptor. 

p,p’-DDT, p,p’-DDD, and p,p’-DDE bound the androgen receptor 14, 11, and 200 times more effectively 

than the estrogen receptor, respectively. A control experiment was done to demonstrate that the 

antiandrogenic effects of the DDT isomers were not due to their ability to inhibit the conversion of 

testosterone by 5α-reductase to 5α-dihydrotestosterone and 5α-androstan-3α-17β-diol. In microsomes 

isolated from the adult rat caput and corpus epididymis, the four DDT isomers were all 20-fold less 

effective in inhibiting this reaction than DES or 17β-estradiol. Danzo (1997) also confirmed that 

p,p’-DDE (IC50=6.8 μM) could compete with dihydrotestosterone (IC50=1.1 nM) for binding to androgen 

receptors in rat prostate cytosol. In contrast to Kelce, Danzo found that o,p’-DDT and p,p’-DDT were 

only slightly less effective than p,p’-DDE in competing with dihydrotestosterone for androgen receptor 

binding. 

The binding experiments described above cannot distinguish between an agonist and an antagonist as 

both might bind equally well to the androgen receptor. p,p’-DDE’s action as an androgen receptor 

antagonist has been demonstrated by its ability to inhibit the androgen receptor from either appropriately 

inducing or repressing transcription from androgen responsive genes containing androgen receptor 

binding sites in their regulatory sequences. Kelce et al. (1995) transiently transvected monkey kidney 

cells with an androgen receptor expression vector and a reporter gene containing the MMTV promoter, 

which contains binding sites for the androgen receptor. In a 5-hour expression assay, both 0.2 μM 

p,p’-DDE and 1 μM hydroxyflutamide were able to inhibit 5α-dihydrotestosterone (0.1 nM) induced 

transcription by about 50%. In contrast, in lysates of rat ventral prostate, the androgen antagonist 

hydroxyflutamide was 10 times more effective than p,p’-DDE in inhibiting binding to the androgen 

receptor. In the transient tranvection assay, levels of 5α-dihydrotestosterone were not measured in the 

wells, so there was no control for whether the DDE could have affected its metabolism. p,p’-DDT and 

o,p’-DDT also showed some ability to inhibit the transcription of androgen responsive genes in this assay. 

Kelce et al. (1997) performed an in vivo experiment on castrated adult male rats, which were maintained 

on a constant dose of testosterone from implanted Silastic capsules; this experiment was designed to 

eliminate possible variations in testosterone levels from feedback at the hypothalamus or pituitary. After 

5 days gavage with 200 mg/kg/day of p,p’-DDE, northern blots of mRNA isolated from the prostate 

showed significant increases in steady state levels for mRNA for testosterone repressed prostatic message 

and decreases in testosterone induced C3 (third subunit of prostatic specific binding protein). The 

p,p’-DDE did not affect testosterone metabolism in these animals. Kelce et al. (1995) suggested that their 

findings raised the possibility that the androgen receptor, rather than the estrogen receptor, is the site of 

hormonal blockade by persistent environmental pollutants such as p,p’-DDE. 
DDT, DDE, and DDD 167 

3. HEALTH EFFECTS 

Other investigators have confirmed Kelce’s results showing the antiandrogenic effects of p,p’-DDE on 

androgen receptor regulated genes. Maness et al. (1998) investigated how isomers of DDT, DDE, and 

DDD inhibited androgen receptor regulated gene expression in vitro in HEPG2 human hepatoma cells 

transiently transvected with the human androgen receptor and a reporter gene linked to an androgen 

responsive promoter. These cell were exposed for 24 hours to a maximally inducing dose (0.1 μM) of 

dihydrotestosterone and various doses of either p,p’-DDT, p,p’-DDE, p,p’-DDD, o,p’-DDT, o,p’-DDE, or 

o,p’-DDD. Details of the data were not presented, but from the graphical presentation of the data, both 

isomers of DDT, DDE, and DDD appear to be equipotent (IC50.1–10 μM) in inhibiting dihydrotestosterone induced gene transcription. The authors state that p,p’-DDE was the most potent inhibitor 

(IC50=1.86 μM), but no evidence of statistical significance is discussed. No sampling of the cells or 

media was done to determine whether DDT isomers might have been affecting the metabolism of 

dihydrotestosterone in these experiments on liver cells. 

Sohoni and Sumpter (1998) investigated antiandrogenic effects in yeast expressing the human androgen 

receptor plus a secreted reporter gene controlled by androgen response elements. Both o,p’-DDT and 

p,p’-DDE were approximately equipotent in inhibiting dihydrotestosterone (DHT) responsive gene 

expression. Both were about as effective as the antiandrogen flutamide. From graphs, they appeared to 

have IC50 values on the order of 10 μM from 4-day incubations with a submaximally inducing dose of 

DHT (1.25 nM) and DDT or DDE; however, the authors felt that explicit calculations of IC50 values were 

inappropriate, because if incubation times were extended to 5 or 6 days, DHT seemed to partially 

overcome the inhibition by p,p’-DDE (o,p’-DDT was not tested in these tests). Again, no sampling of 

yeast media was done to determine if the DDT or DDE was affecting metabolism of DHT. 

Collectively, the results of these studies suggest that DDT-related compounds might have estrogenic or 

antiandrogenic activities if sufficient doses are used. The androgen antagonist mechanism demonstrated 

in these studies would explain a number of reproductive and developmental effects seen in male rats of 

various ages exposed to p,p’-DDE. These include reduced anogenital distance and retention of thoracic 

nipples in pups exposed during gestation and lactation (Kelce et al. 1995; Loeffler and Peterson 1999; 

You et al. 1998;); delayed puberty in rats exposed either during juvenile development (Kelce et al. 1995) 

or gestation and lactation (Loeffler and Peterson 1999), and reduced accessory sex organ weights in 

exposed adult males (Kelce et al. 1995, 1997). The experimental details of these studies are discussed 

extensively in Section 3.2.2.5 Reproductive Effects and Section 3.2.2.6 Developmental Effects. 
DDT, DDE, and DDD 168 

3. HEALTH EFFECTS 

Alterations in development and reproduction can also be induced by altering the metabolism of sex 

hormones. For example, in perinatally exposed rats, p,p’-DDE was shown to increase liver CYP enzymes 

systems that hydroxylate testosterone (You et al. 1999c). p,p’-DDE was also shown to increase liver 

aromatase activity in adult male rats (You et al. 2001); aromatase plays a critical role in steroidogenesis 

by catalyzing the conversion of C19 steroids to estrogens. 

DDT and metabolites apparently also exhibit additional estrogen or androgen receptor-independent 

effects as shown for example in a study by Juberg and Loch-Caruso (1992) in which the estrogenic 

o,p’-DDT increased contractions of rat uterus strips ex vivo, but 17β-estradiol did not. The estrogen 

antagonist, tamoxifen, failed to block the stimulatory effect of o,p’-DDT. Furthermore, the nonestrogenic 

DDT analogue, p,p’-DDD, significantly stimulated uterine contraction. The authors also demonstrated 

that the increase in contractility was not due to release of prostaglandin E2 (PGE2) from the uterine strips 

since PGE2 levels in the muscle bath showed no significant differences between control and DDT-treated 

strips. A more recent study from the same group directly showed that p,p’-DDD causes a dramatic 

increase in intracellular free calcium in rat myometrial smooth muscle cells (Juberg et al. 1995). This 

increase resulted primarily from an increase in calcium influx from the extracellular medium and release 

from intracellular calcium stores. The elevation in intracellular calcium is consistent with an increase in 

uterine contractility. 

Mechanistic Studies of DDT in Fish. There have been a number of intriguing mechanistic studies of 

DDT isomers and metabolites in fish that relate to reproductive and developmental effects. Two subtypes 

of nuclear androgen receptors have been identified in both kelp bass and Atlantic croaker; one subtype is 

expressed in the brain and the other in ovarian tissue (Sperry and Thomas 1999). The brain receptor 

subtype seems more similar to mammalian androgen receptors in binding characteristics; the ovarian 

receptor is unique to teleost fish. Both p,p’ and o,p’ isomers of DDT, DDD, and DDE can displace 

dihydrotestosterone from the ovarian androgen receptor subtype. Analogous to ERα and ERβ in humans, 

there are also different estrogen receptor subtypes in Atlantic croaker fish; two slightly different estrogen 

receptors have been identified in the testis and liver of these fish (Loomis and Thomas 1999). o,p’-DDT, 

as well as o,p’-DDE and o,p’-DDD, can bind to both of these receptors with low affinity 

(50,000–500,000 times lower ) relative to DES; binding of these isomers was somewhat better to the 

testicular receptor than to the hepatic receptor. There has been one report in Atlantic croaker that 

o,p’-DDD can competitively inhibit binding of maturation-inducing steroid to its plasma membrane 

receptor (Das and Thomas 1999; Thomas 1999). Similar to estrogen, o,p’-DDT can stimulate 
 

 

DDT, DDE, and DDD 169 

3. HEALTH EFFECTS 

gonadotropin release in Atlantic croaker (Khan and Thomas 1998). Another report reveals that o,p’-DDT 

impairs survival related behavior in larval offspring of exposed Atlantic croaker (Faulk et al. 1999). 

Metabolic Effects. DDT and DDT-analogues were found to be phenobarbital-type (CYP2B and to a 

lesser extent CYP3B) inducers in rat liver (Nims et al. 1998). Since some of these isozymes metabolize 

endogenous steroids and other hormones, it is possible DDT could increase the turnover or change serum 

levels of endogenous macromolecules metabolized by these P-450 isozymes. This would be consistent 

with the increases in microsomal catabolism of cortisol observed in DDT-exposed humans. These 

subjects excreted significantly increased amounts of the cortisol metabolite 17-hydroxycortisone in their 

urine (Nhachi and Loewenson 1989; Poland et al. 1970). Thus, hormonal homeostasis could be indirectly 

affected and effects would be seen on several systems including the reproductive system. Evidence that 

the latter may happen was provided by You et al. (1999c) who examined the potential of in utero DDE 

exposure to affect the developmental expression of hepatic CYPs enzymes responsible for testosterone 

hydroxylations. They treated pregnant Sprague-Dawley rats on gestation days 14–18 and also adult male 

rats with p,p’-DDE and found that the responses of CYP2C11 and CYP2A1 were development-regulated. 

DDE induced 2A1 in males on postnatal day (PND) 10 but not PND 21; pronounced induction of 2B1 

was seen in males and females on both PND 10 and 21 and to a lesser extent of 3A1; there was no 

induction of 2C11. In adult rats, DDE induced 2B1, 3A1, and 2C11, but not 2A1. Similar results, also in 

adult Sprague-Dawley rats, but with technical DDT were published by Sierra-Santoyo et al. (2000). 

Another enzyme involved in steroid metabolism that was affected by p,p’-DDE was aromatase (catalyzes 

the conversion of C19 steroids to estrogens), whose activity was significantly increased in hepatic 

microsomes from treated adult male rats, as were the levels of aromatase protein in the liver (You et al. 

2001). 

Hepatic Effects. Hepatic effects in animals associated with DDT, DDE, and DDD exposure include 

increased liver weights, hypertrophy, hyperplasia, microsomal enzyme and cytochrome P-450 induction 

(CYP2B, phenobarbital-type induction), cell necrosis, and increased levels of SGPT (ALT) and SGOT 

(AST) enzymes released from damaged liver cells. DDT and its metabolites have been reported to disrupt 

the ultrastructure of mitochondrial membranes (Byczkowski 1977). This disruption may result in cell 

damage and some cell death. Consequently, the liver could begin cell regeneration to compensate for this 

loss of cells. This regenerative process often leads to hyperplasia and hypertrophy, which combine to 

increase the weight of the liver, and may contribute to the promotion of liver tumors (Fitzhugh and 

Nelson 1947; Schulte-Hermann 1974). 
 

DDT, DDE, and DDD 170 

3. HEALTH EFFECTS 

Adrenal Gland Effects. The adrenal gland consists of the adrenal medulla, which secretes 

epinephrine and norepinephrine, and the adrenal cortex, which is composed of three zones. These zones 

are: (1) the zona glomerulosa, producing mineralocorticoids involved in controlling electrolyte 

homeostasis; (2) the zona fasciculate, the primary source of glucocorticoids involved in controlling 

carbohydrate metabolism and stress responses, and downregulating some immune responses; and (3) the 

zona reticularis, the primary adrenal source of very small quantities of sex steroids (Ross and Reith 

1985). o,p’-DDD has been used to treat adrenocortical carcinoma and Cushing’s disease in humans for 

almost 4 decades (Bergenstal et al. 1960; Wooten and King 1993). The therapeutic action is based on the 

induction of selective necrosis of the zona fasciculata and zona reticularis of the adrenal cortex and 

inhibition of cortisol synthesis. The latter is accomplished by inhibiting the intramitochondrial 

conversion of pregnenolone and the conversion of 11-deoxycortisol to cortisol (Hart and Straw 1971; 

Hart et al. 1971). Damage to the adrenal gland was first observed in dogs given DDD orally in earlier 

studies by Nelson and Woodard (1949). Many additional studies in dogs have confirmed the original 

observations after acute and long-term administration of o,p’-DDD to the animals (Kirk and Jensen 1975; 

Kirk et al. 1974; Powers et al. 1974). Studies in vitro have shown a high correlation between 

adrenocorticolytic activity and metabolic activation by adrenocortical mitochondria. Dog adrenal 

mitochondria had significantly greater metabolism of o,p’-DDD and covalent binding activities than did 

adrenal mitochondria from rabbits, rats, or guinea pigs (Martz and Straw 1980). Levels of metabolism 

and covalent binding measured in human adrenal mitochondria were intermediate between levels 

measured in dogs and rabbits and those measured in rats and guinea pigs (Martz and Straw 1980). A 

more recent study with dog adrenal cortical homogenates showed that the majority of the 14C-o,p’-DDDderived radioactivity was covalently bound to proteins, and that no radioactivity was associated with 

DNA (Cai et al. 1995). The rank order of species regarding metabolism and protein binding was cow > 

dog > rat adrenal homogenates > human normal adrenal or tumor homogenates. According to Cai et al. 

(1995), their results are consistent with an acyl chloride being the reactive intermediate. This reactive 

intermediate is thought to be formed by hydroxylation of o,p’-DDD at the β-carbon and quick 

transformation by dehydrochlorination into the acyl chloride. The latter either binds covalently to 

nucleophiles, or by losing water, is transformed to DDA for renal excretion (Schteingart 2000). 

As previously mentioned in Section 3.5.3 Metabolism, methylsulfonyl metabolites of DDT 

(3-MeSO2-DDE and 2-MeSO2-DDE; all studies have been conducted with the p,p’-isomers, and 

therefore, the chlorine position is not specified in the abbreviation) have been identified in humans and 

animals (Haraguchi et al. 1989; Norén et al. 1996; Weistrand and Norén 1997). 3-MeSO2-DDE is a 

potent toxicant in the adrenal cortex, particularly of mice, and possibly in humans after local activation by 
DDT, DDE, and DDD 171 

3. HEALTH EFFECTS 

cytochrome P-450. For example, whole body autoradiography of 14C-labeled 3-MeSO2-DDE in female 

mice injected with the test material intravenously showed heavy accumulation of radioactivity confined to 

the zona fasciculata of the adrenal cortex (Lund et al. 1988). Morphological evaluation of the adrenals 

showed extensive vacuolation and necrosis of the zona fasciculata 1–12 days after a single dose of 

25 mg/kg; degenerative changes were seen at 12.5 mg/kg. Studies in vitro showed a dose- and timedependent covalent irreversible binding to protein and formation of water soluble metabolites (Lund et al. 

1988). This was blocked by the cytochrome P-450 inhibitor, metyrapone, and the addition of reduced 

glutathione decreased binding. A subsequent study in mice administered 3-MeSO2-DDE intraperitoneally 

identified mitochondria in the zona fasciculata as the primary targets for 3-MeSO2-DDE-induced toxicity 

(Jönsson et al. 1991). Disorganization and disappearance of central cristae were observed as soon as 

6 hours after a dose of 3 mg/kg. Higher doses showed mitochondrial vacuolization, followed by 

disappearance of mitochondria or cellular necrosis. Jönsson et al. (1991) further suggested that of the two 

cytochrome P-450s expressed predominantly in the adrenal cortex, P-45011β and P-450SCC, the former 

(which converts 11-deoxycorticosterone to cortisone) is involved in the metabolic activation of 

3-MeSO2-DDE in mice. P-450SCC is also expressed in the ovary and testes, but no binding or toxicity 

was seen in these organs. Lund and Lund (1995) later provided direct evidence for an involvement of 

adrenocortical mitochondrial P-450s in the bioactivation of 3-MeSO2-DDE. Intraperitoneal 

administration of radioactive 3-MeSO2-DDE to pregnant or lactating mice resulted in specific 

accumulation and binding of 3-MeSO2-DDE-derived radioactivity in the zona fasciculata of adrenals 

from 16- to 18-day-old fetuses or suckling pups, showing ready transplacental passage of the metabolite 

and transfer via maternal milk (Jönsson et al. 1992). The results also suggested that cytochrome 

P-45011β was involved in the activation of 3-MeSO2-DDE in the fetal adrenal cortex. Quantitative 

measurements showed that 7 days after dosing, the amount of radioactive label of the pups adrenals was 

2 and 3.6 times higher than maternal adrenals at the 1.5 and 25 mg/kg dose levels, respectively, 

suggesting that mother’s milk may be an important route of exposure of 3-MeSO2-DDE in DDT-exposed 

animals. A later study showed that P-45011β seems to be expressed during gestation days 10–12 in the 

adrenal cortex in the mouse fetus (Jönsson et al. 1995). 3-MeSO2-DDE was also found to reduce the 

capacity of pups and maternal adrenals to secrete corticosterone (Jönsson 1994) by a mechanism possibly 

involving competitive inhibition of adrenocortical CYP11B1 (Johansson et al. 1998). Human adrenal 

glands in vitro were found also to bioactivate 3-MeSO2-DDE to a metabolite that bound irreversibly to 

mitochondria (Jönsson and Lund 1994). Binding was inhibited by metyrapone indicating the involvement 

of cytochrome P-450, but the specific isozyme was not elucidated. The bioactivation and mitochondrial 

toxicity of 3-MeSO2-DDE observed in vivo has recently been reproduced in an ex vivo tissue-slice culture 

preparation (Lindhe et al. 2001). 
 

DDT, DDE, and DDD 172 

3. HEALTH EFFECTS 

Comparative studies by Brandt et al. (1992) showed that adrenal cortical cytochrome P-450 from mink 

and otter metabolically activate and bind both p,p’-DDD and o,p’-DDD, whereas 3-MeSO2-DDE is not 

activated by these species in vitro. Both isomers induced necrosis and focal bleeding in the zona 

fascicularis/zona reticularis in vivo in mink. Adrenals from chickens also activated and covalently bound 

3-MeSO2-DDE and o,p’-DDD in vitro, and both compounds were adrenotoxic in vivo. 

Cancer. Although the evidence regarding the carcinogenicity of DDT in humans is inconclusive, DDT 

and related compounds have been shown to be carcinogenic in some laboratory animals. Several 

mechanisms have been proposed: direct mutagenicity through formation of covalent DNA adducts; 

promotion of pre-existing abnormal cells; or cytotoxicity leading to hyperplasia and promotional tumor 

development. These mechanisms may not be mutually exclusive. DDMU epoxide and DDA-Cl are two 

electrophiles that potentially can be produced during DDT metabolism. These could form covalent 

adducts with DNA, and thereby contribute to the cytotoxicity as well as the carcinogenicity of DDT. 

DDMU epoxide is a minor side-product by both of the predominant models (I and II) for DDT 

biotransformation. DDA-Cl is a major DDT metabolite in the Model II catabolism scheme (Gold and 

Brunk 1983). Electrophiles, as a class, are frequently mutagenic. Consistent with this hypothesis, 

DDMU-epoxide has been found to be mutagenic in the S. typhimurium Ames assay (Gold and Brunk 

1983). DDA-Cl is not known to have been similarly assayed, so there is no direct biological evidence of 

its mutagenicity. There are apparently species differences in carcinogenicity of DDT. In the hamster and 

mouse, metabolism of DDT, DDD, and DDMU is similar (Gold and Brunk 1983). Therefore, it is 

unlikely that the observed differences in response between the mouse and hamster are due to relative 

differences in the production of DDMU-epoxide or DDA-Cl. There is a species difference in the 

conversion of DDT to DDE in that the hamster was less efficient than the mouse in the conversion of 

DDT to DDE (Gingell 1976; Gold and Brunk 1983). 

A possible mechanism of carcinogenicity may be a phenobarbital-like promotion of pre-existing abnormal 

cells that develop into types of tumors commonly seen in aging rats (Williams and Weisburger 1991). 

DDT has demonstrated promoting activity in several initiation-promotion assays (see Section 3.10). 

Mechanisms by which DDT promotes tumors may include direct liver injury leading to cell death and cell 

regeneration that could subsequently result in excess cell proliferation, hyperplasia, and tumor 

development (Schulte-Herman 1974); inhibition of apoptosis, which is programmed cell death 

characterized by DNA fragmentation (Wright et al. 1994); and reduction of gap junctional intercellular 

communication (GJIC) (Ren et al. 1998). DDT reduced DNA fragmentation in seven different cell types 

with chemical- or UV-induced apoptosis (Wright et al. 1994), supporting the hypothesis that DDT 
DDT, DDE, and DDD 173 

3. HEALTH EFFECTS 

promotes tumors in part by inhibiting the death of initiated cells. Another possible mechanism of 

carcinogenicity may be hormonal promotion of initiated cells via estrogenic action. The estrogenic 

activity of DDT isomers is discussed above under the subheading Reproductive and Developmental 

Effects. 

A tumor promoter may reduce GJIC by reducing gap junction channel permeability, gap junction number, 

or connexin expression (Ren et al. 1998), by increasing Ca+2 concentration in the cell (Adenuga et al. 

1992), or by disrupting gap junction plaque stability by producing reactive oxygen radicals that affect 

plasma membrane fluidity (Jansen and Jongen 1996). The effects of a specific tumor promoter are 

frequently cell- or connexin-specific. DDT has induced GJIC inhibition in vitro in rat liver epithelial cells 

(Guan and Ruch 1996; Ruch et al. 1994), primary hepatocytes (Budunova et al. 1993), Syrian hamster 

embryo cells (Rivedal et al. 1994, 2000; Roseng et al. 1994), rat precision-cut liver slices (de Graaf et al. 

2000), mouse skin cells (Jansen and Jongen 1996), bovine oviductal cells (Tiemann and Pöhland 1999), 

human breast cells (Kang et al. 1996), and human urothelial cells (Morimoto 1996), and in vivo in rat 

liver cells (Tateno et al. 1994). Mechanisms and determinant factors of DDT-induced GJIC inhibition 

have been investigated extensively. 

Using several combinations of cell lines and connexin types, Ren et al. (1998) showed that DDT-induced 

GJIC inhibition was more consistent with a cell type-dependent mechanism of GJIC inhibition than a 

connexin-dependent mechanism. 

In “normal” human breast epithelial cells that did not express the estrogen receptor, GJIC inhibition was 

strongly DDT dose-dependent after a 90-minute in vitro exposure to $25 μM (Kang et al. 1996). Levels 

of phosphorylated connexin43 were significantly reduced in cells with DDT-induced GJIC inhibition, 

suggesting that intercellular communication in human breast tissue may be modulated by 

dephosphorylation of the connexin proteins. The same levels of DDT did not alter steady-state levels of 

connexin43 mRNA, suggesting that DDT alters connexin43 proteins after translation, rather than during 

transcription (Kang et al. 1996). Human urothelial cell lines were susceptible to DDT-induced GJIC 

inhibition at DDT concentrations that were not cytotoxic for up to 7 days (Morimoto 1996). For 48-hour 

exposures, the degree of inhibition was not dose-dependent, but dose dependence was evident in 96-hour 

exposures. 

Jansen and Jongen (1996) investigated whether tumor stage affected the degree of DDT-induced GJIC 

inhibition in mouse skin cells by conducting parallel fluorescent dye transfer assays in primary mouse 
DDT, DDE, and DDD 174 

3. HEALTH EFFECTS 

keratinocytes, DMBA-initiated mouse epidermal 3PC cells, and mouse skin carcinoma-derived CA3/7 

cells. The investigators found that the order of GJIC percent inhibition in cell types was primary 

keratinocytes > initiated 3PC cells > carcinoma-derived cells, where percent inhibition of GJIC was 

measured by comparison to cell type-specific controls. DDT-induced GJIC inhibition in the 3PC initiated 

cell line was both dose- and time-dependent within a 90-minute exposure. GJIC inhibition by DDT was 

also dose-dependent in liver cells of rats orally administered DDT in vivo for 2 weeks (Tateno et al. 

1994). The number and size of immunostained connexin32 spots were reduced in liver cells of rats 

administered $25 mg/kg/day, while the area of dye spread was significantly reduced only at 

50 mg/kg/day. The time course of GJIC inhibition showed significantly reduced area of dye spread after 

1, 2, and 4 weeks of exposure, with recovery to control levels at 6 weeks. 

As mentioned in Section 3.2.2.8, numerous studies have examined the role of environmental estrogens, 

DDT and related compounds among them, as possible contributors to the increased incidence of breast 

cancer. Several mechanisms for DDT-induced increased breast cancer risk have been proposed including: 

(1) by binding to the ER, which induces a series of biochemical reactions ultimately resulting in protein 

synthesis and cell proliferation in estrogen-sensitive tissues including the breast; (2) by shifting the 

balance of endogenous 17β-estradiol metabolites in favor of the genotoxic 16α-hydroxyestrone, at the 

expense of the nongenotoxic 2-hydroxyestrone; and (3) by mimicking epidermal growth factor in human 

mammary epithelial cells, thus stimulating cell proliferation. 

In addition to the studies that characterized the binding of DDT to the estrogen receptor already 

mentioned in the discussion of reproductive effects, many additional similar studies have specifically used 

breast cancer cells to demonstrate the involvement of the estrogen receptor in breast cancer cell 

proliferation (Dees et al. 1996, 1997a, 1997b; Shekhar et al. 1997; Zava et al. 1997). These and other 

studies clearly demonstrated that DDT analogues increase cell proliferation in human breast cancer cells 

by an ER-mediated mechanism that can be blocked by antiestrogens, and that their estrogenic potencies 

are several orders of magnitude lower than that of 17β-estradiol. 

Another mechanism by which DDT compounds may increase cancer risk is by increasing the production 

of genotoxic metabolites from the breakdown of 17β-estradiol, at the expense of producing nongenotoxic 

metabolites. Two 17β-estradiol metabolites that have been studied extensively are 16α-hydroxyestrone 

(16α-OHE1) and 2-hydroxysterone (2-OHE1), which are formed via two mutually exclusive metabolic 

pathways. 16α-OHE1 binds covalently to the estrogen receptor in human breast cancer cells (Swaneck 

and Fishman 1988), and has induced genetic and proliferative changes in mouse mammary epithelial cells 
DDT, DDE, and DDD 175 

3. HEALTH EFFECTS 

(Telang et al. 1992). 2-OHE1, in contrast, inhibited both basal and estradiol-induced growth of MCF-7 

human breast cancer cells, acting as an anticarcinogen (Schneider et al. 1984). Several studies addressed 

the hypothesis that the balance of the two metabolites, expressed as the ratio 16α-OHE1/2-OHE1, may be 

predictive of xenoestrogen-induced mammary cancer in humans; as the ratio of 16α-OHE1/2-OHE1 

increases, so does the risk of breast cancer. 

The ratio of estradiol metabolites 16α-OHE1/2-OHE1 was increased over controls in cultures of 

ER-positive MCF-7 human breast cancer cells by o,p’-DDT, o,p’-DDE, and p,p’-DDE (Bradlow et al. 

1995). McDougal and Safe (1998) also found that the 16α-OHE1/2-OHE1 ratio in a MCF-7 human 

cancer cell assay was elevated when the cells were exposed to o,p’-DDE, and E2 2-hydroxylation was 

reduced compared to controls in cells exposed to o,p’-DDE and o,p’-DDT; however, the ratio did not 

consistently predict known mammary carcinogens in the same assay. In ER-negative MCF-10 and 

MDA-MB-231 cell lines, on the other hand, neither o,p’-DDE nor o,p’-DDT induced significant changes 

in either C2- or 16α-hydroxylation of estradiol (Bradlow et al. 1997), suggesting an estrogenic 

mechanistic pathway. In the MCF-7 cell line, C2-hydroxylation was reduced to 67% of the control level 

by o,p’-DDT, but addition of the antiestrogen indole-3-carbinol blocked the effect, again suggesting that 

the effect is ER-mediated. 

The 16α-OHE1/2-OHE1 ratio hypothesis is controversial, however. Epidemiology studies of the 

association of 16α-OHE1 with breast cancer have provided mixed results (Safe 1998). In addition, 

although 16α-OHE1 induced cell proliferation, a closely related compound, 16α-OHE2, did not 

(Schneider et al. 1984, as cited in Safe 1998). Also, while 2-OHE1 inhibited cell growth, the isomer 

2-OHE2 enhanced cell proliferation in MCF-7 human breast cancer cells (Schneider et al. 1984, as cited 

in Safe 1998). Safe (1998) did not elaborate on the relative metabolic production of the different isomers. 

A third mechanism by which DDT compounds may increase the risk of breast cancer is by producing 

proliferative changes in ER-negative breast cells via a nonestrogenic pathway by mimicking epithelial 

growth factor (EGF). p,p’-DDT stimulated two growth factor receptors associated with malignant breast 

lesions and increased STATS-mediated gene expression, ultimately increasing proliferation of 

untransformed ER-negative MCF-10A human breast epithelial cells (Shen and Novak 1997a). The study 

results indicated that p,p’-DDT may modulate breast cell proliferation by multiple mechanisms not 

regulated by the ER, while the o,p’-isomer was inactive in this test system. 
DDT, DDE, and DDD 176 

3. HEALTH EFFECTS 

When endogenous growth factors bind to growth factor receptors that are receptor tyrosine kinases, a 

chain of biochemical reactions is initiated that leads to transcription of genes involved in cellular 

proliferation and differentiation (Shen and Novak 1997b). p,p’-DDT mimicked EGF in human mammary 

epithelial cells by increasing EGF receptor tyrosine kinase activity at 0.01 and 0.1 μM concentrations, but 

appeared to impede the activity at 1 and 10 μM concentrations (Shen and Novak 1997b). 

3.6.3 Animal-to-Human Extrapolations 

In humans and in animals, acute high doses of DDT primarily affect the nervous system. However, it is 

uncertain whether effects on other systems and/or organs seen in animals exposed to lower doses for 

prolonged periods of time would also manifest in humans exposed under similar exposure conditions. In 

general, the metabolism of DDT in animals is similar to that in humans; however, there are also 

differences between species, and within species, differences between tissues. Comparisons of elimination 

rates of DDT from fat showed that the process is faster in rats followed by dogs and monkeys and slowest 

in humans (Morgan and Roan 1974). Rats eliminate DDT 10 to 100 times faster than humans. Morgan 

and Roan (1974) suggested that the differences in elimination rats could be due to differences in liver 

metabolism, gut bacterial metabolism, enterohepatic recirculation, or factors related to the accessibility of 

plasma-transported pesticide to the excretory cells of the liver. For specific effects, such as adrenal gland 

toxicity, in vivo and in vitro susceptibility to o,p’-DDD and 3-methylsulfonyl-p,p’-DDE varied among 

animal species (Brandt et al. 1992), and human adrenals in vitro showed potentially different 

susceptibility than animal glands in vitro. This is most likely due to differences in metabolism and 

covalent binding activities. Thus far, the endocrine-disrupting activity of DDT and analogues observed in 

experimental animals and in wildlife has not been observed in humans exposed to DDT. 

3.7 TOXICITIES MEDIATED THROUGH THE NEUROENDOCRINE AXIS 

Recently, attention has focused on the potential hazardous effects of certain chemicals on the endocrine 

system because of the ability of these chemicals to mimic or block endogenous hormones. Chemicals 

with this type of activity are most commonly referred to as endocrine disruptors. However, appropriate 

terminology to describe such effects remains controversial. The terminology endocrine disruptors, 

initially used by Colborn and Thomas (1992) and again by Colborn (1993), was also used in 1996 when 

Congress mandated the Environmental Protection Agency (EPA) to develop a screening program for 

“...certain substances [which] may have an effect produced by a naturally occurring estrogen, or other 

such endocrine effect[s]...”. To meet this mandate, EPA convened a panel called the Endocrine 
DDT, DDE, and DDD 177 

3. HEALTH EFFECTS 

Disruptors Screening and Testing Advisory Committee (EDSTAC), which in 1998 completed its 

deliberations and made recommendations to EPA concerning endocrine disruptors. In 1999, the National 

Academy of Sciences released a report that referred to these same types of chemicals as hormonally 

active agents. The terminology endocrine modulators has also been used to convey the fact that effects 

caused by such chemicals may not necessarily be adverse. Many scientists agree that chemicals with the 

ability to disrupt or modulate the endocrine system are a potential threat to the health of humans, aquatic 

animals, and wildlife. However, others think that endocrine-active chemicals do not pose a significant 

health risk, particularly in view of the fact that hormone mimics exist in the natural environment. 

Examples of natural hormone mimics are the isoflavinoid phytoestrogens (Adlercreutz 1995; Livingston 

1978; Mayr et al. 1992). These chemicals are derived from plants and are similar in structure and action 

to endogenous estrogen. Although the public health significance and descriptive terminology of 

substances capable of affecting the endocrine system remains controversial, scientists agree that these 

chemicals may affect the synthesis, secretion, transport, binding, action, or elimination of natural 

hormones in the body responsible for maintaining homeostasis, reproduction, development, and/or 

behavior (EPA 1997). Stated differently, such compounds may cause toxicities that are mediated through 

the neuroendocrine axis. As a result, these chemicals may play a role in altering, for example, metabolic, 

reproductive, growth, immune, and neurobehavioral functions. Such chemicals are also thought to be 

involved in inducing breast, testicular, and prostate cancers, as well as endometriosis (Berger 1994; 

Giwercman et al. 1993; Hoel et al. 1992). 

Numerous reviews have been written regarding the endocrine disrupting capabilities of DDT and related 

compounds (Bulger and Kupfer 1983; Chapin et al. 1996; Colborn et al. 1993; Crisp et al.1998; Gillesby 

and Zacharewski 1998; Golden et al. 1998; Gray et al. 1997; and many others). Much of the historical 

information summarized below, particularly the early data, has been extracted from these reviews. The 

focus of this section is on effects in humans and in animal species traditionally used in laboratory 

experiments; information on endocrine effects, as well as other health effects, on wildlife is presented in 

Appendix D, Health Effects in Wildlife Potentially Relevant to Human Health. 

It has been known for several decades that DDT and related compounds have weak estrogenic action in 

experimental animals and wildlife, but there is insufficient information to show that these chemicals have 

estrogenic action in humans. Estrogen influences the growth, differentiation, and functioning of many 

target tissues, including male and female reproductive systems such as mammary gland, uterus, vagina, 

ovary, testes, epididymis, and prostate. Several studies in humans have examined possible associations 

between body burdens of DDT and analogues and the incidence of alterations in these systems and tissues 
DDT, DDE, and DDD 178 

3. HEALTH EFFECTS 

and in other relevant end points; the results have been mixed. For instance, no association could be 

shown between the incidence of endometriosis and plasma concentrations of DDT (DDT plus DDE) in a 

case-control study by Lebel et al. (1998). Other studies reported that levels of DDT, DDE, and DDD 

were higher in maternal blood and in placental tissue in mothers who gave birth to premature infants or 

who spontaneously aborted fetuses compared to mothers who gave birth to full-term infants (Saxena et al. 

1980, 1981; Wassermann et al. 1982). However, other potentially endocrine-disrupting chemicals, such 

as PCBs and other chlorinated pesticides, were also increased in the maternal blood of these subjects, and 

the specific contribution of DDT, DDE, or DDD could not be determined. In contrast, a similar study did 

not show the same association between serum DDE levels and pre-term delivery (Berkowitz et al. 1996). 

An earlier study by Procianoy and Schvartsman (1981) found no differences in DDT levels in maternal 

blood at delivery full-term babies and cases of pre-term delivery; however, they found that umbilical cord 

blood from pre-term newborns had higher DDT than full-term newborns. A recent study of 361 pre-term 

infants and 221 small-for-gestational-age cases found that the odds for both outcomes increased steadily 

with increasing maternal blood concentrations of DDE in samples taken in the third trimester of 

pregnancy (Longnecker et al. 2001). The association was evident at DDE concentration $10 ppb, but 

there was essentially no relation at lower DDE concentrations. With regard to cancers of the reproductive 

tissues, no association was found between risk of endometrial cancer and lipid-corrected blood serum 

concentrations p,p’-DDE and o,p’-DDT in a multicenter case-control study of women in the United 

States, but there was a slight increased risk associated with p,p’-DDT (Sturgeon et al. 1998). Also, there 

was no association found between p,p’-DDE concentration in subcutaneous fat and incidence of prostate 

and testicular cancer mortality (Cocco and Benichou 1998) or between DDE and death from uterine 

cancer in a study of women in the United States (Cocco et al. 2000). 

Numerous studies have been conducted in the United States and elsewhere that evaluate the association 

between DDT and related compounds and breast cancer. Although a few studies have found positive 

associations among some subgroups of women (Aronson et al. 2000; Dewailly et al. 1994; Falck et al. 

1992; Güttes et al. 1998; Romieu et al. 2000; Wasserman et al. 1976; Wolff et al. 1993), others have 

found no association between these chemicals and breast cancer (Demers et al. 2000; Dorgan et al. 1999; 

Helzlsouer et al. 1999; Laden et al. 2001a; Krieger et al. 1994; Liljegren et al. 1998; Lopez-Carrillo et al. 

1997; Mendonca et al. 1999; Moysich et al. 1998; Schecter et al. 1997; Unger et al. 1984; van’t Veer et 

al. 1997; Ward et al. 2000; Wolff et al. 2000a, 2000b; Zheng et al. 1999, 2000). A recently published 

combined analysis of five U.S. studies that included 1,400 cases and 1,642 controls found no evidence 

linking breast cancer and DDE body burdens (Laden et al. 2001a). Details of the individual studies about 

breast, endometrial, testicular, and prostate cancer are presented in Section 3.2.2.8 Cancer. 
DDT, DDE, and DDD 179 

3. HEALTH EFFECTS 

Early studies in experimental animals administered the chemicals orally or by parenteral routes, whereas 

in recent years, much research has focused on elucidating the mechanisms of action involved using in 

vitro test systems. Such tests can be used to assay for estrogenic, antiestrogenic, androgenic, and 

antiandrogenic activity. In general, results from in vivo and in vitro studies indicate that DDT and 

analogues have much lower estrogenic potency than the endogenous hormone, 17β-estradiol. 

Experiments conducted by Gellert et al. (1972) showed that young rats administered 5 mg 

o,p’-DDT/kg/day for 27 days had delayed vaginal opening and increased uterine and ovarian weights. In 

another study, using the uterine glycogen response assay, Bitman and Cecil (1970) observed that 

o,p’-DDT was the most potent among several isomers; o,p’-DDE and p,p’-DDT were approximately 

16 times less active than o,p’-DDT and o,p’-DDD was inactive. Clement and Okey (1972) showed that 

o,p’-DDT (200 ppm in the diet) induced premature vaginal opening in Wistar rats. An in vivo assay with 

rats and mink showed that o,p’-DDT had uterotropic activity, whereas p,p’-DDT had only slight activity, 

and the activity of technical DDT was dependent on the level of o,p’-DDT that it contained (Duby et al. 

1971). However, no reproductive effects were observed in two successive generations of rats fed 

technical DDT, p,p’-DDT, or o,p’-DDT (Duby et al. 1971). Many other in vivo studies have shown the 

estrogenic potential of DDT and related analogues (Gellert and Heinrichs 1975; Gellert et al. 1974; 

Heinrichs et al. 1971; Welch et al. 1969). 

More recent in vitro experiments have examined binding of DDT, DDE, and DDD isomers to estrogen 

and androgen receptors and subsequent steroid regulated gene transcription. The experimental details of 

these studies are discussed in Section 3.6.2, Mechanisms of Toxicity: Reproductive and Developmental 

Effects. Although the results have varied somewhat between different investigators, in general, o,p’-DDT 

and o,p’-DDE appear to act as weak estrogen agonists, while p,p’-DDE can function as an androgen 

antagonist. In particular, p,p’-DDE functions as an antagonist after it has bound to the androgen receptor 

(Kelce et al. 1995). This androgen antagonism or antiandrogenic activity can explain a number of 

reproductive and developmental effects seen in male rats of various ages exposed to p,p’-DDE. These 

effects include reduced anogenital distance and retention of thoracic nipples in pups exposed during 

gestation and lactation (Gray et al. 1999; Kelce et al. 1995; Loeffler and Peterson 1999; You et al. 1998); 

delayed puberty in rats exposed either during juvenile development (Kelce et al. 1995) or at very high 

doses during gestation and lactation (Loeffler and Peterson 1999); and reduced accessory sex organ 

weights in exposed adult males (Gray et al. 1999; Kelce et al. 1995, 1997; You et al. 1999a). The 

experimental details of those studies are discussed extensively in Section 3.2.2.5, Reproductive Effects 

and Section 3.2.2.6, Developmental Effects. 
DDT, DDE, and DDD 180 

3. HEALTH EFFECTS 

The results of these and other studies suggest that DDT analogues can produce both agonistic and 

antagonistic responses by interfering with the binding of endogenous androgens and estrogens to their 

receptors. The type and magnitude of the response varied between assays and was dependent on the 

concentration of test material, as well as the temporal aspects of the exposure. Reviews of published data 

suggest that the amount of naturally occurring estrogens ingested daily through a normal diet may be far 

greater than the daily intake of estrogenic pesticides (Safe 1995). In addition, many experiments 

(Balaguer et al. 1999; Danzo 1997; Gaido et al. 1997; Kelce et al. 1995; Shelby et al. 1996; Soto et al. 

1997) show that DDT isomers and metabolites have orders of magnitude less estrogenic activity than 

17β-estradiol. This does not imply a lack of related risk, since estrogenic pesticides such as DDT and 

analogues bioconcentrate in the food chain and accumulate in the body and many have additive effects. 

Moreover, key endocrine processes can be profoundly affected by exposure to extremely small amounts 

of active chemicals during critical windows of embryonic, fetal, and neonatal development. As for 

antiandrogenic effects, Kelce et al. (1995) stated that the concentration of p,p’-DDE required to inhibit 

androgen receptor transcriptional activity in cell culture (64 ppb) is less than levels that accumulate from 

the environment such as in eggs of demasculinized male alligators in Florida’s Lake Apopka (5,800 ppb) 

and in humans in areas where DDT remains in use or is present in contaminated ecosystems. 

The adrenal gland may also be the target of some DDT isomers and metabolites. The metabolism of DDT 

can produce methylsulfonyl metabolites, such as methylsulfonyl-DDE, which are potent adrenal toxicants 

(Bakke et al. 1982; Brandt et al. 1992; Preston et al. 1984). o,p’-DDD is used therapeutically to treat 

adrenocortical carcinomas in humans. Furthermore, in some wild birds, but not others, adrenal weights or 

cortico-medullary ratios were affected by p,p’-DDT, p,p’-DDE, or technical-grade DDT (Hurst et al. 

1974; Jefferies and French 1972; Jefferies et al. 1971; Lehman et al. 1974; Peterleef et al. 1973). Adrenal 

toxicity is discussed more extensively in Section 3.6.2, Mechanisms of Toxicity. 

3.8 CHILDREN’S SUSCEPTIBILITY 

This section discusses potential health effects from exposures during the period from conception to 

maturity at 18 years of age in humans, when all biological systems will have fully developed. Potential 

effects on offspring resulting from exposures of parental germ cells are considered, as well as any indirect 

effects on the fetus and neonate resulting from maternal exposure during gestation and lactation. 

Relevant animal and in vitro models are also discussed. 
DDT, DDE, and DDD 181 

3. HEALTH EFFECTS 

Children are not small adults. They differ from adults in their exposures and may differ in their 

susceptibility to hazardous chemicals. Children’s unique physiology and behavior can influence the 

extent of their exposure. Exposures of children are discussed in Section 6.6 Exposures of Children. 

Children are particularly susceptible to hazardous chemicals due to the fact that many physiological 

systems, such as the nervous system, are being organized during the childhood years. 

Children sometimes differ from adults in their susceptibility to hazardous chemicals, but whether there is 

a difference depends on the chemical (Guzelian et al. 1992; NRC 1993). Children may be more or less 

susceptible than adults to health effects, and the relationship may change with developmental age 

(Guzelian et al. 1992; NRC 1993). Vulnerability often depends on developmental stage. There are 

critical periods of structural and functional development during both prenatal and postnatal life and a 

particular structure or function will be most sensitive to disruption during its critical period(s). Damage 

may not be evident until a later stage of development. There are often differences in pharmacokinetics 

and metabolism between children and adults. For example, absorption may be different in neonates 

because of the immaturity of their gastrointestinal tract and their larger skin surface area in proportion to 

body weight (Morselli et al. 1980; NRC 1993); the gastrointestinal absorption of lead is greatest in infants 

and young children (Ziegler et al. 1978). Distribution of xenobiotics may be different; for example, 

infants have a larger proportion of their bodies as extracellular water and their brains and livers are 

proportionately larger (Altman and Dittmer 1974; Fomon 1966; Fomon et al. 1982; Owen and Brozek 

1966; Widdowson and Dickerson 1964). The infant also has an immature blood-brain barrier (Adinolfi 

1985; Johanson 1980) and probably an immature blood-testis barrier (Setchell and Waites 1975). Many 

xenobiotic metabolizing enzymes have distinctive developmental patterns. At various stages of growth 

and development, levels of particular enzymes may be higher or lower than those of adults, and 

sometimes unique enzymes may exist at particular developmental stages (Komori et al. 1990; Leeder and 

Kearns 1997; NRC 1993; Vieira et al. 1996). Whether differences in xenobiotic metabolism make the 

child more or less susceptible also depends on whether the relevant enzymes are involved in activation of 

the parent compound to its toxic form or in detoxification. There may also be differences in excretion, 

particularly in newborns who all have a low glomerular filtration rate and have not developed efficient 

tubular secretion and resorption capacities (Altman and Dittmer 1974; NRC 1993; West et al. 1948). 

Children and adults may differ in their capacity to repair damage from chemical insults. Children also 

have a longer remaining lifetime in which to express damage from chemicals; this potential is particularly 

relevant to cancer. 
DDT, DDE, and DDD 182 

3. HEALTH EFFECTS 

Certain characteristics of the developing human may increase exposure or susceptibility while others may 

decrease susceptibility to the same chemical. For example, although infants breathe more air per 

kilogram of body weight than adults breathe, this difference might be somewhat counterbalanced by their 

alveoli being less developed, which results in a disproportionately smaller surface area for alveolar 

absorption (NRC 1993). 

No studies were located that specifically addressed effects of exposure to DDT in children, with the 

exception of a report by Hill and Robinson (1945) that described the case of a 1-year-old child who 

ingested 1 ounce of 5% DDT in kerosene that led to coughing, vomiting, tremors, convulsions, and 

eventually, death. The contribution of the kerosene solvent to the DDT toxicity was not clear. Data in 

adults are derived mostly from cases of accidental and/or intentional acute exposure (ingestion) of large 

amounts of DDT and from controlled studies in volunteers. Information from these studies indicate that 

the primary target of DDT toxicity is the nervous system. The effects are manifested as hyperexcitability, 

tremors, and convulsions (Francone et al. 1952; Garrett 1947; Hayes 1982; Hsieh 1954). It is reasonable 

to assume that the same effects would be seen in children similarly exposed. Additional information is 

derived from early occupational studies with limitations including lack of precise exposure data and 

presence of other compounds. Some of the long-term exposure reports provided suggestive evidence of 

adverse liver effects (Hayes 1956; Morgan and Lin 1978). None of these exposure scenarios appear 

likely for children in the United States at the present time. 

Results from a few animal studies suggest that young and older animals exhibit different susceptibility to 

DDT toxicity, at least regarding neurotoxicity in response to relatively high doses of DDT. For example, 

the LD50 values for DDT in newborn, preweanling, weanling, and adult rats were $4,000, 438, 355, and 

195 mg/kg, respectively (Lu et al. 1965). However, when one-quarter of the daily LD50 dose was 

administered daily for 4 days to preweanling and adult rats, both groups had similar 4-day LD50 values. 

Lu et al. (1965) suggested that the elimination mechanisms in the preweaning rats is less developed than 

in the adult rats, thus making them more susceptible to repeated small doses. In another study, 10-dayold rats were more resistant to the acute lethal toxicity of purified p,p’-DDT than 60-day-old rats 

(Henderson and Woolley 1970). In both groups, respiratory failure was the cause of death; however, the 

time course of DDT poisoning in the young rats was prolonged considerably as compared to the adults. 

Furthermore, the immature rats did not exhibit seizures nor the hyperthermia that preceded death in the 

older animals. The decreased sensitivity of the younger rats was attributed to an incomplete development 

of the neural pathways involved in seizure activity and in thermoregulation. The relevance of these 

findings to human health is unknown. 
DDT, DDE, and DDD 183 

3. HEALTH EFFECTS 

A decrease in testis weight was observed in juvenile male rats dosed by oral gavage on days 4 and 5 of 

life with 500 mg/kg/day or from day 4 to day 23 with 200 mg DDT/kg/day (Krause et al. 1975). After 

two doses, significant decreases were seen at 34 days, and after repeated lower doses, decreases were 

significant at days 18, 26, and 34. Treated males were mated with healthy females on days 60 and 90. 

The number of fetuses and implantations was decreased 30% at the 60-day mating but not at the 90-day 

mating of rats dosed on days 4 and 5. For rats receiving multiple doses, the decreases were 95% and 35% 

after mating at days 60 and 90, respectively. Treatment of weanling rats from either day 21 or 25 until 

day 57 of age with 100 mg p,p’-DDE/kg/day did not significantly alter serum levels of testosterone, but 

this treatment significantly delayed onset of puberty (see below) and was likely due to the antiandrogenic 

effects of DDE (Kelce et al. 1995). 

Limited information exists regarding developmental effects of DDT on humans (Reproductive effects are 

discussed in Section 3.2.2.5). A study of 594 children from North Carolina found that height and weight 

adjusted for height of boys at puberty increased with transplacental exposure to DDE, but there was no 

association between any measure of DDE exposure (DDE in breast milk, maternal blood, cord blood, 

placenta) and age at which pubertal stages were attained (Gladen et al. 2000). Children from this cohort 

has normal birth weight and normal growth in the first year of life. A study of German children found an 

inverse association between serum DDE levels and height in girls (Karmaus et al. 2002). A study of Inuit 

infants exposed to p,p’-DDE (and other organochlorines) in utero revealed no association between 

immunological parameters (white blood cell counts, lymphocyte subsets, immunoglobulins) and prenatal 

exposure to organochlorines (Dewailly et al. 2000). However, the authors found that organochlorine 

exposure was associated with the incidence of infectious diseases and that the risk of contracting otitis 

media at 7 months of age increased with prenatal exposure to p,p’-DDE, hexachlorobenzene, and dieldrin. 

They also found that the relative risk for 4- to 7-month-old infants in the highest tertile of p,p’-DDE 

exposure was 1.87 (95% CI, 1.07–3.26) compared to infants in the lowest tertile. In addition, the relative 

risk of otitis media over the entire first year of life also increased with prenatal exposure to p,p’-DDE and 

hexachlorobenzene. 

In animals, DDT can cause abnormal development o sex organs, embryotoxicity and fetotoxicity in the 

absence of maternal toxicity (Clement and Okey 1974; Fabro et al. 1984; Hart et al. 1971, 1972). 

Developmental effects, including preweanling mortality and premature puberty, have been reported in 

animals in multigeneration studies (Del Pup et al. 1978; Green 1969; Ottoboni et al. 1969, 1977; Tomatis 

et al. 1972, Turusov et al. 1973). DDT has shown estrogenic properties in animals administered the 

pesticide orally or parenterally (Bitman and Cecil 1970; Clement and Okey 1972; Fabro et al. 1984; 
DDT, DDE, and DDD 184 

3. HEALTH EFFECTS 

Gellert et al. 1972, 1974; Singhal et al. 1970). In female neonates injected subcutaneously with o,p’-DDT 

or o,p’-DDD, there were significant alterations in the estrous cycle, decreases in ovary weight, and 

decreases in corpora lutea when the animals were evaluated as adults (Gellert et al. 1972, 1974). In 

general, the estrogenic potency of DDT is orders of magnitude lower than that of estradiol. 

p,p’-DDE, a persistent metabolite of DDT, was an androgen receptor antagonist in male rats exposed in 

utero, and also as juveniles (Gray et al. 1999; Kelce et al. 1995, 1997; Loeffler and Peterson 1999; You et 

al. 1998, 1999a). Kelce and co-workers showed that pups from dams exposed during gestation 

days 14–18 to 100 mg p,p’-DDE/kg/day and then exposed indirectly to maternally stored p,p’-DDE via 

breast milk had significantly reduced anogenital distance at birth and retained thoracic nipples on 

postnatal day 13. Treatment of weanling male rats from either day 21 or 25 (specific day unclear in text) 

until day 57 of age with 100 mg p,p’-DDE/kg/day resulted in a statistically significant delayed onset of 

puberty (measured by the age of preputial separation) by 5 days. Gray et al. (1999) and You et al. (1998) 

reported that anogenital distance was not affected in male Sprague-Dawley rats on postnatal day 2 after 

treating the dams with up to 100 mg p,p’-DDE/kg on gestation days 14–18, but was significantly reduced 

in similarly exposed Long-Evans pups. A 10 mg/kg dose to the dams was without effect in the LongEvans pups. Anogenital distance was not affected in female pups from either strain. Treatment of the 

dams with 10 mg p,p’-DDE/kg resulted in retention of thoracic nipples in Sprague-Dawley pups, but only 

the higher dose (100 mg/kg) had this effect in Long-Evans pups. Treatment with p,p’-DDE also resulted 

in an apparent reduction of androgen receptor expression in male sex organs from mainly high-dose 

Sprague-Dawley pups, as shown by immunochemical staining; however, there were no changes in 

androgen receptor steady state mRNA levels in the high-dose Sprague-Dawley rats, but androgen receptor 

mRNA were increased 2-fold in the high-dose Long-Evans rats. Exposure of the pups to p,p’-DDE 

during gestation and lactation had no significant effect on the onset of puberty. An additional study from 

the same group showed that prenatal exposure to p,p’-DDE was associated with expression of TRPM-2, 

an androgen-repressed gene (You et al. 1999a). A similar study in Holtzman rats exposed during 

gestation days 14–18 to doses between 1 and 200 mg p,p’-DDE/kg (offspring were exposed to p,p’-DDE 

in utero and via breast milk) found reduced anogenital distance in males on postnatal day 1 and reduced 

relative ventral prostate weight on postnatal day 21 at 50 mg p,p’-DDE/kg, but not at 10 mg p,p’-DDE/kg 

(Loeffler and Peterson 1999). Doses up to 100 mg/kg/day to the dams had no effect on onset of puberty, 

but 200 mg/kg/day did significantly delay puberty in males by less than 2 days. Androgen receptor 

staining in the ventral prostate was also reduced on postnatal 21. Serum levels of testosterone or 3α-diol 

androgens were not significantly altered at any time. This study also reported that at the 100 mg/kg dose 

level, cauda epididymal sperm number was reduced by 17% on postnatal day 63 relative to controls. No 
DDT, DDE, and DDD 185 

3. HEALTH EFFECTS 

measurement of DDE body burden were made in the 200 mg/kg/day offspring postnatally, so it is difficult 

to determine whether effects on puberty were due to the previous gestational plus lactational exposures or 

directly due to the effects of DDE present near the time of puberty. More details about these studies can 

be found in Section 3.2.2.6, Developmental Effects. Kelce et al. (1995) also showed that treatment of 

adult male rats with 200 mg p,p’-DDE/kg for 4 days significantly reduced androgen-dependent seminal 

vesicle and ventral prostate weight relative to controls. 

There are limited data suggesting that if mice exposed to DDT in utero and during lactation are further 

exposed to DDT postnatally, responses in both immunological plaque forming assays and lymphoproliferative assays are reduced (Rehana and Rao 1992). The results of this study are difficult to interpret 

because of experimental design issues, including the lack of a comparison unexposed group and 

ambiguity about whether statistical testing was done. 

Alterations in learning processes and in other behavioral patterns have also been described in adult mice 

exposed to DDT perinatally (Craig and Ogilvie 1974; Palanza et al. 1999b; vom Saal et al. 1995) or as 

neonates (Eriksson et al. 1990a, 1990b, 1992, 1993; Johansson et al. 1995, 1996); this end point is the 

basis of an acute oral MRL, which is discussed in detail in Section 2.3, Minimal Risk Levels. These 

studies suggest that exposure of the developing fetus or newborn to DDT during critical stages in nervous 

system development can cause developmental toxicity manifested later in life. Eriksson et al. (1990a, 

1990b) pointed out that the dose levels that caused behavioral alterations in mice are comparable to those 

levels to which human neonates might be exposed in areas where DDT is still being used. Behavioral 

neurotoxicity has been described in rats treated with DDT as adults (Sobotka 1971), but only at doses at 

least 50 times those that produced learning deficits in neonates. 

There is no information regarding transgenerational effects associated with DDT exposure in humans, 

despite some limited evidence of genotoxicity in humans. Chromosomal aberrations have been reported 

in blood cells from subjects occupationally exposed to DDT (Rabello et al. 1975; Rupa et al. 1988; Yoder 

et al. 1973) and in human lymphocytes exposed in vitro to DDT (Lessa et al. 1976). These occupational 

exposures, however, also included exposure to many other pesticides. Two studies, one in rats (Palmer et 

al. 1973) and one in mice (Clark 1974), reported an increase in dominant lethality after exposure to high 

doses of DDT (Clark 1974). A study in rabbits administered 50 mg p,p’-DDT/kg by gavage on Gd 7–9 

found no significant alterations in the distribution of chromosomes in liver samples from fetuses or in the 

percentage of chromosomal aberrations relative to controls (Hart et al. 1972). 
DDT, DDE, and DDD 186 

3. HEALTH EFFECTS 

There is no information regarding the pharmacokinetics of DDT in children or regarding the nutritional 

factors that may influence the absorption of DDT. Analysis of urine samples from humans exposed to 

DDT suggest the involvement of both phase I and phase II metabolic enzymes in the biotransformation 

and elimination of DDT and metabolites. The specific P-450 isoenzymes involved in phase I metabolism 

and the particular phase II conjugating enzymes are not known with certainty, so no conclusions can be 

drawn based on general differences in isozyme activities between adults and children. It is not known 

whether the metabolism of DDT in children might be different than in adults. Recent studies have 

suggested that enzymes involved in the metabolism of steroids are development-regulated and can be 

affected by DDE. For example, You et al. (1999c) examined the effects of p,p’-DDE on CYPs enzymes 

responsible for testosterone hydroxylations. They found that exposure of pregnant Sprague-Dawley rats 

to p,p’-DDE during gestation days 14–18 resulted in induction of CYP2A1 in males on PND 10 but not 

PND 21; pronounced induction of CYP2B1 was seen in males and females on both PND 10 and 21 and to 

a lesser extent of CYP3A1; there was no induction of CYP2C11. In contrast, in adult rats, DDE induced 

2B1, 3A1, and 2C11, but not 2A1. 

DDT and DDT-related compounds, particularly DDE, accumulate in fatty tissues and have been found in 

human milk (Dewailly et al. 1996; Scheele et al. 1995; Smith 1999; Torres-Arreola et al. 1999), amniotic 

fluid (Foster et al. 2000), placenta (Gladen and Rogan 1995; Procianoy and Schvartsman 1981; 

Wassermann et al. 1982), and in most organs from stillborn infants (Curley et al. 1969). Although, there 

is no direct evidence of adverse health outcomes in human infants exposed in such a manner, mobilization 

of adipose fat for lactation might provide an increased amount of DDT to a breast-feeding infant. Model 

estimates of increasing body burden of DDE in infants from breast-feeding milk with a DDE 

concentration of 222 ppb (lipid basis) predict that maximum body burden (approximately 70 μg/kg) is 

achieved after approximately 5–6 months of nursing, even if nursing continues, and then gradually 

decreases to <10 μg/kg by 2 years postpartum (LaKind et al. 2000). Levels of DDT and related 

compounds in human tissues from recent studies are presented in Table 6-4. Studies in animals have 

demonstrated placental transfer of DDT and DDE to fetuses and also to newborns via mother’s milk 

(Fang et al. 1977; Seiler et al. 1994; Wooley and Talens 1971; You et al. 1999b). The results of these 

studies indicate that the amounts of chemical transferred via mother’s milk are much greater than the 

amounts that reach the fetus through the placenta. PBPK models for the transplacental and lactational 

transfer of p,p’-DDE in rats were proposed by You et al. (1999b). The models provide an approach to 

estimating tissue doses in fetuses and pups associated with maternal exposures to p,p’-DDE, and can be 

used to explore dose-response relationships for the developmental effects of p,p’-DDE in the rat. In a 

study in rabbits administered DDT by gavage before artificial insemination and throughout gestation, 
DDT, DDE, and DDD 187 

3. HEALTH EFFECTS 

Seiler et al. (1994) found higher concentrations of DDT residues in fetuses (day 11 postcoitum) than in 

blastocytes (day 6 postcoitum), suggesting that transplacental passage may be more easily accomplished 

than passage into blastocytes via uterine secretions. 

Intraperitoneal administration of the DDT metabolite, 3-MeSO2-DDE, to pregnant or lactating mice 

resulted in specific accumulation and binding of 3-MeSO2-DDE-derived radioactivity in the zona 

fasciculata of adrenals from 16- to 18-day-old fetuses or suckling pups, showing ready transplacental 

passage of the metabolite and transfer via maternal milk (Jönsson et al. 1992). The results also suggested 

that cytochrome P-45011β was involved in the activation of 3-MeSO2-DDE in the fetal adrenal cortex. 

Quantitative measurements showed that 7 days after dosing, the labeling of the pups adrenals was 2 and 

3.6 times higher than maternal adrenals at 1.5 and 25 mg/kg dose levels, respectively, suggesting that 

mother’s milk may be an important route of exposure of 3-MeSO2-DDE in DDT-exposed animals. A 

later study showed that P-45011β seems to be expressed during Gd 10–12 in the adrenal cortex in the 

mouse fetus (Jönsson et al. 1995). 3-MeSO2-DDE was also found to reduce the capacity of pups and 

maternal adrenals to secrete corticosterone (Jönsson 1994) by a mechanism possibly involving 

competitive inhibition of adrenocortical CYP11B1 (Johansson et al. 1998). 

There are no biomarkers of exposure or effect for DDT or DDT-related compounds that have been 

validated in children or in adults exposed as children. There are no biomarkers in adults that identify 

previous childhood exposure. No studies were located regarding interactions of DDT with other 

chemicals in children or adults. No information was located regarding pediatric-specific methods for 

reducing peak absorption following exposure to DDT or DDT-related compounds, reducing body burden, 

or interfering with the mechanism of action for toxic effects. In addition, no data were located regarding 

whether methods for reducing toxic effects in adults might be contraindicated in children. 

3.9 BIOMARKERS OF EXPOSURE AND EFFECT 

Biomarkers are broadly defined as indicators signaling events in biologic systems or samples. They have 

been classified as markers of exposure, markers of effect, and markers of susceptibility (NAS/NRC 1989). 

Due to a nascent understanding of the use and interpretation of biomarkers, implementation of biomarkers 

as tools of exposure in the general population is very limited. A biomarker of exposure is a xenobiotic 

substance or its metabolite(s) or the product of an interaction between a xenobiotic agent and some target 

molecule(s) or cell(s) that is measured within a compartment of an organism (NAS/NRC 1989). The 
DDT, DDE, and DDD 188 

3. HEALTH EFFECTS 

preferred biomarkers of exposure are generally the substance itself or substance-specific metabolites in 

readily obtainable body fluid(s), or excreta. However, several factors can confound the use and 

interpretation of biomarkers of exposure. The body burden of a substance may be the result of exposures 

from more than one source. The substance being measured may be a metabolite of another xenobiotic 

substance (e.g., high urinary levels of phenol can result from exposure to several different aromatic 

compounds). Depending on the properties of the substance (e.g., biologic half-life) and environmental 

conditions (e.g., duration and route of exposure), the substance and all of its metabolites may have left the 

body by the time samples can be taken. It may be difficult to identify individuals exposed to hazardous 

substances that are commonly found in body tissues and fluids (e.g., essential mineral nutrients such as 

copper, zinc, and selenium). Biomarkers of exposure to DDT, DDE, and DDD are discussed in 

Section 3.9.1. 

Biomarkers of effect are defined as any measurable biochemical, physiologic, or other alteration within an 

organism that, depending on magnitude, can be recognized as an established or potential health 

impairment or disease (NAS/NRC 1989). This definition encompasses biochemical or cellular signals of 

tissue dysfunction (e.g., increased liver enzyme activity or pathologic changes in female genital epithelial 

cells), as well as physiologic signs of dysfunction such as increased blood pressure or decreased lung 

capacity. Note that these markers are not often substance specific. They also may not be directly 

adverse, but can indicate potential health impairment (e.g., DNA adducts). Biomarkers of effects caused 

by DDT, DDE, and DDD are discussed in Section 3.9.2. 

A biomarker of susceptibility is an indicator of an inherent or acquired limitation of an organism's ability 

to respond to the challenge of exposure to a specific xenobiotic substance. It can be an intrinsic genetic 

or other characteristic or a preexisting disease that results in an increase in absorbed dose, a decrease in 

the biologically effective dose, or a target tissue response. If biomarkers of susceptibility exist, they are 

discussed in Section 3.11, Populations That Are Unusually Susceptible. 

3.9.1 Biomarkers Used to Identify or Quantify Exposure to DDT, DDE, and DDD 

In general, biomarkers of exposure to DDT can be classified as specific, such as DDT itself and its 

metabolites, and nonspecific, such as changes in endogenous chemicals that might indicate exposure to 

DDT, but also to other unrelated chemicals as well. DDT, DDE, and DDD have been detected and 

measured in adipose tissue, blood, serum, urine, feces, semen, and breast milk using several analytical 

techniques (see Chapter 7). Metabolites of DDT have also been measured in body fluids. The major 
DDT, DDE, and DDD 189 

3. HEALTH EFFECTS 

urinary metabolite identified in humans is DDA (Gingell 1976), while the p,p’-isomers of DDT and DDE 

as well as o,p’-DDT have been detected in breast milk (Takei et al. 1983; Torres-Arreola et al. 1999; see 

also Table 6-4). Nair and Pillai (1992) detected o,p’-DDE, p,p’-DDE, o,p’-DDT, p,p’-DDT, and 

p,p’-DDD in human adipose tissue and breast milk, and o,p’-DDE, p,p’-DDE, and p,p’-DDT in whole 

blood. An earlier study had identified p,p’-DDE, p,p’-DDD, and o,p’-DDT in whole blood (Agarwal et 

al. 1976). p,p’-DDE has also been detected in amniotic fluid (Foster et al. 2000). Smith (1999) compiled 

values for DDT in human milk from 130 published studies since 1951 and observed that population 

means have declined in much of the world from 5,000–10,000 ppb (lipid-based) to around 1,000 ppb 

currently in many areas. Using a set of 13 studies from the United States and Canada, Smith (1999) 

estimated that since 1975, there has been an 11–21% decline in average DDT concentrations in breast 

milk. He also estimated a half-life of 4.2–5.6 years for DDT on a population basis. Levels of DDT in 

breast milk from some recent studies are presented in Table 6-4. Because DDT/DDE/DDD are fat 

soluble, they tend to concentrate in the fat portion of the specific biological medium (i.e., milk, serum, 

etc.) and therefore, their levels in the biological medium may vary with the fat content of that medium. 

This variability in the amount of DDT/DDE/DDD in the various media, is often accounted for by using 

lipid-adjusted measurements. 

However, there are no quantitative data available that allow correlation of DDT/DDD/DDE levels in 

human tissue or fluids and exposure to particular levels of environmental contamination. Studies of 

pesticide production workers reported that blood levels of these compounds are generally higher in 

persons exposed in the workplace. Since the biological half-lives for elimination of these compounds are 

ranked as follows: DDE > DDT > DDD, detection of higher ratios of DDD or DDT to DDE is believed to 

indicate more recent exposure while lower ratios are believed to correlate with long-term exposure and 

storage capacity (Morgan and Roan 1971). There is a direct correlation between DDT and DDE levels in 

blood and adipose tissue when concentrations are expressed on a lipid basis (Hayes et al. 1971; Morgan 

and Roan 1971; Mussalo-Rauhamaa 1991). On a wet tissue basis, concentrations of DDT in adipose 

tissue are approximately 280 times higher than those of blood (Anderson 1985). However, because DDT 

and DDE are extensively stored in fatty tissue and slowly released from storage sites, there is no 

correlation between levels in tissues and the time course of exposure in short time spans. Analysis of 

residue levels of p,p’-DDT in skin lipids collected by wiping the skin with cotton and purifying lipids by 

gel permeation chromatography has been investigated by Sasaki et al. (1991b) as a noninvasive method 

for monitoring DDT burdens in pesticide-exposed persons; a correlation was found between skin levels 

and adipose tissue levels; however, these results reflect both body burden and dermal exposure. Levels of 

DDT (total DDT) in adipose tissue from the U.S. population have been continuously declining over the 
DDT, DDE, and DDD 190 

3. HEALTH EFFECTS 

last decades from about 8 ppm (lipid adjusted) in 1970 to about 2 ppm in 1983 (Kutz et al. 1991). A 

review by Kutz et al. (1991) also lists levels of DDT, DDE, and DDD in samples of adipose tissue 

collected during the 1960s through the 1980s from populations around the world. Table 6-4 presents 

additional information on DDT levels in blood and adipose tissue from some recent studies. 

Changes in plasma concentrations of endogenous chemicals that might be consistent with DDT exposure 

include increased vitamin A plasma levels, which have been shown to increase with increasing plasma 

levels of DDE in humans (Nhachi and Kasilo 1990). In rats, DDT administration (oral gavage of 

40 mg/kg) decreased hepatic vitamin A storage (deWaziers and Azais 1987). Increased levels of urinary 

17-hydroxycortisone have also been reported as indicators of DDT exposure in humans (Nhachi and 

Loewenson 1989; Poland et al. 1970). This is consistent with DDT induction of hepatic P-450 enzymes 

increasing the catabolism of cortisol to 17-hydroxycortisol (see Section 3.5.3). A potential biomarker of 

exposure, which has been identified from studies in laboratory animals, is an increase in GGTP. In rats, 

after acute oral exposure to DDT, serum levels of GGTP doubled and remained elevated for 48 hours 

(Garcia and Mourelle 1984). However, none of these potential biomarkers are specific to DDT, DDE, or 

DDD exposure, and not all the body compartments in which these changes occur are accessible for 

sampling in living humans. 

3.9.2 Biomarkers Used to Characterize Effects Caused by DDT, DDE, and DDD 

The primary target organs for DDT, DDE, and DDD toxicity include the nervous system, the reproductive 

system, and the liver. No biomarkers of effect specific for DDT, DDE, or DDD exposure alone were 

identified in the literature. Tremors and convulsions have been observed in both humans and laboratory 

animals after DDT exposure (Hsieh 1954; Hwang and Van Woert 1978; Matin et al. 1981). Exposure to 

DDT has been shown to induce hepatic microsomal enzymes in both humans and laboratory animals 

(Kolmodin et al. 1969; Morgan and Lin 1978; Pasha 1981; Street and Chadwick 1967). However, these 

biomarkers of effect are not specific for DDT, DDE, or DDD exposure, and not all the body 

compartments in which these changes occur are accessible for sampling in living humans. 

For more general information on biomarkers of effect for the immune, renal, and hepatic systems, see 

CDC/ATSDR Subcommittee Report on Biological Indicators of Organ Damage (CDC/ATSDR 1990), 

and for biomarkers of effect for the neurological system, see the Office of Technology Assessment report 

(OTA 1990). For more information on the health effects after exposure to DDT, DDE, and DDD see 

Section 3.2. 
DDT, DDE, and DDD 191 

3. HEALTH EFFECTS 

3.10 INTERACTIONS WITH OTHER CHEMICALS 

This section discusses the potential for DDT to act synergistically and/or antagonistically with other 

chemicals to cause physiological harm. DDT may have broad effects by changing the metabolism of 

other chemicals, both xenobiotics and endogenous macromolecules. DDT induces microsomal mixed 

function oxidases that are involved in the catabolism of both xenobiotics and many endogenous 

hormones, such as cortisol. DDE has also caused the induction of hepatic enzymes (Conney 1971), 

including cytochrome P-450 microsomal enzymes (Pasha 1981). Its effects on the latter are discussed in 

Section 3.5.3. In most cases, this biotransformation results in compounds that are less toxic than the 

parent compound and more readily excreted from the body. For some chemicals, this metabolism results 

in the production of metabolites that are more toxic than the parent compound and that may be 

carcinogenic. One interaction of concern is the enhanced conversion of other chemicals to active, 

carcinogenic forms mediated by microsomal enzymes induced by DDT. Several investigations indicate 

that DDT administered to animals along with a known carcinogen may result in either increase or 

decrease in tumor production relative to the carcinogen tested without DDT. A study by Walker et al. 

(1972) suggested that the liver enlargement was greater and the time to palpability of liver masses was 

earlier in mice fed dieldrin and DDT than those fed either pesticide separately. A potentiation of 

carcinogenic activity of dieldrin was suggested but not conclusively shown. DDT alone is thought to 

produce hepatic tumors both through the formation of DNA adducts and through promotional 

mechanisms involving cytotoxicity and compensatory cell proliferation (see Section 3.5.2). It is possible 

that DDT could also promote the formation of hepatic tumors initiated by other carcinogens. 

DDT is reported to promote the tumorigenic effects of several known carcinogens, such as 3-methyl­

(4-dimethylamine)-azobenzene (Kitagawa et al. 1984), 2-acetylaminofluorene (2-AAF) (Peraino et al. 

1975), diethyl-nitrosamine (DEN) (Diwan et al. 1994; Nishizumi 1979), and carbon tetrachloride (CCl4) 

(Preat et al. 1986) when given after the putative carcinogen. The promoting effect of DDT in rats is 

reported to act in a dose-dependent fashion, with DDT decreasing the latency period of tumor 

development and increasing the incidence and yield of hepatic tumors, mainly hepatocellular carcinomas. 

DDT acted as a hepatocellular tumor promoter in D2B6F1 mice when administered in the diet at 300 ppm 

(53 mg/kg/day) for 53 weeks, beginning 2 weeks following initiation with an intraperitoneal injection of 

N-nitrosodiethylamine (Diwan et al. 1994). The incidence of mice with tumors was 22/22 in the DDTtreated group, compared to 12/30 in the group receiving only N-nitrosodiethylamine; the difference was 

statistically significant. 
DDT, DDE, and DDD 192 

3. HEALTH EFFECTS 

Pretreatment of animals with DDT was also reported to decrease the tumorigenic effects of some 

previously determined carcinogens. For example, pretreatment of rats with DDT significantly lowered 

the incidence of mammary tumors per rat after treatment with 7,12-dimethylbenz[a]anthracene (DMBA), 

versus DMBA-treated controls (Silinskas and Okey 1975). The authors suggested that DDT may inhibit 

DMBA-induced mammary tumors by stimulating hepatic metabolism and accelerating the excretion of 

DMBA, so that less carcinogen is available to peripheral tissues. Other studies also have reported the 

DDT induction of hepatic microsomal enzymes, which reduced the carcinogenicity of azo dyes and 

similar carcinogens (Williams and Weisburger 1991). 

Similarly, the hepatocarcinogenicity of aflatoxic B1 in mice was inhibited by pretreatment with DDT and 

by co-treatment with DDT when given throughout aflatoxin B1 dosing (Rojanapo et al. 1988, 1993). 

However, DDT acted as a hepatocarcinogenic promoter to aflatoxin B1 initiation when a 14-week DDT 

administration followed an 8-week aflatoxic B1 treatment, or when the DDT administration began 

halfway through aflatoxin B1 treatment (Rojanapo et al. 1988, 1993). Also, in groups receiving both 

aflatoxic B1 and DDT, in any order, absolute and relative liver weights were significantly increased over 

both the vehicle control and the group receiving just aflatoxin B1; treatment with aflatoxin B1 alone 

increased liver weights, while treatment with DDT alone did not (Rojanapo et al. 1993). The proposed 

mechanisms of tumor promotion are discussed in Section 3.5.2. 

The effects of DDT on the nervous system are altered when DDT is given in combination with certain 

neurologically-active pharmacological agents. Some pharmacological agents (hydantoin, phenobarbital), 

prevent some or all of the neurological effects seen in animals treated with DDT (see Section 3.2.2.4), 

while other agents (trihexyphenidyl, haloperidol, propranolol) enhance DDT-induced neurotoxicity (Herr 

et al. 1985; Hong et al. 1986; Matin et al. 1981). One of the effects of DDT is to hold sodium channels 

open, which probably contributes to DDT-induced neurological effects (tremors and hyperexcitability). 

Studies by Rubin et al. (1993) have shown that DDT analogues and metabolites as well as several 

pyrethroids modify radioligand binding of batrachotoxinin to sodium channels in mouse brain 

synaptosomes. DDT and pyrethroids do not by themselves stimulate Na+

 uptake, but they enhance 

activator-dependent uptake. DDT is more efficacious than the pyrethroids tested. Eriksson et al. (1993) 

have shown that the pyrethroid bioallethrin and DDT can interact in vivo in rats. 

As previously mentioned, in the United States, exposure of the general population to DDT/DDE/DDD 

occurs mainly through consumption of contaminated food (i.e., meat, fish, dairy products). Because 

organochlorines, in general, bioaccumulate in the food chain, it is very likely that there will be also 
DDT, DDE, and DDD 193 

3. HEALTH EFFECTS 

simultaneous exposure to several other persistent organochlorine chemicals, many also with hormonal 

activity. Therefore, interactions at the level of the estrogen and/or androgen receptor may occur, and the 

final outcome probably represents the summation of effects (positive or negative; linear or non-linear) of 

chemicals with similar modes of action. 

3.11 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE 

A susceptible population will exhibit a different or enhanced response to DDT, DDE, and DDD than will 

most persons exposed to the same level of DDT, DDE, and DDD in the environment. Reasons may 

include genetic makeup, age, health and nutritional status, and exposure to other toxic substances (e.g., 

cigarette smoke). These parameters result in reduced detoxification or excretion of DDT, DDE, and 

DDD, or compromised function of organs affected by DDT, DDE, and DDD. Populations who are at 

greater risk due to their unusually high exposure to DDT, DDE, and DDD are discussed in Section 6.7, 

Populations With Potentially High Exposures. 

No data are available on human differences in susceptibility to DDT, DDE, or DDD. Groups who might 

be particularly susceptible to the toxic effects of DDT are individuals with diseases of the nervous system 

or liver. Persons with nervous system disorders in which normal function is altered due to physiological 

changes, such as changes in neurotransmitter balance or impaired neuronal conduction, might be more 

susceptible to DDT neurotoxicity. Persons with diseases of the liver might be more sensitive to the 

hepatotoxic effects of DDT since normal repair function may already be compromised. 

The susceptibility of children is discussed in detail in Section 3.8, Children’s Susceptibility. 

3.12 METHODS FOR REDUCING TOXIC EFFECTS 

This section will describe clinical practice and research concerning methods for reducing toxic effects of 

exposure to DDT, DDE, and DDD. However, because some of the treatments discussed may be 

experimental and unproven, this section should not be used as a guide for treatment of exposures to DDT, 

DDE, and DDD. When specific exposures have occurred, poison control centers and medical 

toxicologists should be consulted for medical advice. 

Most of the strategies discussed in the following sections apply to high-dose exposures. The balance 

between the benefits and detriments of mitigation for low-dose chronic exposures might differ from those 
DDT, DDE, and DDD 194 

3. HEALTH EFFECTS 

for high-dose exposures. Methods to reduce toxic effects should not be applied indiscriminately to all 

individuals exposed to DDT, DDE, or DDD; good clinical judgement should be used. 

3.12.1 Reducing Peak Absorption Following Exposure 

A number of strategies have been suggested to minimize absorption from the gastrointestinal tract. 

Ipecac-induced emesis has been suggested unless there is a risk of lung aspiration due to unconsciousness 

or convulsions (HSDB 2002a, 2002b, 2002c, 2002d). Gastric lavage with mannitol may also be useful 

for limiting absorption, particularly when emesis is contraindicated (Dreisbach 1983). Since activated 

charcoal can absorb DDT, it has also been commonly used as a method for reducing intestinal uptake 

(Dreisbach 1983). Another method of reducing absorption is the use of a cathartic, and in practice, 

activated charcoal is frequently given in a slurry of one of the saline cathartics (Dreisbach 1983; HSDB 

2002a, 2002b, 2002c, 2002d). Vegetable oils should not be used as cathartics because they have been 

shown to promote the absorption of DDT, DDE, and DDD in animals (Keller and Yeary 1980). At this 

point, it is unknown whether there are any specific binding or reactive agents that might prevent 

absorption, but this might be a strategy for future research. 

Dermal absorption of DDT, DDE, and DDD is less efficient than absorption by the oral route. After 

dermal or ocular exposure, absorption may be reduced by decontaminating the exposed area; the 

generally used method is washing (HSDB 2002a, 2002b, 2002c, 2002d). 

3.12.2 Reducing Body Burden 

In humans, DDT and its metabolites distribute in the general circulation, but are eventually selectively 

concentrated in adipose tissue where they are retained for long periods of time. After the pesticide is 

sequestered in adipose tissue, common methods to reduce body burden such as dialysis, exchange 

transfusion, and hemoperfusion are probably ineffective because only small amounts are present in the 

blood. However, the interval immediately after absorption may be a window of opportunity for removing 

the xenobiotic from the circulation before it partitions into adipose tissue. Potential strategies that might 

be investigated include hemodialysis and hemoperfusion (Klaassen 1990). 

Absorbed DDT is primarily excreted in the urine (mostly as conjugated DDA), with minor amounts 

excreted in feces (via biliary excretion), semen, and breast milk. There are several potential strategies for 

enhancing fecal DDT excretion that might be worth investigation. Studies of DDT excretion in 
DDT, DDE, and DDD 195 

3. HEALTH EFFECTS 

laboratory rats indicated that oral administration of sodium muconate resulted in fecal excretion of 

injected DDT that was 22 times greater during the first day than in animals that had not received sodium 

muconate. The trans, trans muconate isomer was given by oral gavage (75 mg/kg) 2 hours after 

intraperitoneal injection of 14C-p,p’-DDT (6.7 mg/kg). Over the first 10 days following sodium muconate 

treatment, fecal excretion was 2.34 times greater in rats receiving sodium muconate compared to rats that 

did not receive this treatment (Boileau et al. 1985). Ten days posttreatment, the experimental rats had 

significantly lower DDT in abdominal fat, liver, and brain showing effective reduction of body burden 

(Boileau 1985). The mechanism of trans, trans muconate in enhancing fecal excretion is unknown. 

Therefore, it is possible that administration of sodium muconate to humans could result in greater fecal 

excretion. Muconate may only be effective in acute poisonings, and it has not been shown if a similar 

effect in rats occurs after acute oral exposure to DDT. 

Fecal metabolites have been measured, and at least in one case compared with biliary excretion straight 

from the bile duct (see Section 3.4.4). If significant enterohepatic circulation could be demonstrated, then 

methods to interfere might be effective in accelerating the excretion of DDT metabolites. There are 

several possible strategies for reducing intestinal resorption of metabolites excreted in the bile; the 

simplest is repeat doses of activated charcoal (Levy 1982). Another strategy that has been effective with 

other lipophilic xenobiotics has been the oral administration of the anion exchange resin, cholestryamine 

(Boylan 1978). Daily administration of mineral oil, a cathartic, in the diet of monkeys beginning 7 days 

after a single oral dose of DDT resulted in a reduction of DDT in adipose tissue and greater elimination in 

feces (Rozman et al. 1983). 

3.12.3 Interfering with the Mechanism of Action for Toxic Effects 

DDT, DDE, and DDD are stimulants that can cause tremors and convulsions by several postulated 

mechanisms (Section 3.5.2). The most well accepted mechanism is interference with membrane ion 

fluxes, which leads to prolongation of the action potential and repetitive firing (Narahashi and Haas 

1967). Other contributory mechanisms that may be secondary to inferences with ion fluxes may include 

decreases in brain serotonin and increases in levels of aspartate and glutamate. A number of drugs can 

alleviate this type of central nervous system excitation on experimental animals. These include the 

anticonvulsant barbiturate phenobarbital, the sedative-antianxiety drug diazepam, and the anticonvulsant 

phenytoin (Herr et al. 1985; HSDB 2002a, 2002b, 2002c, 2002d; Matin et al. 1981; Tilson et al. 1987). A 

reduction in the serotonergic activity in the brain has also been postulated to be responsible for the 

neurotoxic syndrome of myoclonus associated with DDT. Agents that enhance the action of serotonin, 
DDT, DDE, and DDD 196 

3. HEALTH EFFECTS 

such as L-5-hydroxytryptophan, a serotonin precursor; chlorimipramine, a serotonin uptake blocker; and 

phenoxybenzomine or trazodone, two α-receptor blockers, can reduce the neurotoxic effects (myoclonus) 

of DDT in mice (Hwang and Van Woert 1978). Caution should be used in extrapolating from animal 

therapeutics to human applicability. Myoclonus in the rat induced by p,p’-DDT has been studied by Pratt 

et al. (1986), and from electrophysiological and pharmacological analysis it was concluded that the rat 

was not a good model for studying 5-hydroxytryptamine-sensitive myoclonus in humans. 

The reproductive system in animals is another sensitive target organ for DDT toxicity. There are no 

medically proven-methods for reducing DDT, DDD, and DDE reproductive toxicity by interfering with 

the mechanism of action in this organ. 

Mitigation strategies developed in the future for other lipophilic pesticides should be investigated for their 

applicability to DDT, DDE, and DDD. 

3.13 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of DDT, DDE, and DDD is available. Where adequate 

information is not available, ATSDR, in conjunction with the National Toxicology Program (NTP), is 

required to assure the initiation of a program of research designed to determine the health effects (and 

techniques for developing methods to determine such health effects) of DDT, DDE, and DDD. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 

3.13.1 Existing Information on Health Effects of DDT, DDE, and DDD 

The existing data on health effects of inhalation, oral, and dermal exposure of humans and animals to 

DDT are summarized in Figure 3-8. The purpose of this figure is to illustrate the existing information 

concerning the health effects of DDT, DDE, and DDD. Each dot in the figure indicates that one or more 
DDT, DDE, and DDD 197 

3. HEALTH EFFECTS 

studies provide information associated with that particular effect. The dot does not necessarily imply 

anything about the quality of the study or studies, nor should missing information in this figure be 

interpreted as a “data need”. A data need, as defined in ATSDR’s Decision Guide for Identifying 

Substance-Specific Data Needs Related to Toxicological Profiles (Agency for Toxic Substances and 

Disease Registry 1989), is substance-specific information necessary to conduct comprehensive public 

health assessments. Generally, ATSDR defines a data gap more broadly as any substance-specific 

information missing from the scientific literature. 
DDT, DDE, and DDD 198 

3. HEALTH EFFECTS 

Figure 3-8. Existing Information of Health Effects of DDT, DDE, and DDD 
 

 

DDT, DDE, and DDD 199 

3. HEALTH EFFECTS 

3.13.2 Identification of Data Needs 

Acute-Duration Exposure. With acute oral exposure to high doses, the nervous system appears to 

be the major target in both humans and animals. Acute oral exposure has been associated with tremors or 

convulsions in humans (Hsieh 1954; Velbinger 1947a, 1947b) and animals (Hong et al. 1986; Matin et al. 

1981). An acute MRL has been derived based on neurobehavioral effects observed in adult mice 

following acute perinatal exposure to DDT (Eriksson et al. 1990a, 1990b, 1992, 1993; Johansson et al. 

1995, 1996; Talts et al. 1998). Studies of single oral or injection exposures of rats, guinea pigs, and mice 

have provided information on lethal and nonlethal levels of DDT. Further acute oral exposure studies 

during critical windows of embryonic, fetal, or neonatal development may be very informative. 

Information on health effects following acute inhalation of DDT, DDE, or DDD in humans (Neal et al. 

1944) or dermal exposure in animals (Cameron and Burgess 1945) was limited. Because of the lack of 

inhalation data in animals, an acute inhalation MRL could not be derived. Exposure via inhalation at the 

ambient levels in air (Whitmore et al. 1994) is thought to be insignificant compared with dietary uptake 

(see Section 6.5, General Population and Occupational Exposure). Also, in the atmosphere, about 50% of 

DDT is adsorbed to particulate matter and 50% exists in the vapor phase (Bidleman 1988); it is likely that 

particulate-absorbed DDT will be deposited in the upper respiratory tract and swallowed after mucociliary 

transport upward (Hayes 1982). 

Intermediate-Duration. Intermediate-duration exposures in humans and animals have been reported. 

In most human studies, the exact duration and level of exposure cannot be quantified because the 

information is derived from case reports or epidemiological studies that do not adequately characterize 

exposure. Studies on volunteers have been performed in the past that provide useful information (Hayes 

et al. 1956). The animal studies describe predominantly neurological (Cranmer et al. 1972a; Sanyal et al. 

1986), hepatic (Gupta et al. 1989; Laug et al. 1950; Ortega 1956), immunological (Banerjee et al. 1987a, 

1995, 1996, 1997a, 1997b), and reproductive/developmental (Clement and Okey 1974; Jonsson et al. 

1976; Kelce et al. 1995; Krause 1977; Krause et al. 1977; Lundberg 1974; Wrenn et al. 1971) end points. 

Little or no information on respiratory, cardiovascular, gastrointestinal, hematological, musculoskeletal, 

or dermal/ocular effects in animals exists. An intermediate oral MRL was calculated for DDT, based on 

cytoplasmic eosinophilia and cellular hypertrophy of hepatocytes in rats described in the Laug et al. 

(1950) study. Because this study was conducted several decades ago, it would be desirable to replicate 

the findings using modern histopathological techniques and current views on the interpretation of specific 

liver changes. The immunological studies on DDT are not considered to be adequate for MRL 
 

DDT, DDE, and DDD 200 

3. HEALTH EFFECTS 

development. The 1992 study reported by Rehana and Rao has several deficiencies in reporting that don't 

allow adequate data interpretation. The animal data indicate that the liver is the main target organ 

following intermediate or chronic oral exposure, but there is little evidence that liver function has been 

impaired in humans occupationally exposed, although the data are limited. The importance of subtle 

biochemical changes in humans, such as the induction of microsomal enzymes in the liver and decreases 

in biogenic amines in the nervous system, is not known with certainty. This information would be helpful 

in evaluating the toxic effects of DDT, DDE, and DDD following intermediate exposure. Levels in 

environmental media are not expected to be high enough to result in high inhalation or dermal exposures. 

Chronic-Duration Exposure and Cancer. Studies have been conducted in animals in which oral 

exposure was for a chronic duration. Reproductive, neurological, and hepatic effects have been observed 

in animals following chronic oral exposure (Deichmann et al. 1967, 1971; Durham et al. 1963; Fitzhugh 

1948; NCI 1978; Rossi et al. 1977; Takayama et al. 1999). No chronic-duration toxicity studies in which 

animals were exposed dermally or by inhalation were located. However, the inhalation and dermal routes 

are considered minor routes of entry with regard to absorption. Inhaled DDT is largely deposited in the 

upper respiratory tract and then swallowed (Hayes 1982). Pharmacokinetic data using lung dosimetry 

models would provide useful information as to the contribution of the inhalation route to total intake; 

however, direct absorption in the lung is probably minimal. An oral MRL for chronic-duration exposure 

was not derived since the most sensitive noncancer (hepatic) effects were observed at doses higher than 

the doses at which the most sensitive acute- and intermediate-duration effects occurred. This dietary level 

was the lowest level tested in the study (Fitzhugh 1948). Because few experimental details were given in 

the Fitzhugh (1948) report, it may be desirable to replicate the findings using modern histopathological 

techniques and current views on the physiological meaning of specific liver changes. 

The possible association between exposure to DDT and various types of cancers in humans has been 

studied extensively (Aronson et al. 2000; Cocco and Benichou 1998; Cocco et al. 1997a, 1997b, 2000; 

Demers et al. 2000; Dewailly et al. 1994; Dorgan et al. 1999; Garabrant et al. 1992; Hardell et al. 1996; 

Helzlsouer et al. 1999; Hoppin et al. 2000; Krieger et al. 1994; Laden et al. 2001a; Liljegren et al. 1998; 

Lopez-Carrillo et al. 1997; Moysich et al. 1998; Romieu et al. 2000; Schecter et al. 1997; Sturgeon et al. 

1998; van’t Veer et al. 1997; Ward et al. 2000; Wasserman et al. 1976; Wolff et al. 1993, 2000a; Zheng et 

al. 1999, 2000). Nearly all of the human studies evaluated the carcinogenicity of either p,p’-DDT or 

p,p’-DDE. Cancer end points that have been investigated in humans include respiratory system, 

pancreatic, endometrial, breast, prostate, liver, and testicular cancers, Hodgkin’s and non-Hodgkin’s 

lymphomas, and multiple myeloma. Thus far, there is no conclusive evidence linking DDT and related 
 

DDT, DDE, and DDD 201 

3. HEALTH EFFECTS 

compounds to cancer in humans. It is expected that some of these cohorts will continue to be monitored 

to ensure that latency is appropriately accounted for. Other areas where data gaps exist include exposure 

assessment in epidemiological studies; studies on validity and reliability of recall of pesticide use for 

occupationally and nonoccupationally exposed populations; and information on inert ingredients of 

pesticide formulations, where specific pesticides are still used, since inert components are not necessarily 

biologically inert. The finding of Demers et al. (2000) of a potential association between blood DDE 

levels in breast cancer patients and aggressiveness of the disease requires replication in additional studies 

to rule it out as artifactual. In addition, further attention should be paid to physiological changes that may 

occur in cancer patients that could alter the pharmacokinetics of the biomarkers of exposure. This issue 

was brought up by Hoppin et al. (2000) in their study of patients with pancreatic cancer, a condition in 

which wasting may greatly diminish the lipid pool and thus give artificially-elevated values of body 

burden for lipid-soluble chemicals. 

Animal data provide sufficient evidence of carcinogenicity via oral exposure. DDT has been shown to be 

carcinogenic in a number of studies in rodents (Cabral et al. 1982b; Innes et al. 1969; Kashyap et al. 

1977; Rossi et al. 1977; Thorpe and Walker 1973). DDT and related compounds caused primarily liver 

cancer. Information on the mechanism of action for cancer induction by DDT in these susceptible species 

and on whether or not species-specific biomarkers exist would be helpful. For example, mice and 

hamsters metabolize DDT by similar metabolic pathways. Therefore, it is unlikely that the species 

difference is due to differences in the production of DDMU-epoxide or DDA-Cl (Gold and Brunk 1983). 

However, the hamster is less effective in the conversion of DDT to DDE (Gingell 1976; Gold and Brunk 

1983). DDE has been shown to cause liver tumors in hamsters (Rossi et al. 1983). DDT has been shown 

to promote the carcinogenicity of other substances (Diwan et al. 1994; Kitagawa et al. 1984; Nishizumi 

1979; Peraino et al. 1975; Preat et al. 1986; Rojanapo et al. 1993). On the other hand, some studies have 

observed an inhibitory action of DDT on the carcinogenicity of other chemicals (Rojanapo et al. 1993; 

Silinskas and Okey 1975). Virtually no information was located regarding the carcinogenicity of DDT by 

the inhalation and dermal routes of exposure, but additional bioassays do not seem necessary at this time. 

Genotoxicity. There is no conclusive evidence that DDT and related compounds are genotoxic in 

humans. While some studies of humans exposed occupationally have reported chromosomal aberrations 

(Rabello et al. 1975; Rupa et al. 1988, 1989; Yoder et al. 1975), the subjects were also exposed to other 

pesticides and it is unclear whether additional potential confounders were adequately controlled. A recent 

study of subjects residing near a waste site found no association between serum levels of DDE and the 

incidence of micronuclei in peripheral lymphocytes (Vine et al. 2001). Two high-dose oral studies in 
 

DDT, DDE, and DDD 202 

3. HEALTH EFFECTS 

rodents, one in rats (Palmer et al. 1973) and the other in mice (Clark 1974), reported increases in 

dominant lethality. In general, the results from standard gene mutation tests in bacteria did not show 

mutagenicity (Fahrig 1974; Fluck et al. 1976; McCann et al. 1975; Shirasu et al. 1976) and conflicting 

results have been obtained in tests for chromosomal aberrations in cultured rodent and human cells 

(Kelly-Garvert and Legator 1973; Palmer et al. 1972; Probst et al. 1981; Tong et al. 1981; Tsuchimoto et 

al. 1983). It is unlikely that further studies will provide new key information. 

Reproductive Toxicity. Several studies in humans have examined possible associations between 

body burdens of DDT and analogues and alterations of some aspects of reproduction. Duration of 

lactation was found to be inversely related to the concentration of p,p’-DDE in breast milk in two studies 

(Gladen and Rogan 1995; Rogan et al. 1987). Some studies also found weak associations between blood 

DDT/DDE levels and miscarriages (Gerhardt et al. 1998) and infertility (Gerhardt et al. 1999), but these 

findings need to be confirmed. Additional studies have examined associations between body burdens of 

DDT, DDE, and DDD and premature infants and abortions (Saxena et al. 1980, 1981; Wassermann et al. 

1982), but the possible role of other chemicals, such as PCBs and other chlorinated pesticides, could not 

be ruled out. A recent study in the United States found increased odds of having pre-term infants and 

small-for-gestational-age infants in women having blood concentrations of DDE $10 ppb in samples 

taken in the third trimester of pregnancy (Longnecker et al. 2001). There was essentially no relation at 

lower DDE concentrations. Overall, there is no conclusive evidence that body burdens of 

DDT/DDE/DDD currently found in the general U.S. population represent a reproductive hazard, but 

adverse reproductive outcomes may be a concern for women in countries where these chemicals are still 

used. 

Acute exposure to DDT by the oral route in animals has been associated with reproductive effects, 

including a decrease in fertility in male rats (Krause et al. 1975). Reproductive effects observed in 

animals following DDT exposures of intermediate duration are similar to those effects observed with 

excess estrogen, including infertility (Green 1969; Jonsson et al. 1976), and decreases in implanted ova 

(Lundberg 1973). Adult male rats showed a decrease in seminal vesicle and ventral prostate weight after 

short-term treatment with p,p’-DDE, which is thought to be due to the antiandrogenic effects of p,p’-DDE 

(Kelce et al. 1995, 1997). One study in vitro showed that all DDT/DDE/DDD isomers can exhibit some 

degree of antiandrogenicity (Maness et al. 1998), but it is not known whether isomers other than 

p,p’-DDE have antiandrogenic properties in vivo. Also, it would be useful to determine whether the 

persistent DDT metabolite, 3-MeSO2-DDE, functions as an androgen antagonist in both in vitro binding 

assays and gene expression assays, and how its potency compares to that of DDE itself. 
 

DDT, DDE, and DDD 203 

3. HEALTH EFFECTS 

Multigeneration studies have been conducted in rats (Duby et al. 1971; Green 1969; Ottoboni 1969, 1972; 

Treon et al. 1954), mice (Keplinger 1970), and dogs (Deichman et al. 1971; Ottoboni et al. 1977). The 

infertility observed in the Green (1969) study at the very low dose (0.56 mg/kg/day) raises potential 

concerns about human reproductive effects. However, Keplinger (1970) saw no effect at 3.2 mg/kg/day, 

although there was a decrease in fertility (as defined by decreased viability through the lactation period) 

at 32 mg/kg/day. No effects on fertility were seen in 2 or 3 generations of rats exposed to 1.25 mg/kg/day 

(Treon et al. 1954; Ottoboni 1969) or 1.5 mg/kg/day of technical-grade DDT (Duby et al. 1971), nor in 

3 generations of beagles exposed to 10 mg/kg/day (Ottoboni et al. 1977). The Deichman et al. (1971) was 

a mixed exposure to aldrin and DDT from which no conclusions can be drawn. Unfortunately, none of 

the multigeneration studies mentioned above was conducted according to current scientific standards; 

flaws exist in the areas of adequate sample sizes, statistical testing (an lack thereof), and reporting of 

sufficient experimental detail. Thus, there is a critical data need for a well designed multigeneration 

animal study to reduce the scientific uncertainties about the effects of DDT on fertility. 

Well-designed experiments to identify sensitive time periods of exposure and to clarify dose-response 

relationships for these effects would be useful when deriving an MRL or assessing the potential hazard 

resulting from environmental exposures. Because exposure to DDT from hazardous waste sites is of 

concern, additional studies are needed to assess the reproductive toxicity following exposure to doses 

similar to those estimated at hazardous waste sites. However, DDT-induced effects on fertility may be 

difficult to detect since the baseline human infertility rate is so high. Levels in environmental media are 

not expected to be high enough to result in high inhalation or dermal exposures, and therefore, additional 

inhalation or dermal exposure reproductive studies do not seem to be warranted at this time. 

Developmental Toxicity. There are limited data on developmental effects of DDT and analogues in 

humans. A study of children from North Carolina found that transplacental exposure of DDE was 

associated with increased height and weight adjusted for height in boys at puberty (Gladen et al. 2000). A 

study of German children found an inverse association between serum DDE levels and height in girls 

(Karmaus et al. 2002). Also, a study of Inuit infants found that prenatal exposure to DDE and other 

chemicals may have been associated with increased risk of otitis media during the first year of life 

(Dewailly et al. 2000). 

While the limited data in humans do not allow for definite conclusions to be drawn, data in animals 

indicate that perinatal exposure to DDT/DDE/DDD (Craig and Ogilvie 1974; Fabro et al. 1984; Gellert 

and Heinricks 1975; Gray et al. 1999; Hart et al. 1971, 1972; Kelce et al. 1995; Loeffler and Peterson 
 

DDT, DDE, and DDD 204 

3. HEALTH EFFECTS 

1999; You et al. 1998, 1999a) or neonatal exposure to DDT (Eriksson et al. 1990a, 1990b, 1992, 1993; 

Johansson et al. 1995, 1996; Talts et al. 1998) can cause adverse developmental effects in the offspring. 

These effects depend on the dose administered, the timing of exposure during or after gestation, and the 

specific isomer administered. The o,p’-isomers of DDT, DDE, and DDD have been associated with 

estrogen-like effects in the reproductive system and p,p’-DDE has been associated with antiandrogenic 

effects. Further information concerning these factors and their impact on the developmental toxicity of 

DDT, DDE, and DDD would be helpful. For example, cross-fostering studies could help in determining 

the relative impact of gestational versus lactational exposure to p,p’-DDE in relation to delays in the onset 

of puberty. Standardization of both in vivo (pre and postnatal exposures) and in vitro tests for 

estrogenicity and antiandrogenicity of DDT and related compounds would be helpful to compare results 

between research groups. There are limited data suggesting that if mice exposed to DDT in utero and 

during lactation are further exposed to DDT postnatally, responses in both immunological plaque forming 

assays and lymphoproliferative assays are reduced (Rehana and Rao 1992). The results of this study are 

difficult to interpret because of experimental design issues, including the lack of a comparison unexposed 

group and ambiguity about whether statistical testing was done. It would be useful to try to replicate 

these results using a better study design. The results from the series of studies by Eriksson et al. (1990a, 

1990b, 1992, 1993) and Johansson et al. (1995, 1996) on mice exposed perinatally were used as the basis 

for deriving an acute oral MRL. Duplication of these results by other laboratories would greatly increase 

the confidence in these findings. Developmental effects have also been observed in animals following 

intermediate- and chronic-duration oral exposures to DDT. These effects included slowed development 

and premature puberty (Clement and Okey 1974; Craig and Ogilvie 1974; Naishtein and Leibovich 1971; 

Ottoboni et al. 1977; Tomatis et al. 1972; Turusov et al. 1973). Additional studies to assess the 

mechanism of the developmental toxicity, the critical stages in perinatal development affected by DDT, 

and the dose-response relationships would be helpful. More data on prenatal and postnatal exposures and 

postnatal developmental effects might also be useful. No developmental studies by the inhalation and 

dermal routes of exposure were located. However, since levels in environmental media are not expected 

to be high enough to result in high inhalation or dermal exposures, additional inhalation or dermal 

exposure studies assessing developmental effects do not seem to be warranted at this time. 

Immunotoxicity. Evidence of immunotoxicity in humans is inconclusive and is limited to a study 

with only 3 volunteers (Shiplov et al. 1972), a study of 12 fish-consumers in Sweden (Svensson et al. 

1994) and a more recent study of subjects who resided near a waste site (Vine et al. 2001). Acute-, 

intermediate-, and chronic-duration oral studies in animals provide evidence that DDT may cause 

immunological effects. Effects reported included decreases in antibody titers and plaque-forming cells 
 

DDT, DDE, and DDD 205 

3. HEALTH EFFECTS 

(Banerjee 1987a; Banerjee et al. 1986), increases in gamma globulin and serum immunoglobulin and a 

decreased tuberculin skin reaction (Street and Sharma 1975), alterations in the spleen (Deichmann et al. 

1967), and increased growth of the leprosy bacterium (Banerjee et al. 1997a). The immunotoxicity of 

DDT compounds is enhanced by a low-protein diet (Banerjee et al. 1995) and physical/emotional stress 

(Banerjee et al. 1997b), while concurrent treatment with ascorbic acid appears to attenuate DDT-induced 

immunotoxicity (Koner et al. 1998). No clear pattern of relative humoral or cell-mediated 

immunotoxicity was seen from either DDT, DDE, or DDD (Banerjee et al. 1996). In view of the 

complexity of the immune system, a multiple assay battery would be helpful in order to evaluate the 

effects of DDT on major components of the immune system. No immunological studies by the inhalation 

and dermal routes of exposure were located. However, since levels in environmental media are not 

expected to be high enough to result in high inhalation or dermal exposures, additional inhalation or 

dermal exposure studies assessing immunological effects do not seem to be warranted at this time. 

Neurotoxicity. While there are several studies that indicate overt clinical signs of neurotoxicity to 

humans acutely exposed to relatively high oral doses (Hayes 1982; Hsieh 1954; Velbinger 1947a, 1947b), 

there is limited information on effects of chronic exposure. Two studies, one in workers (Ortelee 1958) 

and one in volunteers (Hayes et al. 1956), found no evidence of adverse neurological effects among the 

subjects studied. However, a recent small study of retired malaria-control workers in Costa Rica found 

mildly impaired neurological functions among the workers relative to a control group (van Wendel de 

Joode et al. 2001). Evaluation of further cohorts from countries where DDT is still used may provide 

relevant information. 

Studies in animals indicate that DDT may affect the level of neurotransmitters and the amount of lipids in 

the brain (Eriksson and Nordberg 1986; Herr et al. 1986; Hong et al. 1986; Hudson et al. 1985; Sanyal et 

al. 1986). Clinical observations of overt neurotoxicity such as tremors or convulsions have been reported 

(Herr and Tilson 1987; Hong et al. 1986; Matin et al. 1981). Neurotoxicity can be caused in adult 

animals and can be manifested as tremors or convulsions. Developmental neurotoxicity has also been 

reported in a number of studies (Eriksson and Nordberg 1986; Eriksson et al. 1990a, 1990b). This 

neurotoxicity that occurred during neonatal brain developmental stages resulted in behavioral deficits in 

the adult mouse and correlated with changes in muscarinic acetylcholine receptors in the cerebrum. 

Behavioral deficits in treated adult animals have also been reported (Sobotka 1971); however, doses were 

50-fold higher than those given to neonates. Information clarifying the mechanism of action in the 

neonate and in the adult mouse as well as data to describe the dose-response relationship for these effects 

would be extremely useful in further identifying sensitive subpopulations. A battery of neurotoxicity 
 

 

DDT, DDE, and DDD 206 

3. HEALTH EFFECTS 

tests would provide additional information on the neurotoxicity in animals, which might then be related to 

possible neurotoxic effects in humans. No data were located on neurological effects after inhalation 

exposure in humans or animals or after dermal exposure in humans. A single study reported neurological 

effects in animals after dermal exposure to DDT (Cameron and Burgess 1945). However, since levels in 

environmental media are not expected to be high enough to result in high inhalation or dermal exposures, 

additional inhalation or dermal exposure studies assessing neurological effects do not seem to be 

warranted at this time. 

Epidemiological and Human Dosimetry Studies. Known acute health effects in humans at high 

exposure levels of DDT, DDE, or DDD are irritation of the eyes, nose, and throat, sweating, nausea, 

headache, tremors, and convulsions. This information comes from clinical studies and from studies in 

which volunteers ingested measured amounts of DDT and DDE (Hayes 1982; Hsieh 1954; Velbinger 

1947a, 1947b). Effects in animals include liver alterations, developmental and reproductive effects, and 

neurological effects. More information on the effects of DDT, DDE, or DDD could be obtained from 

epidemiological studies of people who, because of proximity to areas where high concentrations of DDT, 

DDE, or DDD have been found, may have higher exposure to DDT, DDE, or DDD. Also, more insight 

could be gained through future monitoring at National Priorities List (NPL) sites. Because of the virtually 

ubiquitous distribution of DDT in the environment, some DDT would be expected to be detected in 

tissues of the majority of the general population. Studies have monitored human tissue and blood for 

DDT and its metabolites, but no correlation has been made between the levels found in these tissues and 

specific disease states. Based on the known estrogenic effects of the o,p’-isomers and antiandrogenic 

properties of p,p’-DDE in animals, monitoring these chemicals in human tissues seems important. 

Identification and follow-up of cohorts who may have been exposed to DDT before its ban would be 

useful for evaluation of health conditions that have long latency, such as cancer. Also of interest is the 

assessment of long-term low level exposure as well as potential delayed effects of acute high exposure. 

Pharmacokinetic studies to characterize the appropriate measurement of absorbed dose would be useful in 

future epidemiological studies. 

Biomarkers of Exposure and Effect. 

Exposure. DDT, DDE, and DDD can be measured in numerous body tissues and fluids including blood, 

serum, urine, feces, adipose tissue, amniotic fluid, breast milk, and semen. The presence of these 

compounds in blood or urine can be used to determine the relative amount of exposure of an individual, 

but total exposure cannot be quantitated. Methylsulfonyl metabolites of DDT have also been identified in 
 

DDT, DDE, and DDD 207 

3. HEALTH EFFECTS 

samples of human adipose and breast milk. Other potential biomarkers of exposure have been identified, 

but again, exposure cannot be quantified. More information could be provided by studies designed to 

correlate biomarkers of exposure with the temporal aspect of exposure. 

Effect. No biomarkers of effect specific for DDT, DDE, or DDD have been identified in the literature. 

Nonspecific biomarkers of effect include tremors, convulsions, and an increase in hepatic microsomal 

enzymes. Studies designed to identify specific biomarkers of effect for DDT, DDE, and DDD would be 

useful. 

Absorption, Distribution, Metabolism, and Excretion. Qualitative information from 

occupational and ingestion studies indicates that humans absorb DDT via inhalation, dermal, and oral 

routes of administration (Laws et al. 1967; Morgan and Roan 1971, 1974; Morgan et al. 1980). No 

quantitative information exists concerning the rate or extent of absorption following inhalation or dermal 

exposure, although some information exists that quantifies the extent of absorption following oral 

administration. The bioavailability of DDT from environmental media, such as soil or food, is not well 

characterized. Quantitative data (e.g., absorption rates) from animals exposed by oral and dermal routes 

in different environmental media would be useful in providing information on absorption of DDT, DDE, 

and DDD to be used in estimating absorption in humans following exposure by these routes in these 

environmental media. 

Information exists on the distribution of DDT, DDE, and DDD and on the storage and release from 

storage of these compounds (Hayes et al. 1971; Morgan and Roan 1971, 1974). However, there is limited 

information on the long-term release rates from adipose tissue. This information would be helpful in 

determining the retention time of DDT in humans. 

The ultimate metabolites of DDT, which can be isolated from animals, have been well described. 

However, two models have been proposed for the intermediate products between parent and ultimate 

metabolites. These models differ in the quantity of potential electrophilic intermediates (an epoxide and 

an acylating agent) produced. These electrophilic metabolites have not been isolated, but their presence 

may be confirmed if DNA adducts are found. Some species-specific metabolic differences, especially in 

the area of efficiency of the conversion of DDT to DDE in hamsters relative to other species, have been 

identified (Gold and Brunk 1982, 1983). However, the role of these metabolic differences in speciesspecific sensitivity to toxicity, especially carcinogenicity, is not well characterized. Further information 
 

 

DDT, DDE, and DDD 208 

3. HEALTH EFFECTS 

concerning the species-specific metabolic differences would be useful to provide more specific 

information when comparing the toxicokinetics of these substances in humans and animals. 

Comparative Toxicokinetics. Metabolism studies indicate that the metabolism of DDT is 

qualitatively similar among several species, but that the efficiency of formation of certain metabolites and 

the proportion of metabolites excreted may be quantitatively different (Gingell 1976; Gold and Brunk 

1982, 1983, 1984; Peterson and Robison 1964). Comparisons of elimination rates of DDT from fat 

showed that the process is the slowest in humans, followed by monkeys, dogs, and rats (Morgan and 

Roan 1974). Rats eliminate DDT 10 to 100 times faster than humans. Morgan and Roan (1974) 

suggested that the differences in elimination in rats could be due to differences in liver metabolism, gut 

bacterial metabolism, enterohepatic recirculation, or factors related to the accessibility of plasmatransported pesticide to the excretory cells of the liver. Differences in the metabolism of DDT among 

species may account for differences in toxic responses, especially cancer. The potential for DDT to 

produce toxic effects has been investigated in rats, dogs, mice, guinea pigs, and nonhuman primates, but 

the animal species that serves as the best model for extrapolating results to humans has not been 

determined. Ethical considerations limit the amount of information that can be obtained in humans, but 

analysis of urine of persons with known exposure to DDT to determine levels of parent compound and 

metabolites could provide more information on the metabolic pathways in humans. This information 

could help to identify the most appropriate animal model for extrapolation to humans. 

Methods of Reducing Toxic Effects. The available data indicate some ways in which peak 

absorption of DDT, DDE, and DDD might be reduced following oral or dermal exposure (Dreisbach 

1983; HSDB 2002a, 2002b, 2002c, 2002d). Studies that examine the efficacy of gastric lavage with 

mannitol versus other cathartics or of combinations of cathartics would be useful. Also, studies that 

evaluate the effectiveness of intestinal absorbants such as sodium muconate or cholestyramine would 

provide valuable information. No data were located regarding methods for reducing absorption following 

inhalation exposure. 

In humans, DDT and its fat soluble metabolites are stored in adipose tissue, where they are retained for 

long periods of time. During that time, short term toxicity may be expected to be minimal since exposure 

to target organs would be via the circulation and the equilibrium between blood and adipose tissue is such 

that blood levels are about 1:280 of fat levels; however, there is still the potential for chronic toxicity and 

interactions with other environmental chemicals. If very large amounts of DDT or its metabolites are 

stored in adipose tissue, then rapid release of DDT or its metabolites into the circulation during 
 

DDT, DDE, and DDD 209 

3. HEALTH EFFECTS 

mobilization of fat stores, such as during rapid weight loss, may result in certain types of toxic effects. 

Release of DDT into the blood of a pregnant woman during gestation-related fat mobilization might result 

in toxicity to the developing fetus. Also, mobilization of adipose fat for lactation might provide an 

increased amount of DDT to a breast-feeding infant. Development of methods to enhance excretion of 

adipose-sequestered DDT, while minimizing toxicity would be beneficial in reducing the body burden. 

DDT induces neurotoxicity (tremors and convulsions) by various mechanisms (see Section 3.5). No 

method is currently available for directly reducing the neurotoxic effects of DDT, DDE, or DDD by 

interfering with the mechanisms of action, although there are a variety of pharmacological methods for 

alleviating nonspecific symptoms. 

Children’s Susceptibility. The information on health effects of DDT and analogues in humans is 

derived mainly from accidental exposure and from controlled studies in volunteers, and the main adverse 

effect is neurotoxicity. No reports on exposed children were found, but it is reasonable to assume that 

children will exhibit signs and symptoms similar to those in adults under similar exposure conditions. 

There is no information on whether the developmental process is altered in humans exposed to DDT. 

Studies in animals have shown that DDT and analogues can alter the development and maturation of the 

male and female reproductive system (Bitman and Cecil 1970; Clement and Okey 1972; Duby et al. 1971; 

Gellert et al. 1972; Gray et al. 1999; Kelce et al. 1995, 1997; Loeffler and Peterson 1999; Singhal et al. 

1970; You et al. 1998, 1999a). Hormone-like effects have been reproduced in many types of in vitro test 

systems, but there is a need for standardized operating procedures. In vivo tests are preferred over in vitro 

assays because they take into account pharmacokinetic and pharmacodynamic interactions. 

There are no adequate data to evaluate whether pharmacokinetics of DDT in children are different from 

adults. DDT and analogues can cross the placenta and are transferred to offspring via breast milk. It is 

unknown whether the efficiency of gastrointestinal absorption of DDT and analogues in nursing neonates 

differs from adults and what influence the fat content of human milk might make. Further information on 

the dynamics of DDT and analogues during pregnancy and lactation, such as further refinement of PBPK 

models to include humans would be useful. Important information was recently published on estimates of 

body burden of DDE that result from nursing by using a model that incorporates a wide array of variables 

(LaKind et al. 2000). The only existing PBPK model for DDT is that of You et al. (1999b), which 

focuses on pregnant and lactating Sprague-Dawley rats. There is no information to evaluate whether 

metabolism of DDT is different in children than in adults since the specific phase I and II enzymes 
DDT, DDE, and DDD 210 

3. HEALTH EFFECTS 

involved in DDT metabolism have not been identified; it is unknown which phase I P-450 isozymes 

metabolize DDT. 

There is little evidence about whether children or young animals differ in their susceptibility to the health 

effects from DDT from adults. In fact, some animal studies found that young rats are less susceptible 

than older ones to the acute neurotoxic effects produced by a single dose of DDT (Lu et al. 1965). 

However, the relevance of this information to human health is unknown. There is evidence that acute 

perinatal exposure to DDT in mice results in altered behavioral responses in the mice tested as adults 

(Eriksson et al. 1992, 1993; Johansson et al. 1995, 1996; Talts et al. 1998). Research efforts should focus 

on the possible underlying mechanism(s) that are responsible for such long-lasting postexposure 

alterations. 

Continued research into the development of sensitive and specific biomarkers of exposure and effect for 

DDT and analogues would be valuable for both adults and children. There are no data on the interactions 

of DDT with other chemicals in children or adults. There are no pediatric-specific methods to reduce 

peak absorption for DDT following exposure, to reduce body burdens, or to interfere with the mechanism 

of action. Based on the information available, it is reasonable to assume that methods recommended for 

treating adults will also be applicable to children; however, these methods need to be validated in 

children. 

Child health data needs relating to exposure are discussed in 6.8.1 Identification of Data Needs: 

Exposures of Children. 

3.13.3 Ongoing Studies 

A number of studies concerning health effects associated with DDT, DDE, and DDD have been identified 

in the Federal Research in Progress (FEDRIP 2001) database and are listed in Table 3-7. 
DDT, DDE, and DDD 211 

3. HEALTH EFFECTS 

Table 3-7. Ongoing Studies on the Health Effects 

of DDT and DDT Analogues 

Investigator Affiliationa Research description Sponsor 

Bhatia, R Public Health 

Institute, Berkeley, 

CA 

DDT and other organochlorine pesticides 

and male genital anomalies 

NIEHS 

Charles, MJ University of 

California, 

Environmental 

Toxicology Center, 

Berkeley, CA 

Exploration of linkages among 

organochlorines (DDT, DDE), oxidative 

DNA damage and breast cancer 

USDA 

Clark, DR Columbia 

Environmental 

Ecotoxicological investigations of effects 

of DDT on native 

USGS 

Research Center 

Columbia, MO 

wildlife species 

Cohn, BA Public Health 

Institute, Berkeley, 

CA 

Breast cancer and organochlorines NCI 

Cohn, BA Public Health 

Institute, Berkeley, 

CA 

Prenatal exposure to organochlorine and 

fecundity 

NIEHS 

Darvill, T SUNY at Oswego, 

Oswego, NY 

Prenatal exposure to PCBs, DDE and 

other contaminants on cognitive 

development in school-age children 

NIEHS 

Denslow, ND University of Florida, 

Gainesville, FL 

Chlorinated hydrocarbons and non ionic 

surfactants effect on endocrine functions 

NIEHS 

Fraser, PA Center for Blood 

Research, 

Boston, MA 

Genes and chemical exposure associated 

with SLE risk 

NIEHS 

Gammon, MD Columbia University 

Health Sciences, 

New York, NY 

Breast cancer and the environment on 

Long Island 

NCI 

Gross, TS University of Florida, 

Gainesville, FL 

Endocrine, reproductive, and 

developmental disrupting effects of 

chlorinated hydrocarbons on wildlife 

USGS 
DDT, DDE, and DDD 212 

3. HEALTH EFFECTS 

Table 3-7. Ongoing Studies on the Health Effects 

of DDT and DDT Analogues (continued) 

Investigator Affiliation Research description Sponsor 

Gross, TS Florida Caribbean 

Science Center 

Gainesville, FL 

Immuno-toxicity and endocrine disruption 

in Lake Apopka alligator neonates 

USGS 

Gurpide, E Mount Sinai School 

of Medicine of 

CUNY, New York, 

NY 

Hormonal activity of chlorinated 

hydrocarbons from New York harbor 

sediments 

NIEHS 

Hauser, RB Harvard University, 

School of Public 

Health, 

Boston, MA 

Environmental organochlorines and 

semen quality 

NIEHS 

Henny, CJ Forest and 

Rangeland 

Ecosystem Science 

Center 

Effects of persistent organochlorines on 

nesting success of Osprey along the 

Columbia River system 

USGS 

Corvallis, OR 

Hothem, R Western Ecological 

Research Center 

Sacramento, CA 

Contaminant residues and their effects on 

reproductive success of aquatic 

birds nesting at Edwards Air Force Base 

USGS 

Hothem, R Western Ecological 

Research Center 

Sacramento, CA 

Reproduction by black-crowned 

night-herons and snowy egrets on 

Alcatraz Island, California 

USGS 

Korrick, SA Harvard University, 

School of Public 

Health, 

Boston, MA 

In utero PCB, pesticide and metal 

exposure and childhood cognition 

NIEHS 

Landrigan, PJ Mount Sinai School 

of Medicine of 

Environmental distribution and toxic 

effects on human health of PCBs and DDT 

NIEHS 

CUNY, New York, 

NY 

in New York City 

Longnecker, MP NIEHS Human health effects of exposure to 

organochlorine compounds 

NIEHS 

Matsumura, F University of 

California, 

Davis, CA 

Mechanism of insecticides on the 

development of prostate cancer 

USDA 

Remington, P University of 

Wisconsin Madison, 

Madison, WI 

Regional variation of breast cancer rates 

in Wisconsin 

NCI 

Rice, RH University of 

California, 

Environmental 

Toxicology Center, 

Davis, CA 

Toxic and mutagenic responses of 

keratinocytes to heterocyclic amines and 

polycyclic aromatic hydrocarbons 

USDA 
DDT, DDE, and DDD 213 

3. HEALTH EFFECTS 

Table 3-7. Ongoing Studies on the Health Effects 

of DDT and DDT Analogues (continued) 

Investigator Affiliation Research description Sponsor 

Rogan, WJ NIH, NIEHS Human exposure to halogenated aromatic 

compounds 

NIEHS 

Schwartz, S Fred Hutchinson 

Cancer Research 

Center, 

Seattle, WA 

Risk associated with exposure to 

phytoestrogens, organochlorines, and 

fibroids 

NIEHS 

Seegal, RF Wadsworth Center, 

Albany, NY 

Developmental effects of fish borne 

toxicants in the rat 

NIEHS 

Stellman, SD American Health 

Foundation, Valhalla, 

NY 

Epidemiology of breast cancer NCI 

Terranova, PF University of Kansas 

Medical Center, 

Kansas City, KS 

Effects of dioxin, xenoestrogen, and 

estradiol on reproductive processes 

NSF 

Trosko, JE Michigan State 

University, 

East Lansing, MI 

Cell-cell communication carcinogenesis NCI 

Trosko, JE Michigan State 

University, 

East Lansing, MI 

Epigenic effects of environmental 

toxicants on cellular communication 

pathways 

NIEHS 

Weston, A Mount Sinai School 

of Medicine of 

CUNY, New York 

Effects of PCB-containing river sediments 

on carcinogen metabolism 

NIEHS 

Windham, GC Public Health 

Institute, Berkeley, 

CA 

Organochlorine compounds and menstrual 

cycle function 

NIEHS 

Wolff, MS Mount Sinai School 

of Medicine of 

PBC and DDE body burden and breast 

cancer 

NIEHS 

CUNY, New York 

a

Post office state abbreviations used 

DNA = deoxyribonucleic acid; NCI = National Cancer Institute; NIEHS = National Institute of Environmental Health 

Sciences; NSF =National Science Foundation; PCB = polychlorinated biphenyl; USDA = United States Department 

of Agriculture; USGS = United States Geological Survey 

DDT, DDE, and DDD 215 

4. CHEMICAL AND PHYSICAL INFORMATION 

4.1 CHEMICAL IDENTITY 

When we refer to DDT, we are generally referring to p,p’-DDT, which was produced and used for its 

insecticidal properties. However, technical grade DDT, the grade that was generally used as an 

insecticide, was composed of up to fourteen chemical compounds, of which only 65–80% was the active 

ingredient, p,p’-DDT. The other components included 15–21% of the nearly inactive o,p’-DDT, up to 

4% of p,p’-DDD, and up to 1.5% of 1-(p-chlorophenyl)-2,2,2-trichloroethanol (Metcalf 1995). 

The chemical formulas, structures, and identification numbers for p,p’-DDT, p,p’-DDE, p,p’-DDD, 

o,p’-DDT, o,p’-DDE, and o,p’-DDD are listed in Table 4-1. The latter five compounds are either 

impurities or metabolites of technical DDT. 

4.2 PHYSICAL AND CHEMICAL PROPERTIES 

Technical DDT is a white amorphous powder that melts over the range of 80–94 EC (Metcalf 1995). 

Physical and chemical properties of p,p’-DDT, p,p’-DDE, p,p’-DDD, o,p’-DDT, o,p’-DDE, and 

o,p’-DDD are listed in Table 4-2. 
Table 4-1. Chemical Identity of p,p'- and o,p'-DDT, DDE, and DDDa 

Characteristic p,p'-DDT p,p'-DDE p,p'-DDD Synonym(s) Registered trade name(s) Chemical formula Chemical structure Identification numbers: 

CAS registry NIOSH RTECS EPA hazardous waste OHM/TADS DOT/UN/NA/IMCO shipping HSDB NCI 

4,4'-DDT; 1,1,1-trichloro-2,2-bis 

(p-chlorophenyl)ethane; dichlorodiphenyl trichloroethane; DDT; 1,1'­(2,2,2-trichloroethylidene)bis(4-chlorobenzene); α-α-bis(p-chlorophenyl)­β, β, β -trichloroethane 

Genitox, Anofex, Detoxan, Neocid, 

Gesarol, Pentachlorin, Dicophane, Chlorophenothaneb 

C14H9Cl5 

H 

Cl Cl 

CCl3 

50-29-3 KJ3325000 UO61 7216510 IMCO 6.1; UN2761 200 C00465 

4,4'-DDE; dichlorodiphenyldichloroethane; 1,1-dichloro2,2-bis(p-chlorophenyl) ethylene; 1,1'-(2,2-dichloroethylidene)bis(4­chlorobenzene); DDE No data C14H8Cl4 

Cl Cl 

CCl2 

72-55-9 KV9450000 No data No data No data 1625 C00555 

4,4'-DDD; DDbis(p-chlorop(4-chloropheTDE; tetrach

DDD; Rothan

C14H10Cl4 

Cl 

72-54-8 KI0700000 U060 7215098 NA2761; TD285 C00475 
Table 4-1. Chemical Identity of p,p'-and o,p'-DDT, DDE, and DDDa (continued) 

Characteristic o,p'-DDT o,p'-DDE o,p'-DDD Synonym(s) Registered trade name(s) Chemical formula Chemical structure Identification numbers: 

CAS registry NIOSH RTECS EPA hazardous waste OHM/TADS DOT/UN/NA/IMCO shipping HSDB NCI 

4,4'-DDT; 1,1,1-trichloro-2­(o-chlorophenyl)-2-(p-chlorophenyl)­

ethane; o,p'-dichlorodiphenyltrichloroethane 

No data 

C14H9Cl5 

H 

CCl3 

Cl 

Cl 

789-02-6 No data No data No data No data No data No data 

2,4'-DDE; 1,1-dichloro-2­(o-chlorophenyl)-2-(p-chlorophenyl) ethylene; 1-chloro-2-(2,2-dichloro-1­(4-chlorophenyl)ethenylbenzene 

No data 

C14H8Cl4 

CCl2 

Cl 

Cl 

3424-82-6 No data No data No data No data No data No data 

2,4'-DDD; M1,1-dichloro-2-(p-chloropChoditane; 22-(p-chloropLysodren 

C14H10Cl4 

Cl 

53-19-0 KH7880000 No data No data No data 3240 C04933 

aAll information obtained from HSDB 1999a, 1999b, 1999c, 1999d, or Howard and Neal 1992 except where noted.bKlassen et al. 1991

CAS = Chemical Abstracts Service; DOT/UN/NA/IMCO = Department of Transportation/United Nations/North America/International Maritime Dangerous G

EPA = Environmental Protection Agency; HSDB = Hazardous Substances Data Bank; NCI = National Cancer Institute; NIOSH = National Institute for OccOHM/TADS = Oil and Hazardous Materials/Technical Assistance Data System; RTECS = Registry of Toxic Effects of Chemical Substances
DDT, DDE, and DDD 218 

4. CHEMICAL AND PHYSICAL INFORMATION 

Table 4-2. Physical and Chemical Properties of p,p'- and o,p'-DDT, DDE, 

and DDDa 

Property p,p'-DDT p,p'-DDE p,p'-DDD 

Molecular weight 

Color 

Physical state 

Melting point 

Boiling point 

Density 

Odor 

Odor threshold: 

Water 

Air 

Solubility: 

Water 

Organic solvents 

Partition coefficients: 

Log Kow 

Log Koc 

Vapor pressure 

Henry's law constant 

Autoignition temperature 

Flashpoint 

Flammability limits 

Conversion factors 

ppm(v/v) to mg/m3

 in air at 20 EC 

mg/m3

 to ppm(v/v) in air at 20 EC 

Explosive limits 

354.49b 

Colorless crystals, 

white powderc 

Solidd 

109 ECb 

Decomposes 

0.98–0.99 g/cm3 

Odorless or weak 

aromatic odore 

0.35 mg/kgc 

No data 

0.025 mg/L at 25 ECb 

Slightly soluble in 

ethanol, very soluble 

in ethyl ether and 

acetonef 

6.91b 

5.18h 

1.60x10-7 at 20 EC, 

torrb 

8.3x10-6 atm-m3/molb 

No data 

72.2–77.2 EC 

No data 

Not applicablek 

Not applicable 

No data 

318.03b 

White 

Crystalline solid 

89 ECb 

336 ECb 

No data 

No data 

No data 

No data 

0.12 mg/L at 25 ECb 

Lipids and most 

organic solvents 

6.51b 

4.70i 

6.0x10-6 at 25 EC, torrb 

2.1x10-5 atm-m3/molb 

No data 

No data 

No data 

Not applicablek 

Not applicable 

No data 

320.05b 

Colorless crystals, white 

powder 

Solid 

109–110 ECb 

350 ECb 

1.385 g/cm3 

Odorless 

No data 

No data 

0.090 mg/L at 25 ECb 

No datag 

6.02b 

5.18j 

1.35x10-6 at 25 EC, torrb 

4.0x10-6 atm-m3/molb 

No data 

No data 

No data 

Not applicablek 

Not applicable 

No data 
DDT, DDE, and DDD 219 

4. CHEMICAL AND PHYSICAL INFORMATION 

Table 4-2. Physical and Chemical Properties of p,p'- and o,p'-DDT, DDE, 

and DDDa (continued) 

Property o,p'-DDT o,p'-DDE o,p'-DDD 

Molecular weight 354.49b 318.03b 320.05b 

Color White crystalline No data No data 

powderc 

Physical state Solidd No data Solid 

Melting point 74.2 ECe No data 76–78 EC 

Boiling point No data No data No data 

Density 0.98–0.99 g/cm3 No data No data 

Odor Odorless or weak No data No data 

aromatic odore 

Odor threshold: 

Water No data No data No data 

Air No data No data No data 

Solubility: 

Water 0.085 mg/L at 25 ECb 0.14 mg/L at 25 ECb 0.1 mg/L at 25 ECb 

Organic solvents No datag No datag Soluble in ethanol, 

isooctane, carbon 

tetrachloridej 

Partition coefficients: 

Log Kow 

Log Koc 

6.79h 

5.35j 

6.00b 

5.19j 

5.87b 

5.19j 

Vapor pressure 1.1x10-7 at 20 EC, 

torrb 

6.2x10-6 at 25 EC, torrb 1.94x10-6 at 30 EC, torrb 

Henry's law constant 5.9x10-7 atm-m3/molb 1.8x10-5 atm-m3/molb 8.17x10-6 atm-m3/molb 

Autoignition temperature No data No data No data 

Flashpoint No data No data No data 

Flammability limits No data No data No data 

Conversion factors 

ppm(v/v) to mg/m3

 in air at 20 EC 

mg/m3

 to ppm(v/v) in air at 20 EC 

Not applicablek 

Not applicable 

Not applicablek 

Not applicable 

Not applicablek 

Not applicable 

Explosive limits No data No data No data 

a

All information obtained from HSDB 1999a, 1999b, 1999c, 1999d unless otherwise noted 

b

Howard and Meylan 1997 

c

Verschueren 1988 

d

NIOSH 1985 

e

Sax 1979 

f

Lide 1998 

g

Chemical is expected to be soluble in most organic compounds. h

Swann et al. 1981 

i

Sablejic 1984

j

Meylan et al. 1992 (values estimated from a fragment constant method) k

Exists partially in particulate form in air. Conversion factors are only applicable for compounds that are entirely in the vapor phase. 

DDT, DDE, and DDD 221 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL
 

Figures relating to the production, import/export, use, and disposal of a pesticide generally refer to those 

of the active ingredient. In the case of DDT, the active ingredient is p,p’-DDT. Most DDT production 

can be assumed to have been technical grade material that included 15–21% of the nearly inactive 

o,p’-DDT, up to 4% of p,p’-DDD, and up to 1.5% of 1-(p-chlorophenyl)-2,2,2-trichloroethanol (Metcalf 

1995). 

5.1 PRODUCTION 

Technical DDT is made by condensing chloral hydrate with chlorobenzene in concentrated sulfuric acid. 

It was first synthesized in 1874, but it was not until 1939 that Müller and his coworkers discovered its 

insecticidal properties. Production of DDT in 1971 in the United States was estimated to be 2 million kg. 

This represented a sharp decline from the 82 million kg produced in 1962, and from the 56 million kg 

produced in 1960. At the peak of its popularity in 1962, DDT was registered for use on 334 agricultural 

commodities and about 85,000 tons were produced (Metcalf 1995). Production then declined and by 

1971, shortly before it was banned in the United States, production had dipped to about 2,000 tons. The 

cumulative world production of DDT has been estimated as 2 million tons. As of January 1, 1973, all 

uses of DDT in the United States were canceled except emergency public health uses and a few other uses 

permitted on a case-by-case basis (Meister and Sine 1999). Currently, no companies in the United States 

manufacture DDT (Meister and Sine 1999). DDT is presently produced by companies in Mexico and 

China (Meister and Sine 1999). p,p’-DDD has been used as an insecticide, but is no longer produced 

commercially. It was prepared by condensing dichloroacetaldehyde with chlorobenzene (Budavari et al. 

1996). No past production figures are available for this chemical, and there are no indications that 

production was ever very high. o,p’-DDD (Mitotane) is produced by Bristol, Meyer, Squibb under the 

brand name Lysodren® for use as a chemotherapy drug for adrenal gland cancer (PDR 1999). DDD and 

DDE are degradation products formed by dehydrohalogenation of DDT. 

Analytical studies have shown that DDT compounds, including p,p’-DDT and p,p’-DDE, may be 

contaminants in technical grades of the insecticide, dicofol (Risebrough et al. 1986). In addition, another 

DDT-related impurity in dicofol, 1,1,1,2-tetrachloro-2,2-bis(p-chlorophenyl)ethane, has been shown to 

degrade to p,p’-DDE. 
DDT, DDE, and DDD 222 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

5.2 IMPORT/EXPORT 

DDT was last imported into the United States in 1972, when imports amounted to 200 tons. Although the 

use of DDT was banned in the United States after 1972, it was still manufactured for export. In 1985, 

there were two producers of DDT in the United States, and in that year, 303,000 kg of DDT were 

exported (HSDB 1988). Presently, there are no producers of DDT in the United States, and therefore, 

there are no exports of DDT. 

5.3 USE 

DDT is a broad spectrum insecticide that was very popular due its effectiveness, long residual persistence, 

low acute mammalian toxicity, and low cost (Metcalf 1989). DDT was first used as an insecticide starting 

in 1939 and widely used until about 1970 (Van Metre et al. 1997). Its usage peaked in the United States 

in the early 1960s. During World War II, it was extensively employed for the control of malaria, typhus, 

and other insect-transmitted diseases. DDT has been widely used in agriculture to control insects, such as 

the pink boll worm on cotton, codling moth on deciduous fruit, Colorado potato beetle, and the European 

corn borer. In 1972, 67–90% of the total United States consumption of DDT was on cotton; the 

remainder was primarily used on peanuts and soybeans. DDT has been used extensively to eradicate 

forest pests, such as the gypsy moth and spruce budworm. It was used in the home as a mothproofing 

agent and to control lice. The amount of DDT used in U.S. agriculture was 27 million pounds in 1966 

and 14 million pounds in 1971 (Gianessi and Puffer 1992). Since 1973, use of DDT in the United States 

has been limited to the control of emergency public health problems. In some regions of the world where 

malaria is endemic, such as South Africa, Swaziland, and Madagascar, DDT is sprayed onto the interior 

surfaces of homes to decrease the incidence and spread of the disease by controlling mosquitoes (Attaran 

et al. 2000; Roberts et al. 1997). Not only is DDT a contact toxin for mosquitoes, it is also a contact 

irritant and repellent. As such, DDT has been shown to be effective in controlling malaria by not only 

limiting the survival of the mosquito, but also decreasing the odds of an individual being bitten within the 

sprayed homes. p,p’-DDD was also used as an insecticide. o,p'-DDD (Mitotane) is used medically in the 

treatment of cancer of the adrenal gland (PDR 1999). DDE has no commercial use. 
DDT, DDE, and DDD 223 

5. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

5.4 DISPOSAL 

Under current federal guidelines, DDT and DDD are potential candidates for incineration in a rotary kiln 

at 820–1,600 EC. Disposal of DDT formulated in 5% oil solution or other solutions is mainly by using 

liquid injection incineration at 878–1,260 EC, with a residence time of 0.16–1.30 seconds and 26–70% 

excess air. Destruction efficiency with this method is reported to be >99.99%. Multiple-chamber 

incineration is also used for 10% DDT dust and 90% inert ingredients at a temperature range of 

930–1,210 EC, a residence time of 1.2–2.5 seconds, and 58–164% excess air. DDT powder may be 

disposed of by molten salt combustion at 900 EC (no residence time or excess air conditions specified). A 

low temperature destruction method involving milling DDT with Mg, Ca, or CaO is under development 

on a laboratory scale (Rowlands et al. 1994). Landfill disposal methods are rarely used at the present 

time. 

DDT, DDE, and DDD 225 

6. POTENTIAL FOR HUMAN EXPOSURE
 

While this document is specifically focused on the primary forms or isomers of DDT, DDE, and DDD 

(namely p,p’-DDT, p,p’-DDE, and p,p’-DDD), other isomers of these compounds will be discussed when 

appropriate. It should be noted that DDT, DDE, and DDD are also synonyms for p,p’-DDT, p,p’-DDE, 

and p,p’-DDD, respectively, and it is usually understood that when DDT, for example, is mentioned 

p,p’-DDT is being referred to and not both o,p’- and p,p’-DDT. Technical-grade DDT, the grade that was 

generally used as an insecticide was composed of up to 14 chemical compounds, of which only 65–80% 

was the active ingredient, p,p’-DDT. The other components included 15–21% of the nearly inactive 

o,p’-DDT, up to 4% of p,p’-DDD, and up to 1.5% of 1-(p-chlorophenyl)-2,2,2-trichloroethanol (Metcalf 

1995). In some cases, the term DDT will be used to refer to the collection of all forms of DDT, DDE, and 

DDD. Should this not be clear from the context, the term ΣDDT (Σ is used to mean sum of) will be used. 

6.1 OVERVIEW 

DDT and its primary metabolites, DDE and DDD, are manufactured chemicals and are not known to 

occur naturally in the environment (WHO 1979). Historically, DDT was released to the environment 

during its production, formulation, and extensive use as a pesticide in agriculture and vector control 

applications. DDD was also used as a pesticide, but to a far lesser extent than was DDT. Although it was 

banned for use in the United States after 1972, DDT is still being used in some areas of the world. DDT 

and its metabolites are very persistent and bioaccumulate in the environment. 

DDT gets into the atmosphere as a result of spraying operations in areas of the world where it is still used. 

DDT and its metabolites also enter the atmosphere through the volatilization of residues in soil and 

surface water, much of it a result of past use. These chemicals will be deposited on land and in surface 

water as a result of dry and wet deposition. The process of volatilization and deposition may be repeated 

many times, and results in what has been referred to as a ‘global distillation’ from warm source areas to 

cold polar regions. As a result, DDT and its metabolites are transported to the Arctic and Antarctic 

regions where they are found in the air, sediment, and snow and accumulate in biota. 

When in the atmosphere, about 50% of DDT will be found adsorbed to particulate matter and 50% will 

exist in the vapor phase (Bidleman 1988). A smaller proportion of DDE and DDD are adsorbed to 

particulate matter than DDT. Vapor-phase DDT, DDE and DDD react with photochemically-produced 

hydroxyl radicals in the atmosphere; their estimated half-lives are 37, 17, and 30 hours, respectively. 
DDT, DDE, and DDD 226 

6. POTENTIAL FOR HUMAN EXPOSURE 

However, based on the ability of DDT, DDE, and DDD to undergo long range global transport, these 

estimated half-lives do not adequately reflect the actual lifetimes of these chemicals in the atmosphere. 

The dominant fate processes in the aquatic environment are volatilization and adsorption to biota, 

suspended particulate matter, and sediments. Transformation includes biotransformation and photolysis 

in surface waters. The fate of DDT in the aquatic environment is illustrated by a microcosm study in 

which DDT was applied to a pond, and a material balance was performed after 30–40 days. At this time, 

DDT concentrations in the water column had declined to below the detectable limit (EPA 1979c). It was 

found that 90% of the initial DDT was not present in the water, sediment, algae, invertebrates, or fish, and 

was presumed to have volatilized. ΣDDT was present in water mainly as DDT during the first 30 days, as 

DDT and DDD during the next 30 days, and as DDD in the last 30 days. ΣDDT levels rapidly rose in 

invertebrates, reaching equilibrium in 5 days and then declining as the ΣDDT content of the water 

declined. Degradation of DDT is altered by invertebrates, with the conversion of DDT to DDMU. ΣDDT 

levels in fish rose rapidly and reached a high equilibrium level. In a study of a fresh water lake, DDT was 

found to accumulate to higher concentrations in fattier fish occupying high trophic levels than in leaner 

species occupying lower trophic levels (Kidd et al. 2001). Also, accumulation of DDT was significantly 

higher in the pelagic food web than in the benthic food web. 

When deposited on soil, DDT, DDE, and DDD are strongly adsorbed. However, they may also 

revolatilize into the air, which is more likely to occur from moist soils than dry soils. They may 

photodegrade on the soil surface and biodegrade. DDT biodegrades primarily to DDE under unflooded 

conditions (e.g., aerobic) and to DDD under flooded (e.g., anaerobic) conditions. As a result of their 

strong binding to soil, DDT, DDE, and DDT mostly remain on the surface layers of soil; there is little 

leaching into the lower soil layers and groundwater. DDT may be taken up by plants that are eaten by 

animals and accumulate to high levels, primarily in adipose tissue and milk of the animals. 

In discussing DDT and other pesticides in soil, agricultural chemists generally speak of persistence and 

degradation, but it is not always clear what mechanisms are responsible for the loss or dissipation of the 

chemical. This issue is further complicated in the case of DDT because what is often reported is the 

disappearance of ΣDDT residues rather than just p,p’-DDT. Many studies use first-order kinetics to 

model the dissipation of DDT in soils because a half-life for the chemical can be defined. The half-life 

represents the calculated time for loss of the first 50% of the substance, but the time required for the loss 

of half of that which remains may be substantially longer, and the rate of disappearance may decline 

further as time progresses. The rate and extent of disappearance may result from transport processes as 
DDT, DDE, and DDD 227 

6. POTENTIAL FOR HUMAN EXPOSURE 

well as degradation or transformation processes. Initially, much of the disappearance of DDT is a result 

of volatilization losses, after which biodegradation becomes more important. When more than one 

process is responsible for loss, the decrease in the amount of substance remaining will be nonlinear. 

Recent assessments of long-term monitoring studies have indicated that even DDT biodegradation does 

not follow first-order kinetics (Alexander 1995, 1997). The reason is that over long periods of time, DDT 

may become sequestered in soil particles and become less available to microorganisms. The term half-life 

in this document is used to indicate the estimated time for the initial disappearance of 50% of the 

compound, and does not necessarily imply that first-order kinetics were observed throughout the 

experiment unless otherwise noted. The persistence of DDT in soil is highly variable. Dissipation is 

much greater in tropical than in temperate regions. In tropical and subtropical regions, most of the DDT 

is lost within a year; the half-life of ΣDDT in 13 countries ranged from 22 to 327 days. The half-life of 

DDE, the primary degradation product of DDT, ranged from 151 to 271 days. In another country where 

the soil was extremely acidic, the half-life was >672 days. Comparable half-lives in temperate regions 

have been reported to range from 837 to 6,087 days. One investigator concluded that the mean lifetime of 

DDT in temperate U.S. soils was about 5.3 years. In a study of sprayed forest soils in Maine, the halftime for the disappearance of DDT residues was noted to be 20–30 years (Dimond and Owen 1996). 

Highest residues are found in muck soils and in deeply plowed, unflooded fields (Aigner et al. 1998; 

Spencer et al. 1996). Significant concentrations of DDT have been found in the atmosphere over 

agricultural plots. Irrigating the soil dramatically increased the volatilization flux of DDT, which is 

probably related to the amount of DDT in the soil solution. Volatilization, air transport, and redeposition 

were found to be the main avenues of contaminating forage eaten by cows. 

When deposited in water, DDT will adsorb strongly to particulate matter in the water column and 

primarily partition into the sediment. Some of the DDT may revolatilize. DDT bioconcentrates in 

aquatic organisms and bioaccumulates in the food chain. Marine mammals in the Arctic often contain 

very high levels of DDT and DDE. 

Concentrations of DDT in all media have been declining since DDT was banned in the United States and 

most of the world (Arthur et al. 1977; Boul et al. 1994; Van Metre and Callender 1997; Van Metre et al. 

1997; Ware et al. 1978). For example, the concentration of DDT in lake sediments has decreased by 93% 

from 1965 to 1994 and declined by 70% in silt loam between 1960 and 1980 (Boul et al. 1994; Van Metre 

and Callender 1997; Van Metre et al. 1997). GDDT levels in sea lions have decreased by 2 orders of 

magnitude between 1970 and 1992 (Lieberg-Clark et al. 1995). The Market Basket Surveys have shown 

a 86% decline in DDT levels measured in all classes of food from 1965 to 1975 (EPA 1980a). However, 
DDT, DDE, and DDD 228 

6. POTENTIAL FOR HUMAN EXPOSURE 

because of the extensive past use of DDT worldwide and the persistence of DDT and its metabolites, 

these chemicals are virtually ubiquitous and are continually being transformed and redistributed in the 

environment. 

Human exposure to DDT is primarily through the diet. Exposure via inhalation at the ambient levels in 

air (Whitmore et al. 1994) is thought to be insignificant compared with dietary intake. The main source 

of DDT in food is meat, fish, poultry, and dairy products. DDT residues in food have declined since it 

was banned. Residues are more likely to occur in food imported from countries where DDT is still used. 

People eating fish from the Great Lakes were found to consume greater amounts of DDT in their diets 

(Hanrahan et al. 1999; Laden et al. 1999), but as DDT levels in Great Lakes fish continue to decline, 

exposure from consuming fish should also decline (Anderson et al. 1998; Hanrahan et al. 1999; Hovinga 

et al. 1993). The populations having the greatest exposure to DDT are indigenous people in the Arctic 

who eat traditional foods (e.g., seals, cariboo, narwhal whales, etc.) (Kuhnlein et al. 1995). 

DDT, DDE, or DDD have been identified in at least 441 of the 1,613 hazardous waste sites that have been 

proposed for inclusion on the EPA National Priorities List (NPL) (HazDat 2002). However, the number 

of sites evaluated for DDT, DDE, and DDD is not known. The frequency of these sites can be seen in 

Figure 6-1. Of these sites, 438 are located in the United States, 1 is located in Guam (not shown), 1 is 

located in Puerto Rico (not shown), and 1 is located in the U.S. Virgin Islands (not shown). p,p’-DDT, 

p,p’-DDD, and p,p’-DDE, the only DDT-related compounds targeted for analysis at NPL sites (EPA 

1994), were respectively found at 326, 276, and 219 sites. While not targeted, o,p’-DDT, o,p’-DDD, and 

o,p’-DDE were nevertheless listed as detected at 4, 3, and 4 sites, respectively. Releases of DDT, DDE, 

or DDD are not required to be reported in the Toxics Release Inventory (TRI) database (EPA 1999). 
Figure 6-1. Frequency of NPL Sites with DDT, DDE, and DDD Contamination 

Frequen

NPL Si

Derived from HazDat 2002 
DDT, DDE, and DDD 230 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.2 RELEASES TO THE ENVIRONMENT 

6.2.1 Air 

During the period when DDT was extensively used, a large source of DDT release to air occurred during 

agricultural or vector control applications. Emissions could also have resulted during production, 

transport, and disposal. Because use of DDT was banned in the United States after 1972, release of DDT 

in recent years should be negligible in this country. 

Nevertheless, DDT residues in bogs or peat lands across the midlatitudes of North America indicate that 

DDT was still released, even after it was banned for use in the United States (Rapaport et al. 1985). 

These areas are unique in that they receive all of their pollutant input from the atmosphere, and therefore, 

peat cores are important indicators of the atmospheric deposition of a substance and also of its 

atmospheric levels in the present and the past. An analysis of peat cores, as well as rain and snow 

samples, indicated that DDT was still present in the atmosphere, although levels were lower compared to 

those in the 1960s. The implication is that DDT is still being released to the atmosphere either from its 

current production and use in other countries and transport to the United States or from the volatilization 

of residues resulting from previous use. The estimated release of DDT into the atmosphere from the 

Great Lakes in 1994, excluding Lake Huron, was 14.3 kg (Hoff et al. 1996). DDT, DDE, or DDD have 

been identified in air samples collected from 32 of the 441 NPL hazardous waste sites where it was 

detected in some in some environmental media (HazDat 2002). 

6.2.2 Water 

Historically, DDT was released to surface water when it was used for vector control in the vicinity of 

open waters. This source of release may still be occurring in countries that rely on DDT in insect pest 

control near open waters. DDT also enters surface water as a result of dry and wet deposition from the 

atmosphere and direct gas transfer. Atmospheric DDT deposited into tributaries will contribute to the 

loading in rivers, lakes, and oceans. In 1994, the estimated loading of ΣDDT into the Great Lakes as a 

result of dry and wet deposition was estimated as 148 kg, down from 278 kg in 1988 (Hoff et al. 1996). 

Fluvial sources and erosion also contribute to the DDT burden, and they were the predominant source of 

DDT in many areas in the past. This was clearly shown in a United States Geological Survey (USGS) 

study of sediment in reservoirs and lakes in Georgia and Texas compared with DDT levels in nearby peat 

bogs (Van Metre et al. 1997). 
DDT, DDE, and DDD 231 

6. POTENTIAL FOR HUMAN EXPOSURE 

Contaminated sediment near an outfall can act as a source of contamination in distant parts of a body of 

water. This was clearly illustrated in a Norwegian lake that received insecticidal wastes. Nineteen years 

after closing the outfall, DDT concentrations in pike and perch were 5–10 times those in uncontaminated 

lakes (Brevik et al. 1996). DDT was disposed to the Joint Water Pollution Control Plant, Los Angeles 

County by Montrose Chemical Company from about 1950 to 1970, and eventually to the Palos Verdes 

Shelf via sewer pipes. The distribution of DDT with respect to the outfall diffusers and the fact that the 

DDT concentration in the overlying water column exponentially decreased with increasing distance from 

the sea floor indicated that the main source of DDT in the water column was contaminated sediments 

(Zeng et al. 1999). Studies have shown that a variety of organisms live in sediments at the Palos Verde 

site to depths of at least 35 cm and disturb the reservoir of contaminants there (Stull et al. 1996). 

Sediment-bound DDT is being biodiffused up from the subsurface to upper sediments, where they 

undergo resuspension and redistribution. 

DDT, DDE or DDD have been identified in surface water collected at 101 sites and groundwater 

collected at 247 of the 441 NPL hazardous waste sites where it was detected in some environmental 

(HazDat 2002). 

6.2.3 Soil 

In the United States, large amounts of DDT were released to the soil during spraying operations or from 

direct or indirect releases during manufacturing, formulation, storage, or disposal. Since almost all of the 

DDT produced was used to control insects damaging crops and trees or responsible for insect-transmitted 

diseases, we can assume that a large fraction of the DDT produced was released to soil during spraying 

operations. The largest amounts of DDT released to soil were those used in agriculture which amounted 

to 27 million pounds in 1966 and 14 million pounds in 1971, shortly before it was banned (Gianessi 

1992). 

DDT, DDE, or DDD have been identified in soil samples collected from 776 sites and sediment samples 

at 305 of the 441 NPL hazardous waste sites where it was detected in some environmental media (HazDat 

2002). 
DDT, DDE, and DDD 232 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.3 ENVIRONMENTAL FATE 

A large proportion of the environmental fate studies on pesticides such as p,p’-DDT are performed in 

laboratory or field studies by agricultural chemists interested in the persistence of the active ingredient of 

the pesticide in the tilled layer of soil. Therefore, studies may not reveal whether the loss of active 

ingredient is a result of volatilization, leaching, or microbial degradation. Field studies may also report 

the occurrence of obvious metabolites remaining in surface soil months or years after a pesticide was 

applied. Clearly, it is not possible to separate these studies into ‘Transport and Partitioning’ 

(Section 6.3.1) and ‘Transformation and Degradation’ (Section 6.3.2). These studies are discussed in 

Section 6.3.2 with the understanding that ‘degradation’ may only account for part of the reported loss. 

6.3.1 Transport and Partitioning 

DDT and its metabolites may be transported from one medium to another by the processes of 

solubilization, adsorption, remobilization, bioaccumulation, and volatilization. In addition, DDT can be 

transported within a medium by currents, wind, and diffusion. These processes will be discussed in the 

following paragraphs. 

Organic carbon partition coefficients (Koc) of 1.5x105 (Swann et al. 1981), 5.0x104 (Sablejic 1984), and 

1.5x105 (Meylan et al. 1992) reported for p,p’-DDT, p,p’-DDE, and p,p’-DDD, respectively, suggest that 

these compounds adsorb strongly to soil. These chemicals are only slightly soluble in water, with 

solubilities of 0.025, 0.12, and 0.090 mg/L for p,p’-DDT, p,p’-DDE, and p,p’-DDD, respectively 

(Howard and Meylan 1997). Therefore, loss of these compounds in runoff is primarily due to transport of 

particulate matter to which these compounds are bound. For example, DDT and its metabolites have been 

found to fractionate and concentrate on the organic material that is transported with the clay fraction of 

the washload in runoff (Masters and Inman 2000). The amount of DDT transported into streams as runoff 

is dependent on the methods of irrigation used (USGS 1999). In the Western United States, DDT 

concentrations in streambed sediment increased as the percentage of furrow irrigation, as opposed to 

sprinkler or drip irrigation, increased. In the San Joaquin River Basin, more DDT was transported during 

winter runoff than during the irrigation season (Kratzer 1999). Since the compounds are bound strongly 

to soil, DDT would remain in the surface layers of soil and not leach into groundwater. However, DDT 

can adsorb to free-moving dissolved organic carbon, a soluble humic material that may occur in the soil 

solution. This material behaves as a carrier and facilitates transport of DDT into subsurface soil (Ding 

and Wu 1997). DDT released into water adsorbs to particulate matter in the water column and sediment. 
DDT, DDE, and DDD 233 

6. POTENTIAL FOR HUMAN EXPOSURE 

Sediment is the sink for DDT released into water. There it is available for ingestion by organisms, such 

as bottom feeders. Reich et al. (1986) reported that DDT, DDE, and DDD were still bioavailable to 

aquatic biota in a northern Alabama river 14 years after 432,000–8,000,000 kg of DDT was discharged 

into the river. DDT in the water column above the outfall of Los Angeles County’s Joint Water Pollution 

Control Plant’s outfall was present both in the dissolved phase and the particulate phase (defined as 

particles size >0.7 μm) (Zeng et al. 1999). It is interesting to note that more of the DDT was present in 

the dissolved phase than in the particulate phase despite its high hydrophobicity. 

There is evidence that DDT, as well as other molecules, undergoes an aging process in soil whereby the 

DDT is sequestered in the soil so as to decrease its bioavailability to microorganisms, extractability with 

solvents, and toxicity to some organisms (Alexander 1995, 1997; Peterson et al. 1971; Robertson and 

Alexander 1998). At the same time, analytical methods using vigorous extractions do not show 

significant decreases in the DDT concentration in soil. In one such study, DDT was added to sterile soil 

at various concentrations and allowed to age (Robertson and Alexander 1998). At intervals, the toxicity 

of the soil was tested using the house fly, fruit fly, and German cockroach. After 180 days, 84.7% of the 

insecticide remained in the soil, although more than half of the toxicity had disappeared when the fruit fly 

was the test species, and 90% had disappeared when the house fly was the test species. The effect with 

the German cockroach was not as marked. Recently, a study was conducted to determine the 

bioavailability of DDT, DDE, and DDD to earthworms (Morrison et al. 1999). It was shown that the 

concentrations of DDT, DDE, DDD, and ΣDDT were consistently lower in earthworms exposed to these 

compounds that had persisted in soil for 49 years than in earthworms exposed to soil containing freshly 

added insecticides at the same concentration. The uptake percentages of DDT and its metabolites by 

earthworms were in the range of 1.30–1.75% for the 49-year-aged soil, but were 4.00–15.2% for the fresh 

soil (Morrison et al. 1999). Long term monitoring data have also indicated that aged and sequestered 

DDT are not subject to significant volatilization, leaching, or degradation (Boul et al. 1994). The 

concentrations of DDT, DDE, and DDD monitored at two sites in a silt loam in New Zealand declined 

from 1960 to 1980, but very little loss was evident from 1980 to 1989 (Boul et al. 1994). The lack of 

appreciable biodegradation as DDT ages in soil suggests that the compound is not bioavailable to 

microorganisms. Aging is thought to be associated with the continuous diffusion of a chemical into 

micropores within soil particles where it is sequestered or trapped, and is therefore unavailable to 

microorganisms, plants, and animals (Alexander 1995). In the case of biodegradation, the aging process 

results in the gradual unavailability of substrate that makes the reaction kinetics appear to be nonlinear. 
DDT, DDE, and DDD 234 

6. POTENTIAL FOR HUMAN EXPOSURE 

There is abundant evidence that DDT gets into the atmosphere as a result of emissions or volatilization. 

The process of volatilization from soil and water may be repeated many times and, consequently, DDT 

may be transported long distances in the atmosphere by what has been referred to as a ‘global distillation’ 

from warm source areas to cold polar regions (Bard 1999; Bidleman et al. 1992; Goldberg 1975; Ottar 

1981; Wania and MacKay 1993). As a result, DDT and its metabolites are found in arctic air, sediment, 

and snow with substantial accumulations in animals, marine mammals, and humans residing in these 

regions (Anthony et al. 1999; Harner 1997). An analysis of sediment cores from eight remote lakes in 

Canada indicated that ΣDDT concentrations in surface sediments (0–1.3 cm depth) declined significantly 

with latitude (Muir et al. 1995). The maximum ΣDDT concentrations in core slices in midcontinent lakes 

date from the late 1970s to 1980s, which is about 5–10 years later than the maximum for Lake Ontario. 

Volatilization of DDT, DDE, and DDD is known to account for considerable losses of these compounds 

from soil surfaces and water. Their tendency to volatilize from water can be predicted by their respective 

Henry's law constants, which for the respective p,p’- and o,p’- isomers are 8.3x10-6, 2.1x10-5, 4.0x10-6, 

5.9x10-7, 1.8x10-5, and 8.2x10-6 atm-m3

/mol (Howard and Meylan 1997). The predicted volatilization 

half-lives from a model river 1 m deep, flowing at 1 m/sec, with a wind of 3 m/sec are 8.2, 3.3, 10.5, 6.3, 

3.7, and 8.2 days, respectively. Laboratory studies of the air/water partition coefficient of DDE indicate 

that it will volatilize from seawater 10–20 times faster than from freshwater (Atlas et al. 1982). The 

authors suggest that this process may be related to interaction at the bubble-water surface. 

Volatilization from moist soil surfaces can be estimated from the Henry’s law constant divided by the 

adsorptivity to soil (Dow Method) (Thomas 1990). The predicted half-life for DDT volatilizing from soil 

with a Koc of 240,000 is 23 days, compared to an experimental half-life of 42 days. Sleicher and Hopcraft 

(1984) estimated a volatilization half-life of 110 days for DDT from soil in Kenya based on mass transfer 

through the boundary layers, and claimed that volatilization of DDT was sufficient to account for its rapid 

disappearance from soil. However, laboratory experiments in which 14C-p,p’-DDT was incubated in an 

acidic (pH 4.5–4.8), sandy loam soil maintained at 45 EC for 6 hours/day for 6 weeks resulted in neither 

volatilization of DDT or its metabolites nor mineralization (Andrea et al. 1994). Other studies using a 

latosol soil (pH 5.7) found that 5.9% of the radioactivity was lost through volatilization during a 6 week 

incubation at 45 EC (Sjoeib et al. 1994). The volatilization rate of DDT from soil is significantly 

enhanced by temperature, sunlight, and flooding of the soil (Samuel and Pillai 1990). 

Transport of DDT in the atmosphere of central and eastern North America is facilitated by a circulation 

pattern that brings moisture from the Gulf of Mexico into the Midwest and the airflow patterns across the 
DDT, DDE, and DDD 235 

6. POTENTIAL FOR HUMAN EXPOSURE 

eastern seaboard (Rapaport et al. 1985). DDT is removed from the atmosphere by wet and dry deposition 

and diffusion into bodies of water. The largest amount of DDT is believed to be removed from the 

atmosphere in precipitation (Woodwell et al. 1971). 

DDT, DDE, and DDD are highly lipid soluble, as reflected by their log octanol-water partition 

coefficients (log Kow) of 6.91, 6.51, and 6.02, respectively for the p,p’- isomers and 6.79, 6.00, and 5.87, 

respectively, for the o,p’- isomers (Howard and Meylan 1997). This lipophilic property, combined with 

an extremely long half-life is responsible for its high bioconcentration in aquatic organisms (i.e., levels in 

organisms exceed those levels occurring in the surrounding water). Organisms also feed on other animals 

at lower trophic levels. The result is a progressive biomagnification of DDT in organisms at the top of the 

food chain. Biomagnification is the cumulative increase in the concentration of a persistent contaminant 

in successively higher trophic levels of the food chain (i.e., from algae to zooplankton to fish to birds). 

Ford and Hill (1991) reported increased biomagnification of DDT, DDE, and DDD from soil sediment to 

mosquito fish, a secondary consumer. No distinct pattern of biomagnification was evident in other 

secondary consumers such as carp and small mouth buffalo fish. The biomagnification of DDT is 

exemplified by the increase in DDT concentration in organisms representing four trophic levels sampled 

from a Long Island estuary. The concentrations in plankton, invertebrates, fish, and fish-eating birds 

were 0.04, 0.3, 4.1, and 24 mg/kg, whole body basis (Leblanc 1995). Evans et al. (1991) reported that 

DDE biomagnified 28.7 times in average concentrations from plankton to fish and 21 times from 

sediment to amphipods in Lake Michigan. In some cases, humans may be the ultimate consumer of these 

contaminated organisms. 

The bioconcentration factor (BCF) is defined as the ratio of the equilibrium concentration of contaminant 

in tissue compared to the concentration in ambient water, soil, or sediment to which the organism is 

exposed. There are numerous measurements and estimates of BCF values for DDT in fish. Oliver and 

Niimi (1985) estimated the steady-state BCF in rainbow trout as 12,000. Other BCF values that have 

been reported include 51,000–100,000 in fish, 4,550–690,000 in mussels, and 36,000 in snails (Davies 

and Dobbs 1984; Geyer et al. 1982; Metcalf 1973a; Reish et al. 1978; Veith et al. 1979). DDT 

bioconcentration studies in aquatic environments with representatives of various trophic levels 

demonstrate that bioconcentration increases with increasing trophic level (LeBlanc 1995). Trophic level 

differences in bioconcentration are largely due to increased lipid content and decreased elimination 

efficiency among higher level organisms. However, biomagnification also contributes to the increased 

concentration of DDT in higher trophic organisms (LeBlanc 1995). 
DDT, DDE, and DDD 236 

6. POTENTIAL FOR HUMAN EXPOSURE 

The BCF values of p,p’-DDT in brine shrimp (Artemia nauplii) exposed to a mixture containing 0.5 or 

1.0 ng/mL of four DDT analogs for 24 hours were significantly higher than for the three other chemicals. 

The BCF values were 41, 54, 128, and 248 for p,p’-DDE, p,p’-DDD, o,p’-DDT, and p,p’-DDT, 

respectively (Wang and Simpson 1996). The differences in BCF values are due to the different lipid 

solubility and selectivity of the compounds partitioned in the zooplanktonic organisms. p,p’-DDT, which 

has the greatest polarity of the 4 tested analogs, may have been adsorbed to a greater extent to the surface 

of the shrimp. In addition to absorbing DDT directly from the water, fish obtain DDT from their diet 

(Miller 1994; Wang and Simpson 1996). Wang and Simpson (1996) fed brook trout contaminated 

A. nauplii for 24 days followed by depuration for another 24 days during which the trout were fed 

uncontaminated A. nauplii. Although the concentration of p,p’-DDE was the lowest of the four analogs in 

the contaminated brine shrimp, the concentration of this compound in the trout at day 24 was 42.5 ng/g 

which was roughly 5 times more than the other analogs. The levels of the p,p’-isomers initially ranged 

from 1.0 to 2.7 ng/g, while o,p’-DDT was absent. The abnormal accumulation of p,p’-DDE in the fish 

suggests that mixed-function oxidases may have induced the dechlorination of p,p’-DDT to p,p’-DDE. 

This may account for the fact that about 70% of ΣDDT in fish is p,p’-DDE (Schmitt et al. 1990). After 

the fish were fed uncontaminated food, p,p’-DDE had the lowest percentage depuration. After feeding 

the trout for 24 days with the more highly contaminated brine shrimp, 14, 62, 17, and 32% depuration 

were observed for p,p’-DDE, p,p’-DDD, o,p’-DDT, and p,p’-DDT, respectively. 

Fish move from the Great Lakes or other bodies of water with elevated DDT levels to rivers that feed into 

these lakes. In doing so, they transport DDT, which may represent a risk to wildlife along the tributaries 

(Giesy et al. 1994). 

Despite being strongly bound to soil, at least a portion of DDT, DDE, and DDD is bioavailable to plants 

and soil invertebrates. Nash and Beall (1970) studied the DDT residues in soybean plants resulting from 

the application of [14C]DDT to the surface or subsurface soil. They found that the major source of DDT 

contamination was due to sorption of volatilized residues from surface-treated soil. This was 6.8 times 

greater than that obtained through root uptake and translocation after subsurface treatment. In other 

experiments with oats and peas, root uptake of DDT was low and there was little or no evidence of 

translocation of the insecticide (Fuhremann and Lichtenstein 1980; Lichtenstein and Schultz 1980). 

Verma and Pillai (1991a) reported that grain, maize, and rice plants accumulate DDT adsorbed to soil. 

Most of the residues were found in the roots of the plant, and the lowest concentration of DDT residues 

was found in the shoots, indicating low translocation of DDT. Earthworms are capable of aiding the 
DDT, DDE, and DDD 237 

6. POTENTIAL FOR HUMAN EXPOSURE 

mobilization of soil-bound DDT residues to readily bioavailable forms (Verma and Pillai 1991b). DDT 

may collect on the leafy part of plants from the deposition of DDT-containing dust. 

6.3.2 Transformation and Degradation 

6.3.2.1 Air 

In the atmosphere, about 50% of DDT is adsorbed to particulate matter and 50% exists in the vapor phase 

(Bidleman 1988). In the vapor phase, DDT reacts with photochemically produced hydroxyl radicals with 

an estimated rate constant of 3.44x10-12 cm3

/molecule-sec determined from a fragment constant estimation 

method (Meylan and Howard 1993). Assuming an average hydroxyl radical concentration of 1.5x106

 per 

cm3

, its half-life is estimated to be 37 hours. Both DDE and DDD have higher vapor pressures than DDT, 

and a smaller fraction of these compounds will be adsorbed to particulate matter. The estimated half-lives 

of vapor-phase DDE and DDD are 17 and 30 hours, respectively. Direct photolysis may also occur in the 

atmosphere. 

It should be noted that the estimated half-lives for vapor-phase DDT, DDE, and DDD do not necessarily 

reflect the lifetimes of these compounds in air. DDT, DDE, and DDD can be adsorbed on particulate 

matter, where they are not expected to undergo rapid photooxidation, and therefore, may be subject to 

long-range transport. Indeed, long-range transport through the atmosphere has been demonstrated for 

DDT and several of its metabolites (Bard 1999; Bidleman et al. 1992; Goldberg 1975; Ottar 1981; Wania 

and MacKay 1993). The work of Bidleman (1988) suggests that 50% of DDT in the atmosphere is 

adsorbed to particulate matter. Further, when atmospheric sampling of pesticides was performed at nine 

localities in the United States during a time of high DDT usage, DDT was mostly present in the 

particulate phase (Stanley et al. 1971). 

6.3.2.2 Water 

DDT, DDE, and DDD present in water may be transformed by both photodegradation and 

biodegradation. Since the shorter wave radiation does not penetrate far into a body of water, photolysis 

primarily occurs in surface water and is dependent on the clarity of the water. Direct photolysis of DDT 

and DDD are very slow in aquatic systems, with estimated half-lives of >150 years (EPA 1979c). Direct 

photolysis of DDE will vary as a function of photoperiod and brightness, resulting in different half-lives 

depending on the season and latitude. Over the United States, the direct photolysis of DDE results in a 
DDT, DDE, and DDD 238 

6. POTENTIAL FOR HUMAN EXPOSURE 

half-life of about 1 day in summer and 6 days in winter. DDE also undergoes photoisomerization when 

exposed to sunlight. Photolysis of DDE photoisomers is slower by at least one order of magnitude 

compared to DDE. Studies with DDT at shorter wavelengths suggest that the initial reaction results in the 

dissociation of the Cl2C–Cl bond. Some information exists on the indirect photolysis of DDT; no 

information on the indirect photolysis of DDE or DDD was located (Coulston 1985; EPA 1979c; Zepp et 

al. 1977). 

Photoinduced 1,2 addition of DDT to a model lipid, methyl oleate, indicates that light-induced additions 

of DDT to unsaturated fatty acids of plant waxes and cutins may occur on a large scale (Schwack 1988). 

DDT undergoes hydrolysis by a base-catalyzed reaction resulting in a half-life of 81 days at pH 9. The 

product formed in the hydrolysis is DDE. Hydrolysis of DDE and DDD is not a significant fate process 

(EPA 1979c). 

Biodegradation of DDT in water is reported to be a minor mechanism of transformation (Johnsen 1976). 

Biodegradation of DDE and DDD in the aquatic environment is slower than that of DDT (EPA 1979c). 

6.3.2.3 Sediment and Soil 

Four mechanisms have been suggested to account for most losses of DDT residues from soils: 

volatilization, removal by harvest (e.g., plants that have absorbed the residue), water runoff, and chemical 

transformation (Fishbein 1973). Three of these are transport processes, and the fourth, chemical 

transformation, may occur by abiotic and biotic processes. Photooxidation of DDT and DDE is known to 

occur on soil surfaces or when adsorbed to sediment (Baker and Applegate 1970; Lichtenstein and 

Schultz 1959; Miller and Zepp 1979). The conversion of DDT to DDE in soil was enhanced by exposure 

to sunlight in a 90-day experiment with 91% of the initial concentration of DDT remaining in the soil for 

an unexposed dark control and 65% remaining for the sample exposed to light (Racke et al. 1997). 

However, UV-irradiation of 14C-p,p’-DDT on soil for 10 hours mineralized less than 0.1% of the initial 

amount (Vollner and Klotz 1994). (Mineralization is the complete degradation of a chemical, generally to 

carbon dioxide and water for an organic chemical containing carbon, hydrogen, and oxygen.) The 

amount of DDT that may have been converted to DDE was not reported. Biodegradation may occur 

under both aerobic and anaerobic conditions due to soil microorganisms including bacteria, fungi, and 

algae (Arisoy 1998; EPA 1979c; Lichtenstein and Schulz 1959; Menzie 1980; Stewart and Chisholm 

1971; Verma and Pillai 1991b). Since biodegradation studies generally focus on the loss of the parent 
DDT, DDE, and DDD 239 

6. POTENTIAL FOR HUMAN EXPOSURE 

compound rather than complete degradation or mineralization, and since DDT initially biodegrades to 

DDD or DDE, there still may be dangerous compounds remaining after almost all of the DDT that was 

originally present has biodegraded. 

During biodegradation of DDT, both DDE and DDD are formed in soils. Both metabolites may undergo 

further transformation but the extent and rate are dependent on soil conditions and, possibly, microbial 

populations present in soil. The degradation pathways of DDT under aerobic and anaerobic conditions 

have been reviewed by Zook and Feng (1999) and Aislabie et al. (1997). Ligninolytic or lignindegrading fungi have been shown to possess the biodegradative capabilities for metabolizing a large 

variety of persistent compounds, including DDT. Mineralization of DDT and DDE was even observed in 

laboratory experiments using a member of this group of fungi, Phanerochaete chrysosporium (a white rot 

fungus) (Aislabie et al. 1997; Singh et al. 1999). Other soil microorganisms, such as Aerobacter 

aerogenes, Pseudomonas fluorescens, E. coli, and Klebsiella pneumoniae, have also been shown to have 

the capability to degrade DDT under both aerobic and anaerobic conditions, forming 4-chlorobenzoic 

acid and DDE, respectively (Singh et al. 1999). Biodegradation of DDT and its metabolites involves 

cometabolism, a process in which the microbes derive nutrients for growth and energy from sources other 

than the compound of concern. DDE, the dominant DDT metabolite found, is often resistant to 

biodegradation under aerobic and anaerobic conditions (Strompl and Thiele 1997). In laboratory 

experiments with marine sediments, DDT has been shown to degrade to DDE and DDD under anaerobic 

and anaerobic conditions, respectively (Kale et al. 1999). In these same experiments, it was shown that 

extensive degradation of DDT occurred in clams, converting DDT to DDMU. Recent laboratory 

experiments in marine sediment showed that DDE is dechlorinated to DDMU (1-chloro-2,2-bis[p-chlorophenyl]ethylene) under methanogenic or sulfidogenic conditions (Quensen et al. 1998). The rate of DDE 

dechlorination to DDMU was found to be dependent on the presence of sulfate and temperature (Quensen 

et al. 2001). DDD is also converted to DDMU, but at a much slower rate. DDMU degrades further under 

anaerobic conditions to 2,2-bis(chlorophenyl)acetonitrile (DDNU) and other subsequent degradation 

species, such as 2,2-bis(chlorophenyl)ethanol (DDOH) and 2,2-bis(chlorophenyl)acetic acid (DDA), 

through chemical action (Heberer and Dünnbier 1999; Ware et al. 1980). No evidence was found that 

methylsulfonyl metabolites of DDT are formed as a result of microbial metabolism. The rate at which 

DDT is converted to DDD in flooded soils is dependent on the organic content of the soil (Racke et al. 

1997). In a laboratory study, Hitch and Day (1992) found that soils with a low metal content (e.g., Al, 

Ba, Cd, Co, Cr, Fe, and K were the major metals examined) degrade DDT to DDE much more slowly 

than do soils with high metal content. 
DDT, DDE, and DDD 240 

6. POTENTIAL FOR HUMAN EXPOSURE 

As mentioned earlier, the half-life represents the estimated time for the initial disappearance of 50% of the 

compound in question and does not necessarily imply that first-order kinetics were observed throughout 

the experiment unless otherwise noted. In the case of DDT, the disappearance rate slows considerably so 

that after the initial concentration is reduced by half, the time required for the loss of half of that which 

remains is substantially longer. This is largely because much of the initial loss of compound is due to 

volatilization, rather than biodegradation. However, the biodegradation rate also slows in time as 

discussed in Section 6.3.1. This is because DDT migrates into micropores in soil particles where it 

becomes sequestered and unavailable to soil microorganisms (Alexander 1995, 1997). In addition, the 

disappearance of DDT is often reported as the disappearance of ΣDDT residues, and therefore, the 

reported rate of loss is a summation of the component DDT-related chemicals. DDT breaks down into 

DDE and DDD in soil, and the parent-to-metabolite ratio (DDT to DDE or DDD) decreases with time. 

However, this ratio may vary considerably with soil type. In a 1995–1996 study of agricultural soils in 

the corn belt of the central United States, the ratio of p,p’-DDT/p,p’-DDE varied from 0.5 to 6.6 with 

three-quarters of the soils having ratios above 1 (Aigner et al. 1998). In a study of forest soils in Maine, 

the half-life for the disappearance of DDT residues was noted to be 20–30 years (Dimond and Owen 

1996). DDT was much more persistent in muck soils than in dry forest soils. A study of DDT in 

agricultural soils in British Colombia, Canada reported that over a 19-year period, there was a 70% 

reduction of DDT in muck soils and a virtual disappearance of DDT from loamy sand soils (Aigner et al. 

1998). 

Land management practices also affect the persistence of DDT. In 1971, an experiment was conducted in 

a field containing high amounts of DDT to evaluate the effect of various management tools in the 

disappearance of the insecticide (Spencer et al. 1996). The site was revisited in 1994 to determine the 

residual concentrations of DDT and its metabolites and to measure volatilization fluxes. Concentrations 

of DDT were reduced in all plots and the major residue was p,p’-DDE. The highest concentrations of 

residues were found in deep plowed and unflooded plots. Deep plowing places the DDT deeper into the 

soil profile, possibly reducing volatilization. As was noted in Section 6.3.1, the volatilization rate of 

DDT is enhanced by flooding the soil (Samuel and Pillai 1990). Under flooded, reducing conditions, 

DDD was a more common degradation product of DDT than DDE. Significant concentrations of both 

o,p’- and p,p’-DDE and p,p’-DDT were detected in the atmosphere over the plots. Irrigating the soil 

dramatically increased the volatilization flux of all DDT analogs, especially p,p’-DDE. This is probably 

related to the amount of DDT in the soil solution. Volatilization, air transport, and redeposition were 

found to be the main avenues of contaminating forage eaten by cows. In microcosm experiments, Boul 

(1996) found that increasing soil water content enhanced DDT loss from generally aerobic soil. His 
DDT, DDE, and DDD 241 

6. POTENTIAL FOR HUMAN EXPOSURE 

results suggested that increased biodegradation contributed to these effects. Boule et al. (1994) analyzed 

DDT residues in pasture soil as they were affected by long-term irrigation and superphosphate fertilizer 

application. They found that ΣDDT residues in irrigated soil were about 40% that of unirrigated soil. 

The predominant residue was p,p’-DDE, and these residues were much higher in unirrigated than in 

irrigated soil. p,p’-DDE is lost at a lower rate than p,p’-DDT. p,p’-DDD residues were very low in both 

irrigated and unirrigated soil indicating that loss of p,p’-DDD must occur at a rate at least as great as it is 

generated from p,p’-DDT. Superphosphate treatment, which is known to increase microbial biomass, also 

resulted in lower levels of p,p’-DDT and ΣDDT than in unfertilized controls. The distribution of ΣDDT 

with depth suggests that irrigation did not cause increased leaching of the insecticide. 

A set of experiments was conducted during 1982–1987 and 1989–1993 in 14 countries under the auspices 

of the International Atomic Energy Agency (IAEA) on the dissipation of 14C-DDT from soil under field 

conditions in tropical and subtropical areas (Racke et al. 1997). After 12 months, the quantity of DDT 

and metabolites remaining in soil at tropical sites ranged from 5% of applied in Tanzania to 15% in 

Indonesia. The half-life of ΣDDT ranged from 22 days in Sudan to 365 days in China. One exception 

was in an extremely acidic soil (pH 4.5) in Brazil in which the half-life was >672 days. The conclusion 

of the study was that DDT dissipated much more rapidly under tropical conditions than under temperate 

conditions. The major mechanisms of dissipation under tropical conditions were volatilization, biological 

and chemical degradation, and to a lesser extent, adsorption. Comparable half-lives in temperate regions 

that have been reported range from 837 to 6,087 days (Lichtenstein and Schulz 1959; Racke et al. 1997; 

Stewart and Chisholm 1971). One investigator concluded that the mean lifetime of DDT in temperate 

U.S. soils was about 5.3 years (Racke et al. 1997). The primary metabolite detected in tropical soil was 

DDE. With the exception of highly acidic soil from Brazil, the half-lives for DDE ranged from 151 to 

271 days, much less than the >20 years reported for DDE in temperate areas. The increased dissipation of 

DDT in the tropics compared with that in temperate zones is believed to be largely due to increased 

volatility under tropical conditions (Racke et al. 1997). 

6.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT 

6.4.1 Air 

DDT is transported long distances from source areas to the Arctic and Antarctic. Mean ΣDDT levels in 

air over a period of 17 weeks at Signy Island, Antarctica in 1992 and over the ocean separating New 

Zealand and Ross Island, Antarctica between January and March, 1990 were 0.07–0.40 and 0.81 pg/m3

, 
DDT, DDE, and DDD 242 

6. POTENTIAL FOR HUMAN EXPOSURE 

respectively (Bidleman et al. 1993; Kallenborn et al. 1998). The concentration declined with increasing 

latitudes. 

Ten samples taken over the Gulf of Mexico in 1977 contained an average of 34 pg/m3

 of DDT, with a 

range of 10–78 pg/m3

 (Bidleman et al. 1981). Iwata et al. (1993) collected and analyzed 71 samples of air 

over several oceans (18 sampling locations) from April 1989 to August 1990. The range of mean and 

maximum concentrations of DDTs were (substance, range of means, maximum concentration): p,p’-DDE, 

0.3–180 pg/m3

, 180 pg/m3; o,p’-DDT, 0.3–180 pg/m3, 420 pg/m3; p,p’-DDT, 1.2–220 pg/m3, 590 pg/m3; 

and ΣDDT, 2.4–580 pg/m3

, 1,000 pg/m3. The highest concentrations of DDT were found at locations 

near areas where DDT is still used, such as the Arabian Sea off the west coast of India. Other locations 

with high air concentrations of DDT were the Strait of Malacca, South China Sea, and the Gulf of 

Mexico. p,p'-DDT concentrations obtained from monthly air samples collected from Saginaw Bay, Sault 

Ste. Marie, and Traverse City, Michigan between November 1990 and October 1991 were below the 

detection limit during most of the winter months at Saginaw and Traverse City, and were above the 

detection limit at Sault Ste. Marie only in March, May, July, and August (Monosmith and Hermanson 

1996). The highest monthly p,p'-DDT concentrations were 35 pg/m3

 in Saginaw (August), 31 pg/m3 in 

Sault Ste. Marie (May), and 21 pg/m3

 in Traverse City (July). The corresponding highs for p,p'-DDE 

were 63 pg/m3

 (August), 119 pg/m3 (May), and 92 pg/m3 (July). An analysis of the results suggests that 

higher DDT and DDE levels correlated with air mass movement from the south, perhaps from areas 

where DDT is still used (i.e., Central America or Mexico). The fact that the ratio of DDT to DDE was 

<1 in each instance suggests that there is no new DDT use in Michigan. DDT and DDE levels over Green 

Bay, Wisconsin in 1989 were 8.7 and 15 pg/m3

, and those over the four lower Great Lakes obtained 

during a cruise were 38 and 59 pg/m3

 (McConnell et al. 1998). An analysis of air masses indicated that 

the atmospheric sources were not long-range transport, but rather local or regional volatilization. 

Stanley et al. (1971) measured atmospheric levels of pesticides in the United States during a time of high 

DDT usage. Nine localities were sampled representing both urban and agricultural areas. Of 

12 pesticides evaluated, only DDT was detected at all localities. Maximum levels of p,p’-DDT ranged 

from 2.7 ng/m3

 in Iowa City, Iowa to 1,560 ng/m3 in Orlando Florida. Maximum levels of o,p’-DDT and 

p,p’-DDE ranged from 2.4 to 131 ng/m3

. The highest levels were found in the agricultural areas of the 

South. The pesticides were predominantly detected in the particulate phase. Some agricultural areas in 

which DDT was extensively used have been monitored periodically since usage was halted. Atmospheric 

conditions in the Mississippi Delta were monitored intermittently from 1972 to 1975 (Arthur et al. 1977). 

Air samples taken in 1975 from an area with extensive cotton acreage had a mean ΣDDT concentration of 
DDT, DDE, and DDD 243 

6. POTENTIAL FOR HUMAN EXPOSURE 

7.5 ng/m3, compared to 11.9 ng/m3 in 1974. This represents a 36% decline in ΣDDT levels in 1 year. 

Between 1972 and 1974, the first 2 years after the use of DDT was banned, the atmospheric ΣDDT levels 

had declined by 88%. In 3 years, the decrease in ΣDDT air levels was 92%, representing a much more 

rapid decline than had been expected. In a comparison of the results from a 1995 study of the occurrence 

and temporal distribution of pesticides in Mississippi with the results obtained in 1967, a decline in the 

concentration of p,p’-DDE in air over agricultural lands was also noted (Coupe et al. 2000). 

Concentrations of p,p’-DDE in air were lower in the 1995 measurements, ranging from 0.13 to 1.1 ng/m3

, 

as compared to a range of 2.6–7.1 ng/m3

 obtained in 1967. However, these results also attest to the 

persistence of p,p’-DDT degradation products after >2 decades since the ban on DDT use in the United 

States (Coupe et al. 2000). 

p,p'-DDT, p,p'-DDE, and p,p'-DDD have all been detected in the dissolved and particulate phases of 

fogwater and air and in rainwater (Millet et al. 1997). Fogwater samples were 1.5–30 times higher in 

DDT, DDE, and DDD concentration than rainwater samples, and the distribution between dissolved and 

particulate phase appeared to be governed by the solubility of the chemical. The site of the measurements 

was a rural area in France between 1991 and 1993. DDT had not been used in the area since the 1970s. 

Ligocki et al. (1985) conducted concurrent rain and air sampling for rain events in Portland, Oregon, in 

1984. In rain samples, no p,p'-DDT, p,p'-DDE, or p,p'-DDD were detected. However, in the gas phase 

associated with this rainfall, p,p'-DDE was detected in five of seven samples. Levels detected in the 

samples ranged from nondetected to 420 pg/m3

. In another study, Poissant et al. (1997) reported the mean 

concentration of p,p’-DDT in precipitation over a rural site near the St. Lawrence river was 500 pg/L with 

a 75% frequency of detection. Rapaport et al. (1985) measured DDT residues in rain and snow samples 

in Minnesota. Samples of snow taken in 1981–1982, 1982–1983, and 1983–1984 contained an average 

of 0.32, 0.60, and 0.18 ng/L of p,p'-DDT, respectively. Two rain samples taken in 1983 contained 

0.2 and 0.3 ng/L of p,p'-DDT. In rainwater samples taken from a forested region in northeast Bavaria in 

1999, p,p’-DDT and p,p’-DDE were detected in three of six and four of six rainwater samples, 

respectively, ranging in concentration from not detected to 12.9 ng/L and not detected to 13.3 ng/L, 

respectively. In the vapour phase, p,p’-DDT and p,p’-DDE were detected in two of five and three of five 

air samples, respectively, ranging in concentration from not detected to 0.03 ng/m3

 and from not detected 

to 0.055 ng/m3

, respectively (Streck and Herrmann 2000). 
DDT, DDE, and DDD 244 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.4.2 Water 

Although there are numerous reports in the literature of DDT levels in specific bodies of water throughout 

the United States, there is little information providing evidence of trends in the DDT levels over time. 

EPA operates STORET (STOrage and RETrieval), a computerized water quality database. Staples et al. 

(1985) reported limited data on priority pollutants from STORET. Information from data collected from 

1980 to 1983 indicated that 3,500–5,700 ambient water samples were analyzed for DDT, DDE, and DDD 

with approximately 45% of the samples containing one of these compounds. The median level reported 

for both DDT and DDE was 0.001 μg/L, while the median level reported for DDD was 0.000 μg/L. 

Approximately 50 samples of industrial effluents were sampled and showed median levels of 0.010 μg/L 

for all three compounds. DDT was monitored in surface water and sediment as part of the National 

Surface Water Monitoring Program in 1976–1980. The percent occurrence and maximum concentrations 

of the reported DDT-related compounds in surface water were: p,p'-DDT, 0.5%, 0.70 μg/L (ppb); 

o,p'-DDT, 0.1%, 0.42 μg/L; p,p'-DDE, 0.7%, 0.55 μg/L; and o,p'-DDE, 0.3%, 0.54 μg/L (Carey and Kutz 

1985). The USGS and EPA cooperatively monitored levels of pesticides in water and sediment at 

Pesticide Monitoring Network stations between 1975 and 1980 (Gilliom 1984). Of the 177 stations 

(approximately 2,700 samples) monitored, 2.8, 0.6, and 4.0% contained detectible levels of DDT, DDE, 

and DDD in water, respectively. Fewer than 0.4% of the samples contained detectable DDT-related 

residues. The levels detected in water were not reported, but the limit of detection was 0.05 μg/L for 

DDT and DDD, and 0.3 μg/L for DDE. The percentage of sites having detectable levels of DDT-related 

residues in sediment was much higher (see Section 6.4.3). 

Johnson et al. (1988b) reported DDT and metabolite levels in the Yakima River basin in Washington 

State. Use of DDT was halted in this area when the 1972 ban was initiated; however, considerable 

residues are present in the river and sediments. Whole unfiltered water samples, collected mainly from 

the tributaries between May and October 1985, were reported to contain between not detectable to 

0.06 μg/L of DDT-related compounds. Concentrations of p,p'-DDT in water equaled or exceeded those 

of p,p'-DDE; an unexpected finding in light of what is believed concerning biological half-lives of DDT 

and its normal environmental degradation (Singh et al. 1999; Wolfe and Seiber 1993). The authors have 

suggested an unusually long half-life for DDT in Yakima basin soils, which would enter the river through 

runoff to explain the higher than expected p,p’-DDT/p,p’-DDE ratios. 

In the Malheur watershed, DDT was found to be persistent in the watershed, with estimated 

concentrations for ΣDDT ranging from 0.13–0.34 ng/L in rural regions of the watershed to 3.0–4.7 ng/L 
DDT, DDE, and DDD 245 

6. POTENTIAL FOR HUMAN EXPOSURE 

in urbanized areas of the Malheur River near Ontario, Oregon (Anderson and Johnson 2001). Unlike the 

relative concentrations of DDT and DDE in the Yakima River, the concentrations of DDE in water 

samples were higher than those measured for DDT (ranges from 0.03–0.25 and 0.07–0.14 ng/L for DDE 

and DDT, respectively, in rural areas and 1.9–4.3 and 0.25–0.61 ng/L near Ontario, respectively), 

indicating that although DDT was still persistent in this watershed, it was undergoing the expected 

environmental degradation. 

A summary of pesticide levels in surface waters of the United States during 1967 and 1968 was reported 

by Lichtenberg et al. (1970). During these 2 years (which were prior to the ban of DDT use), a total of 

224 samples (unfiltered) were analyzed from various sites in all regions of the country. DDT was found 

in 27 samples at levels ranging from 0.005 to 0.316 μg/L; DDE was found in 3 samples at levels of 

0.02–0.05 μg/L; and DDD was found in 6 samples at levels of 0.015–0.840 μg/L. 

According to the U.S. Geological Survey’s National Water Quality Assessment Plan initiated in 1991, 

that focuses on the water quality in more than 50 major river basins and aquifer systems, the frequency of 

detection of DDT and its metabolites in streams and groundwater was very low (USGS 1999). The top 

15 pesticides found in water were those with high current use. 

Only a few studies report levels of DDT in drinking water. Drinking water in Oahu, Hawaii, was found 

to contain p,p'-DDT at an average level of 0.001 μg/L in 1971 (Bevenue et al. 1972). In a study of 

Maryland drinking water during the September 1995, p,p’-DDE was detected in 22 out of 394 (5.6%) 

water samples, ranging in concentration from 0.039 to 0.133 μg/L (MacIntosh et al. 1999). 

Concentrations of p,p’-DDT and p,p’-DDD could not be measured above their limits of detection of 

0.021 and 0.028 ppb, respectively. Keith et al. (1979) reported that DDE was found in 2-month 

equivalent (the amount of water a person would theoretically consume over a 2 month period) samples 

collected over 2 days from two of three drinking water plants in New Orleans in 1974; the DDE 

concentration was 0.05 μg/L in both samples. 

Iwata et al. (1993) collected and analyzed 68 samples of surface water from several oceans (18 sampling 

locations) mainly affected by atmospheric deposition from April 1989 to August 1990. The range of 

mean and maximum concentrations of DDTs were (substance, range of means, maximum concentration): 

p,p’-DDE, 0.2–3.0 pg/L, 7.9 pg/L; o,p’-DDT, <0.1–5.8 pg/L, 14 pg/L; p,p’-DDT, 0.1–7.5 pg/L, 19 pg/L; 

and ΣDDT, 0.3–16 pg/L, 41 pg/L. The highest concentrations of DDT-related compounds were in the 
DDT, DDE, and DDD 246 

6. POTENTIAL FOR HUMAN EXPOSURE 

East China Sea. Other seas with high concentrations of DDT were the Bay of Bengal, Arabian Sea, and 

South China Sea. 

Canter and Sabatini (1994) reviewed Records of Decision at 450 Superfund Sites and found 49 cases in 

which contaminated groundwater threatened local public water supply wells. However, chlorinated 

organic pesticides were not found to be a major class of contaminants in these cases. In only one of the 

six sites in which the findings were presented in any detail was a DDT analog found at detectable levels. 

p,p’-DDD was found in monitoring wells from the upper aquifer at Pristine, Inc., an industrial site in 

Reading, Ohio at 0–0.14 μg/L but not in the lower aquifer or in water supply samples that were taken 

from the lower aquifer. No p,p’-DDT, or p,p’-DDE was detected in groundwater samples. Surface water 

samples contained levels of p,p’-DDE, p,p’-DDD, and p,p’-DDE at ranges of 0–0.86, 0–0.78, and 

0–1.82 μg/L, respectively. Even at sites where the surficial soil concentrations of p,p’-DDT are 

extremely high (29–959 mg/kg), the concentration of p,p’-DDT and its metabolites, p,p’-DDE and 

p,p’-DDD, in groundwater were close to their detection limits (#0.05 μg/L) (Vine et al. 2000). p,p’-DDE 

was detected in monitoring wells (depth range of 20–110 feet) set up around orchards and row crop fields 

in the Columbia Basin Irrigation Project, but was not detected in shallow domestic wells (depth range of 

80–250 feet) that were within 100 feet of these agricultural sites (Jones and Roberts 1999). 

6.4.3 Sediment and Soil 

Gilliom (1984) presented results of pesticide monitoring in sediment at USGS/EPA Pesticide Monitoring 

Network stations between 1975 and 1980. Of the 171 stations (approximately 900 samples) monitored, 

26, 42, and 31 contained detectible levels of DDT, DDE, and DDD, respectively. Fewer than 17% of the 

samples contained detectable DDT-related residues (limit of detection was 0.5 μg/kg for DDT and DDD, 

and 3 μg/kg for DDE). The percentage of sites with detectable levels of DDT-related residues in 

sediment was much higher than in water, reflecting the preferential partitioning of DDT to sediment (see 

Section 6.4.2). From 1980 to 1983, approximately 1,100 samples of sediments in EPA’s STORET 

database were analyzed for DDT, DDE, and DDD (Staples et al. 1985). The median levels for DDT, 

DDE, and DDD were 0.1, 0.1, and 0.2 μg/kg dry weight, respectively. In order to investigate 

circumstances contributing to the high level of DDT in fish and wildlife, soil and sediment samples 

(n=28) were collected in 1987 from the Upper Steele Bayou Watershed in west-central Mississippi at two 

depths (2.54–7.62 cm and 25.40–30.48) (Ford and Hill 1991). The results are given below: 
DDT, DDE, and DDD 247 

6. POTENTIAL FOR HUMAN EXPOSURE 

Compound Depth (cm) Occurrence Concentration (μg/kg) 

Mean Range 

p,p’-DDD 2.54–7.62 86% 40 ND–410 

p,p’-DDD 25.40–30.48 64% 20 ND–390 

p,p’-DDE 2.54–7.62 93% 100 ND–660 

p,p’-DDE 25.40–30.48 79% 40 ND–560 

p,p’-DDT 2.54–7.62 79% 30 ND–600 

p,p’-DDT 25.40–30.48 64% 20 ND–860 

ND = not detected 

River bed sediment samples collected in 1985 from the Yakima River basin in Washington contained 

0.1–234 ug/kg (dry weight) of ΣDDT and its metabolites (Johnson et al. 1988b). Use of DDT was halted 

in this area in 1972 when the ban was initiated. 

The concentrations of DDE, DDD, DDT, and ΣDDT in bed sediment from the San Joaquin River and its 

tributaries in California (7 sites) in 1992 were 1.4–115, 0.7–14, 0.4–39, and 2.2–170 ng/L, respectively 

(Pereira et al. 1996). One of the seven sites, Orestimba Creek, had DDT levels far higher than the other 

sites. Land use along this creek was dominated by orchards and a variety of row crops. Runoff that 

occurs during winter storms and the irrigation season contributes significant amounts of DDT-ladened 

sediment into the San Joaquin River and its tributaries (Kratzer 1999). For example, during the 1994 

irrigation season, it was estimated that approximately 136 g/day of ΣDDT entered the river and its 

tributaries from runoff, totaling around 5,190–8,920 g of ΣDDT for the season. Additionally, winter 

storm runoff can input large amounts of DDT within sediments into these surface waters in short periods 

of time. For example, a winter storm in January 1995 sent sediment-ladened runoff into the river and its 

tributaries, carrying upwards of 4,500 g/day of ΣDDT into these surface waters for a total contribution of 

1,750–2,620 g of ΣDDT from this one storm alone. 

Total DDT in surface sediment collected in eight remote lakes in Canada along a midcontinental transect 

from 49 EN to 82 EN declined significantly with latitude from 9.7 μg/kg (dry weight) to 0.10 μg/kg (Muir 

et al. 1995). The pattern of DDT deposition in lake sediment in the continental United States is 

exemplified by that in White Rock Lake in Dallas. Total DDT concentrations in the lake sediment 

increased from the mid-1940s to a maximum of 27 μg/kg in about 1965 when DDT usage peaked in the 

United States and have decreased by 93% to 2 μg/kg in the most recent samples collected in 1994 (Van 
DDT, DDE, and DDD 248 

6. POTENTIAL FOR HUMAN EXPOSURE 

Metre and Callender 1997; Van Metre et al. 1997). On the average, DDE accounted for 58% of the total 

DDT in the lake. DDD levels were about half those of DDE. The mean concentration of ΣDDT in 

sediment in the Newark Bay Estuary, New Jersey collected between February 1990 and March 1993 

ranged from about 100 to 300 μg/kg except for the Arthur Kill, where the mean concentrations exceeded 

700 μg/kg (Gillis et al. 1995). These levels may pose a potential threat to aquatic organisms. The 

maximum concentrations of p,p’-DDD, p,p’-DDE, and p,p’-DDT in sediment from 168 sites sampled 

along the southeastern coast of the United States as part of the Environmental Monitoring and Trends 

Program (EMAP) in 1994–1995 were 150.9, 34.2, and 35.0 μg/kg, respectively (Hyland et al. 1998). The 

median concentrations of these compounds were below the detection limit. DDT was monitored in 

surface water and sediment as part of the National Surface Water Monitoring Program in 1976–1980. 

The percent occurrence and maximum concentrations of the reported DDT analogs in sediments were: 

p,p'-DDT, 13.2%, 110.6 μg/kg; o,p'-DDT, 2.9%, 7.2 μg/kg; p,p'-DDE, 22.7%, 163.0 μg/kg; and 

o,p'-DDE, 0.5%, 1.3 μg/kg (Carey and Kutz 1985). Results were not presented for DDD. In 1983–1984, 

quarterly samples of bottom sediment were taken from six sites on tributaries of the Tennessee River near 

Huntsville, Alabama, and were analyzed for ΣDDT (Webber et al. 1989). From 1947 to 1970, DDT was 

manufactured along the tributary, and DDT-contaminated waste water was discharged into the river. The 

concentration of ΣDDT in sediment above the discharge point averaged less than 1 mg/kg dry weight. 

Remaining stations showed a decreasing gradient of ΣDDT with annual means ranging from 2,730 mg/kg 

at the closed site to the point of discharge to 12 mg/kg where the tributary empties into the Tennessee 

River 18 km away. 

According to the U.S. Geological Survey’s National Water Quality Assessment Plan initiated in 1991, 

which focuses on the water quality in more than 50 major river basins and aquifer systems, the frequency 

of detection of DDT and its metabolites in bed sediment in the 1990s remains high (USGS 1999). The 

metabolite with the highest frequency of detection was p,p’-DDE which was approximately 60% in urban 

areas, 48% in agricultural areas, and 46% in mixed land use areas followed by p,p’-DDD, p,p’-DDT, 

o,p’-DDD, o,p’-DDT, and o,p’-DDE. The frequency of detection of o,p’-DDT and o,p’-DDE was below 

5%. Sediments can act as repositories for DDT and its metabolites, serving as sources for these 

compounds for long periods of time, given the long half-lives of these compounds and their resistance to 

biodegradation (Sanger et al. 1999). Because DDT and its metabolites will fractionate and concentrate in 

organic material, the sediments of some waterways, such as salt marshes, that receive a large amount of 

organic content in washloads discharged from sources of water originating from urban and agricultural 

areas can act as potential DDT repositories (Masters and Inman 2000). Also, the concentrations of DDT 

and its metabolites are high enough in some sediments to exceed the threshold effects level (TEL), 
DDT, DDE, and DDD 249 

6. POTENTIAL FOR HUMAN EXPOSURE 

probable effects level (PEL), and the effects range low and median (ER-L, ER-M) for specific biota in 

marine and estuarine environments (Carr et al. 2000; Long et al. 1995). 

The U.S. National Soils Monitoring Program has provided valuable information on the overall pattern of 

DDT residues in soil during the years following the DDT ban. Each year since the ban of DDT, 

approximately 1,500 samples were taken. The results for 1970 are tabulated below (Crockett et al. 1974): 

Substance Occurrence Concentration ()g/kg) 

Mean Range 

p,p'-DDT 20% 180 10–69,300 

o,p'-DDT 14% 40 10–11,700

 p,p'-DDE 31% 50 10–6,820 

o,p'-DDE 3% <10 ND–510 

ΣDDT 23% 300 10–113,090
 

ND = not detectable
 

The mean ΣDDT level in five U.S. cities ranged from 120 to 560 μg/kg in 1971 (Carey et al. 1979a). 

Urban areas generally had higher pesticide levels than did nearby agricultural areas except in some 

southern cities near which the agricultural use of pesticides was traditionally heavy. 

DDT was heavily used in the corn belt in the mid-central United States. In a 1995–1996 sampling of 

38 soils in this region, ΣDDT varied from below quantitation to 11,846 μg/kg, with a geometric mean 

value of 9.63 μg/kg (Aigner et al. 1998). The geometric mean concentrations for p,p’-DDE, p,p’-DDT, 

p,p’-DDD, and o,p’-DDT were 3.75, 4.67, 1.20, and 1.79 μg/kg, respectively. At least one DDT analog 

was found in 33 of the soils. Nine of the samples contained ΣDDT above 200 μg/kg, while the 

concentrations in the rest of the samples were below 40 μg/kg. Two garden soils had ΣDDT levels of 

30 and 1.07 μg/kg. The soil with the high ΣDDT level was a muck soil with a concentration that was 

10 times higher than the sample next highest in concentration and 1,000 times higher than most sample 

concentrations. o,p’-DDD was not found in any of the samples. The DDT/DDE ratio was determined in 

21 of the samples and ranged from 0.5 to 6.6. It is interesting to note that the geometric mean o,p’-DDT 

concentration is 38% of the p,p’-DDT concentration. Since o,p’-DDT comprises between 15 and 21% of 

technical-grade DDT and 5.5% is comprised of other compounds, it would appear that o,p’-DDT 

degrades more slowly than p,p’-DDT. It was shown that the residue level of p,p’-DDT decreased about 

70% in a silt loam in New Zealand over a 30-year period (1960–1989), while the o,p’-DDT level only 
DDT, DDE, and DDD 250 

6. POTENTIAL FOR HUMAN EXPOSURE 

decreased by about 50% in the same time frame (Boul et al. 1994). Most of the degradation occurred 

during the time frame of 1960–1980, with very little loss occurring from 1980–1989. Forest soils in 

Maine that had been subject to aerial spraying with DDT had ΣDDT levels ranging from 270 to 

1,898 μg/kg compared with a maximum concentration of 11 μg/kg in unsprayed locations. A study of 

DDT in agricultural soils in British Colombia, Canada report that ΣDDT levels ranged from 194 to 

763 μg/kg in silt loam soils and from 2,984 to 7,162 μg/kg in muck soils (Aigner et al. 1998). The 

difference in residue levels reflects DDT’s longer persistence in muck soil. 

Hitch and Day (1992) reported that three soil samples taken near Dell City, Texas in 1980 contained an 

average of 4.94 and 0.46 mg/kg (dry weight) of DDT and DDE, respectively. It was suspected that the 

higher DDT concentrations indicated the possible illegal use of DDT. However, further analysis 

indicated that the "suspect" soil degraded DDT much slower than most soils and the high levels originally 

detected in soil were attributed to DDT persistence for many years. DDD was not measured in this study. 

DDT was extensively used in Arizona for 18 years, after which agricultural residues were closely 

monitored following a statewide moratorium on DDT use in January 1969. Levels of DDT plus 

metabolites in green alfalfa fell steadily from an average level of 0.22 mg/kg at the time of the ban to a 

level of 0.057 mg/kg 18 months later, and a level of 0.027 mg/kg after almost 7 years (Ware et al. 1978). 

After 3 years, residues in agricultural soils had decreased 23%. Furthermore, the ratio of DDE to DDT 

was increasing, indicating a transformation of DDT to DDE. Buck et al. (1983) reported similar results 

from monitoring these same sites over 12 years following the ban on DDT use. After 12 years, residues 

in green alfalfa averaged 0.020 mg/kg. At the end of the same period, combined DDT and DDE residues 

in agricultural soils had fallen from 1.2 to 0.39 mg/kg, while those in surrounding desert soil had fallen 

from 0.40 to 0.09 mg/kg. 

In 1985, DDT, DDE, and DDD levels were measured at the Baird and McGuire Superfund Site in 

Holbrook, Massachusetts. Contamination was due to 60 years of mixing and batching of insecticides. In 

the highly contaminated areas, the average concentrations of DDT, DDE, and DDD were 61, 10, and 

70 mg/kg, respectively. DDT, DDE, and DDD levels in leaf litter and leaf litter invertebrates ranged from 

0.2 to 8.4, nondetected to 60, and 0.4 to 25 mg/kg, respectively (Menzie et al. 1992). The high levels of 

DDT relative to DDE probably indicate that the Superfund Site is largely anaerobic, and that DDT is 

largely degrading to DDD. In the Palos Verdes Shelf off of Los Angeles where waste from a large DDT 

manufacturer was discharged via a sewer outfall, sediments contain high levels of DDT isomers and 

metabolites. The levels of these compounds in surface sediment (0–2 cm) at five sites in the area were 

(chemical, concentration range): o,p’-DDE, 6–45 mg/kg; p,p’-DDE, 10–327 mg/kg; p,p’-DDD, 
DDT, DDE, and DDD 251 

6. POTENTIAL FOR HUMAN EXPOSURE 

1–13 mg/kg; p,p’-DDD, 9–25 mg/kg; o,p’-DDT, not detectible–2 mg/kg; and p,p’-DDT, not 

detectible–6 mg/kg (Venkatesan et al. 1996). 

In summary, DDT, DDE, and DDD have been detected in many soil and sediment surfaces throughout the 

world. Concentrations are highest in areas with a history of extensive DDT use and are often detected at 

concentrations close to 1 mg/kg (ppm) or more. Even though concentrations of DDT, DDE, and DDD in 

soils are declining due to the discontinued production and use of DDT in most countries, detectable levels 

will probably exist for decades to come because of the long persistence time of these compounds. 

6.4.4 Other Environmental Media 

From 1980 to 1983, 19, 59, and 14 samples of aquatic biota in EPA’s water quality STORET database 

were analyzed for DDT, DDE, and DDD, respectively (Staples et al. 1985). The median levels for DDT, 

DDE, and DDD were 14, 26, and 15 μg/kg (wet weight), respectively. All samples tested had detectable 

levels of these chemicals. 

According to the U.S. Geological Survey’s National Water Quality Assessment Plan initiated in 1991, 

which focuses on the water quality in more than 50 major river basins and aquifer systems, DDT and its 

metabolites were detected in 94% of whole fish samples analyzed in the 1990s even though the total DDT 

concentration in fish continues to decline (USGS 1999). This is attributed to the presence of DDT in 

stream beds and continued inputs of DDT to streams as contaminated soils erode. The metabolite with 

the highest frequency of detection was p,p’-DDE followed by p,p’-DDD, p,p’-DDT, o,p’-DDD, 

o,p’-DDE, and o,p’-DDT. The frequency of detection of the o,p’-isomers was below 15%. 

ΣDDT concentrations in fish (8 species, 23 samples) collected in August and September 1990 from 

3 rivers in Michigan ranged from 4.71 to 976.92 μmol/kg, wet weight with a median of 82.1 μmol/kg 

(Giesy et al. 1994). The range of concentrations of DDT and metabolites were (chemical, range in μg/kg 

wet weight): p,p’-DDE, 3.54–627.13; o,p’-DDE, 0.15–37.95; p,p’-DDD, 0.43–58.82; o,p’-DDD, 

0.13–81.70; and p,p’-DDT, <0.42–89.58. The mean ΣDDT concentrations in samples taken below dams 

that separated the rivers from the Great Lakes, 0.5–1.6 μmol/kg, were higher than those taken above, 

0.05–0.35 μmol/kg. The relative contribution of DDE to ΣDDT was fairly constant in all three rivers 

both above and below the dams. The ratio of DDE:DDT ranged from 5 to 758, which suggests that the 

accumulation of DDE resulted from direct exposure to DDE in the diet rather than from recent exposure 
DDT, DDE, and DDD 252 

6. POTENTIAL FOR HUMAN EXPOSURE 

to parent DDT. The fact that DDT is still observed in the fish was ascribed to long-range transport and 

deposition. 

From 1986 to 1988, elements of the arctic marine food web near the Canadian Ice Island in the Arctic 

Ocean were sampled for DDT, DDE, and DDD (Hargrave et al. 1992). The average concentration of 

ΣDDT in plankton was 11.8 ng/g dry weight (43.5 μg/kg lipid), and the level increased with decreasing 

size of the plankton. Amphipods collected under pack ice in the open sea, over the Canadian continental 

shelf (190–315 m depth), and near the bottom of the Alpha Ridge (2,075 m depth), had mean ΣDDT 

concentrations of <57, 299, and 3,769 μg/kg dry weight (<347, 1,594, and 12,511 μg/kg lipid), 

respectively. Pelagic fish contained a mean ΣDDT of 200 μg/kg lipid, while abyssal fish (2,075 m) 

contained 819 μg/kg dry weight (1,465 μg/kg lipid). Similar comparisons have also been conducted on 

surface and deep-sea fish caught in the North and South Atlantic oceans and northwest Pacific ocean off 

California, showing higher concentrations of ΣDDT in deep-sea fish (Atlantic 175–1,090 μg/kg lipid; 

Pacific 2,380–2,420 μg/kg lipid) in comparison to surface fish (Atlantic 59–125 μg/kg lipid; Pacific 

1,260–1,875 μg/kg lipid) (Looser et al. 2000). The DDT levels in Arctic plankton are generally lower 

than those reported elsewhere. It is not clear why the DDT levels are higher in organisms living at greater 

depths since DDT appears to be evenly distributed in the water column. Since DDT adsorbs to particulate 

matter that sinks into the sediment, as with detritus from aquatic organisms, fish and other organisms 

living at the bottom of the sea may accumulate higher levels of DDT than organisms living at the surface 

because their food chain is associated with benthic feeders. Regional differences in DDT levels in biota 

may be associated with the productivity of the ocean and greater sedimentation of detritus from aquatic 

organisms. Arctic mammals feeding on DDT-contaminated fish bioaccumulate the chemical in their fat 

(Bard 1999). The ringed neck seal (n=19) and polar bear (n=10) had mean ΣDDT concentrations of 

1,482 and 266 μg/kg (lipid basis) (Muir et al. 1988). Beluga whales, ringed neck seals, and walruses near 

Baffin Island in the eastern Arctic had mean ΣDDT levels (wet weight) of 3.16, 0.33, and 1.42 μg/g, 

respectively (Kuhnlein et al. 1995). 

Exposure to DDT could occur to populations that consume fish from DDT-contaminated marine 

environments. DDT in white croaker and Dover sole of the Southern California Bight, especially the 

Palos Verdes shelf area, are the highest in the United States. This is due to the fact that this area received 

1,000,000 kg of DDT discharged into the Bight from the Montrose Chemical Company and also receives 

a large amount of sewage outfall from the southern California region (Zeng et al. 1999). Historically, 

DDT levels in these fish exceeded the Food and Drug Administration (FDA) action level of 5 mg/kg wet 
DDT, DDE, and DDD 253 

6. POTENTIAL FOR HUMAN EXPOSURE 

weight of fish tissue, and fish intended for human consumption were confiscated to prevent human 

exposure to DDT (NOAA 1988). 

Levels of ΣDDT have declined markedly since the early 1970s in fish, shellfish, and aquatic mammals 

(Addison and Stobo 2001; Bard 1999; Lauenstein 1995; Lieberg-Clark et al. 1995; Odsjo et al. 1997; 

Schmitt et al. 1990). Levels of DDT in fish were determined at 112 locations across the United States by 

the National Contaminant Biomonitoring Program in 1976 and 1984 (Schmitt et al. 1990). The mean 

concentrations of p,p'-DDT, p,p'-DDE, p,p'-DDD, and ΣDDT decreased from 50, 260, 80, and 370 μg/kg, 

respectively, in 1976 to 30, 190, 60, and 260 μg/kg, respectively, in 1984. A follow-up study of DDT in 

California sea lions reported a decrease in ΣDDT and DDE of over two orders of magnitude between 

1970 and 1992 (Lieberg-Clark et al. 1995). ΣDDT concentrations in maternal grey seals decreased from 

12 μg/g lipid in 1974 to 0.5 μg/g lipid in 1994; ΣDDT concentrations in seal pups were lower (60% of 

maternal concentrations) and decreased at similar rates over the same 20-year period (Addison and Stobo 

2001). ΣDDT for mussels and oysters analyzed as part of the National Oceanic and Atmospheric 

Administration’s National Status and Trends Mussel Watch Project in 1992 reported a geometric mean 

ΣDDT concentration for mussels and oysters at 51 sites of 20 μg/kg dry weight, down from a high of 

53 μg/kg in 1977 (Lauenstein 1995). Over 90% of the ΣDDT present was as metabolites rather then the 

parent compounds (p,p’- and o,p’-DDT). 

Among the metabolites of DDT are two methylsulfonyl metabolites of DDE, 2-methylsulfonyl-DDE 

(2-MeSO2-DDE) and 3-methylsulfonyl-DDE (3-MeSO2-DDE). These DDE metabolites are known to be 

persistent and have been measured in several species of mammals, including humans (for example, 

Bergman et al. 1994). The methylsulfonyl derivatives of DDE are formed through the action of phase I 

and II enzymes in the liver and the mercapturic acid pathway, as shown in Figure 3-4 of Chapter 3 

(Letcher et al. 1998; Weistrand and Norén 1997). DDE is converted to an arene oxide through the action 

of the phase I cytochrome (CYP) P450 2B-type enzymes, followed by the conjugation of the arene oxide 

with glutathione as part of the phase II reactions. As part of the mercapturic acid pathway, the 

glutathione function is converted to a cysteine residue, which is then cleaved by C-S lyase to form the 

thiol-substituted intermediates, 2-SH-DDE or 3-SH-DDE. These thiolated DDE derivatives are 

methylated by adenosyl-methionine and then oxidized to the methylsufonyl derivatives of DDE. The 

ratio of 2-MeSO2-DDE to 3-MeSO2-DDE varies between species and tissue site (Bergman et al. 1994). 

The two sulfonyl DDE metabolites have been measured in fat and various tissues of arctic mammals and 

in humans (Bergman et al. 1994; Haraguchi et al. 1989; Letcher et al. 1998; Norén et al. 1996; 
DDT, DDE, and DDD 254 

6. POTENTIAL FOR HUMAN EXPOSURE 

Weinstrand and Norén 1997). In pooled adipose tissue of polar bears from 12 arctic regions, the 

concentration of 3-MeSO2-DDE ranged from 0.60 to 11 μg/kg lipids, and the ratio of methylsulfone to 

DDE ranged from 0.009 to 0.056 with a mean of 0.033 (Letcher et al. 1995). These ratios may be the 

result of both biotransformation of DDE to methylsulfonyl-DDE in the animal and bioaccumulation 

(Letcher et al. 1998). In the polar bear food chain, lipid adjusted concentrations of 3-MeSO2-DDE in 

arctic cod (<0.01 ng/g, in whole body pools), ringed seal (0.4 ng/g, in blubber) and polar bear (2.0 ng/g, 

in fat tissue) were measured, showing an increase in the concentration of 3-MeSO2-DDE as a function of 

the trophic level (Letcher et al. 1998). In humans, methylsulfonyl-DDE has been measured in liver, lung, 

and adipose tissue at respective concentrations of 1.1, 0.3, and 6.8 ppb, wet weight (Haraguchi et al. 

1989). In plasma, the concentration of 3-MeSO2-DDE (0.1–2 ng/g lipid) was 2–3 orders of magnitude 

lower than the concentration of DDE (0.11–0.88 μg/g lipid) (Norén et al.1999). In a comparison of the 

concentrations of the two methylsulfonyl-DDE isomers in paired human liver and adipose tissues, 

3-MeSO2-DDE is the most abundant of the two isomers in these tissues (Weistrand and Norén 1997). In 

human breast milk, the concentration of 3-MeSO2-DDE has been found to range between 0.4 and 5 ng/g 

lipid (Norén et al. 1996). 

From 1979 to 1983, a study was conducted on the presence of DDT and metabolites in wildlife, 

predominantly birds, in orchards in central Washington State (Blus et al. 1987). Technical DDT was 

applied at very high rates to orchards in Washington between 1946 and 1970 with some areas probably 

receiving more than 1,000 kg/ha over this period. High levels of DDE, DDT, and DDD were found in the 

wildlife. Ninety-six percent of the wildlife samples (n=552) contained >0.01 μg/g of DDE, and 70% 

contained levels > 0.1 μg/g. In addition, many samples contained unusually low (#10:1) DDE:DDT 

ratios. The study attempted to identify whether the residues resulted from past legal use of DDT, ongoing 

illegal use, use of dicofol and related compounds, or foreign sources. While this matter wasn’t 

completely resolved, it was suspected that residues were from several sources. However, residues in 

certain samples, particularly resident wildlife, apparently originated from past legal use of the insecticide. 

High concentrations have been noted in animals from areas of historically high DDT use. Mean GDDT 

concentrations were 1,362 μg/kg in spring peeper frogs living in southern Ontario, Canada (Russell et al. 

1995). These concentrations exceed the suggested maximum concentration of 1,000 μg/kg proposed by 

the Great Lakes Water Quality Agreement of 1978. DDT was also applied at very high rates in the Delta 

region of Mississippi. The geometric mean concentration of p,p'-DDE residues in resident wood ducks 

decreased from 0.75 mg/kg in 1984 to 0.21 mg/kg in 1988 (Ford and Hill 1990). This decrease also 

corresponded with the reduction of residue levels in wood duck eggshells. Recent studies reporting 

concentrations of DDT and its metabolites in various biota is shown in Table 6-1. 
DDT, DDE, and DDD 255 

6. POTENTIAL FOR HUMAN EXPOSURE 

Even with the reduction in the levels of DDT in the environment, there are still areas of concern where 

heavy applications of DDT during past legal uses of the pesticide have resulted in high concentrations of 

residual DDT and its metabolites that can, in turn, have potentially adverse effects on wildlife. For 

example, the transfer of DDT, DDE, and DDD from fruit orchard soils to American robins had been 

investigated in Okanagan, British Columbia, and Ontario, Canada, showing increasing concentrations of 

these compounds in soil6earthworm6robin eggs (Harris et al. 2000). In Okanagan, high average 

concentrations (mg/kg, dry weight) of DDE and DDT in soil (5.5 and 9.2), earthworms (52 and 21) and 

robin eggs (484 and 73) were consistent with the recorded contamination of this area. These 

concentrations are comparable to those where mortality or reproductive effects have been observed to 

occur in field studies. These results also illustrate one way in which DDT and its metabolites in soil can 
Table 6-1. Concentrations of DDT and Metabolites in Biota 

Species Location Year Concentration Type Marine Mammals Pilot whale (n=7) North Atlantic Since 3,847 (942–7,118) ng/g (f.w.) [DDE] mean (range) 

1987 7,748 (1,708–13,035) ng/g (f.w.) [

ΣDDT] 

Harbor Porpoise (n=5) North Atlantic Since 3,260 (1,880–4,900) ng/g (f.w.) [DDE] mean (range) 

1987 7,280 (4,690–11,200) ng/g (f.w.) [

ΣDDT] 

Beluga whale (n=12) Arctic Since 1,415 (142–2,230) ng/g (f.w.) [DDE] mean (range) 

1987 2,492 (332–3,820) ng/g (f.w.) [

ΣDDT] 

Beluga whale (n=12) Cook Inlet Since 624 (65.9–1,630) ng/g (f.w.) [DDE] mean (range) 

1987 1,050 (133–2,350) ng/g (f.w.) [

ΣDDT] 

Northern fur seal North Pacific Since 1,190 (1,050–1,330) ng/g (f.w.) [DDE] mean (range) 

(n=2) 1987 1,280 (1,090–1,480) ng/g (f.w.) [

ΣDDT] 

Ringed seal (n=4) Arctic Since 198 (27–350) ng/g (f.w.) [DDE] mean (range) 

1987 543 (35–1,430) ng/g (f.w.) [

ΣDDT] 

Harbour seals (n=18) Northern Sea 1987 3,161 (355–6,598) μg/kg (f.w.) [

ΣDDT] mean (range) 

Harbour seals (n=32) Northern Sea 1988 3,903 (1,501–11,475) μg/kg (f.w.) [

ΣDDT] mean (range) 

Beluga whale Canadian 1988 3.16 μg/g (w.w.) [

ΣDDT] mean 

Arctic 

Narwhal whale Canadian 1988 2.73 μg/g (w.w.) [

ΣDDT] mean 

(n=unspecified) Arctic Walrus Canadian 1988 1.42 μg/g (w.w.) [

ΣDDT] mean 

(n=unspecified) Arctic Ringed seal Canadian 1988 0.33 μg/g (w.w.) [

ΣDDT] mean 

(n=unspecified) Arctic 
Table 6-1. Concentrations of DDT and Metabolites in Biota (continued) 

Species Location Year Concentration Type Beluga whale (neonate) (n=1) Terrestrial mammals Polar bear (n=320) Arctic ground squirrel (n=13)Arctic ground squirrel(n=6)Arctic ground squirrel(n=17)BirdsBald eagle chicks(n=51)Bald eagle eggs (n=6)Bald eagle eggs (n=6) Bald eagle eggs (n=7) 

St. Lawrence estuary near Quebec 

Arctic (16 

regions) Elusive Lake Feniak Lake Schrader Lake Great Lakes region 

Lake Erie Lake Erie 

Lake of the 

Woods, Canada 

1991 1989– 1993 

1991– 

1993 

1991– 

1992 1992– 1993 1990– 1996 

1974– 

1980 

1989– 

1994 1993– 1996 

702 ng/g (brain); 2,332 ng/g (kidney); 3,467 ng/g

(liver); 2,230 ng/g (fat) [

ΣDDT] 

689 ng/g (brain); 2,289 ng/g (kidney); 3,370 ng/g(liver); 2,106 ng/g (fat) [DDE] ND (brain); ND (kidney); 15 ng/g (liver); 17 ng/g(fat) [DDD] 219 μg/kg (f.w.) [DDE]52–560 μg/kg (f.w.) [DDE]6.13 (0.34–34.08) μg/kg (w.w.) (liver) [

ΣDDT] 

1.51 (0.33–5.57) μg/kg (w.w.) (liver) [DDE] 1.43 (0.19–5.16) μg/kg (w.w.) (liver) [

ΣDDT] 

0.86 (0.19–3.10) μg/kg (w.w.) (liver) [DDE] 12.25 (0.12–39.76) μg/kg (w.w.) (liver) [

ΣDDT] 

4.47 (0.12–13.63) μg/kg (w.w.) (liver) [DDE] ND–0.0171 mg/kg (plasma) [DDT] 0.0036–0.1484 mg/kg (plasma) [DDE] 24.4 (13.8–35.8) mg/kg [DDE] 10.8 (2.7–22.2) mg/kg [DDE] 3.3 (0.9–12.6) mg/kg [DDE] 

median range of geomeans mean (range) mean (range) mean (range) range 

mean (range) mean (range) mean (range) 
Table 6-1. Concentrations of DDT and Metabolites in Biota (continued) 

Species Location Year Concentration Type Blue heron eggs (n=10) Fish and shellfish Mussels and oysters Clams Clams Clams Clams Mountain whitefish (10composites from 7sites)

Southern Lake Michigan 

United States 

(51 sites) 

San Joaquin 

River (Orestimba Creek) 

San Joaquin 

River (Dry Creek) 

San Joaquin 

River (Mokelumne River) 

San Joaquin 

River (Stanislaus River) Yakima River Basin, Washington 

1993199219921992199219921989–1991

0.02 (ND–0.12) (μg/g) (w.w.) [DDT] 

1.58 (0.23–13.00) (μg/g) (w.w.) [DDE] 0.03 (ND–0.12) (μg/g) (w.w.) [DDD] 

0.51–1,400 ng/g (d.w.) [

ΣDDT] 

20 ng/g (d.w.) [

ΣDDT] 

4,350 ng/g (w.w.) [

ΣDDT] 

3,300 ng/g (w.w.) [DDE] 390 ng/g (w.w.) [DDD] 

29 ng/g (w.w.) [

ΣDDT] 

25 ng/g (w.w.) [DDE] 0.5 ng/g (w.w.) [DDD] 

15 ng/g (w.w.) [

ΣDDT] 

13 ng/g (w.w.) [DDE] 0.5 ng/g (w.w.) [DDD] 

24 ng/g (w.w.)[

ΣDDT] 

22 ng/g (w.w.) [DDE] <0.5 ng/g (w.w.)[DDD] 0.10–1.7 mg/kg (w.w.) (whole fish) [

ΣDDT] 

mean (range) range of sites geomean mean mean mean mean 

range of 

composites 
Table 6-1. Concentrations of DDT and Metabolites in Biota (continued) 

Species Location Year Concentration Type Largescale sucker (18 composites from 13 sites) 

Yakima River Basin, Washington 

1989– 1991 

0.05–4.37 mg/kg (w.w.) (whole fish) [

ΣDDT] range of 

composites 

Perch (n=5) Lake Ørsjøen, 

Norway Mid-lake 

1994 1.15 ng/g (w.w.), 1,643 ng/g (f.w.) [

ΣDDT] 

0.53 ng/g (w.w.), 757 ng/g (f.w.) [DDE] 0.26 ng/g (w.w.), 371 ng/g (f.w.) [DDD] 0.28 ng/g (w.w.), 400 ng/g (f.w.) [DDT] 

mean 

Perch (n=5) Lake Ørsjøen, 

Norway. 

1994 5.59 ng/g (w.w.), 11,180 ng/g (f.w.) [

ΣDDT] 

2.56 ng/g (w.w.), 5,120 ng/g (f.w.) [DDE] 1.48 ng/g (w.w.), 2,960 ng/g (f.w.) [DDD] 1.15 ng/g (w.w.), 2,300 ng/g (f.w.) [DDT] 

mean 

Pike (n=5) Lake Ørsjøen, 

Norway Mid-lake 

1994 7.3 ng/g (w.w.), 8,111 ng/g (f.w.) [

ΣDDT] 

3.5 ng/g (w.w.), 3,888 ng/g (f.w.) [DDE] 1.5 ng/g (w.w.), 1,667 ng/g (f.w.) [DDD] 1.8 ng/g (w.w.), 2,000 ng/g (f.w.) [DDT] 

mean 

Lake trout (n=59) Lake Ontario 1992 1.159 μg/g (w.w.) [DDE] mean 

Rainbow smelt (n=8) Lake Ontario 1992 0.256 μg/g (w.w.) [DDE] mean 

aU.S. National Biomonitoring Specimen Bank 

d.w. = dry weight; f.w. = fat weight basis; n = number; ND = not detected; geomean = geometric mean; w.w.= wet weight 
DDT, DDE, and DDD 260 

6. POTENTIAL FOR HUMAN EXPOSURE 

be mobilized and bioaccumulated by soil organisms which, in turn, are further accumulated in higher 

trophic levels. 

Market Basket Surveys indicated that there were decreases in the overall residue levels on a lipid basis of 

DDT and DDE in all classes of food tested from 1965 to 1975 (EPA 1980a). Between 1970 and 1973, 

DDE residues decreased only 27% compared to decreases of 86 and 89% for DDT and DDD, respectively 

(EPA 1980a). A study by Duggan et al. (1983) reported the following average residues of p,p'-DDT and 

o,p'-DDT in grocery items from 1969 to 1976: domestic cheese, 3 ppb; ready-to-eat meat, fish, and 

poultry, 5 ppb; eggs, 4 ppb; domestic fruits, 13 ppb; domestic leaf and stem vegetables, 24 ppb; domestic 

grains, 7 ppb; corn and corn products, 0.7 ppb; and peanuts and peanut products, 11 ppb. 

Mean DDT residues by food group have been reported by Gartrell et al. (1985, 1986a, 1986b) as part of 

the FDA Total Diet Studies for October 1979–September 1980 and October 1980–March 1982. The 

average DDE and DDT residues for 12 food groups and the daily intake for each of these groups obtained 

from the Total Diet Studies are shown in Table 6-2. The highest intake of DDE is shown to come from 

meat, fish, and poultry. More recent Total Diet Studies have only reported the number of occurrences of 

a pesticide and not the concentration levels. In the survey for 1984–1986, there were 433 findings of 

DDE out of 1,872 samples analyzed (Gunderson 1995b). In the Total Diet Study for 1993–1994, 

p,p’-DDE was found in 115 out of 783 (15%) items analyzed (FDA 1995). In the 1999 FDA Total Diet 

Study, DDT was found in 255 out of 1,040 (22%) items analyzed (FDA 1999). In a recent FDA study, 

the mean concentrations of p,p-DDT and o,p-DDT ranged from 0.0002 to 0.005 ppm. The mean 

concentrations of p,p-DDE ranged from 0.0001 to 0.0257 ppm, with the highest values found in dairy, 

fish and vegetable products (FDA 2001). Analyses of samples from 10 states taken during fiscal years 

(FY) 1988 (n=13,980) and 1989 (n=13,085) resulted in a frequency of detection of 0.028 and 0.12%, 

respectively, for p,p'-DDT. DDE (any isomer) was detected in 1.5 and 0.99% of samples and p,p'-DDE in 

0.18 and 0.25% of samples in 1988 and 1989, respectively (Minyard and Roberts 1991). Overall, these 

surveys indicate that DDT and DDE levels are very low in food commodities. However, with continued 

use of DDT in other countries, imported foods may continue to contribute small amounts of DDT and 

DDE to the daily diet of consumers. From 1981 to 1986, the FDA analyzed 13,283 imported agricultural 

commodities for pesticide residues. No commodities exceeded the EPA tolerance levels for DDT or 

DDE; however, 3.1% of the samples had detectable levels of DDT or DDE (Hundley et al. 1988). A 

pesticide residue survey of produce from 1989 to 1991 found that 41 out of 6,970 samples analyzed 

contained p,p’-DDE (Schattenberg and Hsu 1992). 
DDT, DDE, and DDD 261 

6. POTENTIAL FOR HUMAN EXPOSURE 

Table 6-2. Average Residues in Food Groups and Average Daily Intake from U.S.
 

FDA Total Diet Studies
 

October 1979 – September 1980a October 1980 – March 1982b 

DDE DDT DDE DDT 

Food Group 

Residue 

(ppb) 

Intake 

(μg/day) 

Residue 

(ppb) 

Intake 

(μg/day) 

Residue 

(ppb) 

Intake 

(μg/day) 

Residue 

(ppb) 

Intake 

(μg/day) 

Dairy products 0.9 0.626 0 0 1.5 1.05 0 0 

Meat, fish, and 

poultry 

4.8 1.28 0.8 0.219 3.0 0.77 

7 

0 0 

Grains and cereal 0 0 0 0 0 0 0 0 

Potatoes 0.5 0.0847 <0.1 0.0079 0.5 0.08 

64 

0 0 

Leafy vegetables 1.7 0.0954 0.2 0.0137 2.4 0.13 

2 

0 0.01 

95 

Legumes 0 0 0 0 <0.1 0.00 

14 

0.4 0 

Root vegetables 1.0 0.0309 0 0 4.6 0.14 

6 

0 0.01 

92 

Garden vegetables 0.2 0.0185 0 0 0.1 0.00 

95 

0.6 0 

Fruits 0 0 0 0 <0.1 0.00 

81 

0 0 

Oils and fats <0.1 0.0028 0 0 <0.1 0.00 

18 

0 0 

Sugar <0.1 0.0042 0 0 <0.1 0.00 

18 

0 0 

Beverages 0 0 0 0 0 0 0 0 

a

Adapted from Gartrell et al. 1985

b

Adapted from Gartrell et al. 1986b 
DDT, DDE, and DDD 262 

6. POTENTIAL FOR HUMAN EXPOSURE 

Baking, frying, broiling, smoking, and microwaving all effectively reduce the total DDT concentration in 

fish and meat tissue (Bayarri et al. 1994; Khanna et al. 1997; Wilson et al. 1998). The average reduction 

in fish ranged from 16 to 82% and in lamb from 37 to 56% depending on cooking method. It is not clear 

whether residues are lost as a result of volatilization or decomposition or carried away in fat runoff. 

p,p’-DDT (but not p,p’-DDE or p,p’-DDD) decomposes on heating (see Table 4-2). Concentrations of 

p,p’-DDT in tomatoes could be reduced by between 11.5 and 33.7% by washing the fruit with acetic acid 

and sodium chloride solutions; the concentrations of p,p’-DDT, o,p’-DDE, p,p’-DDD, and o,p’-DDD 

residues in tomatoes could be reduced by up to approximately 80% from the fruit by simply removing the 

peel (Abou-Arab 1999). Production of tomato paste through home-canning methods reduced p,p’-DDT 

concentrations by 30.7%. 

Djordjevik et al. (1995) assessed the chlorinated pesticide residues in U.S. and foreign cigarettes 

manufactured from the 1960s to the 1990s. Since 1970, the concentration of DDT analogs decreased by 

>98%. Concentration ranges of DDT-related compounds in samples of cigarettes manufactured between 

1961 and 1979 and between 1983 and 1994 were (chemical, 1961–1979 levels, 1983–1994 levels): 

p,p’-DDD, 1,540–30,100 ng/g, 12.6–99.7 ng/g; o,p’-DDD, 396–7,150 ng/g, ND-19.0 ng/g; p,p’-DDT, 

720–13,390 ng/g, 19.7–145 ng/g; o,p’-DDT, 105–1,940 ng/g; ND-88 ng/g; p,p’-DDE, 58–959 ng/g, 

6.6–15.8 ng/g; and p,p’-DDMU (1-chloro-2,2-bis(p-chlorophenyl)ethylene), 92.7–2,110 ng/g, 

ND-27.5 ng/g. The transfer rate from tobacco into mainstream smoke amounts to 22% for DDD, 19% for 

DDT, and 27% for DDE. 

Monitoring in older homes reveal that carpeting in these homes may have high levels of DDT, DDE, and 

DDD (Lewis et al. 1994). In one house built in 1930, the carpeting, which was believed to be at least 

25 years old, contained up to 10.8 μg/m2

 or 5.7 μg/g of ΣDDT (p,p’-DDT, DDD, and DDE). 

Organochlorine pesticides have been detected and quantified in composting feedstocks and finished 

compost (Büyüksönmez et al. 2000). Although banned for several decades, DDT and its metabolites have 

been detected in lawn trimmings and municipal waste compost as recently as 1990–1996, with 

concentrations of DDT, DDE, and DDD at 0.01–0.21, 0.01–0.11, and 0.007–0.13 ppm, respectively. The 

concentrations of the DDT, DDE, and DDD have been found to typically decrease as composting 

feedstocks are converted to finished compost. For example, DDT and DDE concentrations in lawn 

trimmings were found to decrease from 0.0466 and 0.0143 ppm to 0.0159 and 0.0108 ppm, respectively, 

after 90 days of composting. However, under some compositing conditions, DDE concentrations have 

been observed to increase in the finished compost (mean concentration of 0.0807 ppm, maximum value of 
DDT, DDE, and DDD 263 

6. POTENTIAL FOR HUMAN EXPOSURE 

0.483 ppm) compared to the initial feedstock (mean concentration of 0.0516 ppm, maximum value of 

0.201 ppm) (Strom 2000). 

6.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE 

The general population is currently exposed to DDT and its metabolites primarily in food. As indicated 

in the previous section, although residue levels in food continue to slowly decline, there are measurable 

quantities in many commodities. A 1989 pesticide screening program of produce delivered to 

supermarkets in Texas, for example, found p,p’-DDE residues in 41 of the 6,970 produce samples tested 

(Schattenberg and Hsu 1992). An FDA study of residues in infant foods and adult food eaten by infants 

and children in which over 10,000 samples of domestic and imported foods were analyzed during 

1985–1991 was published (Yess et al. 1993). ΣDDT was detected in 2 of 2,464 apples at a maximum 

concentration of 0.08 ppm; 312 of 2,464 plain milk samples at a maximum concentration of 0.92 ppm; 

8 of 180 vitamin D fortified milk samples at a maximum concentration of 0.10 ppm; and 1 of 

735 imported apple juice samples at 0.18 ppm (Yess et al. 1993). A similar 1992–1994 Canadian survey 

found DDE or DDT residues in 1 of 380 domestic heads of lettuce; 1 of 769 domestic potatoes; 36 of 

612 imported carrots; 4 of 721 imported cucumbers; 1 of 702 imported heads of lettuce; 14 of 

121 imported green onions; 7 of 17 imported parsnips; 1 of 933 imported peppers; 5 of 264 imported 

spinach; 1 of 155 imported tomato pastes; and 1 of 1,153 imported tomatoes (Neidert and Saschenbrecker 

1996). In a U.S. Market Basket study of ready to eat foods, o,p’-DDE was detected 8 times in 4 different 

food items at an average concentration of 0.0025 μg/g; p,p’-DDE was detected 1,700 times in 

142 different food items at an average concentration of 0.0026 μg/g; o,p’-DDT was detected 5 times in 

4 different food items at an average concentration of 0.0053 μg/g; p,p’-DDT was detected 98 times in 

31 different food items at an average concentration of 0.0045 μg/g (KAN-DO Office and Pesticide Team 

1995). 

Because of the extreme persistence of DDT and DDE, it is anticipated that low levels of residues will be 

present in commodities for decades. In fact, depending on use and export patterns in other countries, 

levels in the diet may even increase (Coulston 1985). Even in domestic commodities, the potential for 

low levels of dietary exposure of consumers may result from residues bioaccumulated in some food items, 

including fish. 

The estimated dietary intake of DDT and metabolites in the United States was 62 μg/person/day in 1965, 

240 μg/person/day in 1970, and 8 μg/person/day in 1974 (Coulston 1985). The FDA Adult Total Diet 
DDT, DDE, and DDD 264 

6. POTENTIAL FOR HUMAN EXPOSURE 

Study for October 1979–September 1980 (FY 1980) found that the intakes of ΣDDT, DDE, DDT, and 

DDD were 0.034, 0.003, 0.031, and <0.001 μg/kg body weight/day, respectively, down from highs of 

0.093, 0.004, 0.087, and 0.002, respectively, in FY 1979 (Gartrell et al. 1986a). The adult intake was 

assumed to be the diet of a 16- to 19-year-old male. Analogous studies for infants and toddlers for FY 

1980 reported daily intakes of the respective DDTs as 0.034, 0.034, ND (not determined), and ND μg/kg 

body weight/day for infants and 0.049, 0.045, 0.002, and 0.002 μg/kg body weight/day for toddlers 

(Gartrell et al. 1986b). Estimated dietary intakes of DDT determined from the FDA Total Diet Studies 

for June 1984–April 1986 and July 1986–April 1991 for eight population groups appear in Table 6-3 

(Gunderson 1995a, 1995b). To facilitate comparisons of DDT intakes from Gunderson (1995a, 1995b) 

with those of earlier estimates (Coulston 1985), the daily intake of ΣDDT for a 70 kg 16-year-old male as 

reported by Gunderson (1995a, 1995b) would have been 6.51, 2.38, 1.49, and 0.97 μg/day for 

1978–1979, 1979–1980, 1984–1986, and 1986–1991, respectively. The acceptable daily intake of DDT 

established by WHO/FAO is 20 μg/kg/day (WHO 1991). 

Exposure to DDT, DDE, and DDD in imported foods is minimized due to FDA enforcement programs. 

FDA randomly collects and analyzes a wide variety of imported commodities (e.g., coffee, tropical fruits) 

to determine if pesticide residues are above EPA tolerances. Pesticide tolerances established by EPA 

apply equally to domestic and imported food (Wessel and Yess 1991). 

Arctic indigenous people ingest high levels of DDT from traditional foods. A study covering three age 

groups in communities in the eastern and western Canadian Arctic found the average daily ΣDDT intake 

of 24.2 to 27.8 μg/day for the eastern Arctic community and 0.51 to 1.0 μg/day for the western Arctic 

communities (Kuhnlein et al. 1995). The foods with the highest ΣDDT concentrations were raw Beluga 

whale blubber (316 μg/g wet weight) and aged Narwhal whale blubber (273 μg/g wet weight) in the 

eastern Arctic, and baked Loch (species of fish) liver (1.85 μg/g wet weight) and smoked Canada goose 

meat (1.47 μg/g wet weight) in the western Arctic. 

In 1986–1988, EPA collected data at two sites, Jacksonville, Florida and Springfield/Chicopee, 

Massachusetts, to assess the nonoccupational exposure to pesticides (NOPES) for residents of these cities 

(Whitmore et al. 1994). Indoor p,p’-DDE and p,p’-DDT levels in air were higher than outdoor levels in 

these communities, and the highest number of indoor air samples with detectable DDT was observed in 
Table 6-3. Mean Daily Intake of DDT Per Unit Body Weight (μg/kg body weight/d

for Various Age Groups in the United States 

Analyte 6–11 mo 2 yr 14–16 yr F 14–16 yr M 25–30 yr F 25–30 yr M 60–65 yr 1984

–1986 

ΣDDT 0.0485 0.0499 0.0154 0.0213 0.0128 0.0155 0.0

o,p'-DDE 0.0002 0.0001 <0.0001 <0.0001 <0.0001 <0.0001 0.0p,p'-DDE 0.0468 0.0484 0.0149 0.0207 0.0123 0.0150 0.0p,p'-DDT 0.0004 0.0010 0.0003 0.0004 0.0003 0.0003 0.0p,p'-DDD 0.0011 0.0004 0.0002 0.0002 0.0002 0.0002 0.0

1986

–1991 

ΣDDT1 0.0448 0.0438 0.0138 0.0139 0.0106 0.0127 0.0

o,p'-DDE <0.0001 <0.0001 <0.0001 <0.0001 <0.0001 <0.0001 <0.0p,p'-DDE 0.0441 0.0420 0.0130 0.0151 0.0099 0.0119 0.0o,p'-DDT <0.0001 <0.0001 <0.0001 <0.0001 <0.0001 <0.0001 <0.0p,p'-DDT 0.0004 0.0011 0.0005 0.0005 0.0005 0.0005 0.0p,p'-DDD 0.0003 0.0007 0.0003 0.0003 0.0002 0.0003 0.0Source: Gunderson 1995a, 1995b 1The average daily 

ΣDDT intake of 0.8 )g/day for an adult used in Section 1.3 was derived from the average intakes for 25–30 year old 

assuming a body weight of 70 kg. The data presented in the table were derived from the June 1984 through April 1991 FDA Total Diet F = female; M = male; mo = month; yr = year 
DDT, DDE, and DDD 266 

6. POTENTIAL FOR HUMAN EXPOSURE 

the spring in Jacksonville (14%) and in the winter in Springfield/Chicopee (20%), with estimated mean 

air DDE and DDT concentrations of #1.0 ng/m3

. Mean ΣDDT air exposures were estimated as 22 ng/day 

in Jacksonville and 94 ng/day in Springfield/Chicopee. For comparison, dietary exposures in these two 

communities for 1982–1984 were estimated to be around 1,900 ng/day. Nine of 11 carpets tested in 

Jacksonville contained ΣDDT with median and mean levels of 0.7 and 1.2 μg/g, respectively. 

Until 1970, tobacco smoke contributed significantly to the intake of DDT by people, but since then, the 

amount of DDT in tobacco has dropped markedly and today, cigarette smoke is a minor source of human 

exposure (Djordjevic et al. 1995). 

Because of the extremely low solubility of DDT and DDE in water and the efficiency of standard water 

treatment methods in eliminating DDT-type chemical residues, intake of these compounds via drinking 

water is believed to be negligible. The criterion cited in the EPA Ambient Water Quality Criteria 

document is 0.059 ng/L, based on ingestion of 2L of drinking water per day plus 6.5 g of fish and 

shellfish per person (EPA 1999a). This criterion corresponds to an estimated increased cancer risk level 

of 1x10-7 or 1 in 10 million. 

Data indicate that, even with relatively high doses, there is minimal absorption of DDT through skin 

(Gaines 1969; Wester et al. 1990; Wolfe and Armstrong 1971). Therefore, exposure via dermal 

absorption was considered to be negligible. However, in reviewing the literature and using a dermal 

absorption factor of 15% measured in their laboratory, Moody and Chu (1995) calculated that in the 

worse-case scenario where a swimmer was in contact with 1 ppm of DDT from a water slick or sediment 

for 1 hour, a swimmer would absorb 200 μg of DDT, equivalent to a dose from a meal of contaminated 

fish. 

DDT and its metabolites are ubiquitous in the atmosphere but are present in such low concentrations that 

exposure via inhalation is negligible. Potential inhalation of relatively high levels of DDT should be 

possible only in areas of production or formulation. Wolfe and Armstrong (1971) estimated a respiratory 

exposure potential of 14.1 mg/person/hour for formulating plant workers; however, no current data were 

located on exposure of workers utilizing modern technology in the production and formulation of these 

compounds. 

DDT and DDE elimination from the body is not an efficient process; therefore, tissue levels will increase 

with repeated exposure if the absorbed dose is high enough. For this reason, body burdens of DDT and 
DDT, DDE, and DDD 267 

6. POTENTIAL FOR HUMAN EXPOSURE 

DDE tend to correspond with exposure levels, as indicated in long-term studies. From July 1969 to 1975, 

residues of DDT and its metabolites were measured in human adipose tissue collected through an annual, 

national survey — the National Human Monitoring Program for Pesticides (Kutz et al. 1977). During 

that time, levels of DDT and DDE in tissue samples declined. However, the frequency of occurrence in 

lipid samples did not decline, indicating both a long biological half-life and the ubiquitous occurrence of 

these compounds in the population. For FY 1970–1974, all samples were positive for DDT and 

metabolites (a total of 1,412 samples). Using all age groups sampled, the geometric mean lipid DDT and 

metabolite (combined) levels reported for each year from 1970–1974 were 7.88, 7.95, 6.88, 5.89, and 

5.02 ppm, respectively. Notable trends reported in Kutz et al. (1977) included increasing body burden 

with increasing age as well as a significant increase in residues in blacks when compared to whites. 

Results published for 1975 showed little change compared to 1974 (Kutz et al. 1979). Exposure to DDT 

in nonoccupationally exposed individuals, as manifested by their plasma DDE concentrations, was most 

reliably predicted by age and serum cholesterol concentration (Laden et al. 1999). Kutz et al. (1991) 

contains a listing of studies on DDT, DDE, and DDD levels in human adipose tissue in the general 

population of various countries from the 1950s to the mid 1980s. 

The Second National Health and Nutrition Examination Survey (NHANES II) has served as a 

continuation of the National Human Monitoring Program; however, published results have been few. 

Murphy and Harvey (1985) published selected results from the NHANES II survey for 1976–1980 based 

on data from the Northeast, Midwest, and South. These results are based, not on adipose samples, but on 

serum samples. For the years covered, 3,300 serum specimens were analyzed for DDT and DDE. In 31% 

of those samples p,p'-DDT was detected, with a median quantifiable level of 3.3 ppb (0.0033 ppm). 

However, p,p'-DDE was detected in 99% of samples tested, with a median quantifiable level of 11.8 ppb 

(0.0118 ppm). The limits of detectability was 2 ppb for p,p'-DDT and 1 ppb for p,p'-DDE. These results 

offered further proof of the extensive biological half-life of DDE as compared to DDT. Again, for both 

compounds, serum levels increased with increasing age. These data were not reported for each year, but a 

decreasing trend could be expected based on the data of Murphy and Harvey (1985). A more recent 

report on NHANES II for the period of 1976–1980 confirmed the above results on serum samples from 

5,994 persons. p,p'-DDE was detected in the serum of 99.5% of persons with a median level of 12.6 ppb 

(range: 0–379 ppb) whereas p,p'-DDT was quantifiable (>2 ppb) in only 10% of serum samples (StehrGreen 1989). Levels of p,p'-DDE increased with age and were higher in farm residents and in the South 

and West. 
DDT, DDE, and DDD 268 

6. POTENTIAL FOR HUMAN EXPOSURE 

Results of EPA’s 1986 National Human Adipose Tissue Survey (NHATS) in which 671 adipose tissue 

specimens were pooled into composite samples according to age, census region, sex, and race showed 

significant differences in p,p’-DDT and p,p’-DDE levels depending on age and census region (Lordo et 

al. 1996). The concentration of both compounds increased with age group, and while levels of p,p’-DDT 

were highest in the Northeast and lowest in the South, those of p,p’-DDE were highest in the West and 

lowest in the North Central region. Levels of both compounds had significantly increased from the 1984 

NHATS. The estimated national mean with relative standard error (%) p,p’-DDT concentrations for the 

1982, 1984, and 1986 NHATS were 189 (31%), 123 (11%), and 177 (20%) ng/g, respectively. Those for 

p,p’-DDE were 1,840 (350%), 1,150 (90%), and 2,340 (270%) ng/g, respectively. A 1985 survey of 

108 Canadian autopsy samples resulted in respective mean and maximum levels of p,p’-DDE at 811 and 

6,070 ng/g and p,p’-DDT at 48 and 250 ng/g (Mes et al. 1990). Adeshina and Todd (1990) analyzed 

DDT isomer and metabolite levels in 35 human adipose tissue samples of North Texas residents who 

were not occupationally exposed to DDT. The samples were obtained during autopsy in 1987 and 1988. 

The geometric mean concentrations were (substance, ng/g lipid): o,p’-DDE, 8 ng/g; p,p’-DDE, 679 ng/g; 

o,p’-DDT, 14 ng/g; p,p’-DDT, 294 ng/g; and ΣDDT, 1,031 ng/g. The ΣDDT levels can be compared with 

those from the human adipose tissue survey which were 7,950 ng/g lipid in 1970, 5,150 ng/g lipid in 

1974, and 1,670 ng/g lipid in 1983 (Adeshina and Todd 1990). 

DDT in milk is discussed in Section 6.6. Levels of DDE and DDT in human milk, blood, and tissue 

appear in Table 6-4. Correlations in the concentrations of DDT and its metabolites in human milk, 

adipose tissue, and blood serum have been observed. The lipid adjusted mean concentrations of 

p,p’-DDT, p,p’-DDE, and p,p’-DDD in human milk (0.65, 4.00, and 0.01 mg/kg, respectively) have been 

shown to correlate well (r2

=0.95, 0.89, and 0.75, respectively) with concentrations in adipose tissue (1.22, 

4.36, and 0.02 mg.kg, respectively) (Waliszewski et al. 2001). The lipid adjusted serum concentrations of 

p,p’-DDE were observed to correlate with concentrations of this compound in breast adipose tissue 

(Dorea et al. 2001). The concentrations of p,p’-DDT and p,p’-DDE in blood serum taken from mothers in 

Veracruz, Mexico (1.848 and 4.378 mg/kg fat, respectively) have also been observed to correlate 

(r2

=0.854 and 0.779, respectively) with concentrations in umbilical cord blood (2.800 and 4.676 mg/kg 

fat, respectively) (Waliszewski et al. 2000). 
Table 6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent 

DDT Units as compound Population Tissue Meana concentration reported Re

Milk 

p,p’-DDT Maternity patients in Milk fat 0.71, 1.69, 4.84b mg/kg Elv

Mexico City, Cuernavaca and rural Morelos 

p,p'-DDT Mothers in Sweden Milk fat 14 μg/g No

200

p,p'-DDT Maternity patients in Milk fat 0.162 mg/kg To

Mexico City 

p,p'-DDT Women in Germany Milk fat 0.7 (estimated from graph) mg/kg Scho,p'-DDT Maternity patients in Milk fat 0.138 mg/kg To

Mexico City 

p,p'-DDE Maternity patients in Milk fat 3.85, 6.51, 16.52b mg/kg Elv

Mexico City, Cuernavaca and rural Morelos 

p,p'-DDE Mothers in Sweden Milk fat 129 μg/g No

200

p,p'-DDE Inuit women in Canada Milk fat 962 μg/kg Dep,p'-DDE Quebec women between Milk fat 0.34 mg/kg De

1989 and 1990 (n=536) 

p,p'-DDE Maternity patients in Milk fat 0.594 mg/kg To

Mexico City 

p,p'-DDE Maternity patients in Milk fat 5.302 mg/kg Pa

Veracruz, Mexico 
Table 6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent Studie

DDT Units as compound Population Tissue Meana concentration reported Rep,p'-DDE Mothers of hospitalized Milk fat 0.318 mg/kg Kra

children in Zagreb, Croatia 

ΣDDT Canadian women - 1986 Milk fat 0.385 mg/kg Sm

(n=412) 

ΣDDT Arkansas women - 1986 Milk fat 0.99 mg/kg Sm

(n=536) 

Blood 

p,p'-DDT Workers in Sao Paulo, Serum 13.5 (DDT appliers) μg/L Min

Brazil 1.5 (unexposed) 

o,p'-DDT Workers in Sao Paulo, Serum <0.7–4.7 (range; DDT appliers) μg/L Min

Brazil 

p,p'-DDE Workers in Sao Paulo, Serum 64.3 (DDT appliers) μg/L Min

Brazil 14.3 (unexposed) 

p,p'-DDT Men in southeast Blood plasma 0.11 (lipid adjusted) ng/g Asp

Sweden 

p,p'-DDE Woman hospital patients Serum (lipid- 967 (median), <1.0–2261.5 (range) ng/g Arc

in New Haven, adjusted) (n=36) 199Connecticut 

p,p'-DDE Iowa and North Carolina Serum 0.39–6.51 (range) μg/L Bro

farmers and spouses 

o,p'-DDE Iowa and North Carolina Serum 0.71–2.31 (range) μg/L Bro

farmers and spouses 
Table 6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent Studie

DDT Units as compound Population Tissue Meana concentration reported Rep,p'-DDE Women without breast 

cancer in Long Island 

p,p'-DDE New York University 

Women’s Health Study (1985 to 1991) 

p,p'-DDE Female hospital patients 

in New York City 

p,p'-DDE Female hospital patients 

in New York City 

p,p'-DDE Mothers in Veracruz, 

Mexico 

p,p'-DDE Men in southeast 

Sweden 

p,p'-DDE Men in southeast 

Sweden 

DDE Controls in a casecontrol study nested 

within the Nurses Health Study (n=240) 

p,p'-DDE Four groups of refugees 

from Asia, ‘USSR’, Africa, ‘Yugoslavia’ (n=103); Controls from Germany (n=34) 

Serum Serum Plasma Plasma (lipidadjusted) Serum Plasma Plasma (lipidadjusted) Plasma Plasma 

4.7 

11.0±9.1 in cancer patients (n=58)

 7.7±6.8 controls (n=171) 6.93–7.29 (range of mean values) 0.963–0.997 (range of mean values) 

14.5 

2.4–14 (range of mean values among 

groups of men with different levels of fish consumption) 750–4500 (range of mean values among groups of men with different levels of fish consumption) 7.09 

2.30–16.90 (range of median values) 

12.20–93.00 (range of maximum values) (refugees) 1.14 (median), 4.97 (maximum) (controls) 

μg/L 

μg/L 

μg/L 

μg/mL 

ng/mL 

ng/g ng/g 

ppbc 

μg/L 

Ste

Wo

GaGa

Wa

AspAsp

Lad

Sc
Table 6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent Studie

DDT Units as compound Population Tissue Meana concentration reported Rep,p'-DDT Residents of Nainital, Serum 4.46 (mean), range (0.78–14.29) mg/L Du

India 

p,p'-DDE Residents of Nainital, Serum 1.55 (mean), range (0.14–4.10) mg/L Du

India 

p,p'-DDD Residents of Nainital, Serum 0.91 (mean), range (ND–2.82) mg/L Du

India 

p,p'-DDE New Bedford area infants Cord blood 0.493 ng/g Ko

serum 

p,p'-DDE Maternity patients in Cord blood 6.0 ng/mL Wa

Veracruz, Mexico 

p,p'-DDE Maternity patients in Venous blood 7.12 (mean), range (0–35.23) (n=52) ng/g Do

Nicaragua (lipid-adjusted) 

p,p'-DDE Maternity patients in Cord blood (lipid- 6.39 (mean), range (0–9.35) (n=52) ng/g Do

Nicaragua adjusted) 

p,p'-DDT Great Lakes fishermen Serum 0.3 (median), 0.05–0.8 (range) ppbc An

(n=30); Controls (n=180) ND (controls) 

o,p'-DDT Great Lakes fishermen Serum 0.06 (median), 0.03–0.3 (range) ppbc An

(n=30); Controls (n=180) ND (controls) 

p,p'-DDE Great Lakes fishermen Serum 5.2 (median), 0.6–23.9 (range) ppbc An

(n=30); Controls (n=180) 2.8 (median), ND–38.5 (range) 

(controls) 
Table 6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent Studie

DDT Units as compound Population Tissue Meana concentration reported Re

DDE Frequent GLSCF (Lake Serum 6.9 (males), 2.9 (females), 2.6 ppbc Ha

Michigan) males (n=98); (controls, males), 1.4 (controls, females (n=83); male females) (geometric means) controls (n=23); female controls (n=22) 

DDE Frequent GLSCF (Lake Serum 3.5 (males), 2.3 (females), ppbc Ha

Huron) males (n=65); 2.6 (controls, males), 0.6 (controls, females (n=37); male females) (geometric means) controls (n=3); female controls (n=3) 

DDE Frequent GLSCF (Lake Serum 3.8 (males), 2.0 (females), 2.0 ppbc Ha

Erie) males (n=89); (controls, males), 1.7 (controls, females (n=67); males females) (geometric means) controls (n=31); female controls (n=17) 

ΣDDT 1982 Great Lakes fish Serum 28.8 ppbc Ho

eaters (n=572); controls 10.6 (controls) (n=419) 

ΣDDT 1982 Southern Great Serum 25.8 ppbc Ho

Lakes fish eaters 9.6 (controls) (n=115); controls (n=95) 

ΣDDT 1989 Southern Great Serum 15.6 ppbc Ho

Lakes fish eatersd 6.8 (controls) (n=115); controls (n=95) 

Adipose and other tissue 

p,p'-DDT Children in Germany Adipose 0.6 (estimated from graph) mg/kg Sch
Table 6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent Studie

DDT Units as compound Population Tissue Meana concentration reported Rep,p'-DDT Children in Germany Bone marrow 

(lipid-adjusted) 

1.75 (estimated from graph) mg/kg Sch

p,p'-DDT Woman hospital patients 

in New Haven, Connecticut 

Adipose, breast (lipid-adjusted) 

132.2 (median), 54.0–418.2 (range) (n=36) 

ng/g Arc

199

p,p'-DDT Maternity patients in Adipose, 1.22 (mean), 0.01–9.03 (range) mg/kg Wa

Veracruz, Mexico abdomenal (lipid- (n=60) 

adjusted) 

p,p'-DDE Maternity patients in Adipose, 4.36 (mean), 0.31–16.04 (range) mg/kg Wa

Veracruz, Mexico abdomenal (lipid- (n=60) 

adjusted) 

p,p'-DDD Maternity patients in Adipose, 0.02 (mean), ND–0.25 (range) (n=60) mg/kg Wa

Veracruz, Mexico abdomenal (lipidadjusted) 

p,p'-DDE Woman hospital patients Adipose, breast 970 (median), 240.0–2,644.1 (range) ng/g Arc

in New Haven, (lipid-adjusted) (n=36) 199Connecticut 

p,p'-DDT Adults in Germany Bone marrow (dry 0.364 ppmc Sc

lipid-adjusted) 

p,p'-DDE Adults in Germany Bone marrow (dry 1.689 ppmc Sc

lipid-adjusted) 

p,p'-DDE Women without breast Adipose 546.7 ng/g Ste

cancer in Long Island 

p,p'-DDE Adults in Sweden who Liver 836 ng/g We

suffered sudden death (lipid-adjusted) 199
Table 6-4. Levels of DDT Compounds in Human Milk, Blood, and Tissues — Recent Studie

DDT Units as compound Population Tissue Meana concentration reported Rep,p'-DDE Adults in Sweden who 

suffered sudden death 

Adipose, abdominal 

788 ng/g We

199

p,p'-DDE FY1986 National 

Adipose Tissue Survey Composite samples (n=50, from 671 specimens) 

Adipose 2,340 (SE 12) (nation) 

1,710 (SE 22%) (0–14 years) 2,150 (SE 17%) (15–44 years) 3,080 (SE 13%) (45+ years) 

ng/g Lor

p,p'-DDT FY1986 National 

Adipose Tissue Survey Composite samples (n=50, from 671 specimens) 

Adipose 177 (SE 11%) (nation) 

73.0 (SE 36%) (0–14 years) 177 (SE 16%) (15–44 years) 252 (SE 13%) (45+ years) 

ng/g Lor

p,p'-DDE Women patients at 

Hartford Hospital, Hartford, Connecticut 

Adipose, breast (lipid basis) 

2,200±1,470 cancer patients (n=20) 1,487±842 controls (n=20) 

ng/g Fal

p,p'-DDT Women patients at 

Hartford Hospital, Hartford, Connecticut 

Adipose, breast (lipid-adjusted) 

216±174 cancer patients (n=20) 148±75 controls (n=20) 

ng/g Fal

aArithmetic mean concentrations are reported unless otherwise specified.bGeometric mean concentrationscppm=μg/g; ppb=ng/gdSame Southern Great Lakes fish eaters that participated in the 1982 study.

FY = fiscal year; ND = not detected; SE = standard error 
DDT, DDE, and DDD 276 

6. POTENTIAL FOR HUMAN EXPOSURE 

Comparison of the concentrations of DDT and its metabolites in breast tissues and serum have been 

conducted in women with breast cancer in comparison to control subjects. The mean concentrations 

(unadjusted for age) of p,p’-DDT and p,p’-DDD in breast tissues of U.S. women with breast cancer 

(261.6 and 9.8 ng/g lipid, respectively) did not statistically differ (p=0.23 and 0.79) from those measured 

in control subjects (267.3 and 24.0 ng/g lipid); however, the concentration of p,p’-DDE was statistically 

higher (p=0.006) in breast cancer patients (800.0 ng/g lipid) than in control subjects (709.1 ng/g lipid) 

(Bagga et al. 2000). When the mean concentrations for p,p’-DDT, p,p’-DDE, and p,p’-DDD were 

adjusted for age, there was no statistical difference (p#0.001) between cases and controls for all three 

compounds. Statistically higher concentrations (p<0.05) of p,p’-DDE in serum were observed in both 

premenopausal (2.40 μg/g lipid) and postmenopausal (5.10 μg/g lipid) breast cancer patients from Mexico 

City, Mexico, in comparison to controls (1.93 and 3.12 μg/g lipid, respectively) (Romieu et al. 2000). 

The increased concentrations of p,p’-DDE in breast tissues of cancer patients has been attributed to 

increased exposures of these women to p,p’-DDE rather than to differences in the metabolism of 

p,p’-DDE by cancer cells (Romieu et al. 2000). However, in a study of breast cancer patients and control 

subjects in the New York University Women’s Health Study, there was no statistical difference between 

the geometric means of the p,p’-DDE concentrations in serum (1,097 ng/g lipid in patients versus 

977 ng/g lipid in controls) obtained from these two groups of women (Wolff et al. 2000b). A number of 

other studies have also found no statistical differences in the mean concentrations of DDT and/or its 

metabolites in serum or adipose tissue between cancer cases and controls (Demers et al. 2000; Dorgan et 

al. 1999; Helzlsouer et al. 1999; Krieger et al. 1994; Laden et al. 2001a; Liljegren et al. 1998; LopezCarrillo et al. 1997; Mendonca et al. 1999; Moysich et al. 1998; Schecter et al. 1997; Unger et al. 1984; 

van’t Veer et al. 1997; Ward et al. 2000; Wolff et al. 2000; Zheng et al. 1999, 2000). 

Methyl sulfonyl metabolites of p,p’-DDE, primarily the 3-methylsulfone isomer, have been found in 

seven surveys of human milk in Sweden between 1972 and 1992 (Norén et al. 1996). In that time, levels 

declined from 5.05 to 0.46 ng/g lipids and the ratio of the 3-methylsulfone metabolite to p,p’-DDE 

remained constant at 0.002. 

Fish from areas like the Great Lakes and Baltic Sea appear to be an important source of exposure to DDT 

and DDE, and human blood levels of these compounds have been found to correlate with the 

consumption of fish containing high levels of DDT and DDE (Anderson et al. 1998; Asplund et al. 1994; 

Hovinga et al. 1992). A Swedish study found that the mean plasma lipid concentrations of p,p’-DDE 

were 750, 1,200, and 4,500 ng/g in groups of men eating no fish, moderate quantities of fish, and large 

quantities of fish, respectively, from the Baltic Sea (Asplund et al. 1994). The respective lipid plasma 
DDT, DDE, and DDD 277 

6. POTENTIAL FOR HUMAN EXPOSURE 

concentrations of p,p’-DDT in these groups of men were 20, 45, and 130 ng/g. The mean serum DDT 

level in individuals eating more than 20 pounds of sport-caught Great Lakes fish dropped from 25.8 to 

15.6 ppb (65% decrease) during the period from 1982 to 1989. Mean serum DDT levels in the controls 

dropped from 9.6 to 6.8 ppb (41% decrease). It was concluded that the decrease in serum DDT 

concentrations was due to lower levels of DDT in the fish and in the environment, rather than to a 

decrease in fish consumption (Hovinga et al. 1992). 

A study of residents in Triana, Alabama, living downstream from a former DDT manufacturing facility 

revealed mean serum levels of total DDT of 76.2 ppb (Kreiss et al. 1981). This was several times higher 

than other reported levels. Kreiss et al. (1981) also found that serum DDT levels increased with 

increasing age. Residents living near a pesticide dump site in Aberdeen, North Carolina, known to 

contain high concentrations of DDT, have been shown to have age-adjusted mean levels of DDE in their 

blood of 4.05 ppb, which is higher than the mean value of 2.95 ppb obtained from residents of 

neighboring communities (Vine et al. 2000). 

Mean levels of total equivalent of DDT, DDE, and DDD in maternal blood in pregnant women in India 

(20 samples) were found to be 25.3 ppb compared to levels in placental tissue of 22.2 ppb (Saxena et al. 

1987). Similar levels (30.8 ppb) were seen in maternal blood in Brazilian women (Procianoy and 

Schvartsman 1981). Saxena et al. (1981, 1983) presented data on a limited number of samples of blood 

and placental tissues of women that aborted or delivered prematurely, which suggested that p,p'-DDE 

concentrations were elevated compared to control groups. 

Adipose tissue from a subgroup of 40 workers engaged in spraying DDT for malaria control in Mexico 

contained the following median and maximum levels of DDT metabolites (μg/g): ΣDDT, 114.60, 665.56; 

p,p’-DDT, 46.96, 344.98; o,p’-DDT, 2.96, 29.74; p,p’-DDE, 64.96, 298.42; and p,p’-DDD, 0.62, 

3.51 (Rivero-Rodriguez et al. 1997). Based on these measurements and a survey of the work habits of 

other workers, a geometric mean p,p’-DDE concentration of 67.41 μg/g was predicted for the population 

of 331 workers, 80% of whom were employed in the sanitation campaign for 20 years. Mean ΣDDT 

serum level in a group of 26 malaria control sprayers in Brazil was 76.9 μg/L and ranged from 7.5 to 

473.5 μg/L, whereas 16 unexposed workers had mean serum levels of 16.1 μg/L (range: 5.1–32.9 μg/L) 

(Minelli and Ribeiro 1996). p,p’-DDT and p,p’-DDE serum levels in the exposed workers ranged from 

1.6 to 62.9 and 5.9 to 405.9 μg/L, respectively. 
DDT, DDE, and DDD 278 

6. POTENTIAL FOR HUMAN EXPOSURE 

6.6 EXPOSURES OF CHILDREN 

This section focuses on exposures from conception to maturity at 18 years in humans. Differences from 

adults in susceptibility to hazardous substances are discussed in Section 3.8 Children’s Susceptibility. 

Children are not small adults. A child’s exposure may differ from an adult’s exposure in many ways. 

Children drink more fluids, eat more food, breathe more air per kilogram of body weight, and have a 

larger skin surface in proportion to their body volume. A child’s diet often differs from that of adults. 

The developing human’s source of nutrition changes with age: from placental nourishment to breast milk 

or formula to the diet of older children who eat more of certain types of foods than adults. A child’s 

behavior and lifestyle also influence exposure. Children crawl on the floor, put things in their mouths, 

sometimes eat inappropriate things (such as dirt or paint chips), and spend more time outdoors. Children 

also are closer to the ground, and they do not use the judgment of adults to avoid hazards (NRC 1993). 

Children are exposed to DDT through their diet. Since the greatest dietary intake of DDT is from meat, 

fish, poultry, and dairy products, infants and young children for whom a substantial part of their food is 

milk may be exposed to DDT. According to the FDA study of 1986–1991, the mean daily intake of DDT 

and its metabolites is 0.0448, and 0.0438 μg/kg body weight/day for a 6–11-month-old infant and 

2-year-old child, respectively (Gunderson 1995b). This is roughly four times the intake per unit body 

weight for an adult (see Table 6-3). 

DDT and DDE selectively partition into fatty tissue and into human breast milk, which has a higher fat 

content than cow's milk. The concentration of DDT, or other hydrophobic pollutants, in milk is often 

expressed on a lipid basis (i.e., μg/g lipid rather than μg/mL milk) as it is a more accurate measure of 

DDT content due to the fluctuating fat content of the milk. Generally, these compounds are found in 

human breast milk in concentrations higher than in cow's milk or other infant foods. As a result, breastfed infants may receive higher dietary exposure than those who are not breast-fed. If a woman has been 

exposed to high levels of DDT in the past, her milk may contain high levels of DDT, which would be 

transferred to her child. Women exposed to high levels of DDT would include Eskimos and Indian 

women in Arctic regions who eat traditional foods as well as women who eat large quantities of fish from 

lakes and rivers known to have high concentrations of DDT in fish, such as the Great Lakes and the 

Yakima River, Washington (Bard 1999; Kuhnlein et al. 1995; Marien and Laflamme 1995). Methods 

have been proposed for estimating breast milk lipid concentrations of DDT from a mother’s daily intake 

(Marien and Laflamme 1995). Mean levels of p,p’-DDT in human breast milk in pooled milk from the 
DDT, DDE, and DDD 279 

6. POTENTIAL FOR HUMAN EXPOSURE 

Mothers’ Milk Center in Stockholm steadily declined from 0.71 μg/g lipid in 1972 to 0.36, 0.18, and 

0.061 μg/g lipid in 1976, 1980, and 1984–1985, respectively (Norén 1988). Mean levels of p,p’-DDE for 

these years were 2.42, 1.53, 0.99, and 0.50 μg/g lipid, respectively. Between 1967 and 1985, the levels of 

p,p’-DDE and p,p’-DDT in human milk in Sweden declined by 75 and 95% (Norén 1993). In another 

study conducted between 1972 and 1992, this same group of investigators noted a similar decline in 

p,p’-DDE and p,p’-DDT concentrations in human milk in Sweden; the rates at which the concentration of 

these two compounds have been declining has also been progressively decreasing with time during this 

same period (Norén et al. 1996). The use of DDT was banned in Sweden in 1970. Mean (maximum) 

p,p’-DDT concentrations in 54 samples of mothers’ milk from Hawaii (1979–1980) were 0.16 (0.52) μg/g 

lipid compared with 0.19 (1.7) μg/g lipid in 102 samples from the U.S. mainland (Takei et al. 1983). 

Mean (maximum) p,p’-DDE levels in Hawaiian and mainland samples were 2.0 (5.7) and 1.9 (11.0) μg/g 

lipid. A 1982 Canadian survey that included 210 samples of breast milk from across the country resulted 

in mean levels of p,p’-DDE, p,p’-DDT, p,p’-DDD, and o,p’-DDT in ng/g milk (ng/g milkfat) of 34 (911), 

3 (80), 1 (27), and trace (12), respectively, down from 103, 33, 4, and 5 ng/g milk, respectively, obtained 

in a 1967 survey (Mes et al. 1986). The maximum p,p’-DDE, p,p’-DDT, p,p’-DDD, and o,p’-DDT levels 

in the 1982 survey were 5,500, 450, 113, and 58 ng/g milkfat. Levels of DDT in breast milk have shown 

a downward trend starting in about 1970. In 28 studies from the United States and Canada, average DDT 

levels in breast milk were about 4,000–5,000 ng/g lipid in the early 1970s, and then steadily declined by 

1975. For 13 studies from 1975 on, there was an 11–21% reduction in mean ΣDDT levels per year. 

Another way of viewing this is that the mean breast milk level in the population is being reduced by onehalf in 4.2–5.6 years. Similar reductions have been observed in Western European countries. While 

exposure of humans by eating fish from the Great Lakes has been a source of concern, in one study, Mes 

and Malcolm (1992) found that levels of DDE and DDT in breast milk were lower in women in the Great 

Lake’s Basin than in women in the rest of Canada. Levels of DDE in cow’s milk have similarly declined. 

The mean level of DDE in milk supplies in Southern Ontario, Canada declined from 96 ng/g lipid in 

1970–1971 to 16 ng/g lipid in 1985–1986, indicating that the levels are being reduced by one-half in 

5.8 years (Frank and Braun 1989). Since levels of DDT in food have been declining, exposure of 

children to DDT through their diet would be much less than in the past. 

Children may be exposed to DDT by ingesting contaminated soil or dust, from dermal contact with the 

soil, or by inhaling in the dust and then swallowing it after mucociliary transport up out of the lungs. 

DDT is extremely persistent in soil and there are soils that still contain high levels of the insecticide. No 

reports have been found, however, concerning childhood exposures to DDT by ingesting dirt. DDT is 

strongly adsorbed to soil, especially when the organic content of the soil is high. No studies were found 
DDT, DDE, and DDD 280 

6. POTENTIAL FOR HUMAN EXPOSURE 

as to how bioavailable DDT-adsorbed soil is when ingested. In addition, no information was found on the 

absorption of ingested DDT in any form in children. Children may also be exposed to DDT improperly 

stored at waste sites. A recent study indicated that old carpeting may contain high levels of DDT (Lewis 

et al. 1994). The DDT may have contaminated the carpet material or may have been tracked in from 

outside. Children may be exposed to this DDT while crawling around or playing on contaminated 

carpeting. 

Since DDT partitions into lipids and is not readily metabolized, levels of DDT in adipose tissue increase 

with age. Levels of DDT and DDE in children aged 0–14 as reported in EPA’s FY 1986 National 

Adipose Tissue Survey appear in Table 6-4 (Lordo et al. 1996). 

6.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES 

Because of the ban on DDT use after 1972, fewer persons in the United States should be exposed to high 

levels of these compounds today than in the past. Only fish and marine mammal consumption in the 

Arctic appear to be significant dietary contributors to human exposure to DDT in the general population 

(Laden et al. 1999). A 1982 study by the Michigan Department of Public Health found that people eating 

large quantities of Great Lakes fish had significantly higher serum DDT levels compared to non-fisheating controls. Furthermore, fish consumption was a major predictor of exposure. A follow-up study in 

1989 found that serum DDT levels were primarily a reflection of historic exposures and previously 

established body burden rather than recent exposure (Hovinga et al. 1993). Other studies confirm these 

findings (Anderson et al. 1998; Hanrahan et al. 1999). The best predictors of serum DDE levels in 

frequent Great Lakes sport fish consumers were found to be age, years of eating sport caught fish, male 

gender, and body mass index (BMI), which respectively accounted for 20, 10, 9, and 9% of the variance 

(Hanrahan et al. 1998). In general, DDT-contaminated fish are caught by sport or subsistence fisherman 

and not purchased at the market (Laden et al. 1999). As the levels of DDT in Great Lakes fish decline, 

fish consumption is less likely to be a source of potentially high exposure. Because of the partitioning of 

DDT and DDE into fatty tissue and fluids, breast-fed infants are likely to receive doses in excess of those 

occurring from ingestion of cow's milk or other infant foods. Monitoring exposure of infants via breast 

milk has been extensive and provides evidence of the persistence of DDT and DDE in fatty tissues. The 

finding that old carpeting may contain high levels of DDT indicates that this may be an important, but 

unevaluated source of exposure, especially in small children crawling on the carpeting (Lewis et al. 

1994). More details about children’s exposures can be found in Section 6.6 Exposures of Children. 
 

 

DDT, DDE, and DDD 281 

6. POTENTIAL FOR HUMAN EXPOSURE 

Workers involved with formulation, packaging, and application of DDT in the past would be expected to 

have been exposed to levels higher than those encountered in the environment. Persons who live near 

NPL sites containing DDT, DDE, or DDD might be exposed to higher levels than the general population 

since DDT has been detected in 441 of the 1,613 hazardous waste sites that have been proposed for 

inclusion on the EPA NPL (HazDat 2002). 

6.8 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of DDT, DDE, and DDD is available. Where adequate 

information is not available, ATSDR, in conjunction with the National Toxicology Program (NTP), is 

required to assure the initiation of a program of research designed to determine the health effects (and 

techniques for developing methods to determine such health effects) of DDT, DDE, and DDD. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 

6.8.1 Identification of Data Needs 

Physical and Chemical Properties. The physical and chemical properties of p,p’-DDT, DDE, and 

DDD are well described in the literature although there are some gaps in data for the o,p’-isomers (see 

Table D-1). The p,p’-isomers are those of primary environmental concerns and the data available are 

sufficient to allow estimation of the environmental fate of DDT, DDE, and DDD. 

Production, Import/Export, Use, Release, and Disposal. Since the banning of DDT in the 

early 1970s in the United States, there has been little information published on the production of DDT. 

DDT is no long produced in the United States or in most countries in the world. The most recent 

information indicates that it is produced in at least two countries, and is used in some underdeveloped 

countries for vector control. However, data would be useful on the production and use of DDT world 

wide. This type of information is important for estimating the potential for environmental releases from 
 

 

DDT, DDE, and DDD 282 

6. POTENTIAL FOR HUMAN EXPOSURE 

various uses, as well as estimating the potential environmental burden. In turn, this would provide a basis 

for estimating potential exposure and public health risk. 

Disposal information is equally important for determining environmental burden and areas where 

environmental exposure may be high. Although disposal methods for DDT and its metabolites are 

reported to a limited extent, no current information on disposal sites and quantity disposed was located. 

Information on how the current users (e.g., hazardous waste clean-up crews) wash DDT equipment and 

dispose of the remaining waste would be helpful for estimating potential environmental and human 

exposure. 

Environmental Fate. DDT, DDE, and DDD released to the environment may be transported from 

one medium to another by the processes of solubilization, adsorption, bioaccumulation, or volatilization. 

The transport of DDT, DDE, and DDD between environmental compartments has been predicted mostly 

from their physical and chemical properties. Volatilization and adsorption account for loss of DDT and 

its metabolites from surface water and soil. Monitoring studies indicate that DDT and its isomers and 

metabolites are extremely persistent in soil (EPA 1986a) and substantiate their predicted environmental 

fate. DDT, DDE, and DDD are highly lipid soluble. This, combined with their extremely long 

persistence, contributes to bioaccumulation of DDT and its metabolites in freshwater and marine life. 

Limited data were located on the soil degradation rates of DDT and its metabolites. Data are available for 

disappearance rates including losses due to transport processes. While adequate data are available on the 

time for the disappearance of 50% of the DDT initially applied to a variety of soils, there is abundant 

evidence that subsequent declines in DDT in soil occur at a much slower rate largely due to an aging 

process. More data on the biodegradation rates of DDT and its metabolites as well as how soil properties 

and aging affect these rates would be useful. Experimental information characterizing the environmental 

fate of DDT, DDE, and DDD, particularly on those properties that govern transport to air, would be 

helpful to further confirm their predicted environmental behavior and potential human exposure. 

Bioavailability from Environmental Media. Limited information was located regarding the 

bioavailability of DDT, DDE, and DDD from environmental media. It has been shown that the 

bioavailability of DDT in soil declines with time (Alexander 1995, 1997; Robertson and Alexander 1998) 

and soil properties that influence the bioavailability of DDT and its toxicity to certain organisms have 

been studied (Peterson et al. 1971). More information regarding the aging process of DDT in soil and its 

affect on bioavailability would be helpful in identifying potential routes of human exposure. It is known 

that fish and some plants bioaccumulate these compounds and that those who consume these fish and 
 

 

DDT, DDE, and DDD 283 

6. POTENTIAL FOR HUMAN EXPOSURE 

plants will incur some exposure to these compounds. However, because of universal body burdens of 

these compounds, the relative contribution of any particular medium, especially soil and sediment, is not 

clearly understood. Even if DDT, DDE, and DDD concentrations in various media are known, the 

difference between the exposure level and the absorbed dose is still unknown. 

Food Chain Bioaccumulation. Information was located regarding food chain biomagnification of 

total DDT in the arctic marine food web (Hargrave et al. 1992). Fairly extensive monitoring of fish 

populations has been performed and a bioconcentration factor in fish is available. The steady-state BCF 

in rainbow trout was reported as 12,000, suggesting that bioconcentration in aquatic organisms is very 

high (Oliver and Niimi 1985). Although DDT has been detected in plants and vegetables, root uptake of 

DDT is considered low (Fuhremann and Lichtenstein 1980; Lichtenstein and Schultz 1980). A clearer 

understanding of the potential for bioaccumulation would aid in determining how levels in the 

environment affect the food chain and potentially impact human exposure levels. This type of 

information could be obtained by studying accumulation of these compounds in organisms from several 

trophic levels. 

Exposure Levels in Environmental Media. Information on environmental levels of DDT, DDE, 

and DDD are abundant for the 1970s and 1980s (Blus et al. 1987; Carey et al. 1979b; Crockett et al. 

1974; Ford and Hill 1990; Hargrave et al. 1992; Lichtenberg et al. 1970; Stanley et al. 1971). More 

recent information has been more limited in scope (Aigner et al. 1998; McConnell et al. 1998; Monosmith 

and Hermanson 1996). Continuation of data collection on environmental levels would contribute to the 

understanding of current worldwide concentrations and trends, especially in regions where DDT is 

currently used in vector control for malaria and as an agricultural pesticide. 

Reliable monitoring data for the levels of DDT, DDE, and DDD in contaminated media at hazardous 

waste sites are needed so that the information obtained on levels of DDT, DDE, and DDD in the 

environment can be used in combination with the known body burden of DDT, DDE, and DDD to assess 

the potential risk of adverse health effects in populations living in the vicinity of hazardous waste sites. 

Estimates of human intake have been limited to dietary intakes based on current market basket surveys 

(EPA 1980a; Gartrell et al. 1985, 1986a, 1986b; Gunderson 1995a). Additional information is needed 

relating to the levels in environmental media to which the general population is exposed, particularly at or 

near hazardous waste sites, and the subsequent development of health effects. 
 

 

 

DDT, DDE, and DDD 284 

6. POTENTIAL FOR HUMAN EXPOSURE 

Exposure Levels in Humans. Data are available on levels of DDT and its metabolites in adipose 

tissue, blood, and milk (Hovinga et al. 1992; Lordo et al. 1996; Smith 1999). Recent monitoring data 

appear in Table 6-4. 

Exposures of Children. More data are needed on the concentrations of DDT in breast milk of 

exposed women and on the DDT intake of breast-fed infants. In addition, the oral availability of DDT 

from soil and dust is lacking. Such data would allow for the estimation of the exposure of children to 

DDT from eating soil and dust. 

Child health data needs relating to susceptibility are discussed in Section 3.13.2, Identification of Data 

Needs: Children’s Susceptibility. 

Exposure Registries. No exposure registries for DDT, DDE, or DDD were located. The ATSDR 

Division of Health Studies will consider these chemicals when primary chemical selection is made for 

future subregistries for the National Exposure Registry. The information that is amassed in the National 

Exposure Registry facilitates the epidemiological research needed to assess adverse health outcomes that 

may be related to exposure to these substances. 

6.8.2 Ongoing Studies 

The bioavailability of DDT, DDE, and DDD in soils to earthworms and the estimation of bioavailability 

by chemical extraction methods is being investigated by Professor Martin Alexander and coworkers at 

Cornell University. These studies will aid in the understanding of how the aging process of chemicals 

affects bioavailability. 

The remediation of soil and sediments contaminated with select organic contaminants, like DDT and its 

metabolites, is being investigated by several groups. A study of the bioremediation of soil through the 

action of coryneform bacteria in metabolizing chlorinated hydrocarbon insecticides, as compared to 

faster-growing pseudomonad bacteria is being investigated by D.D. Focht and colleagues at the 

University of California at Riverside. This group is examining the substrate specificity of a biphenyl-A 

gene (bphA) in the metabolism of chlorinated hydrocarbon insecticides and are also looking at the role of 

surfactants in uptake and catabolism of chlorinated hydrocarbon pesticides. Laboratory investigations 

into the feasibility of remediation of sediments contaminated with select organic contaminants, like DDT, 
DDT, DDE, and DDD 285 

6. POTENTIAL FOR HUMAN EXPOSURE 

are currently being undertaken by W.L. Mauck and colleagues at the Columbia Environmental Research 

Center in Columbia, Missouri. 

An investigation of the phytoremediation of soils to remove persistent organics like p,p’-DDE is being 

conducted by J.C. White and colleagues at the Connecticutt Agriculture Experiment Station in New 

Haven, Connecticutt. In this research, the investigators plan to identify plants that are able to intercept 

and actively remove p,p’-DDE and other compounds from soils through the use of such techniques as 

chelating agents to open up soil and the addition of surfactants to mobilize recalcitrant compounds in soil. 

These investigators also propose to identify and test organisms in rhizosphere and nonrhizosphere zones 

in soil that effectively degrade pesticide-contaminants and then innoculate the most tolerant and effective 

degraders into rhizosphere of plants to stimulate the effectiveness of phytoremediation. 

Exposures of persistent organochlorine compounds, like DDT, to wildlife is in the deep sea, in remote, 

pristine environments and in the Columbia River system are ongoing. D.C. Kadkoto and R.G. Zika at the 

University of Miami, and C.R. Smith at the University of Hawaii, are investigating the exposure of deepsea whale-fall communities in the Santa Catalina Basin to persistent organic compounds, like DDT, 

through the measurement of these compounds in the skeletal remains of whales. J.E. Estes and colleagues 

at the Western Ecological Research Center in Santa Cruz, California, are monitoring endocrine-disrupting 

compounds in Adak Island, Alaska, measuring environmental concentrations of select compounds, like 

DDT, in species (e.g., bald eagles, fish, mussels, seabirds, otters) and aboitic media, gaining both 

temporal and spatial assessments of contamination. Monitoring of persistent compounds (dioxins, furans, 

and PCBs) in Columbia River system and the relationship of the presence of these compounds to the 

nesting success of Osprey is being investigated by C.J. Henny and colleagues at the Forest and Rangeland 

Ecosystem Science Center in Corvallis, Oregon. 

A study of dietary human exposure to select contaminants, such as p,p’-DDE, is being conducted by D.L. 

Macintosh and colleagues in the Department of Environmental Health Sciences, University of Georgia, 

Athens, Georgia. Associations will be drawn between contaminant levels in biological fluids, such as 

blood, and urine, drawn from study participants to concentrations of these contaminants in environmental 

samples such as food products, air, water, and soil, with the purpose of identifying various exposure 

pathways with regard to total exposure. These investigators will also investigate temporal (e.g., time of 

year), spatial (e.g., region of country), and demographic (e.g., socioeconomic status) factors on the 

magnitude and sources of exposure. 
DDT, DDE, and DDD 286 

6. POTENTIAL FOR HUMAN EXPOSURE 

An analytical methodology has been developed to quantitate pesticides in carpet and surface dust and is 

now being used to assess exposure of agricultural workers, and their families, or families in agricultural 

regions, to pesticides, with an emphasis on children’s exposures to select pesticides, like DDT. A.T. 

Lemley, S.M. Snedeker and colleagues at Cornell University will use this information, in conjunction 

with information available in the literature, to assess cancer risk and to identify risk reduction procedures. 

A number of research groups are investigating the levels of DDT and its metabolites in humans and the 

association of these levels with specific human health end points. R. Bhatia and colleagues at the Public 

Health Institute, Berkley, California, are examining DDT concentrations in the cord blood of a subset of 

50 infants from the Children’s Health and Development Studies cohort to determine if maternal levels of 

DDT have an effect on the incidence of male genital anomalies. Another group headed by S. Korrick at 

the School of Public Health, Harvard University, is measuring DDE and other compounds in cord serum 

and/or breast milk in children/women living near a PCB-contaminated harbor (Superfund site) in 

Massachusetts to ascertain the effect that these compounds may have on childhood cognitive function. In 

the Oswego Newborn and Infant Development Project, T. Darvill and colleagues are tracking newborns 

of women who did or did not consume Great Lakes fish during pregnancy, analyzing cord blood and 

placental tissues to determine the effect persistent compounds, such as DDT, have on cognitive function 

in children. Breast cancer risk as a function of exposure to DDT and other xenoestrogens, as measured in 

blood samples, collected from women in Wisconsin is being investigated by P.L. Remington at the 

University of Wisconsin at Madison. 
DDT, DDE, and DDD 287 

7. ANALYTICAL METHODS 

The purpose of this chapter is to describe the analytical methods that are available for detecting, 

measuring, and monitoring DDT, DDE, and DDD, their metabolites, and other biomarkers of exposure 

and effect to DDT, DDE, and DDD. The intent is not to provide an exhaustive list of analytical methods. 

Rather, the intention is to identify well-established methods that are used as the standard methods of 

analysis. Many of the analytical methods used for environmental samples are the methods approved by 

federal agencies and organizations such as EPA and the National Institute for Occupational Safety and 

Health (NIOSH). Other methods presented in this chapter are those that are approved by groups such as 

the Association of Official Analytical Chemists (AOAC) and the American Public Health Association 

(APHA). Additionally, analytical methods are included that modify previously used methods to obtain 

lower detection limits and/or to improve accuracy and precision. 

7.1 BIOLOGICAL MATERIALS 

Table 7-1 lists the analytical methods used for determining DDT in biological fluids and tissues. DDT, 

DDE, and DDD residues have been measured in biological samples such as adipose tissue, skin lipids, 

blood serum, urine, milk, and other samples primarily by gas chromatographic (GC) methods. GC 

methodology provides high resolution and a reproducibility of retention time, which is ideal for 

distinguishing between the p,p’- and o,p’-isomers of the compounds, especially when using GC capillary 

columns (Mukherjee and Gopal 1996). The GC separation method has been historically coupled with 

electron capture detection (ECD) quantitative techniques (Fishbein 1974). For example, the GC/ECD 

methodology proposed by Cranmer et al. (1972b) has detected DDT, DDE, and DDD in human urine at 

levels as low as 50 pg/sample. In human serum, detection limits of between 2 and 7 pg/g serum for DDT, 

DDE, and DDD have been reported for the GC/ECD quantitative method (Atume and Aune 1999). With 

the wider availability of gas chromatography-mass spectrometry (GC/MS) instrumentation in analytical 

laboratories, GC/MS detection methods have been used to quantify DDT and its metabolites (Akiyama et 

al. 2000; Gill et al. 1996). Both the GC/ECD and GC/MS analytical methods are suitable for the analysis 

of DDT, DDE, and DDD. However, the GC/ECD method typically provides greater detection sensitivity, 

whereas the GC/MS method has the advantage of providing qualitative information to determine the 

specificity of the analysis. Various authors cited in Table 7-1 used GC methods to monitor the residues of 

these compounds in blood, serum, semen, liver, human milk, and adipose tissue, which were detectable at 

the ppm and ppb level. Since DDT partitions in fat, analyses are often performed on 
Table 7-1. Analytical Methods for Determining DDT, DDE, and DDD in Biological Sa

Analytical 

Sample matrix Sample preparation method Sample detection limit Percent recovery 

Blood/plasma/serum Extract with hexane GC/ECD 2 ppb (DDT); 1ppb (DDE); 

2 ppb (DDD) 

>90% (DDT); 100–110% (DDE); No data (DDD) 

Blood/plasma/serum Extract with methanol and 

hexane-ethyl ethers; cleanup with Florisil 

GC/ECD 0.8 ppb (DDE); 

No data (DDT, DDD) 

90–100% (DDE); No data (DDT, DDD) 

Blood Extract with hexane; 

concentrate to 5 mL 

GC/ECD HERL_004 

No data No data 

Serum Extract with hexane; cleanup 

with Florisil 

GC/ECD 91 pg/g serum (DDE) 83.1–85.5% (PCB) 

Semen Extract with acetone; cleanup 

with Florisil 

GC/ECD No data 96–97% (DDT); 

91.4% (DDE); 91.4% (DDD) 

Urine Extract with acetic acid in 

hexane followed by methylation 

HPLC/NAA 0.01 mg/mL (DDT); 

No data (DDE, DDD) 

No data 

Urine Extract with hexane GC/ECD 2 pg (DDE); 

No data (DDT, DDD) 

93.2–106.2% (DDE); No data (DDT, DDD) 

Liver, kidney, human milk 

Macerate sample with acetonitrile; cleanup with Florisil 

GC/ECD No data 81% (DDD); 

No data (DDT, DDD) 
Table 7-1. Analytical Methods for Determining DDT, DDE, and DDD in Biological Samples 

Analytical 

Sample matrix Sample preparation method Sample detection limit Percent recovery 

Muscle Homogenized and extracted 

with hexane 

GC/ECD 2 pg (DDE); 

No data (DDT, DDD) 

93.2–106.2% (DDE); No data (DDT, DDD) 

Human milk Triple solvent extraction with 

ethanol, hexane, and hexane-ethyl ether; Florisil cleanup 

GC/ECD 2 ppb (DDE); 

No data (DDT, DDD) 

81–108% (DDE); No data (DDT, DDD) 

Human milk Extract with hexane; cleanup 

with Florisil 

GC/MS 2 ppb (DDT); 1.5 ppb 

(DDD); No data (DDE) 

80–100% (DDD); No data (DDT, DDE) 

Human milk Head-space solid-phase 

extraction; desorption from solid phase in GC injector 

GC/ECD 0.08 μg/L (p,p’-DDT) 

2.79 μg/L (o,p’-DDT) 1.92 μg/L (p,p’-DDE) 1.36 μg/L (o,p’-DDE) 1.62 μg/L (p,p’-DDD) 1.85 μg/L (o,p’-DDD) 

No data 

Milk Remove proteins with 

ethanol; extract with hexane; cleanup with Florisil 

GC/ECD 6 pg/g serum (DDE) 95.1% (PCB) 

Milk/adipose tissue Extract with hexane and 

petroleum ether; cleanup with GPC 

GC/ECD HERL_026 

No data No data 

Adipose tissue Digest with perchloric-acetic 

acid; extract with n-hexane 

GC/ECD 2 pg (DDE); 

No data (DDT, DDD) 

93.2–106.2% (DDE); No data (DDT, DDD) 
Table 7-1. Analytical Methods for Determining DDT, DDE, and DDD in Biological Samples 

Analytical 

Sample matrix Sample preparation method Sample detection limit Percent recovery 

Adipose tissue Extract with petroleum ether, 

cleanup with Florisil 

GC/ECD No data 85–100% (DDT); 

No data (DDD, DDE) 

Feces Hexane extraction; evaporate 

and reconstitute with isooctane 

GC/ECD 20 ppb (DDT, DDE); 

No data (DDD) 

92–111% (DDT); 96–109% (DDE); No data (DDD) 

Lymph Co-extract with ether; final 

extraction with cyclopentanone 

HPLC No data 96.4% (DDT); 

No data (DDE, DDD) 

Skin lipids Purify with GPC, wash with 

sulfuric acid 

GC/ECD No data 96–109% (DDE); 

No data (DDT, DDD) 

Milk (MeSO2-DDEa) Liquid-gel partitioning 

followed by adsorption and GPC cleanup 

capillary GC/MS 

No data 80%, mean 

(MeSO2-DDE) 

Lung, blubber, liver (MeSO2-DDE) 

Extraction with GPC cleanup GC/ECD 

GC/AED 

No data No data 

aDDE methyl sulfone 

AED = atomic emission detection; ECD = electron capture device; GC = gas chromatography; GPC = gel permeation chromatography; 

Environmental Research Laboratory of the Environmental Protection Agency; HPLC = high performance liquid chromatography; MS = m

NAA = neutron activation analysis; PCB = polychlorinated biphenyls 
DDT, DDE, and DDD 291 

7. ANALYTICAL METHODS 

adipose tissue, milk fat, or a lipid extract of serum or other material. In the latter case, the results may be 

reported on a lipid or fat basis (i.e., ng DDT/g lipids). By reporting monitoring studies of DDT on a lipid 

basis, variability in results due to variability in fat content is reduced (McKinney et al. 1984; Phillips et 

al. 1989). 

A methyl sulfonyl metabolite of DDE that has been found in many tissue samples, 3-methyl sulfonyl 

2,2-bis(4-chlorophenyl)-1,1-dichloroethene, can be analyzed by GC/MS or GC/ECD (Janak et al. 1998; 

Norén et al. 1996), although gas chromatography-tandem mass spectrometry is the preferred instrumental 

technique for the analysis of 3-methyl sulfonyl 2,2-bis(4-chlorophenyl)-1,1-dichloroethene in complex 

biological matrices (Letcher and Norstom 1995). However, the use of atomic emission detection 

significantly improves its determination (Janak et al. 1998). While methods exist for measuring DDT, 

DDE, and DDD in liver, breast milk, and adipose tissue, the most frequently used sampling techniques 

utilize samples of blood, urine, and semen because of ease of sample collection. DDT, DDE, and DDD 

can also be measured in skin lipids collected by wiping the face with cotton (Sasaki et al. 1991b). 

Although these methods can detect and quantify levels of DDT, there is no information available to 

quantitatively correlate levels in these fluids with environmental levels or toxic effects. 

7.2 ENVIRONMENTAL SAMPLES 

DDT residues are found in the environment because of its slow transformation. DDT was used as an 

insecticide from the late 1940s until the early 1970s. Well-established analytical test procedures to 

analyze environmental samples use GC and MS (see Table 7-2). EPA methods 608 and 8081B are 

recommended to detect DDT, DDE, and DDD in surface water and municipal and industrial discharges 

(EPA 1982, 1998j). These are required procedures under the Clean Water Act. Behzadi and Lalancette 

(1991) described a modified isotope dilution (MID) GC/MS method to analyze DDT, DDE, and DDD in 

water and soil samples. Sample preparation for MID GC/MS does not require extensive extraction and 

cleanup compared to GC/MS. The detection limits are in the 0.001 μg/L (ppt) range, and recoveries 

range from 73 to 110% for soil and from 90 to 116% for water. EPA methods 8081B and 8270D are 

GC/MS methods used to determine DDT and its metabolites in soils with detection limits of 

0.3–0.4 μg/kg (EPA 1998j, 1998k). GC/ECD and nitrogen-phosphorus detection (NPD) is used for the 

analysis of DDT in foods with a detection limit of 0.5 μg/kg (ppb) (Rodriguez et al. 1991). GC/ECD is 

also used for the analysis of DDT and its metabolites in fish, oysters, and waterfowl. Detection limits 

were reported in the ppb range and recoveries ranged from 66 to 97% (Blus et al. 1987; Ford and Hill 
Table 7-2. Analytical Methods for Determining DDT, DDE, and DDD in Environmental 

Analytical Sample detection 

Sample matrix Sample preparation method limit Percent recovery 

Air Separate with silicic acid 

column 

GC/ECD 0.20 ng/m3 (DDE); 

0.16 ng/m3 (DDT) 

104–106% (DDT); 100% (DDE); No data (DDD) 

Air Filter collection and 

iso-octane extraction 

GC 0.49–2.60 mg/m3 

(DDT); No data (DDD, DDE) 

No data 

Air Sample collection on glass 

fiber filter; Soxhlet extraction; cleanup with alumina 

GC/ECD AREAL Method TO-4 

No data No data 

Water Extract using hexane 

followed by acetonitrile 

GC No data 85% (DDT); 

No data (DDD, DDE) 

Water Extract using methylene 

chloride cleanup with Florisil 

GC/ECD EPA Method 608 

0.012 μg/L (DDT); 0.004 μg/L (DDE); 0.011 μg/L (DDD) 

92% (DDT, DDD); 89% (DDE) 

Water Extract at neutral pH with 

methylene chloride 

GC/ECD or GC/ELCD EPA Method 8081B 

0.081 μg/L (DDT); 0.058 μg/L (DDE); 0.050 μg/L (DDD). 

121.1% (4.4'-DDT); 98.0% (4,4'-DDE); 86.8% (4,4'-DDD) 

Water Digest with chromic acid; 

extract with hexane 

GC No data 100% (DDT, DDE); 

No data (DDD) 

Water Extract with methylene 

chloride 

MID GC/MS 0.012 μg/L (DDT); 

0.007 μg/L (DDE); 0.008 μg/L (DDD) 

93–110% (DDT); 73–110% (DDE); 76–110% (DDD) 

Water Immersion solid-phase 

extraction, desorption from solid-phase in GC injector 

GC/ECD 0.30 ng/L (DDT) 

0.20 ng/L (DDE) No data (DDD) 

32.3% (DDT); 103.8% (DDE); 113.6% (DDD) 
Table 7-2. Analytical Methods for Determining DDT, DDE, and DDD in Environmental Samp

Analytical Sample detection 

Sample matrix Sample preparation method limit Percent recovery 

Water Solid-phase extraction, 

eluted with methylene chloride/methanol (80:20) 

GC/Ion trap MS 

0.16 μg/L (DDT) 0.07 μg/L (DDE) No data (DDD) 

24–77% (DDT); 27–51% (DDE); No data (DDD) 

Finished drinking water and groundwater 

Extract with methylene chloride; solvent exchange to methyl tert-butyl ether 

GC/ECD EPA Method 508 

0.060 μg/L (DDT); 0.010 μg/L (DDE); 0.003 μg/L (DDD) 

No data 

Landfill leachate Head-space solid-phase 

microextraction, desorption from solid-phase in GC injector 

GC/ECD 0.1 μg/L 78.5% (DDT,DDE, 

and DDD) 

Soil Extract with 

hexane/acetone; cleanup with Florisil 

GC/ECD AOAC 970.52 

No data No data 

Soil Extract with 

hexane/acetone; cleanup with Florisil 

GC/ECD No data No data 

Soil Extract with hexane-acetone 

or methylene chlorideacetone, cleanup by appropriate method. 

GC/ECD or GC/ELCD EPA Method 8081B 

0.0036 μg/kg (DDT); 0.0025 μg/kg (DDD); 0.0042 μg/kg (DDE) 

121.1% (DDT); 98.0% (DDE); 86.8% (DDD) 

Soil Extraction with methylene 

chloride 

GC/MS EPA Method 

No Data 111–134% (DDT) 

8270D 

Soil Extraction with methylene 

chloride 

MID GC/MS 0.4 μg/kg (DDT); 

0.3 μg/kg (DDD); 0.3 μg/kg (DDE) 

91–109% (DDT); 90–116% (DDD); 93–104% (DDE) 
Table 7-2. Analytical Methods for Determining DDT, DDE, and DDD in Environmental Samp

Analytical Sample detection 

Sample matrix Sample preparation method limit Percent recovery 

Food Extract with acetonitrile into 

petroleum ether; cleanup with Florisil 

GC/ECD No data >80% 

Food Soxhlet extraction using 

redistilled hexane 

on-line SEC-GC 

10–50 μg/kg (DDE); No data (DDT, DDD) 

No data 

Food Extract with n-hexane; 

cleanup with Florisil 

GCECD/NPDMS 0.50–10 μg/kg (DDT, DDE); No data (DDD) 68–95% (DDT, DDE); No data 

(DDD) 

Food Mixed ether extraction; 

cleanup with Florisil 

DC-GC/ECD 0.05–1.5 ng (DDT, 

DDE); No data (DDD) 

No data 

Food Mix sample with dried 

potassium bromide powder 

IR/UV-SP No data No data 

Milk Solid-phase extraction, 

eluted with hexane, cleanup on neutral alumina, eluted with hexane 

GC/ECD 0.12 μg/L (p,p’-DDT) 

0.12 μg/L (o,p’-DDT) 0.07 μg/L (p,p’-DDE) 0.05 μg/L (o,p’-DDE) 0.07 μg/L (p,p’-DDD) 0.12 μg/L (o,p’-DDD) 

100% (p,p’-DDT); 

104% (o,p’-DDT); 93% (p,p’-DDE); 

97% (o,p’-DDE); 106% (p,p’-DDD); 

83% (o,p’-DDD); 

(spiked at 1 μg/L) 

Animal fat Extract with methylene 

chloride/cyclohexane; separation by GPC 

GC/ECD AOAC 984.21 

No data No data 

Plants Extract with 

hexane/methanol/acetone 

SP No data 92–99% (DDT); 

No data (DDD, DDE) 

Plants Extract with methylene 

chloride; organic phase concentrated 

GC/HECD AOAC 985.22 

No data No data 
Table 7-2. Analytical Methods for Determining DDT, DDE, and DDD in Environmental Samp

Sample matrix Sample preparation 

Analytical method 

Sample detection limit Percent recovery 

Fish Soxhlet extraction with 

hexane; cleanup with Florisil 

GC/ECD 10 μg/kg No data 

Fish Extract with petroleum ether; 

cleanup with Florisil 

GC/ECD AOAC 983.21 

No data No data 

AOAC = Association of Official Analytical Chemists; AREAL = Atmospheric Research and Exposure Laboratory of the Environmental P

DC = dual capacity; ECD = electron capture device; ELCD = electrolytic conductivity detector; EPA = Environmental Protection Agencychromatography; GPC = gel permeation chromatography; HECD = halogen-specific electron capture device; IR/UV infrared/ultraviolet; 

dilution; MS = mass spectrometry; NPD = nitrogenphosphorous detection; SEC = size exclusion chromatography; SP = spectro-photom
DDT, DDE, and DDD 296 

7. ANALYTICAL METHODS 

1991; Long et al. 1991b; Lott and Barker 1993). The use of silica acid column chromatography 

purification methods in conjunction with GC/ECD quantitative techniques has provided a detection limit 

for DDT in air of 0.16 ng/m3

 (Bidleman et al. 1978). Spectrometry with an automatic quench correction 

facility was used for the analysis of DDT in rice, maize, and grain plants. Recoveries ranged from 92 to 

99%, and detection limits were not reported (Verma and Pillai 1991a). Even though analytical methods 

exist for detection of DDT in almost all samples, many references did not state detection limits or 

accuracy of the method. 

Alternative approaches are being developed to improve sample recoveries, speed analysis time, or lower 

detection sensitivities in the analysis of DDT, DDE, and DDD. Sample extraction techniques such as 

supercritical fluid extraction (SFE), pressurized liquid extraction (PLE), and microwave extraction 

techniques have been applied to the analysis of DDT, DDE, and DDD to improve extraction of these 

compounds from soil (de Andrea et al. 2001; Fitzpatrick et al. 2000; Glazkov et al. 1999). SFE has been 

shown to improve recoveries of organochlorines from water samples over standard liquid/liquid 

extraction techniques (Glazkov et al. 1999). Solid-phase microextraction techniques can aid in improving 

sample clean-up and detection sensitivities (Aguilar et al. 1999; Brás et al. 2000; de Jager and Andrews 

2000; Röhrig and Meisch 2000). In this technique, fiber-based solid-phase extractants can either directly 

extract DDT, DDE, and DDD through immersion in water or other aqueous samples (e.g., milk), or 

extract DDT, DDE, and DDD from the head-space over a sample as it is heated (Aguilar et al. 1999). 

Once the compounds have been adsorbed into the solid-phase fibers, the compounds are directly desorbed 

from the fibers in a GC injector and then detected using either electron-capture or mass spectrometric 

detection techniques. In addition to the instrumental analytical techniques, immunoassays offer good 

detection sensitivities (<0.1 μg/L) of DDT compounds in complex matrices, although the assay is not 

specific to one specific DDT compound (Abad et al. 1997; Beasley et al. 1998). These newer analytical 

methods show much promise towards improving the analysis DDT, DDE, and DDD in complex 

environmental and biological samples (Röhrig and Meisch 2000). 

7.3 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with the 

Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of DDT, DDE, and DDD is available. Where adequate 

information is not available, ATSDR, in conjunction with the NTP, is required to assure the initiation of a 
 

DDT, DDE, and DDD 297 

7. ANALYTICAL METHODS 

program of research designed to determine the health effects (and techniques for developing methods to 

determine such health effects) of DDT, DDE, and DDD. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will be 

evaluated and prioritized, and a substance-specific research agenda will be proposed. 

7.3.1 Identification of Data Needs 

Methods for Determining Biomarkers of Exposure and Effect. 

Exposure. Methods for the analysis of DDT, DDE, and DDD in blood/plasma, semen, urine, liver, 

kidney, adipose tissue, skin lipids, human milk, and lymph are described in the literature. These methods 

are helpful in estimating the potential health risk of exposed populations. In certain cases, spike 

recoveries were performed in a variety of biological samples to determine the recovery efficiency and 

analytical sensitivity of the method. In some cases, information was unavailable on the detection limit 

and accuracy of a method. Obtaining detection limits and information on the accuracy of a method is 

important to effectively and precisely quantify the parent compound and metabolites in a biological 

system. Once tissue levels of DDT, DDE, and DDD are obtained, there is no acceptable methodology for 

extrapolating backwards from those tissue levels to the amount of exposure. Even in those studies in 

which volunteers were fed measured doses of DDT, such a relationship could not be determined because 

of bioaccumulation of DDT and its metabolites in adipose tissues and because of individual variability. 

Further research would help in understanding the relationship between exposure and DDT levels 

measured in body compartments. 

Effect. No specific biomarkers of effect have been determined. Until these biomarkers are determined, 

methodology needed to identify them cannot be established. 
 

DDT, DDE, and DDD 298 

7. ANALYTICAL METHODS 

Methods for Determining Parent Compounds and Degradation Products in Environmental 

Media. Human exposure is most likely to occur from ingesting food contaminated with small amounts 

of DDT, DDE, or DDD. Analytical methods are available for measuring DDT, DDE, and DDD in air, 

water, soil, fish, waterfowl, plants, and food. Of the techniques available, MID GC/ECD appears to be 

the most sensitive for measuring background levels of DDT, DDE, and DDD in all environmental media. 

No information was available on background levels of DDT, DDE, or DDD at which health effects occur. 

Although analytical techniques are available for measuring DDT, DDE, and DDD in environmental 

media, further information on the accuracy and precision of these techniques is needed. 

7.3.2 Ongoing Studies 

No ongoing studies were located on the analytical methods of DDT, DDE, or DDD. 
DDT, DDE, and DDD 299 

8. REGULATIONS AND ADVISORIES
 

The international, national, and state regulations and advisories regarding DDT, DDE, and DDD in air, 

water, and other media are summarized in Table 8-1. 

ATSDR has derived an acute-duration oral MRL of 5x10-4 mg/kg/day for DDT based on effects on 

perinatal development of the nervous system in neonatal mice with behavioral neurotoxicity manifested in 

the adult animals (Eriksson et al. 1990a, 1990b, 1992, 1993; Johansson et al. 1995, 1996; Talts et al. 

1998). An intermediate-duration oral MRL of 5x10-4 mg/kg/day was derived based on hepatic histologic 

changes in rats (Laug et al. 1950). A chronic oral MRL was not derived because the most sensitive 

noncancer effects were observed at doses higher than doses for the most sensitive acute- and intermediateduration effects. EPA derived an oral reference dose (RfD) of 5x10-4 mg/kg/day for DDT based on liver 

lesions in rats (Laug et al. 1950); an uncertainty factor of 100 was used (IRIS 2001a). 

EPA assigned DDT, DDE, and DDD a weight-of-evidence classification of B2, probable human 

carcinogens (IRIS 2001a, 2001b, 2001c). An oral slope factor of 0.34 (mg/kg/day)-1 was derived for DDT 

based on increased incidence of liver tumors in rats and mice in several studies (Cabral et al. 1982b; Rossi 

et al. 1977; Terracini et al. 1973; Thorpe and Walker 1973; Tomatis and Turusov 1975; Turusov et al. 

1973). Based on the oral data, EPA (IRIS 2001c) derived an inhalation unit risk of 9.7x10-5 (μg/m3

)

-1 for 

DDT. EPA (IRIS 2001c) derived an oral slope factor of 0.34 (mg/kg/day)-1 for DDE based on observed 

lesions as neoplastic liver nodules or hepatocellular tumors in mice and hamster studies (NCI 1978; Rossi 

et al. 1983; Tomatis et al. 1974). An oral slope factor of 0.24 (mg/kg/day)-1 was derived for DDD based 

on increased incidence of liver tumors in male mice (Tomatis et al. 1974b). 

IARC has assigned a weight-of-evidence classification of 2B to DDT, possibly carcinogenic to humans 

(IARC 2002). The Department of Health and Human Services (DHHS) has determined that DDT may 

reasonably be anticipated to be a human carcinogen (NTP 2002). 

The use of DDT and DDD in the United States was canceled in 1972. Exceptions include use by Public 

Health Service officials and other health officials for control of vector-borne disease, use by the U.S. 

Department of Agriculture or military for health quarantine, and use in drugs for controlling body lice 

(EPA 1972). 
DDT, DDE, and DDD 300 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD 

Agency Description	 Information References 

INTERNATIONAL 

Guidelines: 

IARC	 Carcinogenicity classification— Group 2Ba IARC 2002 

DDT 

WHO	 Acceptable daily intake—DDT 0.02 mg/kg-bw WHO 2002 

Drinking water guideline—DDT 2 μg/L WHO 2002 

and metabolites 

NATIONAL 

Regulations and 

guidelines: 

a. Air 

ACGIH TLV–TWA (8-hour)—DDT 1 mg/m3 ACGIH 2001 

NIOSH REL (10-hour TWA)—DDTb,c 0.5 mg/m3 NIOSH 2002 

IDLH—DDT	 500 mg/m3 

OSHA	 PEL (8-hour TWA) for general 1 mg/m3 OSHA 2001b 

industry—DDTc 29CFR1910.1000 

PEL (8-hour TWA) for 1 mg/m3 OSHA 2001c 

construction industry—DDTc 29CFR1926.55 

PEL (8-hour TWA) for shipyard 1 mg/m3 OSHA 2001a 

industry—DDTc 29CFR1915.1000 

USC Hazardous air pollutant—DDE	 USC 2001
 

42 USC 7412
 

b. Water 

EPA	 Designated as a hazardous EPA 2001e
 

substance in accordance with 40CFR116.4
 

Section 311(b)(2)(A) of the
 

Clean Water Act—DDT and
 

DDD
 

Effluent guidelines and EPA 2002e 

standards; designated as a 40CFR401.15 

toxic pollutant pursuant to 

Section 307(a)(1) of the Clean 

Water Act—DDT and 

metabolites 

Reportable quantities for 1 pound EPA 2002b 

hazardous substances 40CFR117.3 

designated pursuant to 

Section 311 of the Clean Water 

Act—DDT and DDE 
DDT, DDE, and DDD 301 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD (continued) 

Agency Description Information References 

NATIONAL (cont.) 

EPA Groundwater monitoring 

DDT and DDD 

DDE 

Suggested 

method 

8080 

8270 

8080 

8270 

PQL (μg/L) 

0.1 

10 

0.05 

10 

EPA 2001a 

40CFR264, 

Appendix IX 

Human health consumption of: 

DDT/DDE 

Water and organism 

Organism only 

DDD 

Water and organism 

Organism only 

5.9x10-4 μg/Ld,e 

5.9x10-4 μg/Ld,e 

8.3x10-4 μg/Ld,e 

8.4x10-4 μg/Ld,e 

EPA 1999a 

Toxic pollutant effluent 

standard; ambient water 

criterion in navigable waters— 

DDT/DDE/DDD 

0.001 μg/L EPA 2002i 

40CFR129.101 

Toxic pollutant effluent 

standard; toxic pollutant subject 

to regulations under provisions 

of this subpart— 

DDT/DDE/DDD 

EPA 2002d 

40CFR129.4(b) 

c. Food 

FDA Action levels (ppm)— 

DDT/DDE/DDD 

Fat of meat (cattle, goats, 

hogs, horses, sheep), fish 

5 

FDA 2002a 

Carrots 3 

Manufactured dairy products 1.25 

Beans (cocoa, whole raw), 

peppermint oil, potatoes, 

soya bean oil (crude), 

spearment oil, sweet potatoes 

1.0 

Artichokes, asparagus, barley 

grain (food, feed), broccoli, 

Brussels’ sprouts, cabbage, 

cauliflower, celery, collards, 

eggs, endives (escarole), 

hay, kale, kohlrabi, lettuce, 

maize grain (food, feed) 

0.5 
DDT, DDE, and DDD 302 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD (continued) 

Agency Description Information References 

NATIONAL (cont.) 

FDA (cont.) Action levels (ppm)— FDA 2002a 

DDT/DDE/DDD 

Milo sorghum grain (food, 0.5 

feed), mushrooms, mustard 

greens, oat grain (food, feed), 

peppermint hay, rice grain 

(food, feed), rye grain (food, 

feed), spearmint hay, 

spinach, Swiss chard, tomato 

pomace (dried, for use in dog 

and cat food), wheat grain 

(food, feed) 

Apricots, avocadoes, beans, 0.2
 

beans (dried), beets (roots,
 

tops), cherries, guavas,
 

mangoes, nectarines, okra,
 

onions (dry bulb), papayas,
 

parsnips (roots, tops),
 

peaches, peanuts, peas,
 

pineapples, plums (fresh
 

prunes), radishes (roots,
 

tops), rutabagas (roots, tops),
 

soya beans (dry), turnips
 

(roots, tops)
 

Apples, blackberries, 0.1
 

blueberries (huckleberries),
 

boysenberries, citrus fruit,
 

maize (fresh sweet plus cob
 

with husk removed),
 

cottonseed, cranberries,
 

cucumbers, currants,
 

dewberries, eggplant,
 

gooseberries, hops (fresh),
 

loganberries, melons, pears,
 

peppers, pumpkins, quinces,
 

raspberries, squash, squash
 

(summer), strawberries,
 

youngberries
 

Grapes, hops (dried), 0.05
 

tomatoes, lettuces
 
DDT, DDE, and DDD 303 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD (continued) 

Agency Description Information References 

NATIONAL (cont.) 

FDA Drug products containing 

certain active ingredients 

offered over-the-counter for 

certain uses; as there are 

inadequate data to establish 

general recognition of the 

safety and effectiveness of 

these ingredients for their 

specified uses 

Pediculicide (lice) drug 

products 

FDA 2002b 

21CFR310.545 

(a)(25) 

USDA Labeling treated seed; 

commonly accepted 

abbreviated chemical names— 

DDT 

USDA 2001 

7CFR201.31a(b) 

d. Other 

ACGIH Carcinogenicity classification— 

DDT 

A3f ACGIH 2001 

EPA DDT 

Carcinogenicity classification 

Inhalation unit risk 

Oral slope factor 

RfD 

B2g 

9.7x10-5 (μg/m3)

-1 

0.34 (mg/kg/day)-1 

5x10-4 mg/kg/day 

IRIS 2001c 

DDE 

Carcinogenicity classification 

Inhalation unit risk 

Oral slope factor 

RfD 

B2g 

No data 

0.34 (mg/kg/day)-1 

No data 

IRIS 2001b 

DDD 

Carcinogenicity classification 

Inhalation unit risk 

Oral slope factor 

RfD 

B2g 

No data 

0.24 (mg/kg/day)-1 

No data 

IRIS 2001a 

Criteria for municipal solid 

waste landfills; listed as a 

hazardous constituent 

DDT and DDD 

DDE 

Suggested 

method 

8080 

8270 

8080 

8270 

PQL (μg/L) 

0.1 

10 

0.05 

10 

EPA 2002g 

40CFR258, 

Appendix II 

Health based limits for 

exclusion of waste-derived 

residues; concentration limit for 

residues—DDT 

1x10-3 mg/kg EPA 2001b 

40CFR266, 

Appendix VII 
DDT, DDE, and DDD 304 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD (continued) 

Agency Description Information References 

NATIONAL (cont.) 

EPA 

NTP
 

Identification and listing of 

hazardous waste; hazardous 

waste number 

DDT 

DDD 

Land disposal restrictions; 

universal treatment standards 

Wastewater standard 

DDT 

DDE 

DDD 

Non-wastewater standard 

DDT/DDE/DDD 

Pesticide effluent limitation 

guidelines and standards— 

DDT/DDE/DDD 

Pesticide residue tolerances on 

agricultural commodities— 

DDT/DDD 

Reportable quantity designated 

as a CERCLA hazardous 

substance under Sections 

311(b)(4) and 307(a) of the 

Clean Water Act and Section 

3001 of RCRA—DDT and DDD 

Reportable quantity designated 

as a CERCLA hazardous 

substance under Section 

307(a) of the Clean Water Act 

and Section 112 of the Clean 

Air Act—DDE 

Standards for the management 

of specific hazardous wastes 

Unit risk 

Risk specific dose 

TSCA significant new use 

subject to reporting—DDT 

Reasonably anticipated to be a 

human carcinogen—DDT 

U061 

U060 

0.023 mg/L2 

0.031 mg/L2 

0.0039 mg/L2 

0.087 mg/kg3 

Class of chlorinated 

organic pesticides 

1 pound 

1 poundh 

9.7x10-5 (μg/m3

)

-1 

1.0x10-1 μg/m3 

Manufacture, import, or 

process 10,000 pounds 

or more per year per 

facility for any use 

EPA 2002f 

40CFR261.33(f) 

EPA 2001c 

40CFR268.48(a) 

EPA 2002c 

40CFR455.20(b) 

EPA 2001d 

40CFR180.3(e)(4) 

EPA 2002a 

40CFR302.4 

EPA 2002a 

40CFR302.4 

EPA 2002h 

40CFR266, 

Appendix V 

EPA 2002j 

40CFR721.2287 

NTP 2002 
DDT, DDE, and DDD 305 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD (continued) 

Agency Description	 Information References 

STATE 

Regulations and 

guidelines: 

a. Air 

Hawaii	 Hazardous air pollutant—DDE BNA 2001 

Idaho DDT UATW 1999d 

Unit risk factor 9.7x10-5 (μg/m3

)

-1 

Emissions level 6.8x10-5 pounds/hour 

AACC (annual average) 1x10-2 μg/m3 

Illinois	 Toxic air contaminant— BNA 2001 

DDT/DDE/DDD 

Kansas	 Hazardous air pollutant—DDE BNA 2001 

Kentucky	 Hazardous air pollutant—DDE BNA 2001 

Maryland	 Toxic air pollutant—DDT BNA 2001 

Minnesota	 Hazardous air pollutant 0.01 tons/year BNA 2001 

threshold; de minimis 

level—DDE 

Nebraska	 Hazardous air pollutant—DDE BNA 2001 

New Hampshire	 Regulated toxic air pollutant; 1 mg/m3 BNA 2001 

occupational exposure level— 

DDT 

New Mexico Toxic air pollutant—DDT BNA 2001 

OEL 1.00 mg/m3 

Emissions 0.0667 pounds/hour 

Rhode Island	 Hazardous air pollutant—DDE BNA 2001 

South Carolina Toxic air emissions—DDE	 High toxicity, may cause BNA 2001 

chronic effects that result 

in death or permanent 

injury after very short 

exposure to small 

quantities 

Vermont	 Hazardous air BNA 2001 

contaminant—DDE 

Washington	 Threshold for hazardous air 0.005 tons/year BNA 2001 

pollutants—DDE 
DDT, DDE, and DDD 306 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD (continued) 

Agency Description Information References 

STATE (cont.) 

b. Water 

Alabama Aquatic life criteria—DDT 

Freshwater 

Acute 

Chronic 

Marine 

Acute 

Chronic 

1.1 μg/L 

0.001 μg/L 

0.13 μg/L 

0.001 μg/L 

UATW 1999b 

Arizona Drinking water guideline—DDT 0.1 μg/L HSDB 2001 

Colorado Human health based standards 

DDT/DDE 

Water supply 

Water and fish 

DDD 

Water supply 

Water and fish 

0.1 μg/L 

5.9x10-4 μg/L 

No data 

8.3x10-4 μg/L 

UATW 1999a 

Florida Drinking water guideline— 

DDT/DDD/DDE 

0.1 μg/L HSDB 2001 

Hawaii Water quality criteria—DDT 

and metabolite DDD 

Freshwater 

Acute 

Chronic 

Saltwater 

Acute 

Chronic 

Fish consumption 

1.1 μg/L 

0.001 μg/L 

0.013 μg/L 

0.001 μg/L 

8x10-6 μg/L 

UATW 1999c 

Illinois Drinking water standard—DDT 50 μg/L HSDB 2001 

Maine Drinking water guideline—DDT 0.83 μg/L HSDB 2001 

Minnesota Drinking water guideline— 

DDT/DDD/DDE 

1 μg/L HSDB 2001 

New Hampshire Drinking water guideline—DDT 0.1 μg/L HSDB 2001 

c. Food 

Oklahoma Alert level in fish tissues 

Concern levels in fish tissue 

Water column criteria to protect 

for the consumption of fish 

flesh 

5.0 mg/kg 

2.5 mg/kg 

5.9x10-3 μg/L 

FSTRAC 1999 
DDT, DDE, and DDD 307 

8. REGULATIONS AND ADVISORIES 

Table 8-1. Regulations and Guidelines Applicable to DDT/DDE/DDD (continued) 

Agency Description Information References 

STATE (cont.) 

Alabama 

Alaska 

Arizona 

California 

Georgia 

Massachusetts 

Michigan 

New York 

Oklahoma 

Oregon 

Texas 

Washington 

Fish and Wildlife Advisory for 

DDT/DDE/DDD 

Fish 

Fish 

Fish and turtles 

Fish 

Fish 

Fish 

Fish 

Fish and waterfowl 

Fish 

Fish 

Fish 

Fish 

EPA 1999b 

d. Other 

California Chemicals known to the state BNA 2001 

to cause cancer— 

DDT/DDD/DDE 

Hawaii Restricted use pesticide— 

DDT/DDE 

All concentrations BNA 2001 

a

Group 2B: possibly carcinogenic to humans

b

Potential occupational carcinogen 

c

Skin designation: refers to the potential significant contribution to the overall exposure by the cutaneous route, 

including mucous membranes and the eyes, either by contact with vapors or, of probable greater significance, by 

direct skin contact with the substance 

d

This criterion has been revised to reflect The Environmental Protection Agency’s q1* or RfD, as contained in the 

Integrated Risk Information System (IRIS) as of April 8, 1998. The fish tissue bioconcentration factor (BCF) from 

the 1980 Ambient Water Quality Criteria document was retained in each case. 

e

This criterion is based on carcinogenicity of 10-6 risk. Alternate risk levels may be obtained by moving the decimal 

point (e.g., for a risk level of 10–5, move the decimal point in the recommended criterion one place to the right). f

A3: confirmed animal carcinogen with unknown relevance to humans 

g

B2; probable human carcinogen

h

The 1-pound reportable quantity is a Section 102 CERCLA statutory reportable quantity 

AACC = acceptable ambient concentrations for carcinogens; ACGIH = American Conference of Governmental 

Industrial Hygienists; BNA = Bureau of National Affairs; CERCLA = Comprehensive Environmental Response, 

Compensation, and Liability Act; CFR = Code of Federal Regulations; EPA = Environmental Protection Agency; 

FDA = Food and Drug Administration; FSTRAC = Federal-State Toxicology and Risk Analysis Committee; 

HSDB = Hazardous Substances Data Bank; IARC = International Agency for Research on Cancer; 

IDLH = immediately dangerous to life and health; IRIS = Integrated Risk Information System; NIOSH = National 

Institute of Occupational Safety and Health; NTP = National Toxicology Program; OEL = occupational exposure 

limit; OSHA = Occupational Safety and Health Administration; PEL = permissible exposure limit; PQL = practical 

quantitation limits; RCRA = Resource Conservation Recovery Act; REL = recommended exposure limit; RfD = oral 

reference dose; TLV = threshold limit value; TSCA = Toxic Substances Control Act; TWA = time-weighted 

average; UATW = Unified Air Toxics Website; USC = United States Code; USDA = U.S. Department of 

Agriculture; WHO = World Health Organization 

_______________ 

DDT, DDE, and DDD 309 

9. REFERENCES
 

*Abad A, Manclus JJ, Mojarrad F, et al. 1997. Hapten synthesis and production of monoclonal 

antibodies to DDT and related compounds. J Agric Food Chem 45:3694-3702. 

*Abou-Arab AA. 1999. Behavior of pesticides in tomatoes during commercial and home preparation. 

Food Chem 65:509-514. 

Abraham K, Päpke O, Wahn U, et al. 2000. POP accumulation in infants during breast-feeding. 

Organohalogen Compounds 48:25-26. 

ACGIH. 1998. Threshold limit values for chemical substances and physical agents. 6th ed. American 

Conference of Governmental Industrial Hygienists. Cincinnati, OH 

*ACGIH. 2001. Threshold limit values for chemical substances and physical agents and biological 

exposure indices. American Conference of Governmental Industrial Hygienists. Cincinnati, OH. 

*Adami H-O, Lipworth L, Titus-Ernstoff L, et al. 1995. Organochlorine compounds and estrogenrelated cancers in women. Cancer Causes Control 6:551-566. 

*Adamson R, Sieber S. 1979. The use of nonhuman primates for chemical carcinogenesis studies. 

Ecotoxicol Environ Qual 2:257-296. 

*Adamson R, Sieber S. 1983. Chemical carcinogenesis studies in nonhuman primates. Basic Life Sci 

24:129-156. 

*Addison RF, Stobo WT. 2001. Trends in organochlorine residue concentrations and burdens in grey 

seals (Halichoerus grypus) from Sable Is., NS, Canada, between 1974-1994. Environ Pollut 112(3):505­

513. 

*Adenuga G, Bababunmi E, Hendrickse R. 1992. Depression of the CA2+-ATPase activity of the rat 

liver endoplasmic reticulum by the liver tumour promoters 1,1,1-tricloro-2,2-bis(p-chlorophenyl)ethane 

and phenobarbital. Toxicology 71:1-6. 

*Adeshina F, Todd E. 1990. Organochlorine compounds in human adipose tissue from north Texas. J 

Toxicol Environ Health 29:147-156. 

Adeshina F, Todd EL. 1991. Exposure assessment of chlorinated pesticides in the environment. J 

Environ Sci Health A26(1):139-154. 

*Adinolfi M. 1985. The development of the human blood-CSF-brain barrier. Dev Med Child Neurol 

27:532-537. 

*Adlercreutz H. 1995. Phytoestrogens: Epidemiology and a possible role in cancer protection. Environ 

Health Perspect 103(Suppl 7):103-112. 

*Cited in text 
DDT, DDE, and DDD 310 

9. REFERENCES 

*Agarwal HC, Pillai MKK, Yadav DV, et al. 1976. Residues of DDT and its metabolites in human blood 

samples in Delhi, India. Bull World Health Organ 54:349-351. 

*Agarwal N, Sanyal S, Khuller G, et al. 1978. Effect of acute administration of dichlorodiphenyl 

trichloroethane on certain enzymes of Rhesus monkey. Indian J Med Res 68:1001-1006. 

*Agthe C, Garcia H, Shubik P, et al. 1970. Study of the potential carcinogenicity of DDT in the Syrian 

golden hamster (34740). Proc Soc Exp Biol Med 134:113-116. 

*Aguilar C, Penalver A, Pocurull E, et al. 1999. Optimization of solid-phase microextraction conditions 

using a response surface methodology to determine organochlorine pesticides in water by gas 

chromatography and electron-capture detection. J Chromatogr A 844:425-432 

*Ahlborg U, Lipworth L, Titus-Ernstoff L, et al. 1995. Organochlorine compounds in relation to breast 

cancer, endometrial cancer, and endometriosis: An assessment of the biological and epidemiological 

evidence. Crit Rev Toxicol 25(6):463-531. 

Ahmad N, Bugueno G, Guo L, et al. 1999. Determination of organochlorine and organophosphate 

pesticide residues in fruits, vegetables and sediments. J Environ Sci Health B 34(5):829-848. 

*Ahmed F, Hart R, Lewis N. 1977. Pesticide induced DNA damage and its repair in cultured human 

cells. Mutat Res 42:161-174. 

Ahmed SA. 2000. The immune system as a potential target for environmental estrogens (endocrine 

disruptors): a new emerging field. Toxicology 150:191-206. 

*Aigner E, Leone A, Falconer R. 1998. Concentrations and enantiomeric ratios of organochlorine ­

pesticides in soils from the U.S. corbel. Environ Sci Technol 32:1162-1168 

*Aislabie JM, Richards NK, Boul HL. 1997. Microbial degradation of DDT and its residues—a review. 

N Z J Agric Res 40:269-282. 

*Akiyama Y, Yoshioka N, Tsuji M. 2000. Solid-phase extraction for cleanup of pesticide residues 

suspected as endocrine disruptors in foods. J Health Sci 46(1):49-55. 

Aktay G, Deliorman D, Ergun E, et al. 2000. Hepatoprotective effects of Turkish folk remedies on 

experimental liver injury. J Ethnopharmacol 73:121-129. 

Albone ES, Eglinton G, Evans NC, et al. 1972. Fate of DDT in Severn estuary sediments. Environ Sci 

Technol 6:914-919. 

*Alexander M. 1995. How toxic are toxic chemicals in soil? Environ Sci Technol 28(11):2713-2717. 

*Alexander M. 1997. Sequestration and bioavailability of organic compounds in soil. In: Linz DG, 

Nakles DU, eds. Environmentally acceptable endpoints in soil. Annapolis, MD: American Academy of 

Environmental Engineers, 43-136. 

Al-jamal JH, Dubin NH. 2000. The effect of raloxifine on the uterine weight response in immature mice 

exposed to 17β-estrodiol, 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane, and methoxychlor. Am J Obstet 

Gynecol 182:1099-1102. 
DDT, DDE, and DDD 311 

9. REFERENCES 

*Allen-Gil SM, Gubala CP, Wilson R, et al. 1997. Organochlorine pesticides and polychlorinated 

biphenyls (PCBs) in sediments and biota from four US arctic lakes. Arch Environ Contam Toxicol 

33:378-387. 

*Altman PK, Dittmer DS, eds. 1974. Biological handbooks: Biology data book. Vol. III, 2nd ed. 

Bethesda, MD: Federation of American Societies for Experimental Biology, 1987-2008. 

Altschuh J, Bruggermann R, Santl H, et al. 1999. Henry's law constants for a diverse set of organic 

chemicals Experimental determination and comparison of estimation methods. Chemosphere 

39(11):1871-1887. 

Altshul L, Korrick S, Tolbert P, et al. 1999. Cord blood levels of PCBs, p,p'-DDE and HCB in infants 

born in communities adjacent to a PCB-contaminated hazardous waste site. Organohalogen Compounds 

44:67-70. 

*Amacher D, Zelljadt I. 1984. Mutagenic activity of some clastogenic chemicals at the hypoxanthine 

guanine phosphoribosyl transferase locus of Chinese hamster ovary cells. Mutat Res 136:137-145. 

Ames A, Van Vleet E. 1996. Organochlorine residues in the Florida Manatee, Trichechus manatus 

latirostris. Mar Pollut Bull 32(4):374-377. 

Andersen A, Kasperlik-Zaluska AA, Warren DJ. 1999. Determination of mitotane (o,p'-DDD) and its 

metabolites o,p'-DDA and o,p'-DDE in plasma by high performance liquid chromatography. Ther Drug 

Monit 21:355-359. 

*Andersen ME, Kirshnan K. 1994. Relating in vitro to in vivo exposures with physiologically based 

tissue dosimetry and tissue response models. In: Salem H, ed. Animal test alternatives: Refinement, 

reduction, replacement. New York, NY: Marcel Dekker, Inc., 9-25. 

*Andersen ME, Clewell HJ, Gargas ML, et al. 1987. Physiologically based pharmacokinetics and the 

risk assessment process for methylene chloride. Toxicol Appl Pharmacol 87:185-205. 

*Anderson H. 1985. Utilization of adipose tissue biopsy in characterizing human halogenated 

hydrocarbon exposure. Environ Health Perspect 60:127-131. 

Anderson HA, Wolff MS. 2000. Environmental contaminants in human milk. J Expo Anal Environ 

Epidemiol 10:755-760. 

*Anderson HA, Falk C, Hanrahan L, et al. 1998. Profiles of Great Lakes critical pollutants: A sentinel 

analysis of human blood and urine. Environ Health Perspect 106(5):279-289. 

*Anderson K, Johnson E. 2001. Bioavailable organo chlorine pesticides in a semi-arid region of eastern 

Oregon, USA, as determined by gas chromatography with electron-capture detection. J AOAC Int 

84:1371-1382 

Ando M. 1978. Transfer of 2,4,5,2',4',5'-hexachlorobiphenyl and 2,2-bis(p-chlorophenyl), 

1,1,1-trichloroethane (p,p'-DDT) from maternal to newborn and suckling rats. Arch Toxicol 41:179-186. 

*Ando M. 1979. Effects of cadmium on the metabolism of p,p'-DDT [2,2-bis(p-chlorophenyl)-1,1,1­

trichloroethane] in rats. Environ Res 19:70-78. 
DDT, DDE, and DDD 312 

9. REFERENCES 

Ando M. 1982. Dose-dependent excretion of DDE (1,1-dichloro-2,2-bis(p-chlorophenyl)ethylene) in 

rats. Arch Toxicol 49:139-147. 

*Andrea M, Tomita R, Luchini L, et al. 1994. Laboratory studies on volatilization and mineralization of 

14C-p,p’-DDT in soil, release of bound residues and dissipation from solid surfaces. J Environ Sci Health 

B29(1):133-139. 

Andrews SA, Colvin HJ. 1989. Verification of intoximeter 3000 breath alcohol concentration by 

magnesium perchlorate tube method in long-term field program. J Anal Toxicol 13:113-116. 

*Anthony RG, Miles AK, Estes JA, et al. 1999. Productivity, diets, and environmental contaminants in 

nesting bald eagles from the Aleutian Archipelago. Environ Toxicol Chem 18:2054-2062 

Arbuckle TE, Sever LE. 1998. Pesticide exposures and fetal death: A review of the epidemiologic 

literature. Crit Rev Toxicol 28(3):229-270. 

*Archibeque-Engle S, Tessari J, Winn D, et al. 1997. Comparison of organochlorine pesticide and 

polychlorinated biphenyl residues in human breast adipose tissue and serum. J Toxicol Environ Health 

52:285-293. 

*Ardies C, Dees C. 1998. Xenoestrogens significantly enhance risk for breast cancer during growth and 

adolescence. Med Hypotheses 50:457-464. 

*Arisoy M. 1998. Biodegradation of chlorinated organic compounds by white-rot fungi. Bull Environ 

Contam Toxicol 60:872-876. 

*Aronson KJ, Miller AB, Woolcott CG, et al. 2000. Breast adipose tissue concentrations of 

polychlorinated biphenyls and other organochlorines and breast cancer risk. Cancer Epidemiol 

Biomarkers Prev 9:55-63. 

*Arthur R, Cain J, Barrentine B. 1977. DDT residues in air in the Mississippi delta, 1975. Pestic Monit 

J 10:168. 

*Aschengrau A, Coogan P, Quinn M, et al. 1998. Occupational exposure to estrogenic chemicals and the 

occurrence of breast cancer: An exploratory analysis. Am J Ind Med 34:6-14. 

*Ashari E, Gabliks J. 1973. DDT and immunological responses: II. Altered histamine levels and 

anaphylactic shock in guinea pigs. Arch Environ Health 26:309-312. 

*Asplund L, Svensson B-G, Nilsson A, et al. 1994. Polychlorinated biphenyls, 1,1,1-trichloro-2,2-bis(pchlorophenyl)ethane (p,p'-DDT) and 1,1-dichloro-2,2-bis(p-chlorophenyl)-ethylene (p,p'-DDE) in human 

plasma related to fish consumption. Arch Environ Health 49(6):477-486. 

Astroff AB, Young AD. 1998. The relationship between maternal and fetal effects following maternal 

organophosphate exposure during gestation in the rat. Toxicol Ind Health 14(6):869-889. 

*Atlas E, Foster R, Glam C. 1982. Air-sea exchange of high molecular weight organic pollutants: 

Laboratory studies. Environ Sci Technol 16:283-286. 
DDT, DDE, and DDD 313 

9. REFERENCES 

*ATSDR. 1989. Decision guide for identifying substance-specific data needs related to toxicological 

profiles. Agency for Toxic Substances and Disease Registry, U.S. Public Health Service. Federal 

Register vol 54 no. 174. 

*Attaran A, Roberts DR, Curtis CF, et al. 2000. Balancing risks on the backs of the poor. Nature Med 

6:729-731 

Attia AM, Awney HA, Habib SL, et al. 1999. O6

-alkylguanine-DNA-alkyltransferase and cytochrome P450 activities in mice treated with lindane, DDT, and endrin. Environ Nutr Interact 3:155-165. 

*Atuma SS, Aune M. 1999. Method for the determination of PCB congeners and chlorinated pesticides 

in human blood serum. Bull Environ Contam Toxicol 62:8-15. 

*Austin H, Keil J, Cole P. 1989. A prospective follow-up study of cancer mortality in relation to serum 

DDT. Am J Public Health 79:43-46. 

*Azevedo JA, Hunt EG, Woods LA. 1965. Physiological effects of DDT on pheasants. Calif Fish Game 

51(4):276-293. 

Backstrom J, Hansson E, Ullberg S. 1965. Distribution of C14-DDT and 14C-dieldrin in pregnant mice 

determined by whole-body autoradiography. Toxicol Appl Pharmacol 7:90-96. 

*Bagga D, Anders KH, Wang H-J, et al. 2000. Organochlorine pesticide content of breast adipose tissue 

from women with breast cancer and control subjects. J Natl Cancer Inst 92(9):750-753. 

*Baker RD, Applegate HG. 1970. Effect of temperature and ultraviolet radiation on the persistence of 

methyl parathion and DDT in soils. Agron J 62:509-512. 

*Bakke JE, Bergman AL, Larsen GL. 1982. Metabolism of 2,4',5-trichlorobiphenyl by the mercapturic 

acid pathway. Science 217(13):645-647. 

*Balaguer P, Francois F, Comunale F, et al. 1999. Reporter cell lines to study the estrogenic effects of 

xenoestrogens. Sci Total Environ 233:47-56. 

*Banerjee B. 1987a. Sub-chronic effect of DDT on humoral immune response to a thymus-independent 

antigen (bacterial lipopolysaccharide) in mice. Bull Environ Contam Toxicol 39:822-826. 

*Banerjee B. 1987b. Effects of sub-chronic DDT exposure on humoral and cell-mediated immune 

responses in albino rats. Bull Environ Contam Toxicol 39:827-834. 

Banerjee BD. 1999. The influence of various on immune toxicity assessment of pesticide chemicals. 

Toxicol Lett 107:21-31. 

*Banerjee B, Koner B, Pashsa S. 1997a. Influence of DDT exposure on susceptibility to human leprosy 

bacilli in mice. Int J Lepr 65:97-99. 

*Banerjee B, Koner B, Ray A. 1997b. Influence of stress on DDT-induced humoral immune 

responsiveness in mice. Environ Res 74:43-47. 

*Banerjee B, Pasha ST, Koner BC. 1996. Hapten synthesis and production of rabbit antibodies with 

reactivity to DDT and its metabolites for the development of an immunoassay. Med Sci Res 24:553-555. 
DDT, DDE, and DDD 314 

9. REFERENCES 

*Banerjee B, Ramachandran M, Hussain Q. 1986. Sub-chronic effect of DDT on humoral immune 

response in mice. Bull Environ Contam Toxicol 37:433-440. 

Banerjee BD, Ray A, Pasha ST. 1996. A comparative evaluation of immunotoxicity of DDT and its 

metabolites in rats. Indian J Exp Biol 34:517-522. 

*Banerjee B, Saha S, Mohapatra T, et al. 1995. Influence of dietary protein on DDT-induced immune 

responsiveness in rats. Indian J Exp Biol 33:739-744. 

*Barahona MV, Sánchez-Fortún S. 1996. Comparative sensitivity of three age classes of Artemia salina 

larvae to several phenolic compounds. Bull Environ Contam Toxicol 56:271-278. 

*Bard SM. 1999. Global transport of anthropogenic contaminants and the consequences for the arctic 

marine ecosystem. Mar Pollut Bull 38(5):356-379. 

*Baris D, Zahm S, Cantor K, et al. 1998. Agricultural use of DDT and risk of non-Hodgkin’s 

lymphoma: Pooled analysis of three case control studies in the United States. Bethesda, MD: National 

Cancer Institute. 

*Barnes DG, Dourson M. 1988. Reference dose (RfD): Description and use in health risk assessments. 

Regul Toxicol Pharmacol 8:471-486. 

Barquet A, Morgade C, Pfaffenberger C. 1981. Determination of organochlorine pesticides and 

metabolites in drinking water, human blood serum, and adipose tissue. J Toxicol Environ Health 7:469­

479. 

Barrie L, Gregor D, Hargrave B, et al. 1992. Arctic contaminants: Sources, occurrence and pathways. 

Sci Total Environ 122:1-74. 

*Bathe R, Ullmann L, Sachsse K, et al. 1976. Relationship between toxicity to fish and to mammals: A 

comparative study under defined laboratory conditions. Proc Eur Soc Toxicol 17:351-355. 

*Bayarri S, Conchello P, Arino A, et al. 1994. DDT, DDT metabolites, and other organochlorines as 

affected by thermal processing in three commercial cuts of lamb. Bull Environ Contam Toxicol 52:554­

559. 

*Beard J, Marshall S, Jong K, et al. 2000. 1,1,1-trichloro-2,2-bis(p-chlorophenyl)-ethane (DDT) and 

reduced bone mineral density. Arch Environ Health 55(3):177-180. 

*Beasley HL, Phongkham T, Daunt MH, et al. 1998. Development of a panel of immunoassays for 

monitoring DDT, its metabolites, and analogues in food and environmental matrices. J Agric Food Chem 

46:3339-3352. 

*Becker PR, Mackey EA, Demiralp R, et al. 1997. Concentrations of chlorinated hydrocarbons and 

trace elements in marine mammal tissues archived in the U.S. National Biomonitoring Specimen Bank. 

Chemosphere 34(9/10):2067-2098. 

Beckmen KB, Ylitalo GM, Towell RG, et al. 1999. Factors affecting organochlorine contaminant 

concentrations in milk and blood of northern fur seal (Callorhinus ursinus) dams and pups for St. George 

Island, Alaska. Sci Total Environ 231:183-200. 
DDT, DDE, and DDD 315 

9. REFERENCES 

*Behzadi H, Lalancette RA. 1991. Extending of the use of USEPA method 1625 for the analysis of 

4,4'-DDT, 4,4'-DDD, and 4,4'-DDE. Microchem J 44:122-129. 

*Bend JR, Miller DS, Kinter WB, et al. 1977. DDE-induced microsomal mixed-function oxidases in the 

puffin (Fratercula arctica). Biochem Pharmacol 26:1000-1001. 

*Ben-Dyke R, Sanderson D, Noakes D. 1970. Acute toxicity data for pesticides (1970). World Rev 

Pestic Cont 9:119-127. 

*Bennison B, Mostofi F. 1950. Observations on inbred mice exposed to DDT. J Natl Cancer Inst 

10:989-992. 

*Berganstal D, Hertz R, Lipsett M, et al. 1960. Chemotherapy of adrenocortical cancer with o,p'-DDD. 

Ann Intern Med 53(4):672-682. 

*Berger G. 1994. Epidemiology of endometriosis. In: Modern surgical management of endometriosis. 

New York, NY: Springer-Verlag. 

Bergman A, Brouwer A, Ghosh M, et al. 1997. Risk of endocrine contaminants (RENCO)- aims and a 

summary of initial results. Organohalogen Compounds 34:396-401. 

*Bergman A, Norstrom R, Haraguchi K, et al. 1994. PCB and DDE methyl sulfones in mammals from 

Canada and Sweden. Environ Toxicol Chem 13:121-128. 

*Berkowitz G, Lapinski R, Wolff M. 1996. The role of DDE and polychlorinated biphenyl levels in 

preterm birth. Arch Environ Contam Toxicol 30:139-141. 

*Bernard RF. 1963. Studies on the effects of DDT on birds. Volume 2. Baker R, Ball R, Cantlon J, et 

al. ed. Lansing, Michigan: Michigan State University. 

*Bernard R, Gaertner R. 1964. Some effects of DDT on reproduction in mice. J Mammal 45(2):272­

276.
 

Bevenue A. 1976. The "bioconcentration" aspects of DDT in the environment. Residue Rev 61:37-112.
 

*Bevenue A, Hylin J, Kawano Y, et al. 1972. Organochlorine pesticide residues in water, sediment,
 

algae, and fish: Hawaii, 1970-71. Pestic Monit J 6(1):56-64. 

*Bidleman T. 1988. Atmospheric processes: Wet and dry deposition of organic compounds are 

controlled by their vapor-particle partitioning. Environ Sci Technol 22(4):361-367. 

*Bidleman T, Christensen E, Billings W, et al. 1981. Atmospheric transport of organochlorines in the 

North Atlantic gyre. J Mar Res 39:443-464. 

*Bidleman TF, Cotham WE, Addison RF, et al. 1992. Organic contaminants in the northwest Atlantic 

atmosphere at Sable Island, Nova Scotia 1988-1989. Chemosphere 24:1389-1412. 

*Bidleman T, Matthews J, Olney C, et al. 1978. Separation of polychlorinated biphenyls, chlordane, and 

p,p'-DDD from toxaphene by silicic acid column chromatography. J Assoc Off Anal Chem 61(4):820­

828. 
DDT, DDE, and DDD 316 

9. REFERENCES 

*Bidleman T, Walla M, Roura R, et al. 1993. Organochlorine pesticides in the atmosphere of the 

southern ocean and Antarctica, January-March, 1990. Mar Pollut Bull 26(5):258-262. 

*Bishop CA, Brooks RJ, Carey JH, et al. 1991. The case for a cause-effect linkage between 

environmental contamination and development in eggs of the common snapping turtle (chelydra S. 

Serpentina) from Ontario, Canada. J Toxicol Environ Health 33:521-547. 

Bishop CA, Koster MD, Chek AA, et al. 1995. Chlorinated hydrocarbons and mercury in sediments, 

red-winged blackbirds (Agelaius phoeniceus) and tree swallows (Tachycineta bicolor) from wetlands in 

the Great Lakes-St. Lawrence river basin. Environ Toxicol Chem 14(3):491-501. 

Bishop CA, Ng P, Norstrom RJ, et al. 1996. Temporal and geographic variation of organochlorine 

residues in eggs of the common snapping turtle (Chelydra serpentina serpentina) (1981-1991) and 

comparisons to trends in the Herring Gull (Larus argentatus) in the Great Lakes basin in Ontario, Canada. 

Arch Environ Contam Toxicol 31:512-524. 

*Bitman J, Cecil H. 1970. Estrogenic activity of DDT analogs and polychlorinated biphenyls. J Agric 

Food Chem 18(6):1108-1112. 

*Bitman J, Cecil HC, Harris SJ, et al. 1968. Estrogenic activity of o,p'-DDT in the mammalian uterus 

and avian oviduct. Science 162:371-372. 

*Bitman J, Cecil HC, Harris SJ, et al. 1971. Comparison of DDT effect on pentobarbital metabolism in 

rats and quail. J Agric Food Chem 19(2):333-338. 

Bitman J, Cecil HC, Harris SJ. 1978. Estrogenic activity of o,p'-metabolites and related compounds. J 

Agric Food Chem 26:149-151. 

Bjorseth A, Lunde G, Dybing E. 1977. Residues of persistent chlorinated hydrocarbons in human tissues 

as studied by neutron activation analysis and gas chromatography. Bull Environ Contam Toxicol 

18(5):581-587. 

Blair A, Grauman D, Lubin J, et al. 1983. Lung cancer and other causes of death among licensed 

pesticide applicators. J Natl Cancer Inst 71(1):31-37. 

*Blus LJ. 1978. Short-tailed shrews: Toxicity and residue relationships of DDT, dieldrin, and endrin. 

Arch Environ Contam Toxicol 7:83-98. 

*Blus LJ, Henny CJ, Stafford CJ, et al. 1987. Persistence of DDT and metabolites in wildlife from 

Washington state orchards. Arch Environ Contam Toxicol 16:467-476. 

*BNA. 2001. Environment and Safety Library on the Web. States and Territories. Washington, DC: 

Bureau of National Affairs, Inc. http//www.esweb.bna.com/. February 9, 2001. 

*Bohannon AD, Cooper GS, Wolff MS, et al. 2000. Exposure to 1,1-dichloro-2,2-bis(pchlorophenyl)ethylene (DDT) in relation to bone mineral density and rate of bone loss in menopausal 

women. Arch Environ Health 55(6):386-391. 

*Boileau SR, Simard JJ, Wouters JR, et al. 1985. Enhanced excretion of DDT and metabolites in rats 

treated with trans,trans-muconate. Pestic Biochem Physiol 24:336-342. 
DDT, DDE, and DDD 317 

9. REFERENCES 

Bopp R, Simpson H, Olsen C, et al. 1982. Chlorinated hydrocarbons and radionuclide chronologies in 

sediments of the Hudson River and Estuary, New York. Environ Sci Technol 16:666-676. 

Borgmann U, Whittle DM. 1992a. Bioenergetics and PCB, DDE, and mercury dynamics in Lake 

Ontario lake trout (Salvelinus namaycush): A model based on surveillance data. Can J Fish Aquat Sci 

49(6):1086-1096. 

Borgmann U, Whittle DM. 1992b. DDE, PCB, and mercury concentration trends in Lake Ontario 

rainbow smelt (Osmerus mordax) and slimy sculpin (Cottus cognatus): 1977 to 1988. J Great Lakes Res 

18(2):298-308. 

Boul H. 1996. Effect of soil moisture on the fate of radiolabelled DDT and DDE in vitro. Chemosphere 

32(5):855-866. 

*Boul HL, Garnham ML, Hucker D, et al. 1994. Influence of agricultural practices on the levels of DDT 

and its residues in soil. Environ Sci Technol 28(8):61-66. 

Bouwman H, Schutte CH. 1993. Effect of sibship on DDT residue levels in human serum from a malaria 

endemic area in northern KwaZulu. Bull Environ Contam Toxicol 50:300-307. 

Bouwman H, Becker PJ, Cooppan RM, et al. 1992. Transfer of DDT used in malaria control to infants 

via breast milk. Bull World Health Organ 70(2):241-250. 

*Bouwman H, Reinecke AJ, Cooppan RM, et al. 1990. Factors affecting levels of DDT and metabolites 

in human breast milk from KwaZulu. J Toxicol Environ Health 31:93-115. 

Bowerman W, Best D, Grubb T, et al. 1998. Trends of contaminants and effects in Bald Eagles of the 

Great Lakes basin. Environ Monit Assess 53:197-212. 

*Boykins EA. 1967. The effects of DDT-contaminated earthworms in the diet of birds. Bioscience 

January:37-39. 

*Boylan J, Egle J, Guzelian P. 1978. Cholestryramine: Use as a new therapeutic approach for 

chlordecone (kepone) poisoning. Science 199:893-895. 

*Bradlow H, Davis D, Lin G, et al. 1995. Effects of pesticides on the ratio of 16α/2-hyroxyestrone: A 

biological marker of breast cancer risk. Environ Health Perspect 103(Suppl 7):147-150. 

*Bradlow HL, Davis D, Sepkovic DW, et al. 1997. Role of the estrogen receptor in the action of 

organochlorine pesticides on estrogen metabolism in human breast cancer cell lines. Sci Total Environ 

208:9-14. 

*Bradt P, Herrenkohl R. 1976. DDT in human milk: What determines the levels? Sci Total Environ 

6:161-163. 

*Braham HW, Neal CM. 1974. The effects of DDT on energetics of the short-tailed shrew Blarina 

brevicauda. Bull Environ Contam Toxicol 12(1):32-37. 

*Brandt I, Jonsson C-J, Lund B-O. 1992. Comparative studies on adrenocorticolytic DDT-metabolites. 

Comparative Studies. Ambio 21:602-605. 
DDT, DDE, and DDD 318 

9. REFERENCES 

*Brás I, Santos L, Alves A. 2000. Monitoring organochlorine pesticides from landfill leachates by gas 

chromatography-electron-capture detection after solid-phase microextraction. J Chromatogr A 891:305­

311. 

*Brevik E, Grande M, Knutzen J, et al. 1996. DDT contamination of fish and sediments from Lake 

Orsjoen, Southern Norway: Comparison of data from 1975 and 1994. Chemosphere 33(11):2189-2200. 

Brien SE, Heaton JPW, Racz WJ, et al. 2000. Effects of an environmental anti-androgen on erectile 

function of an animal penile erection model. J Urol 163:1315-1321. 

Brilliant LB, Van Amburg G, Isbister J, et al. 1978. Breast-milk monitoring to measure Michigan's 

contamination with polybrominated biphenyls. Lancet September 23:643-646. 

*Britton WM. 1975. Influence of high levels of DDT in the diet on liver microsomal estrogen 

metabolism in the laying hen. Bull Environ Contam Toxicol 13(6):698-702. 

Brock JW, Melnyk LJ, Caudill SP, et al. 1998. Serum levels of several organochlorine pesticides in 

farmers correspond with dietary exposure and local use history. In: Colborn T, vom Saal F, Short P, eds. 

Advances in modern environmental toxicology. Princeton, NJ: Princeton Scientific Publishing Co., Inc., 

323-340. 

*Brock J, Melnyk L, Caudill S, et al. 1998. Serum levels of several organochlorine pesticides in farmers 

correspond with dietary exposure and local use history. Toxicol Ind Health 14:275-289. 

Brouwer A, Longnecker MP, Birnbaum LS, et al. 1999. Characterization of potential endocrine-related 

health effects at low-dose levels of exposure to PCBs. Environ Health Perspect Suppl 107(4):639-649. 

*Brown DP. 1992. Mortality of workers employed at organochlorine pesticide manufacturing plantsAn update. Scand J Work Environ Health 18:155-161. 

Brown LR. 1997. Concentrations of chlorinated organic compounds in biota and bed sediment in 

streams of the San Joaquin Valley, California. Arch Environ Contam Toxicol 33:357-368. 

*Brown L, Burmeister L, Everett G, et al. 1993. Pesticide exposures and multiple myeloma in Iowa men. 

Cancer Causes Control 4:153-156. 

Brown N, Lamartiniere C. 1995. Xenoestrogens alter mammary gland differentiation and cell 

proliferation in the rat. Environ Health Perspect 103(7-8):708-713. 

Buchert H, Bihler S, Schott P, et al. 1981. Organochlorine pollutant analysis of contaminated and 

uncontaminated lake sediments by high resolution gas chromatography. Chemosphere 10:945-946. 

*Buck N, Estesen B, Ware G. 1983. DDT moratorium in Arizona: Residues in soil and alfalfa after 12 

years. Bull Environ Contam Toxicol 31:66-72. 

Buckley T, Liddle J, Ashley D, et al. 1997. Environmental and biomarker measurements in nine homes 

in the lower Rio Grande Valley: Multimedia results for pesticides, metals, PAHs, and VOCs. Environ Int 

23(5):705-732. 

*Budavari S, O'Neil M, Smith A, et al. 1996. The Merck Index. 12th ed. Whitehouse Station, NJ: 

Merck and Company, Inc.,518. 
DDT, DDE, and DDD 319 

9. REFERENCES 

*Budunova I, Williams G, Spray D. 1993. Effect of tumor promoting stimuli on gap junction 

permeability and connexin 43 expression in ARL 18 rat liver cell line. Arch Toxicol 67:565-572. 

*Bulger WH, Kupfer D. 1983. Estrogenic action of DDT analogs. Am J Ind Med 4:163-173. 

*Bunyan PJ, Page JMJ. 1973. Pesticide-induced changes in hepatic microsomal enzyme systems: Some 

effects of 1,1-di(p-chlorophenyl)-2,2-dichloroethylene (DDE) and 1,1-di(p-chlorophenyl)-2­

chloroethylene (DDMU) in the rat and Japanese quail. Chem Biol Interact 6:249-257. 

*Bunyan PJ, Davidson J, Shorthill MJ. 1970. Hepatic glucose-6-phosphate dehydrogenase and 6­

phosphogluconate dehydrogenase levels in Japanese quail following the ingestion of p,p'-DDT and 

related compounds. Chem Biol Interact 2:175-182. 

*Bunyan PJ, Townsend MG, Taylor A. 1972. Pesticide-induced changes in hepatic microsomal enzyme 

systems. Some effects of 1,1-Di(p-chlorophenyl)-2,2,2-trichloroethane (DDT) and 1,1-Di(pchlorophenyl)-2,2-dichloroethylene (DDE) in the rat and Japanese quail. Chem Biol Interact 5:13-26. 

*Burlington H, Lindeman VF. 1950. Effect of DDT on testes and secondary sex characters of white 

leghorn cockerels. Proc Soc Exp Biol Med 74:48-51. 

*Burlington H, Lindeman VF. 1952. Action of DDT on the blood of the chicken. Proc Fed Am Soc Exp 

Biol 11:20-21. 

Burse VW, Patterson DG, Brock JW, et al. 1996. Selected analytical methods used at the Centers for 

Disease Control and prevention for measuring environmental pollutants in serum. Toxicol Ind Health 

12:481-498. 

*Buselmaier W, Roehrborn G, Propping P. 1973. Comparative investigations on the mutagenicity of 

pesticides in mammalian test systems. Mutat Res 21:25-26. 

Buser H-R, Muller MD. 1997. Determination and environmental behavior of persistent halogenated 

chiral contaminants. Organohalogen Compounds 31:225-228. 

Bustos S, Soto J, Tchernitchin AN. 1996. Estrogenic activity of p,p'-DDT. Environ Toxicol Water Qual 

11:265-271. 

Büyüksönmez F, Rynk R, Hess TF, et al. 1999. Occurrence, degradation and fate of pesticides during 

composting. Compost Sci Util 7(4):66-82. 

*Büyüksönmez F, Rynk R, Hess TF, et al. 2000. Occurrence, degradation and fate of pesticides during 

composting. Part II. Occurrence and fate of pesticides in compost and composting systems. Compost 

Sci Util 8:61-81 

Bunce NJ, Cox BJ, Partridge AW. 2000. Method development and interspecies comparison of estrogen 

receptor binding assays for estrogen agonists and antagonists. ACS Symp Ser 747:11-21. 

Burger M, Mate M, Lavina R, et al. 2000. Rol de los plaguicidas organoclorados en el cáncer de mama. 

Rev Toxicol (Iberica) 17:79-82. 

*Byczkowski JZ. 1977. Biochemical and ultrastructural changes in biological membranes induced by 

p,p'-DDT (chlorophenotanum) and its metabolites. Pol J Pharmacol Pharm 29:411-417. 
DDT, DDE, and DDD 320 

9. REFERENCES 

*Cabral J, Hall R, Rossi L, et al. 1982a. Lack of carcinogenicity of DDT in hamsters. Tumori 68:5-10.
 

*Cabral J, Hall R, Rossi L, et al. 1982b. Effects of long-term intake of DDT on rats. Tumori 68:11-17.
 

*Cai W, Counsell R, Djanegara T, et al. 1995. Metabolic activation and binding of mitotane in adrenal
 

cortex homogenates. J Pharm Sci 84(2):134-138.
 

*Cameron G. 1945. Risks to man and animals from the use of 2,2-bis(p-chlorophenyl),
 

1,1,1-trichloroethane (DDT): With a note on the toxicology of γbenzene hexachloride (666, gammexane). 

Br Med Bull 3:233-235.
 

*Cameron G, Burgess F. 1945. The toxicity of 2,2-bis(p-chlorophenyl) 1,1,1-trichloroethane (DDT). Br
 

Med J 1:865-871.
 

Cameron G, Cheng K. 1951. Failure of oral DDT to induce toxic changes in rats. Br Med J 819-821.
 

Cannon M, Holcomb L. 1968. The effect of DDT on reproduction in mice. Ohio J Sci 68(1):19-24. 

*Canter L, Sabatini D. 1994. Contamination of public ground water supplies by superfund sites. Int J
 

Environ Studies 46:35-57.
 

*Cantor KP, Blair A, Everett G. 1992. Pesticides and other agricultural risk factors for non-Hodgkin's
 

lymphoma among men in Iowa and Minnesota. Cancer Res 52:2447-2455.
 

*Carey AE, Kutz FW. 1985. Trends in ambient concentrations of agrochemicals in humans and the
 

environment of the United States. Environ Monit Assess 5:155-163.
 

*Carey A, Douglas P, Tai H, et al. 1979a. Pesticide residue concentrations in soils of five United States
 

cities, 1971—Urban soils monitoring program. Pestic Monit J 13:17-22.
 

*Carey A, Gowen J, Tai H, et al. 1979b. Pesticide residue levels in soils and crops from 37 states,
 

1972—National Soils Monitoring Program (IV). Pestic Monit J 12:209-229.
 

Carey A, Yang H, Wiersma G, et al. 1980. Residual concentrations of propanil, TCAB, and other
 

pesticides in rice-growing soils in the United States, 1972. Pestic Monit J 14(1):23-25.
 

Carr RS, Long ER, Windom HL, et al. 1996. Sediment quality assessment studies of Tampa Bay,
 

Florida. Environ Toxicol Chem 15(7):1218-1231.
 

Carr RS, Montagna PA, Biedenbach JM, et al. 2000. Impact of storm-water outfalls on sediment quality
 

in Corpus Christi, Texas, USA. Environ Toxicol Chem 19:561-574
 

CDC. 1999. CDC state health department search. Center for Disease Control and Prevention.
 

http://search.cdc.gov/shd/search2.html. May 25, 1999. 

*CDC/ATSDR. 1990. Biomarkers of organ damage or dysfunction for the renal, hepatobiliary and
 

immune systems. Summary report, August 27, 1990. Atlanta, GA: Centers for Disease Control, Agency 

for Toxic Substances and Disease Registry. 

*Cecil HC, Bitman J, Harris SJ. 1971. Effects of dietary p,p'-DDT and p,p'-DDE on egg production and 

egg shell characteristics of Japanese quail receiving an adequate calcium diet. Poult Sci 50(2):657-659. 
DDT, DDE, and DDD 321 

9. REFERENCES 

CELDS. 1992. Computer-environmental legislative data systems. Database. March, 1992. 

*Chapin R, Stevens J, Hughes C, et al. 1996. Symposium overview; endocrine modulation of 

reproduction. Fundam Appl Toxicol 29:1-17. 

Cheek AO, Kow K, Chen J, et al. 1999. Potential mechanisms of thyroid disruption in humans: 

Interaction of organochlorine compounds with thyroid receptor, transthyretin, and thyroid-binding 

globulin. Environ Health Perspect 107(4):273-278. 

*Chen HC, Farese RV Jr. 1999. Steroid hormones: Interactions with membrane-bound receptors. Curr 

Biol 9(13):R478-R481. 

*Chen S-W, Dziuk PJ, Francis BM. 1994. Effects of four environmental toxicants on plasma Ca and 

estradiol 17(beta) ans hepatic P450 in laying hens. Environ Toxicol Chem 13(5):789-796. 

*Chhabra R, Fouts J. 1973. Stimulation of hepatic microsomal drug-metabolizing enzymes in mice by 

1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) and 3,4-benzpyrene. Toxicol Appl Pharmacol 

25:60-70. 

*Chowdhury AR, Gautam AK, Venkatakrishna-Bhatt H. 1990. DDT (2,2-bis(p-chlorophenyl) 

1,1,1-trichloroethane) induced structural changes in adrenal glands of rats. Bull Environ Contam Toxicol 

45:193-196. 

*Chung E, Van Woert M. 1984. DDT myoclonus: Sites and mechanism of action. Exp Neurol 85:273­

282. 

*Chura NJ, Stewart PA. 1967. Care, food consumption, and behavior of bald eagles used in DDT tests. 

The Wilson Bulletin 79(1):441-448. 

*Clark DR. 1981. Death in bats from DDE, DDT or dieldrin: Diagnosis via residues in carcass fat. Bull 

Environ Contam Toxicol 26:367-374. 

*Clark DR, Kroll JC. 1977. Effects of DDE on experimentally poisoned free-tailed bats (Tadarida 

brasiliensis): Lethal brain concentrations. J Toxicol Environ Health 3:893-901. 

*Clark J. 1974. Mutagenicity of DDT in mice, Drosophila melanogaster and Neurospora crassa. Aust J 

Biol Sci 27:427-440. 

*Clement J, Okey A. 1972. Estrogenic and anti-estrogenic effects of DDT administered in the diet to 

immature female rats. Can J Physiol Pharmacol 50:971-975. 

*Clement J, Okey A. 1974. Reproduction in female rats born to DDT-treated parents. Bull Environ 

Contam Toxicol 12(3):373-377. 

*Clewell HJ III, Andersen ME. 1985. Risk assessment extrapolations and physiological modeling. 

Toxicol Ind Health 1(4):111-113. 

*Cocco PL. 1997. Environmental exposure to p,p'-DDE and human fertility. Bull Environ Contam 

Toxicol 59:677-680. 
DDT, DDE, and DDD 322 

9. REFERENCES 

*Cocco P, Benichou J. 1998. Mortality from cancer of the male reproductive tract and the environmental 

exposure to the anti-androgen p,p'-dichlorodiphenyldichloroethylene in the United States. Oncology 

55:334-339. 

*Cocco P, Blair A, Congia P, et al. 1997a. Long-term health effects of the occupational exposure to 

DDT: A preliminary report. Ann N Y Acad Sci 834:246-256. 

*Cocco P, Blair A, Congia P, et al. 1997b. Proportional mortality of dichloro-diphenyl-trichloroethane 

(DDT) workers: a preliminary report. Arch Environ Health 52(4):299-303. 

*Cocco P, Kazerouni N, Zahm SH. 2000. Cancer mortality and environmental exposure to DDE in the 

United States. Environ Health Perspect 108(1):1-4. 

*Colborn T, Clement C, eds. 1992. Chemically-induced alterations in sexual and functional development: 

The wildlife-human connection. In: Advances in modern environmental toxicology, Vol. XXI, Princeton, 

NJ: Princeton Scientific Publishing. 

*Colborn T, vom Saal F, Soto A. 1993. Developmental effects of endocrine-disrupting chemicals in 

wildlife and humans. Environ Health Perspect 101(5):378-384. 

Collier MJ, McAnulty PA, Tesh JM. 1992. Evaluation of methylmercury and DDT in an EPA-style 

developmental neurotoxicity screen [Abstract]. Teratology 45(5):501. 

Committee for Analytical Methods for Residues of Pesticides in Foodstuffs of the Ministry of 

Agriculture, Fisheries and Food. 1992. Report of two co-operative trials of a gel permeation 

chromatographic method for the isolation of pesticide residues from oils and fats. Analyst 

117(9):1451-1455. 

*Conney A. 1971. Environmental factors influencing drug metabolism. In: LaDu B, et al. eds. 

Fundamentals of drug metabolism and drug disposition. Baltimore, MD: Williams & Wilkins Company, 

253-278. 

*Cooke AS. 1970a. The effect of o,p'-DDT on Japanese quail. Bull Environ Contam Toxicol 5(2):152­

157. 

*Cooke AS. 1970b. The effect of p,p'-DDT on tadpoles of the common frog (Rana temporaria). 

Environ Pollut 1:57-71.
 

*Cooke AS. 1972. The effects of DDT, dieldrin and 2,4-D on amphibian spawn and tadpoles. Environ
 

Pollut 3:51-68.
 

*Cooke AS. 1973a. The effects of DDT, when used as a mosquito larvicide, on tadpoles of the frog 

Rana temporaria. Environ Pollut 5:259-273. 

*Cooke AS. 1973b. Response of Rana temporaria tadpoles to chronic doses of pp'-DDT. Copeia 4:647­

652.
 

*Cooke AS. 1973c. Shell thinning in avian eggs by environmental pollutants. Environ Pollut 4:85-152.
 

Cormican M, Collignon P, van den Bogaard T. 2000. The DDT question. Lancet 356(9236):1190-1191.
 
DDT, DDE, and DDD 323 

9. REFERENCES 

Corona-Cruz A, Gold-Bouchot G, Gutierrez-Rojas M, et al. 1999. Anaerobic-aerobic biodegradation of 

DDT (dichlorodiphenyl trichloroethane) in soils. Bull Environ Contam Toxicol 63:219-225. 

Corrigan FM, Weinburg CL, Shore RF, et al. 2000. Organochlorine insecticides in substantia nigra in 

Parkinson's disease. J Toxicol Environ Health A 59:229-234. 

*Coulston F. 1985. Reconsideration of the dilemma of DDT for the establishment of an acceptable daily 

intake. Regul Toxicol Pharmacol 5:332-383. 

*Coupe RH, Manning MA, Foreman WT, et al. 2000. Occurrence of pesticides in rain and air in urban 

and agricultural areas of Mississippi, April-September 1995. Sci Total Environ 248:227-240. 

*Cowley SM, Hoare S, Mosselman S, et al. 1997. Estrogen receptors α and β form heterodimers on 

DNA. J Biol Chem 272(32):19858-19862. 

Craan A, Haines D. 1998. Twenty-five years of surveillance for contaminants in human breast milk. 

Arch Environ Contam Toxicol 35:702-710. 

*Craig G, Ogilvie D. 1974. Alteration of T-maze performance in mice exposed to DDT during 

pregnancy and lactation. Environ Physiol Biochem 4:189-199. 

*Crain DA and Guillette LJ. 1998. Reptiles as models of contaminant-induced endocrine disruption. 

Anim Reprod Sci 53:77-86. 

*Crain DA, Guillette LJ, Rooney AA, et al. 1997. Alterations in steroidogenesis in alligators (Alligator 

mississippiensis) exposed naturally and experimentally to environmental contaminants. Environ Health 

Perspect 105(5):528-533. 

*Cranmer M, Carroll J, Copeland M. 1972a. Determinations of DDT and metabolites including DDA, in 

human urine by gas chromatography. In: Davies J, Edmundson W, eds. Epidemiology of DDT. Mount 

Kisco, New York, NY: Futura Publishing Company, Inc., 147-152. 

*Cranmer M, Peoples A, Chadwick R. 1972b. Biochemical effects of repeated administration of 

p,p'-DDT on the squirrel monkey. Toxicol Appl Pharmacol 21:98-101. 

Crellin NK, Kang HG, Swan CL, et al. 2000. Dichlorodiphenyldichloroethylene (DDE) and 

methoxychlor (MxC) inhibit basal and stimulated progesterone P4 synthesis in stable granulosa cells 

[Abstract]. Biol Reprod 62(5):182-183. 

Crellin NK, Rodway MR, Swan CL, et al. 1999. Dichlorodiphenyldichloroethylene potentiates the effect 

of protein kinase A pathway activators on progesterone synthesis in cultured granulosa cells. Biol Reprod 

61:1099-1103. 

*Crisp T, Clegg E, Cooper R, et al. 1998. Environmental endocrine disruption: An effects assessment 

and analysis. Environ Health Perspect 106(Suppl 1):11-56. 

*Crockett A, Wiersma G, Tai H, et al. 1974. Pesticides in soils. Pesticide residue levels in soils and 

crops, FY-70—national soils monitoring program (II). Pestic Monit J 8(2):69-97. 

CSCORP. 1989. STN International Network, Chemical Abstracts Service, Columbus, OH. 

September 28, 1989. 
DDT, DDE, and DDD 324 

9. REFERENCES 

CSCORP. 1992. STN International Network, Chemical Abstracts Service, Columbus, OH. 

April 20, 1992. 

*Cueto C. 1970. Cardiovascular effects of o,p'-DDD. Ind Med 39:55-56. 

Cullen RT, Tran CL, Buchanan D, et al. 2000. Inhalation of poorly soluble particles. I. Differences in 

inflammatory response and clearance during exposure. Inhal Toxicol 12:1089-1111. 

*Curley A, Copeland M, Kimbrough R. 1969. Chlorinated hydrocarbon insecticides in organs of 

stillborn and blood of newborn babies. Arch Environ Health 19:628-632. 

Curtis CF, Lines JD. 2000. Should DDT be banned by international treaty? Parasitol Today 16(3):119­

121. 

*Custer TW, Hines RK, Stewart PM, et al. 1998. Organochlorines, mercury, and selenium in Great Blue 

Heron eggs from Indiana Dunes National Lakeshore, Indiana. J Great Lakes Res 24(1):3-11. 

Danz M, Urban H. 1980. Elevated mitotic number in the adrenal cortex as a reflection of early events in 

growth induction by carcinogens and promoters—a possible short-term assay. Arch Toxicol (Suppl 

4):19-21. 

*Danzo B. 1997. Environmental xenobiotics may disrupt normal endocrine function by interfering with 

the binding of physiological ligands to steroid receptors and binding proteins. Environ Health Perspect 

105(3):294-301. 

Darvill T, Lonky E, Reihman J, et al. 2000. Prenatal exposure to PCBs and infant performance of the 

Fagan test of infant intelligence. Neurotoxicology 21(6):1029-1038. 

*Das S, Thomas P. 1999. Pesticides interfere with the nongenomic action of a progestogen on meiotic 

maturation by binding to its plasma membrane receptor on fish oocytes. Endocrinology 140: 1953-1956. 

*Daston GP, Gooch JW, Breslin WJ, et al. 1997. Environmental estrogens and reproductive health: A 

discussion of the human and environmental data. Rep Toxicol 11(4):465-481. 

*Datta P. 1970. In vivo detoxication of p,p'-DDT via p,p'-DDE to p,p'-DDA in rats. In: Deichmann W, 

et al., eds. Pesticide symposia. 6th-7th International-American Conference of Toxicology and 

Occupational Medicine, Miami, FL: Haol and Associates, Inc., 41-45. 

*Datta P, Nelson M. 1970. p,p'-DDT detoxication by isolated perfused rat liver and kidney. In: 

Deichmann W, et al., eds. Pesticide symposia. 6th-7th International-American Conference of Toxicology 

and Occupational Medicine, Miami, FL: Halo and Associates, Inc., 47-50. 

*Davies RP, Dobbs AJ. 1984. The prediction of bioconcentration in fish. Water Res 18(10):1253-1262. 

*Davison KL. 1978. Calcium-45 uptake by shell gland, oviduct, plasma and eggshell of DDT-dosed 

ducks and chickens. Arch Environ Contam Toxicol 7:359-367. 

*Davison KL, Sell JL. 1972. Dieldrin and p,p'-DDT effects on egg production and eggshell thickness of 

chickens. Bull Environ Contam Toxicol 7(1):9-18. 
DDT, DDE, and DDD 325 

9. REFERENCES 

*Davison KL, Engebretson KA, Cox JH. 1976. p,p'-DDT and p,p'-DDE effects on egg production, 

eggshell thickness and reproduction of Japanese quail. Bull Environ Contam Toxicol 15(3):265-270. 

Dawson A. 2000. Mechanisms of endocrine disruption with particular reference to occurrence in avian 

wildlife: A review. Ecotoxicology 9:59-69. 

Dean M, Smeaton T, Stock B. 1980. The influence of fetal and neonatal exposure to 

dichlorodiphenyltrichloroethane (DDT) on the testosterone status of neonatal male rat. Toxicol Appl 

Pharmacol 53:315-322. 

*de Andrea MM, Papini S, Nakagawa LE. 2001. Optimizing microwave-assisted solvent extraction 

(mase) of pesticides from soil. J Environ Sci Health B 36(1):87-93. 

*Dees C, Askari M, Foster J, et al. 1997a. DDT mimics estradiol stimulation of breast cancer cells to 

enter the cell cycle. Molec Carcinog 18:107-114. 

*Dees C, Askari M, Garrett S, et al. 1997b. Estrogenic and DNA-damaging activity of red no. 3 in 

human breast cancer cells. Environ Health Perspect 105(Suppl 3):625-632. 

*Dees C, Garrett S, Henley D, et al. 1996. Effects of 60-Hz fields, estradiol and xenoestrogens on 

human breast cancer cells. Radiat Res 146:444-452. 

*DeFranco DB. 1999. Regulation of steroid receptor subcellular trafficking. Cancer Biochem Biophys 

30:1-24. 

*de Graaf IAM, Tajima O, Groten JP, et al. 2000. Intercellular communication and cell proliferation in 

precision-cut rat liver slices: effect of medium composition and DDT. Cancer Lett 154:53-62. 

*Deichmann W, Keplinger M. 1966. Effect of combinations of pesticides on reproduction of mice. 

Toxicol Appl Pharmacol 8:337-338. 

*Deichmann W, Keplinger M, Sala F, et al. 1967. Synergism among oral carcinogens. Toxicol Appl 

Pharmacol 11:88-103. 

*Deichmann W, MacDonald W, Beasley A, et al. 1971. Subnormal reproduction in beagle dogs induced 

by DDT and aldrin. Ind Med 40:10-20. 

Deichmann W, MacDonald W, Cubit D. 1975. Dieldrin and DDT in the tissues of mice fed aldrin and 

DDT for seven generations. Arch Toxicol 34:173-182. 

*de Jager LS, Andrews RJ. 2000. Development of a rapid screening technique for organochlorine 

pesticides using solvent microextraction (SME) and fast gas chromatography (GC). Analyst 125:1943­

1948. 

Dello Iacovo R, Celentano E, Strollo AM, et al. 1999. Organochlorines and breast cancer: A study on 

Neapolitan women. In: Zappia et al, ed. Advances in nutrition and cancer. New York, NY: Kluwer 

Academic/ Plenum Publishers, 57-66. 

*DeLong RL, Gilmartin WG, Simpson JG. 1973. Premature births in California sea lions: Association 

with high organochlorine pollutant residue levels. Science 181:1168-1169. 
DDT, DDE, and DDD 326 

9. REFERENCES 

*Del Pup J, Pasternack B, Harley N, et al. 1978. Effects of DDT on stable laboratory mouse populations. 

J Toxicol Environ Health 4:671-687. 

*Demers A, Ayotte P, Brisson J, et al. 2000. Risk and aggressiveness of breast cancer in relation to 

plasma organochlorine concentrations. Cancer Epidemiol Biomark Prev 9:161-166. 

Desaulniers D, Leingartner K, Cooke G, et al. 1999. Effects of exposure to a human milk PCB-DDTDDE mixture from day 1 to 20, or to TCDD on day 18 of age, in prepubertal females, and on the 

development of methylnitrosourea-induced mammary tumors in the adult rat. Organohalogen 

Compounds 42:221-224. 

*de Solla SR, Bishop CA, Van Der Kraak G, et al. 1998. Impact of organochlorine contamination on 

levels of sex hormones and external morphology of common snapping turtles (Chelydra serpentina 

serpentina) in Ontario, Canada. Environ Health Perspect 106(5):253-260. 

*De Stefani E, Kogevinas M, Boffetta P, et al. 1996. Occupation and the risk of lung cancer in Uruguay. 

Scand J Work Environ Health 22:346-352. 

DeVito MJ, Sparrow BR. 1999. Antiandrogenic effects of 2,2-bis(4-chlorophenyl)-1,1-dichloroethylene 

(p,p'-DDE) in neonatal and fetal rats. Organohalogen Compounds 42:1-4. 

*Dewailly E, Ayotte P, Bruneau S, et al. 2000. Susceptibility to infections and immune status in Inuit 

infants exposed to organochlorines. Environ Health Perspect 108(3):205-211. 

*Dewailly E, Ayotte P, Laliberte C, et al. 1996. Polychlorinated biphenyl (PCB) and dichlorodiphenyl 

dichloroethylene (DDE) concentrations in the breast milk of women in Quebec. Am J Public Health 

86(9):1241-1246. 

*Dewailly E, Ryan JJ, Laliberte C, et al. 1994. Exposure of remote maritime populations to coplanar 

PCBs. Environ Health Perspect Suppl 102(Supplement 1):205-209. 

*De Waziers I, Azais V. 1987. Drug-metabolizing enzyme activities in the liver and intestine of rats 

exposed to DDT: Effects of vitamin A status. Arch Environ Contam Toxicol 16:343-348. 

Dickstein G, Shechner C, Arad E, et al. 1998. Is there a role for low doses of mitotane (o,p'-DDD) as 

adjuvant therapy in adrenocortical carcinoma? J Clin Endocrinol Metab 83(9):3100-3103. 

*Diel P, Schulz T, Smolnikar K, et al. 2000. Ability of xeno-and phytoestrogens to modulate expression 

of estrogen-sensitive genes in rat uterus: estrogenicity profiles and uterotropic activity. J Steroid 

Biochem Mol Biol 73:1-10. 

*Dieter MP. 1974. Plasma enzyme activities in Coturnix quail fed graded doses of DDE, polychlorinated 

biphenyl, malathion and mercuric chloride. Toxicol Appl Pharmacol 27:86-98. 

DiGiulio W, Beierwaltes W. 1968. Tissue localization studies of a DDD analog. J Nucl Med 9:634-637. 

*Dimond JB, Owen RB. 1996. Long-term residue of DDT compounds in forest soils in Maine. Environ 

Pollut 92:227-230 

Ding J-Y, Wu S. 1995. Partition coefficients of organochlorine pesticides on soil and on the dissolved 

organic matter in water. Chemosphere 30(12):2259-2266. 
DDT, DDE, and DDD 327 

9. REFERENCES 

*Ding J-Y, Wu S. 1997. Transport of organocholorine pesticides in soil columns enhanced by dissolved 

organic carbon. Water Sci Technol 35(7):139-145. 

*Diwan B, Ward J, Kurata Y, et al. 1994. Dissimilar frequency of hepatoblastomas and hepatic 

cystadenomas and adenocarcinomas arising in hepatocellular neoplasms of D2B6F1 mice initiated with 

–nitrosodiethylamine and subsequently given aroclor-1254, dichlorodiphenyltrichloroethane, or 

phenobarbital. Toxicol Pathol 22(4):430-439. 

*Djordjevic M, Fan J, Hoffmann D. 1995. Assessment of chlorinated pesticide residues in cigarette 

tobacco based on supercritical fluid extraction and GC-ECD. Carcinogenesis 16(11):2627-2632. 

*Donaldson GM, Schutt JL, Hunter P. 1999. Organochlorine contamination in Bald Eagle eggs and 

nestlings from the Canadian Great Lakes. Arch Environ Contam Toxicol 36:70-80. 

*Dorea JG, Cruz-Granja AC, Lacayo-Romero ML, et al. 2001. Perinatal metabolism of 

dichlorodiphenyldichloroethylene in Nicaraguan mothers. Environ Res A 86:229-237. 

*Dorgan J, Brock J, Rothman N, et al. 1999. Serum organochlorine pesticides and PCBs and breast 

cancer risk: results from a prospective analysis (USA). Cancer Causes Control 10:1-11. 

DOT. 1989a. Hazardous materials table. U.S. Department of Transportation. Code of Federal 

Regulations. 49 CFR 172.101. 

DOT. 1989b. Hazardous materials table. U.S. Department of Transportation. Federal Register 54 

(185):39501-39505. 

Douthwaite RJ. 1992. Effects of DDT on the fish eagle Haliaeetus vocifer population of Lake Kariba in 

Zimbabwe. Ibis 134:250-258. 

*Dreisbach RH, ed. 1983. Halogenated insecticides. In: Handbook of poisoning: Prevention diagnosis 

and treatment. 11th ed. Los Altos, CA: Lange Medical Publications, 107-116. 

*Driscoll MS, Hassett JP, Fish CL, et al. 1991. Extraction efficiencies of organochlorine compounds 

from Niagara River water. Environ Sci Technol 25(8):1432-1439. 

*Dua VK, Kumari R, Sharma VP, et al. 2001. Organochlorine residues in human blood from Nainital 

(U.P.), India. Bull Environ Contam Toxicol 67:42-45. 

*Duby R, Travis H, Terrill C. 1971. Uterotropic activity of DDT in rats and mink and its influence on 

reproduction in the rat. Toxicol Appl Pharmacol 18:348-355. 

Duce RA, Liss PS, Merrill JT, et al. 1991. The atmospheric input of trace species to the world ocean. 

Global Biogeochem Cycles 5:193-260. 

*Duggan R, Corneliussen P, Duggan M, et al. 1983. Pesticide residue levels in foods in the United 

States from July 1, 1969, to June 30, 1976: Summary. J Assoc Off Anal Chem 66(6):1534-1535. 

*Dunstan R, Roberts T, Donohoe M, et al. 1996. Bioaccumulated chlorinated hydrocarbons and 

red/white blood cell parameters. Biochem Molec Med 58:77-84. 
DDT, DDE, and DDD 328 

9. REFERENCES 

*Durham W, Ortega P, Hayes W. 1963. The effect of various dietary levels of DDT on liver function, 

cell morphology, and DDT storage in the Rhesus monkey. Arch Int Pharmacodyn Ther 141:111-129. 

Eagle DJ, Jones JLO, Jewell EJ, et al. 1992. Determination of pesticides by gas chromatography and 

high-pressure liquid chromatography. In: Smith KA, ed. Books in soils, plants, and the environment: 

Soil analysis: Modern instrumental techniques, second edition. New York, NY: Marcel Dekker, Inc., 

547-600. 

*Ecobichon DJ. 1995. Toxic effects of pesticides. In: Klaassen C, et al., eds. Casarett and Doull's 

toxicology. 5th ed. New York, NY: Macmillan Co., Inc., 643-689. 

Eisenreich S, Looney B, Thornton J. 1981. Airborne organic contaminants in the Great Lakes 

ecosystem: Atmospheric fluxes to the Great Lakes are a combination of dry and wet removal processes. 

Dry deposition is 1.5-5.0 times the wet deposition for the trace organics selected-chlorinated pesticides 

and polychlorinated biphenyls. Environ Sci Technol 15(1):30-38. 

*Eitzer BD, Chevalier A. 1999. Landscape care pesticide residues in residential drinking water wells. 

Bull Environ Contam Toxicol 62:420-427. 

Ellenhorn MJ, Barceloux DG. 1988. Organochlorines. In: Medical toxicology: diagnoses and 

treatment of human poisoning. New York, NY: Elsevier, 1078-1080. 

*Elvia LF, Sioban HD, Bernardo HP, et al. 2000. Organochlorine pesticide in rural and urban areas in 

Mexico. J Expo Anal Environ Epidemiol 10:394-399. 

Ema M, Kurosaka R, Amano H, et al. 1996. Comparative developmental toxicity of di- tr- and 

tetrabutylin compounds after administration during late organogenesis in rats. J Appl Toxicol 16(1):71­

76. 

*EMMI. 1999. EPA environmental monitoring methods index: Detail analyte. Version I. PC no. 4082. 

Rockland, MD: Government Institutes (1997). 

*EPA. 1972. U.S. Environmental Protection Agency. Consolidated DDT hearings: Opinion and order 

of the administrator. Federal Register 37(131):13369. 

*EPA. 1975. DDT: A review of scientific and economic aspects of the decision to ban its use as a 

pesticide. Washington, DC: U.S. Environmental Protection Agency. EPA-540/1-75-022. 

EPA. 1978a. Designation of hazardous substances. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 116.4 

EPA. 1978b. Designation of hazardous substances. U.S. Environmental Protection Agency. Federal 

Register 43(49):10474-10487. 

EPA. 1979a. Toxic pollutants. U.S. Environmental Protection Agency. Code of Federal Regulations. 

40 CFR 401.15. 

EPA. 1979b. Toxic pollutants. U.S. Environmental Protection Agency. Federal Register 

44(147):44502-44503. 
DDT, DDE, and DDD 329 

9. REFERENCES 

*EPA. 1979c. Water-related environmental fate of 129 priority pollutants: Volume I: Introduction and 

technical background, metals and inorganics, pesticides and PCBs. Washington, DC: U.S. 

Environmental Protection Agency. EPA-440/4-79-029b, 1-1 - 1-4, 2-1 - 2-16. 

*EPA. 1980a. Ambient Water Quality Criteria for DDT. Washington, DC: U.S. Environmental 

Protection Agency. EPA 440/5-80-038. 

*EPA. 1980b. Analysis of pesticide residues in human and environmental samples: A compilation of 

methods selected for use in pesticide monitoring programs. U.S. Environmental Protection Agency. EPA 

600/8-80-038. 

EPA. 1981. Toxic pollutants. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 

CFR 401.15. 

*EPA. 1982. Test Method-608: Organochlorine pesticides and PCB's. Cincinnati, OH: U.S 

Environmental Protection Agency, Environmental Monitoring and Support Laboratory. 

EPA. 1984. Guidelines establishing test procedures for the analysis of pollutants under the Clean Water 

Act (40 CFR 126). U.S. Environmental Protection Agency. Federal Register 49:43321-43331. 

EPA. 1985. List of hazardous substances and reportable quantities. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 302.4 

*EPA. 1986a. Superfund public health evaluation manual. Washington, DC: U.S. Environmental 

Protection Agency, Office of Emergency and Remedial Response. EPA 540/1-86-060. 

EPA. 1986b. Laboratory Manual, Physical/Chemical Methods: Test methods for evaluating solid waste. 

U.S. Environmental Protection Agency. SW-848 (v. 1B). 

EPA. 1986c. Revocation of DDT and TDE Tolerances. U.S. Environmental Protection Agency. Code 

of Federal Regulations 40 CFR 180. 

*EPA. 1986d. Reference Values for Risk Assessment. Cincinnati, OH: Environmental Protection 

Agency, Office of Research and Development. 

EPA. 1987a. List (phase 1) of hazardous constituents for ground-water monitoring. U.S. Environmental 

Protection Agency. Federal Register 52(131):25942-25953. 

EPA. 1987b. Maximum concentration of constituents for ground-water protection. U.S. Environmental 

Protection Agency. Code of Federal Regulations. 40 CFR 264.94, Appendix IX. 

EPA. 1988a. Integrated Risk Information System (IRIS). Cincinnati, Ohio: U.S. Environmental 

Protection Agency, Environmental Criteria and Assessment Office. 

EPA. 1988b. Summary of state and federal drinking water standards and guidelines. Washington, DC: 

U.S. Environmental Protection Agency, Chemical Communication Subcommittee, Federal-State 

Toxicology, and Registry Alliance Committee (FSTRAC). 

EPA. 1988c. General pretreatment regulations for existing and new sources of pollution. U.S. 

Environmental Protection Agency. Code of Federal Regulations. 40 CFR 403 Appendix B. 
DDT, DDE, and DDD 330 

9. REFERENCES 

EPA. 1988d. General pretreatment regulations for existing and new sources of pollution. U.S.
 

Environmental Protection Agency. Federal Register 53(200):40610-40616.
 

EPA. 1988e. Hazardous waste management system: Identification and listing of hazardous waste. U.S.
 

Environmental Protection Agency. Code of Federal Regulations. 40 CFR 261.
 

EPA. 1988f. Hazardous waste management system: Identification and listing of hazardous waste. U.S.
 

Environmental Protection Agency. Federal Register 53(78):13382-13393.
 

EPA. 1988g. Recommendations for and documentation of biological values for use in risk assessment. 

Cincinnati, OH: Environmental Protection Agency, Office of Research and Development.
 

EPA. 1989a. Designation of hazardous substances. U.S. Environmental Protection Agency. Code of
 

Federal Regulations. 40 CFR 116.4.
 

EPA. 1989b. List of hazardous substances and reportable quantities. U.S. Environmental Protection
 

Agency. Code of Federal Regulations. 40 CFR 302.4.
 

EPA. 1989c. List of hazardous substances and reportable quantities. U.S. Environmental Protection
 

Agency. Federal Register 54(155):33428-33484.
 

*EPA. 1990a. Interim methods for development of inhalation reference doses. Washington, DC: U.S.
 

Environmental Protection Agency, Environmental criteria and assessment office. EPA/600/8-90/066A.
 

EPA. 1990b. U.S. Environmental Protection Agency. Code of federal regulations. 40 CFR 136 A. 

*EPA. 1991. Hazardous constituents: Identification and listing of hazardous waste. Washington, DC: 

U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 261, Appendix VIII. 

*EPA 1994. USEPA Contract Laboratory Program. Statement of work for organics analysis. U.S.
 

Environmental Protection Agency. www.EPA.Gov/region09/QA/OLM032.PDF.
 

*EPA. 1997. Special report on environmental endocrine disruption: An effects assessment and analysis.
 

Washington, DC: U.S. Environmental Protection Agency, Office of Research and Development. 

EPA/630/R-96/012.
 

EPA. 1998a. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 401.15.
 

EPA. 1998b. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 129.101.
 

EPA. 1998c. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 261.33.
 

EPA. 1998d. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 258.
 

EPA. 1998e. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 455.20.
 

EPA. 1998f. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 180.3.
 

EPA. 1998g. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 116.4.
 

EPA. 1998h. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 302.4.
 
DDT, DDE, and DDD 331 

9. REFERENCES 

EPA. 1998i. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 129.4. 

*EPA 1998j. Method 8081B: Organochlorine pesticides by gas chromatography. U.S. Environmental 

Protection Agency. 

*EPA. 1998k. Method 8270D: Semivolatile organic compounds by gas chromatography/mass 

spectrometry (GC/MS). Washington, DC: U.S. Environmental Protection Agency. 

*EPA. 1999a. National Recommended Water Quality Criteria- Correction. Washington, DC: U.S. 

Environmental Protection Agency, Office of Water. EPA 822-Z-99-001. 

*EPA. 1999b. Listing of fish and wildlife advisories. U.S. Environmental Protection Agency. 

http://fish.rti.org/. 

*EPA. 2001a. Groundwater monitoring list. U.S. Environmental Protection Agency. Code of Federal 

Regulations. 40 CFR 264, Appendix IX. http://ecfrback.access.gpo.gov/otcgi/cfr/otfilter.cgi...nd 

&QUERY=49673&RGN=BAPPCT&SUBSET=SUBSET&FROM=1&ITEM=1. September 24, 2001. 

*EPA. 2001b. Health based limits for exclusion of waste-derived residues. U.S. Environmental 

Protection Agency. Code of Federal Regulations. 40 CFR 266, Appendix VII. 

http://ecfrback.access.gpo.gov/otcgi/cfr...0&RGN=BAPPCT&SUBSET=SUBSET&FROM=1&ITEM=1. 

September 24, 2001. 

*EPA. 2001c. Land disposal restrictions. Universal treatment standards. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 268.48(a). http://ecfr.access.gpo.gov/otcgi/cfr/otf... 

1&RGN=BSECCT&SUBSET=SUBSET& FROM=1&ITEM=1. September 24, 2001. 

*EPA. 2001d. Tolerances for related pesticide chemicals. U.S. Environmental Protection Agency. Code 

of Federal Regulations. 40 CFR 180.3(e)(4). http://ecfrback.access.gpo.gov/otcgi/cfr...3 

&RGN=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. September 24, 2001. 

*EPA. 2001e. Water programs—designation of hazardous waste. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 116.4. http://ecfrback.access.gpo.gov/otcgi/cfr/otfilter. 

cgi?DB=...TI&QUERY=66703&RGN=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. 

September 24, 2001. 

*EPA. 2002a. Designation, reportable quantities, and notification. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 302.4. http://ecfrback.access.gpo.gov/otcgi/cfr/otfilter. 

cgi...I&QUERY=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 02, 2002. 

*EPA. 2002b. Determination of reportable quantity. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 117.3. http://ecfrback.access.gpo.gov/otcgi/cfr...8&RGN= 

BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 03, 2002. 

*EPA. 2002c. Effluent guidelines and standards. Pesticides. U.S. Environmental Protection Agency. 

Code of Federal Regulations. 40 CFR 455.20(b). http://ecfrback.access.gpo.gov/otcgi/cfr/otf...1 

&RGN=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 02, 2002. 

*EPA. 2002d. Effluent guidelines and standards. Toxic pollutants. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 129.4. http://ecfrback.access.gpo.gov/otcgi/cfr...6 

&RGN=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 02, 2002. 
DDT, DDE, and DDD 332 

9. REFERENCES 

*EPA. 2002e. Effluent guidelines and standards. Toxic pollutants. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 401.15. http://ecfrback.access.gpo.gov/otcgi/cfr...1 

&RGN=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 02, 2002. 

*EPA. 2002f. Identification and listing of hazardous waste. U.S. Environmental Protection Agency. 

Code of Federal Regulations. 40 CFR 261.33(f). http://ecfrback.access.gpo.gov/otcgi/cfr...8 

&RGN=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 02, 2002. 

*EPA. 2002g. Municipal solid waste landfill. U.S. Environmental Protection Agency. Code of Federal 

Regulations. 40 CFR 258, Appendix II. http://ecfrback.access.gpo.gov/otcgi/cfrotfilter.cgi...nd 

&QUERY=BAPPCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 02, 2002. 

*EPA. 2002h. Risk specific doses. U.S. Environmental Protection Agency. Code of Federal 

Regulations. 40 CFR 266, Appendix V. http://ecfrback.access.gpo.gov/otcgi/cfr/otfilter.cgi?DB=...and 

&QUERY=9974&RGN=BAPPCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 03, 2002. 

*EPA. 2002i. Toxic pollutant effluent standards. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 129.101. http://ecfrback.access.gpo.gov/otcgi/cfr 6&RGN 

=BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 02, 2002. 

*EPA. 2002j. TSCA. Significant new uses. U.S. Environmental Protection Agency. Code of Federal 

Regulations. 40 CFR 721.2287. http://ecfrback.access.gpo.gov/otcgi/cfr/otf...6&RGN= 

BSECCT&SUBSET=SUBSET&FROM=1&ITEM=1. January 03, 2002. 

*Eriksson M, Karlsson M. 1992. Occupational and other environmental factors and multiple myeloma: 

A population based case-control study. Br J Ind Med 49:95-103. 

Eriksson P. 1984. Age-dependent retention of (14C)DDT in the brain of the postnatal mouse. Toxicol 

Lett 22:323. 

Eriksson P, Darnerud P. 1985. Distribution and retention of some chlorinated hydrocarbons and a 

phthalate in the mouse brain during the pre-weaning period. Toxicology 37:189-203. 

*Eriksson P, Nordberg A. 1986. The effects of DDT, DDOH-palmitic acid, and chlorinated paraffin on 

muscarinic receptors and the sodium-dependent choline uptake in the central nervous system of immature 

mice. Toxicol Appl Pharmacol 85:121-127. 

*Eriksson P, Ahlbom J, Fredriksson A. 1992. Exposure to DDT during a defined period in neonatal life 

induces permanent changes in brain muscarinic receptors and behaviour in adult mice. Brain Res 

582:277-281. 

*Eriksson P, Archer T, Fredriksson A. 1990a. Altered behaviour in adult mice exposed to a single low 

dose of DDT and its fatty acid conjugate as neonates. Brain Res 514:141-142. 

*Eriksson P, Johansson U, Ahlbom J, et al. 1993. Neonatal exposure to DDT induces increased 

susceptibility to pyrethroid (bioallethrin) exposure at adult age - Changes in cholinergic muscarinic 

receptor and behavioural variables. Toxicology 77:21-30 

*Eriksson P, Nilsson-Hakansson L, Nordberg A, et al. 1990b. Neonatal exposure to DDT and its fatty 

acid conjugate: Effects on cholinergic and behavioural variables in the adult mouse. Neurotoxicology 

11:345-354. 
DDT, DDE, and DDD 333 

9. REFERENCES 

*Evans MS, Noguchi GE, Rice CP. 1991. The biomagnification of polychlorinated biphenyls, 

toxaphene, and DDT compounds in a Lake Michigan offshore food web. Arch Environ Contam Toxicol 

20:87-93. 

*Fabro S, McLachlan J, Dames N. et al. 1984. Chemical exposure of embryos during the 

preimplantation stages of pregnancy: Mortality rate and intrauterine development. Am J Obstet Gynecol 

148(7):929-938. 

Fahim M, Bennett R, Hall D, et al. 1970. Effect of DDT on the nursing neonate. Nature 228:1222-1223. 

*Fahrig R. 1974. Comparative mutagenicity studies with pesticides. IARC Sci Publ 10:161-181. 

*Falck F, Ricci A, Wolff MS, et al. 1992. Pesticides and polychlorinated biphenyl residues in human 

breast lipids and their relation to breast cancer. Arch Environ Health 47(2):143-146. 

Fang H, Tong W, Perkins R, et al. 2000. Quantitative comparisons of in vitro assays for estrogenic 

activities. Environ Health Perspect 108(8):723-729. 

*Fang S, Fallin E, Freed V. 1977. Maternal transfer of 14C-p,p'-DDT via placenta and milk and its 

metabolism in infant rats. Arch Environ Contam Toxicol 5:427-436. 

*Faulk CK, Fuiman LA, Thomas P. 1999. Parental exposure to ortho, para-dichorodiphenyltrichloroethane impairs survival skills of Atlantic croaker (Micropogonias undulatis) Larvae. Environ 

Toxicol Chem 18: 254-262. 

Fawcett S, Bunyan P, Huson L, et al. 1981. Excretion of radioactivity following the intraperitoneal 

administration of 14C-DDT, 14C-DDD, 14C-DDE and 14C-DDMU to the rat and Japanese quail. Bull 

Environ Contam Toxicol 27:386-392. 

*Fawcett S, King L, Bunyan P, et al. 1987. The metabolism of [14C]-DDT, [14C]-DDD, [14C]-DDE and 

[

14C]-DDMU in rats and Japanese quail. Xenobiotica 17(5):525-538. 

FDA. 1998. U.S. Food and Drug Administration. Code of Federal Regulations. 21 CFR 310.545. 

*FDA. 1999. Food and Drug Administration Pesticide Program: Residue Monitoring 1999. U.S. Food 

and Drug Administration. http://www.cfsan.fda.gov/~dms/pesrpts.html. 

*FDA. 2001. Food and Drug Administration total diet study. College Park, MD: Center for Food 

Safety, U.S. Food and Drug Administration. http://www.cfsan.fda.gov/~comm/lds_toc.html. 

*FDA. 2002a. Action levels for poisonous or deleterious substances in human food and animal feed. 

U.S. Food and Drug Administration. http://www.cfsan.fda.gov/~krd/fdaact.html. January 11, 2002. 

*FDA. 2002b. Drug products containing certain active ingredients offered over-the-counter (OTC) for 

certain uses. U.S. Food and Drug Administration. Code of Federal Regulations. 21 CFR 310.545(a)(25). 

http://frwebgate.access.gpo.gov/cgi-bin/...PART=310&SECTION 

=545&YEAR=2001&TYPE=TEXT. January 02, 2002. 

FEDRIP. 1999. Federal Research in Progress [database]. Dialog Information Retrieval System, CA. 

*FEDRIP. 2001. Federal Research in Progress [database]. Dialog Information Retrieval System, CA. 
DDT, DDE, and DDD 334 

9. REFERENCES 

Field B, Selub M, Hughes CL. 1990. Reproductive effects of environmental agents. Semin Reprod 

Endocrinol 8(1):44-54. 

*Fishbein L. 1973. Mutagens and potential mutagens in the biosphere I. DDT and its metabolites, 

polychlorinated biphenyls, chlorodioxins, polycyclic aromatic hydrocarbons, haloethers. Sci Total 

Environ 4:305-340. 

*Fishbein L. 1974. Chromatographic and biological aspects of DDT and its metabolites. J Chromatogr 

98:177-251. 

Fitzgerald EF, Brix KA, Deres DA, et al. 1996. Polychlorinated biphenyl (PCB) and dichlorodiphenyl 

dichloroethylene (DDE) exposure among Native American men from contaminated Great Lakes fish and 

wildlife. Toxicol Ind Health 12(3/4):361-368. 

*Fitzhugh O. 1948. Use of DDT insecticides on food products. Ind Eng Chem 40:704-705. 

*Fitzhugh O, Nelson A. 1947. The chronic oral toxicity of DDT (2,2-bis(p-chlorophenyl1,1,1-trichloroethane). J Pharmacol Exp Ther 89:18-30. 

*Fitzpatrick LJ, Dean JR, Comber MH, et al. 2000. Extraction of DDT [1,1,1-trichloro-2,2-bis(pchlorophenyl)-ethylene] and its metabolites DDE [1,1-dichloro-2,2-bis(p-chlorophenyl)-ethylene] and 

DDD [1,1-dichloro-2,2-bis(p-chlorophenyl)-ethane]. J Chromatogr A 874:257-264. 

Flesch-Janys D, Berger J, Konietzko J, et al. 1992. Quantification of exposure to dioxins and furans in a 

cohort of workers of a herbicide producing plant in Hamburg, FRG. Chemosphere 25(7-10):1021-1027. 

Flodström S, Hemming H, Wärngard L, et al. 1990. Promotion of altered hepatic foci development in rat 

liver, cytochrome P450 enzyme induction and inhibition of cell-cell communication by DDT and some 

structurally related organohalogen pesticides. Carcinogenesis 11:1413-1417. 

*Fluck E, Poirier L, Ruelius H, et al. 1976. Evaluation of a DNA polymerase-deficient mutant of E. coli 

for the rapid detection of carcinogens. Chem Biol Interact 15:219. 

*Folmar LC, Denslow ND, Rao V, et al. 1996. Vitellogenin induction and reduced serum testosterone 

concentrations in feral male carp (Cyprinus carpio) captured near a major metropolitan sewage treatment 

plant. Environ Health Perspect 104(10):1096-1101. 

*Fomon SJ. 1966. Body composition of the infant: Part I: The male” reference infant”. In: Falkner F, 

ed. Human Development. Philadelphia, PA: WB Saunders, 239-246. 

*Fomon SJ, Haschke F, Ziegler EE, et al. 1982. Body composition of reference children from birth to 

age 10 years. Am J Clin Nutr 35:1169-1175. 

*Ford WM, Hill EP. 1990. Organochlorine contaminants in eggs and tissue of wood ducks from 

Mississippi. Bull Environ Contam Toxicol 45:870-875. 

*Ford WM, Hill EP. 1991. Organochlorine pesticides in soil sediments and aquatic animals in the upper 

Steele bayou watershed of Mississippi. Arch Environ Contam 20:161-167. 

Forster M, Wilder E, Heinrichs W. 1975. Estrogenic behavior of 2(o-chlorophenyl)-2-(p-chlorophenyl)­

1,1,1-trichloroethane and its homologues. Biochem Pharmacol 24:1777-1780. 
DDT, DDE, and DDD 335 

9. REFERENCES 

Foster EPF MS, Feist GW, Schreck CB, et al. 2001. Gonad organochlorine concentrations and plasma 

steroid levels in white sturgeon (Acipenser transmontanus) from the Columbia River, USA. Bull Environ 

Contam Toxicol 67:239-245. 

Foster WG, Desaulniers D, Leingartner K, et al. 1999. Reproductive effects of tris(4­

chlorophenyl)methanol in the rat. Chemosphere 39(5):709-724. 

*Foster W, Chan S, Hughes PL. 2000. Detection of endocrine disrupting chemicals in samples of second 

trimester human amniotic fluid. J Clin Endocrinol Metab 85(8):2954-2957. 

*Fourie FR, Hattingh J. 1979. DDT administration: Haematological effects observed in the crowned 

guinea-fowl (Numida meleagris). J Environ Pathol Toxicol 2:1439-1446. 

*Francone M, Mariani F, Demare Y. 1952. [Clinical signs of poisoning by DDT.] Rev Assoc Med 

Argentina 66:56-59. (Spanish) 

*Frank R, Braun HE. 1989. PCB and DDE residues in milk supplies of Ontario, Canada 1985-1986. 

Bull Environ Contam Toxicol 42:666-669. 

Frank R, Thomas R, Holdrinet M, et al. 1979. Organochlorine insecticides and PCB in the sediments of 

Lake Huron (1969) and Georgian Bay and North Channel (1973). Sci Total Environ 13:101-117. 

Fregly M. 1968. Effect of o,p'-DDD and metyrapone (SU-4885) on development of renal hypertension 

in rats. Toxicol Appl Pharmacol 12:548-559. 

Fregly M, Water I, Straw J. 1968. Effect of isomers of DDD on thyroid and adrenal function in rats. 

Can J Physiol Pharmacol 46:59-66. 

*Friend M, Trainer DO. 1974a. Response of different-age mallards to DDT. Bull Environ Contam 

Toxicol 11(1):49-56. 

*Friend M, Trainer DO. 1974b. Experimental DDT-duck hepatitis virus interaction studies. J Wildl 

Manage 38(4):887-895. 

*Friend M, Haegele MA, Wilson R. 1973. DDE: Interference with extra-renal salt excretion in the 

mallard. Bull Environ Contam Toxicol 9(1):49-53. 

*Fry DM, Toone CK. 1981. DDT-induced feminization of gull embryos. Science 213:93-95. 

*Fry JR, Garle MJ, Lal K. 1992. Differentiation of cytochrome P-450 inducers on the basis of 

7-alkoxycoumarin O-dealkylase activities. Xenobiotica 22(2):211-215. 

*Fry DM, Toone CK, Speich SM, et al. 1987. Sex ratio skew and breeding patterns of gulls: 

Demographic and toxicological considerations. Stud Avian Biol 10:26-43. 

*Fryzek J, Garabrant D, Harlow S, et al. 1997. A case-control study of self-reported exposures to 

pesticides and pancreas cancer in southeastern Michigan. Int J Cancer 72:62-67. 

*FSTRAC. 1999. Links to states. U.S. Environmental Protection Agency. May 21, 1999. 
DDT, DDE, and DDD 336 

9. REFERENCES 

*Fuhremann TW, Lichtenstein EP. 1980. A comparative study of the persistence, movement, and 

metabolism of six carbon-14 insecticides in soils and plants. J Agric Food Chem 28:446-452. 

*Gabliks J, Al-zubaidy T, Askari E. 1975. DDT and immunological responses. 3. Reduced anaphylaxis 

and mast cell population in rats fed DDT. Arch Environ Health 30:81-84. 

*Gabliks J, Askari E, Yolen N. 1973. DDT and immunological responses: I. Serum antibodies and 

anaphylactic shock in guinea pigs. Arch Environ Health 26:305-308. 

Gad SC, Weil CS. 1989. Statistics for toxicologists. In: Hayes AW, ed. Principles and methods of 

toxicology. 2nd ed. New York, NY: Raven Press, 435-475. 

*Gaido K, Leonard L, Lovell S, et al. 1997. Evaluation of chemicals with endocrine modulating activity 

in a yeast-based steroid hormone receptor gene transcription assay. Toxicol Appl Pharmacol 143:205­

212.
 

*Gaines T. 1969. Acute toxicity of pesticides. Toxicol Appl Pharmacol 14:515-534. 

Galadi A, Bitar H, Chanon M, et al. 1995. Photosensitized reductive dechlorination of chloroaromatic
 

pesticides. Chemosphere 30(9):1655-1669.
 

*Galand P, Mairesse N, Degraef C, et al. 1987. o,p'-DDT (1,1,1-trichloro-2 (p-chlorophenyl) 2-(ochlorophenyl) ethane is a purely estrogenic agonist in the rat uteri in vivo and in vitro. Biochem 

Pharmacol 36(3):397-400. 

*Gammon M, Wolff M, Neugut A, et al. 1997. Temporal variation in chlorinated hydrocarbons in 

healthy women. Cancer Epidemiol Biomarkers Prev 6:327-332. 

Gans DA, Kilgore WW, Ito J. 1994. Residues of chlorinated pesticides in processed foods imported into 

Hawaii from western Pacific Rim countries. Bull Environ Contam Toxicol 52:560-567. 

Gao L, Li G, Wang S. 1991. [Concentration of organochlorine pesticides in aqueous solutions using the 

macromolecular porous resin GDX-102.] Chin J Environ Sci 12:14-17. (Chinese) 

*Garabrant DH, Held J, Langholz B, et al. 1992. DDT and related compounds and risk of pancreatic 

cancer. J Natl Cancer Inst 84(10):764-771. 

*Garcia M, Mourelle M. 1984. Gamma-glutamyl transpeptidase: A sensitive marker in DDT and 

toxaphene exposure. J Appl Toxicol 4(5):246-248. 

*Garrett R. 1947. Toxicity of DDT for man. J Med Assoc State Ala 17:74-78. 

*Gartrell M, Craun J, Podrebarac D, et al. 1985. Pesticides, selected elements, and other chemicals in
 

adult total diet samples, October 1979 - September, 1980. J Assoc Off Anal Chem 68(5):1184-1197.
 

*Gartrell M, Craun J, Podrebarac D, et al. 1986a. Pesticides, selected elements, and other chemicals in 

adult total diet samples, October, 1980 - March, 1982. J Assoc Off Anal Chem 69(1):146-161. 

*Gartrell M, Craun J, Podrebarac D, et al. 1986b. Pesticides, selected elements, and other chemicals in 

infant and toddler total diet samples, October 1980-March 1982. J Assoc Off Anal Chem 69(1):123-145. 
DDT, DDE, and DDD 337 

9. REFERENCES 

*Gauthier JM, Pelletier E, Brochu C, et al. 1998. Environmental contaminants in tissues of neonate St 

Lawrence Beluga whale (Delphinapterus leucas). Mar Pollut Bull 36(1):102-108. 

Gazi E, Conolly RB, You L, et al. 1998. Modeling of p,p'-DDE toxicokinetics during gestation and 

lactation [Abstract]. Toxicologist 42(1-S):142-143. 

*Gellert R, Heinrichs W. 1975. Effects of DDT homologs administered to female rats during the 

perinatal period. Biol Neonate 26:283-290. 

Gellert R, Wilson C. 1979. Reproductive function in rats exposed prenatally to pesticides and 

polychlorinated biphenyls (PCB). Environ Res 18:437-443. 

*Gellert R, Heinrichs W, Swerdloff R. 1972. DDT homologues: Estrogen-like effects on the vagina, 

uterus and pituitary of the rat. Endocrinology 91:1095-1100. 

*Gellert R, Heinrichs W, Swerdloff R. 1974. Effects of neonatally-administered DDT homologs on 

reproductive function in male and female rats. Neuroendocrinology 16:84-94. 

*Geluso KN, Altenbach JS, Wilson DE. 1976. Bat mortality: Pesticide poisoning and migratory stress. 

Science 19:184-186. 

*Genelly RE, Rudd RL. 1956. Effects of DDT, toxaphene, and dieldrin on pheasant reproduction. Auk 

73(4):529-539. 

*George VT, Sundararaj A. 1995. Effect of DDT on reproductive performance of white leghorn 

cockerels. Indian Vet J 72:694-697. 

*Gerhard I, Daniel V, Link S, et al. 1998. Chlorinated hydrocarbons in women with repeated 

miscarriages. Environ Health Perspect 106:675-681. 

*Gerhard I, Monga B, Krahe J, et al. 1999. Chlorinated hydrocarbons in infertile women. 

Environmental Research (Section A) 80:299-310. 

Geyer H, Scheunert I, Korte F. 1986. Bioconcentration potential of organic environmental chemicals in 

humans. Regul Toxicol Pharmacol 6:313-347. 

Geyer H, Scheunert I, Korte F. 1987. Correlation between the bioconcentration potential of organic 

environmental chemicals in humans and their–octanol/water partition coefficients. Chemosphere 

16(1):239-252. 

*Geyer H, Sheehan P, Kotzias D, et al. 1982. Prediction of ecotoxicological behaviour of chemicals: 

Relationship between physico-chemical properties and bioaccumulation of organic chemicals in the 

mussel Mytilus edulis. Chemosphere 11(11):1121-1134. 

*Gianessi LP, Puffer CA. 1992. Insecticide use in U.S. crop production. Washington, DC: Resources 

for the Future. USEPA Cooperative Agreement CR 813719. 

*Giesy JP, Verbrugge DA, Othout RA, et al. 1994. Contaminants in fishes from the Great lakesinfluenced sections and above dams of three Michigan rivers. I: Concentrations of organo chlorine 

insecticides, polychlorinated biphenyls, dioxin equivalents, and mercury. Arch Environ Contam Toxicol 

27:202-212. 
DDT, DDE, and DDD 338 

9. REFERENCES 

*Gill US, Schwartz HM, Wheatley B. 1996. Development of a method for the analysis of PCB 

congeners and organochlorine pesticides in blood/serum. Chemosphere 32(6):1055-1061. 

*Gillesby B, Zacharewski T. 1998. Exoestrogens: Mechanisms of action and strategies for identification 

and assessment. Environ Toxicol Chem 17(1):3-14. 

*Gillett JW, Porter RD, Wiemeyer SN, et al. 1970. Induction of liver microsomal activities in Japanese 

quail [with supplemental information on] The induction of microsomal ethylmorphine demethylase 

activity in sparrow hawk livers by diets containing DDT and dieldrin or DDE. In: Gillett JW, ed. The 

biological impact of pesticides in the environment. Environmental health science series no. 1. Corvallis: 

Oregon State University, 59-64. 

*Gilliom R. 1984. Pesticides in rivers of the United States. National Water Summary 1984. Paper 

2275:85-93. 

*Gillis CA, Bonnevie NL, Su SH, et al. 1995. DDT, DDD, and DDE contamination of sediment in the 

Newark Bay Estuary, New Jersey. Arch Environ Contam Toxicol 28:85-92. 

*Gingell R. 1975. Enterohepatic circulation of bis(p-chlorophenyl)acetic acid in the rat. Drug Metab 

Dispos 3:42-46. 

*Gingell R. 1976. Metabolism of [14C]-DDT in the mouse and hamster. Xenobiotica 6(1):15-20. 

*Gingell R, Wallcave L. 1974. Species differences in the acute toxicity and tissue distribution of DDT in 

mice and hamsters. Toxicol Appl Pharmacol 28:385-394. 

*Gish CD, Chura NJ. 1970. Toxicity of DDT to Japanese quail as influenced by body weight, breeding 

condition, and sex. Toxicol Appl Pharmacol 17:740-751. 

*Giwercman A, Carlsen E, Keiding N, et al. 1993. Evidence for increasing incidence of abnormalities of 

the human testis: A review. Environ Health Perspect Suppl 101(2):65-71. 

Gladen BC, Rogan WJ. 1991. Effects of perinatal polychlorinated biphenyls and dichlorodiphenyl 

dichloroethene and PDE on later development. J Pediatr 119(1 Part 1):58-63. 

*Gladen BC, Rogan WJ. 1995. DDE and shortened duration of lactation in a northern Mexican town. 

Am J Public Health 85(4):504-508. 

*Gladen BC, Rogan NB, Rogan WJ. 2000. Pubertal growth and development and prenatal and 

lactational exposure to polychlorinated biphenyls and dichlorodiphenyl dichloroethene. J Pediatr 

136(4):490-496. 

*Glazkov IN, Revelsky IA, Efimov IP, et al. 1999. Supercritical fluid extraction of water samples 

containing ultratrace amounts of organic micropollutants. J Microcolumn Sep 11:729-736 

*Glick B. 1974. Antibody-mediated immunity in the presence of mirex and DDT. Poult Sci 53:1476­

1485. 

*Glynn AW, Michaelsson K, Lind PM, et al. 2000. Organochlorines and bone mineral density in 

Swedish men from the general population. Osteoporos Int 11:1036-1042. 
DDT, DDE, and DDD 339 

9. REFERENCES 

*Gold B, Brunk G. 1982. Metabolism of 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane and 1,1-dichloro2,2-bis(p-chlorophenyl)ethane in the mouse. Chem Biol Interact 41:327-339. 

*Gold B, Brunk G. 1983. Metabolism of 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT), 1,1­

dichloro-2,2-bis(p-chlorophenyl)ethane, and 1-chloro-2,2-bis(p-chlorophenyl)ethene in the hamster. 

Cancer Res 43:2644-2647. 

*Gold B, Brunk G. 1984. A mechanistic study of the metabolism of 1,1-dichloro-2,2-bis(pchlorophenyl)ethane (DDD) to 2,2-bis(p-chlorophenyl)acetic acid (DDA). Biochem Pharmacol 

33(7):979-982. 

*Gold B, Leuschen T, Brunk G, et al. 1981. Metabolism of a DDT metabolite via a chloroepoxide. 

Chem Biol Interact 35:159-176. 

*Goldberg ED. 1975. Synthetic organohalides in the sea. Proc Royal Soc London B 189:277-289 

*Golden RJ, Noller KL, Titus-Ernstoff L, et al. 1998. Environmental endocrine modulators and human 

health: An assessment of the biological evidence. Crit Rev Toxicol 28(2):109-227. 

*Goldman M. 1981. The effect of a single dose of DDT on thyroid function in male rats. Arch Int 

Pharmacodyn 252:327-334. 

*Gomez-Catalan J, To-Figueras J, Rodamilans M, et al. 1991. Transport of organochlorine residues in 

the rat and human blood. Arch Environ Contam Toxicol 20:61-66. 

*Gore RC, Hannah RW, Pattacini SC, et al. 1971. Infrared and ultraviolet spectra of seventy-six 

pesticides. J Assoc Off Anal Chem 54(5):1040-82. 

*Graillot C, Gak J, Lancret C, et al. 1975. [Research on the toxic mechanism of action of organochlorine 

insecticides. II. Study of the long-term toxic effects of DDT on house hamsters.] Eur J Toxicol 

8:653-659. (French) 

Grandjean P, Bjerve KS, Weihe P, et al. 2001. Birthweight in a fishing community: significance of 

essential fatty acids and marine food contaminants. Int J Epidemiol 30:1272-1278. 

*Grasman KA, Scanlon PF, Fox GA. 1998. Reproductive and physiological effects of environmental 

contaminants in fish-eating birds of the Great Lakes: A review of historical trends. Environ Monit 

Assess 53:117-145. 

Grasso P, Hinton RH. 1991. Evidence for and possible mechanisms of non-genotoxic carcinogenesis in 

rodent liver. Mutat Res 248:271-290. 

*Grasso C, Bernardi G, Martinelli F. 1973. [The significance of the presence of chlorinated insecticides 

from the blood of newborns and in respective mothers]. Igiene Moderna 66:362-371. (Italian) 

Gray LE. 1998. Chemical-induced alterations of sexual differentiation: A review of effects in humans 

and rodents. J of Clean Technol Environ Toxicol Occup Med 7(2):121-145. 

Gray LE. 1998. Xenoendocrine disruptors: laboratory studies on male reproductive effects. Toxicol 

Lett 102-103:331-335. 
DDT, DDE, and DDD 340 

9. REFERENCES 

Gray LE, Otsby J. 1998. Effects of pesticides and toxic substances on behavioral and morphological 

reproductive development: Endocrine versus nonendocrine mechanisms. Toxicol Ind Health 14:159-183. 

*Gray L, Kelce W, Wiese T, et al. 1997. Endocrine screening methods workshop report: Detection of 

estrogenic and androgenic hormonal and antihormonal activity for chemicals that act via receptor or 

steroidogenic enzyme mechanisms. Reprod Toxicol 11(5):719-750. 

*Gray LE, Wolf C, Lambright C, et al. 1999. Administration of potentially antiandrogenic pesticides 

(procymidone, linuron, iprodione, chlozolinate, p,p'-DDE, and ketoconazole) and toxic substances 

(dibutyl- and diethylhexyl phthalate, PCB 169, and ethane dimethane sulphonate) during sexual 

differentiation produces diverse profiles of reproductive malformations in the male rat. Toxicol Ind 

Health 15(1-2):94-118. 

Gray R, Dinman B, Bernstein I. 1970. Ultrastructural abnormalities in rat liver after exposure to DDT. J 

Cell Biol 47:78. 

*Green V. 1969. Effects of pesticides on rat and chick embryo. In: Hemphill D, ed. Trace substances 

in environmental health. Proceedings of the University of Missouri 3rd Annual Conference. 2:183-209. 

Greenlee AR, Quail CA, Berg RL. 2000. The antiestrogen ICI 182,780 abolishes development injury for 

murine embryos exposed in vitro to o,p'-DDT. Reprod Toxicol 14(3):225-234. 

*Greichus YA, Hannon MR. 1973. Distribution and biochemical effects of DDT, DDD, and DDE in 

penned double-crested cormorants. Toxicol Appl Pharmacol 26:483-494. 

*Greichus YA, Call DJ, Ammann BM, et al. 1975. Physiological effects of polychlorinated biphenyls or 

a combination of DDT,DDD, and DDE in penned white pelicans. Arch Environ Contam Toxicol 3:330­

343. 

*Greizerstein HB, Gigliotti P, Vena J, et al. 1997. Standardization of a method for the routine analysis of 

polychlorinated biphenyl congeners and selected pesticides in human serum and milk. J Anal Toxicol 

21:558-566. 

*Griffen DE, Hill WE. 1978. In vitro breakage of plasmid DNA by mutagens and pesticides. Mutat Res 

52:161-169. 

*Grob K, Kalin I. 1991. Attempt for an on-line size exclusion chromatography-gas chromatography 

method for analyzing pesticide residues in foods. J Agric Food Chem 39:1950-1953. 

*Guan X, Ruch R. 1996. Gap junction endocytosis and lysosomal degradation of connexin43-p2 in WBF344 rat liver epithelial cells treated with DDT and lindane. Carcinogenesis 17(9):1791-1798. 

Guenthner T, Mannering G. 1977. Induction of hepatic mono-oxygenase systems in fetal and neonatal 

rats with phenobarbital, polycyclic hydrocarbons and other xenobiotics. Biochem Pharmacol 26:567-575. 

*Guillette LJ and Crain DA. 1996. Endocrine-disrupting contaminants and reproductive abnormalities in 

reptiles. Comments Toxicol 5(4-5):381-399. 

*Guillette LJ, Brock JW, Rooney AA, et al. 1999. Serum concentrations of various environmental 

contaminants and their relationship to sex steroid concentrations and phallus size in juvenile American 

alligators. Arch Environ Contam Toxicol 36:447-455. 
 

DDT, DDE, and DDD 341 

9. REFERENCES 

*Guillette LJ, Crain DA, Rooney AA, et al. 1997. Effect of acute stress on plasma concentrations of sex 

and stress hormones in juvenile alligators living in control and contaminated lakes. J Herpetol 31(3):347­

353. 

*Guillette LJ, Gross TS, Gross DA, et al. 1995. Gonadal steroidogenesis in vitro from juvenile alligators 

obtained from contaminated or control lakes. Environ Health Perspect Suppl 103(4):31-36. 

*Guillette LJ, Gross TS, Masson GR, et al. 1994. Developmental abnormalities of the gonad and 

abnormal sex hormone concentrations in juvenile alligators from contaminated and control lakes in 

Florida. Environ Health Perspect 102(8):680-688. 

*Guillette LJ, Pickford DB, Crain DA, et al. 1996. Reduction in penis size and plasma testosterone 

concentrations in juvenile alligators living in a contaminated environment. Gen Comp Endocrinol 

101:32-42. 

*Gunderson EL. 1995a. FDA total diet study, July 1986-April 1991, dietary intakes of pesticides, 

selected elements, and other chemicals. J Assoc Off Anal Chem 78(6):1353. 

*Gunderson EL. 1995b. Dietary intakes of pesticides, selected elements, and other chemicals: FDA total 

diet study, June 1984-April 1986. J Assoc Off Anal Chem 78(4):910-920. 

*Gupta PH, Mehta S, Mehta SK. 1989. Effect of DDT on hepatic mixed-function oxidase activity in 

normal and vitamin A-deficient rats. Biochem Int 19(2):247-256. 

*Guttes S, Failing K, Neumann K, et al. 1998. Chlororganic pesticides and polychlorinated biphenyls in 

breast tissue of women with benign and malignant breast disease. Arch Environ Contam Toxicol 35:140­

147. 

*Guyton AC, Hall JE eds. 2000. Pregnancy and lactation. In: Textbook of medical physiology. 10th ed. 

Philadelphia, PA: W.B. Saunders Company, 47-50. 

*Guzelian PS, Henry CJ, Olin SS, eds. 1992. Similarities and differences between children and adults: 

Implications for risk assessment. Washington, DC: International Life Sciences Institute Press. 

Haake J, Kelley M, Keys B, et al. 1987. The effects of organochlorine pesticides as inducers of 

testosterone and benzo(a)pyrene hydroxylases. Gen Pharmacol 18(2):165-169. 

*Hadley ME, ed. 1984. Endocrinology. Engelwood Cliffs, New Jersey: Prentice-Hall, Inc.,102. 

*Haegele MA, Hudson RH. 1973. DDE effects on reproduction of ring doves. Environ Pollut 4:53-57. 

*Haegele MA, Hudson RH. 1977. Reduction of courtship behavior induced by DDE in male ringed 

turtle doves. Wilson Bull 89(4):593-601. 

*Hamid J, Sayeed A, McFarlane H. 1974. The effect of 1-(o-chlorophenyl)-1-(p-chlorophenyl)-2,2­

dichloroethane (o,p'-DDD) on the immune response in malnutrition. Br J Exp Pathol 55:94. 

*Hanrahan LP, Falk C, Anderson HA, et al. 1999. Serum PCB and DDE levels of frequent Great Lakes 

sport fish consumers—a first look. Environ Res A80:S26-S37. 
DDT, DDE, and DDD 342 

9. REFERENCES 

Hansler RJ, Moisey J, Montie E, et al. 1999. PCBs, DDTs and methylsulfone PCB and 4,4'-DDE 

metabolites in cetaceans from the Atlantic Ocean. Organohalogen Compounds 42:197-2000. 

*Haraguchi K, Kuroki H, Masuda Y. 1989. Occurrence and distribution of chlorinated aromatic 

methylsulfones and sulfoxides in biological samples. Chemosphere 19(1-6):487-492. 

Harbison R. 1975. Comparative toxicity of some selected pesticides in neonatal and adult rats. Toxicol 

Appl Pharmacol 32:443-446. 

*Hardell L, Van Bavel B, Lindstrom G, et al. 1996. Higher concentrations of specific polychlorinated 

biphenyl congeners in adipose tissue from non-Hodgkin's lymphoma patients compared with controls 

without a malignant disease. Int J Oncol 9:603-608. 

Hardy ML. 2000. distribution of decabromodiphenyl oxide in the environment. Organohalogen 

Compounds 47:237-239. 

*Hargrave BT, Harding GC, Vass WP, et al. 1992. Organochlorine pesticides and polychlorinated 

biphenyls in the Arctic Ocean food web. Arch Environ Contam Toxicol 22:41-54. 

*Harner T. 1997. Organochlorine contamination of the Canadian Arctic, and speculation on future 

trends. Int J Environ Pollut 8(1⁄2):51-73. 

*Harri MNE, Laitinen J, Valkama E-L. 1979. Toxicity and retention of DDT in adult frogs, Rana 

temporaria L. Environ Pollut 20(1):45-55. 

*Harris ML, Wilson LK, Elliott JE, et al. 2000. Transfer of DDT and metabolites from fruit orchard soils 

to American robins (Turdus migratorius) twenty years after agricultural use of DDT in Canada. Arch 

Environ Contam Toxicol 39:205-220. 

Hart M, Fouts J. 1963. Effects of acute and chronic DDT administration on hepatic microsomal drug 

metabolism in the rat. Proc Soc Exp Biol Med 114:388-392. 

Hart M, Fouts J. 1965. Further studies on the stimulation of hepatic microsomal drug metabolizing 

enzymes by DDT and its analogs. Naunyn-Schmiedebergs Arch Exp Pathol Pharmacol 249:486-500. 

*Hart M, Straw JA. 1971. Studies on the site of action of o,p’-DDD in the dog adrenal cortex. 1. 

Inhibition of ACTH-mediated pregnenolone synthesis. Steroids 17(5):559-574. 

*Hart M, Adamson R, Fabro S. 1971. Prematurity and intrauterine growth retardation induced by DDT 

in the rabbit. Arch Int Pharmacodyn 192:286-290. 

*Hart M, Whang-Peng J, Sieber S, et al. 1972. Distribution and effects of DDT in the pregnant rabbit. 

Xenobiotica 2(6):567-574. 

Harvey GR, Giam CS. 1976. Polychlorobiphenyls and DDT compounds. In: Goldberg ED. Strategies 

for marine pollution monitoring. New York, NY: John Wiley and Sons, 35-46. 

Hatolkar VS, Pawar SS. 1992. Hepatic mixed function oxidase system and effect of phenobarbital, 

3-methylcholanthrene and 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane in developing chickens. Indian J 

Exp Biol 30:407-409. 
DDT, DDE, and DDD 343 

9. REFERENCES 

*Hayes W, ed. 1982. Chlorinated hydrocarbon insecticides. In: Pesticides studied in man. Baltimore, 

MD: Williams and Wilkins, 180-195. 

*Hayes W, Dale W, Pirkle C. 1971. Evidence of safety of long-term, high, oral doses of DDT for man. 

Arch Environ Health 22:119. 

*Hayes W, Durham W, Cueto C. 1956. The effect of known repeated oral doses of chlorinophenothane 

(DDT) in man. JAMA 162:890-897. 

*Haynes RJ. 1972. Effects of DDT on glycogen and lipid levels in bobwhites. J Wildl Manage 

36(2):518-523. 

HazDat. 1992. Agency for Toxic Substances and Disease Registry (ATSDR), Atlanta, GA. February 13, 

1992. 

*HazDat. 1999. Agency for Toxic Substances and Disease Registry (ATSDR), Atlanta, GA. April 14, 

1999. 

HazDat. 2001. Agency for Toxic Substances and Disease Registry (ATSDR), Atlanta, GA. 

*HazDat. 2002. Agency for Toxic Substances and Disease Registry (ATSDR), Atlanta, GA. 

*Heath RG, Spann JW, Kreitzer JF. 1969. Marked DDE impairment of mallard reproduction in 

controlled studies. Nature 224:47-48. 

*Heberer T, Dünnbier U. 1999. DDT metabolite bis(chlorophenyl)acetic acid: The neglected 

environmental contaminant. Environ Sci Technol 33(4):2346-2351. 

*Heinrichs WL, Gellert RJ, Bakke J. 1971. DDT administered to neonatal rats induces persistent estrus 

syndrome. Science 17:642-643. 

Heinz, G, Erdman T, Haseltine S, et al. 1985. Contaminant levels in colonial waterbirds from Green Bay 

and Lake Michigan, 1975-80. Environ Monit Assess 5:223-236. 

*Heinz GH, Hill EF, Contrera JF. 1980. Dopamine and norepinephrine depletion in ring doves fed DDE, 

dieldrin, and aroclor 1254. Toxicol Appl Pharmacol 53:75-82. 

*Helrich K, ed. 1990. Pesticide and industrial chemical residues. In: Official Methods of Analysis of 

the Association of Official Analytical Chemists. Arlington, VA: Association of Official Analytical 

Chemists, Inc., 274-284. 

*Helzlsouer KJ, Alberg AJ, Huang H-Y, et al. 1999. Serum concentrations of organochlorine 

compounds and the subsequent development of breast cancer. Cancer Epidemiol Biomarkers Prev 8:525­

532. 

Henderson G, Woolley D. 1969a. Tissue concentrations of DDT: Correlation with neurotoxicity in 

young and adult rats. Proc West Pharmacol Soc 12:58-62. 

*Henderson G, Woolley D. 1969b. Studies on the relative insensitivity of the immature rat to the 

neurotoxic effects of 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT). J Pharmacol Exp Ther 

170(1):173-180. 
DDT, DDE, and DDD 344 

9. REFERENCES 

*Henderson GL, Woolley DE. 1970. Mechanisms of neurotoxic action of 1,1,1-trichloro-2,2-bis(pchlorophenyl)ethane (DDT) in immature and adult rats. J Pharmacol Exp Ther 175(1):60-68.
 

*Herr D, Tilson H. 1987. Modulation of p,p'-DDT-induced tremor by catecholaminergic agents. 

Toxicol Appl Pharmacol 91:149-158.
 

*Herr D, Hong J-S, Chen P, et al. 1986. Pharmacological modification of DDT-induced tremor and
 

hyperthermia in rats: Distributional factors. Psychopharmacology 89:278-283.
 

*Herr D, Hong J-S, Tilson H. 1985. DDT-induced tremor in rats: Effects of pharmacological agents. 

Psychopharmacology 86:426-431.
 

*Hietanen E, Vainio H. 1976. Effect of administration route of DDT on acute toxicity and drug 

biotransformation in various rodents. Arch Environ Contam Toxicol 4:201-216. 

Higginson J. 1983. DDT: Epidemiological evidence. IARC Sci Publ 65:107-117.
 

*Hill EF, Dale WE, Miles JW. 1971. DDT intoxication in birds: Subchronic effects and brain residues. 

Toxicol Appl Pharmacol 20:502-514.
 

*Hill K, Robinson G. 1945. A fatal case of DDT poisoning in a child. Br Med J 2:845.
 

*Hilpert DW, Romen W, Newmann HG. 1983. The role of partial hepatectomy and of promoters in the
 

formation of tumors in non-target tissues of trans-4-acetylaminostilbene in rats. Carcinogen 4:1519-1525.
 

*Hitch RK, Day HR. 1992. Unusual persistence of DDT in some western USA soils. Bull Environ
 

Contam Toxicol 48:259-264.
 

Hodgson JT, Jones RD. 1990. Mortality of a cohort of tin miners 1941-1986. Br J Ind Med 47:665-676.
 

*Hoel DG, Davis DL, Miller AB, et al. 1992. Trends in cancer mortality in 15 industrialized countries,
 

1969-1986. J Natl Cancer Inst 84(5):313-320. 

*Hoff R, Strachan W, Sweet C, et al. 1996. Atmospheric deposition of toxic chemicals to the Great
 

Lakes: A review of data through 1994. Atmos Environ 30(20):3505-3527.
 

Holt R, Cruse S, Greer E. 1986. Pesticide and polychlorinated biphenyl residues in human adipose tissue
 

from northeast Louisiana. Bull Environ Contam Toxicol 36:651-655.
 

*Hong J, Herr D, Hudson P, et al. 1986. Neurochemical effects of DDT in rat brain in vivo. Arch
 

Toxicol 9:14-26.
 

*Hopper ML. 1991. Analysis of organochlorine pesticide residues using simultaneous injection of two
 

capillary columns with electron capture and electrolytic conductivity detectors. J Assoc Off Anal Chem
 

74(6):974-981.
 

*Hoppin JA, Tolbert PE, Holly EA, et al. Pancreatic cancer and serum organochlorine levels. Cancer
 

Epidemiol Biomark Prev 9:199-205.
 

*Hovinga M, Sowers M, Humphrey HEB. 1992. Historical changes in serum PCB and DDT levels in an 

environmentally-exposed cohort. Arch Environ Contam Toxicol 22:362-366. 
DDT, DDE, and DDD 345 

9. REFERENCES 

*Hovinga M, Sowers M, Humphrey H. 1993. Environmental exposure and lifestyle predictors of lead, 

cadmium, PCB, and DDT levels in Great Lakes fish eaters. Arch Environ Health 48(2):98-104. 

*Howard P, Meylan W, eds. 1997. Handbook of physical properties of organic chemicals. Boca Raton, 

FL: CRC Press, Lewis Publishers, 3, 18, 49,518,723. 

*Howard P, Neal M, eds. 1992. Dictionary of chemical names and synonyms. Ann Arbor, MI: Lewis 

Publishers, I558,I769,I24-25. 

*Hoyer AP, Grandjean P, Jorgensen T, et al. 1998. Organochlorine exposure and risk of breast cancer. 

Lancet 352:1816-1820. 

*Høyer AP, Jørgensen T, Grandjean P, et al. 2000. Repeated measurements of organochlorine exposure 

and breast cancer risk (Denmark). Cancer Causes Control 11:177-184. 

*Hrdina P, Singhal R. 1972. Neurotoxic effects of DDT: Protection by cycloheximide. J Pharm 

Pharmacol 24:167-169. 

*Hrdina P, Singhal R, Peters D, et al. 1972. Comparison of the chronic effects of p,p'-DDT and 

a-chlordane on brain amines. Eur J Pharmacol 20:114-117. 

*Hrdina P, Singhal R, Peters D, et al. 1973. Some neurochemical alterations during acute DDT 

poisoning. Toxicol Appl Pharmacol 25:276-288. 

*HSDB. 1988. Hazardous Substances Data Bank. National Library of Medicine, National Toxicology 

Information Program, Bethesda, MD. November 11, 1988. 

HSDB. 1999a. p,p’-DDT - 50-29-3. Hazardous Substances Data Bank. National Library of Medicine, 

National Toxicology Information Program, Bethesda, MD. April 16, 1999. 

HSDB. 1999b. p,p’-DDD - 72-54-8. Hazardous Substances Data Bank. National Library of Medicine, 

National Toxicology Information Program, Bethesda, MD. April 16, 1999. 

HSDB. 1999c. o,p’-DDD - 53-19-0. Hazardous Substances Data Bank. National Library of Medicine, 

National Toxicology Information Program, Bethesda, MD. April 16, 1999. 

HSDB. 1999d. p,p’-DDE - 72-55-9. Hazardous Substances Data Bank. National Library of Medicine, 

SRI 1982; National Toxicology Information Program, Bethesda, MD. April 16, 1999. 

*HSDB. 2001a. p,p’-DDT- 50-29-3. Hazardous Substances Data Bank. National Library of Medicine, 

National Toxicology Information Program, Bethesda, MD. http://toxnet.nlm.hih.gov/cgibin/sis/htmlgen?HSDB. September 05, 2001. 

*HSDB. 2001b. p,p’-DDD - 72-54-8. Hazardous Substances Data Bank. National Library of Medicine, 

National Toxicology Information Program, Bethesda, MD. http://toxnet.nlm.hih.gov/cgibin/sis/htmlgen?HSDB. September 05, 2001. 

*HSDB. 2001c. o,p’-DDD - 53-19-0. Hazardous Substances Data Bank. National Library of Medicine, 

National Toxicology Information Program, Bethesda, MD. http://toxnet.nlm.hih.gov/cgibin/sis/htmlgen?HSDB. September 05, 2001. 
DDT, DDE, and DDD 346 

9. REFERENCES 

*HSDB. 2001d. p,p’-DDE - 72-55-9. Hazardous Substances Data Bank. National Library of Medicine, 

National Toxicology Information Program, Bethesda, MD. http://toxnet.nlm.hih.gov/cgibin/sis/htmlgen?HSDB. September 05, 2001. 

*Hsieh H. 1954. DDT intoxication in a family of Southern Taiwan. Am Med Assoc Arch Ind Hyg 

10:344-346. 

Hubert A, Wenzel K-D, Engelwald W, et al. 2001. accelerated solvent extraction - more efficient 

extraction of POPs and PAHs from real contaminated plant and soil samples. 20(2):101-144. 

*Hudson P, Chen P, Tilson H, et al. 1985. Effects of p,p'-DDT on the rat brain concentrations of 

biogenic amine and amino acid neurotransmitters and their association with p,p'-DDT-induced tremor and 

hyperthermia. J Neurochem 45(5):1349-1355. 

*Hundley HK, Cairns T, Luke MA, et al. 1988. Pesticide residue findings by the Luke Method in 

domestic and imported foods and animal feeds for fiscal years 1982-1986. J Assoc Off Anal Chem 

71:875-892. 

*Hunter D, Hankinson S, Laden F, et al. 1997. Plasma organochlorine levels and the risk of breast 

cancer. N Engl J Med 337(18):1253-1257. 

*Hurst JG, Newcomer WS, Morrison JA. 1974. Some effects of DDT, toxaphene, and polychlorinated 

biphenyl on thyroid function in bobwhite quail. Poult Sci 53:125-133. 

Hussain A, Tirmazi S, Maqbool U, et al. 1994. Studies of the effects of temperatures and solar radiation 

on volatilization, mineralization and binding of 14C-DDT in soil under laboratory conditions. J Environ 

Sci Health B29(1):141-151. 

*Hwang E, Van Woert M. 1978. p,p'-DDT-induced neurotoxic syndrome: Experimental myoclonus. 

Neurology 28:1020-1025. 

Hwang E, Van Woert M. 1979. p,p'-DDT-induced myoclonus: Serotonin and alpha noradrenergic 

interaction. Res Commun Chem Pathol Pharmacol 23:257-266. 

*Hyland J, Snoots T, Balthis WL. 1998. Sediment quality of estuaries in the southeastern U.S. Environ 

Monit Assess 51:331-343. 

IARC. 1973. IARC monographs of the evaluation of the carcinogenic risk of chemicals to man: Some 

organochlorine pesticides. Vol 5. Lyon, France: World Health Organization, International Agency for 

Research on Cancer. 

IARC. 1987a. IARC monographs on the evaluation of the carcinogenic risk to humans: Summary table. 

Suppl. 7:66. Lyon, France: World Health Organization, International Agency for Research on Cancer. 

IARC. 1987b. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans: Overall 

Evaluations of Carcinogenicity: An Updating of IARC Monographs. Lyon, France: International 

Agency for Research on Cancer, Supplement 7:186. 

IARC. 1991. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans: Occupational 

exposures in insecticide application, and some pesticides. Lyon, France: International Agency for 

Research on Cancer, Vol 53:193 
DDT, DDE, and DDD 347 

9. REFERENCES 

*IARC. 2002. DDT and associated compounds (Group 2B). International Agency for Research on 

Cancer. http://193.51.164.11/htdocs/Monographs/Vol53/04-DDT.HTM. January 02, 2002. 

*Innes J, Ulland B, Valerio M, et al. 1969. Bioassay of pesticides and industrial chemicals for 

tumorigenicity in mice: A preliminary note. J Natl Cancer Inst 42(6):1101. 

International Labour Office. 1983. Encyclopedia of occupational health and safety. Vols. I & II. 

Geneva, Switzerland, 1636. 

*IRIS. 2001a. DDD. Integrated Risk Information System. U.S. Environmental Protection Agency. 

http://www.epa.gov/iris/subst/0347.htm. September 05, 2001. 

*IRIS. 2001b. DDE. Integrated Risk Information System. U.S. Environmental Protection Agency. 

http://www.epa.gov/iris/subst/0328.htm. September 05, 2001. 

*IRIS. 2001c. DDT. Integrated Risk Information System. U.S. Environmental Protection Agency. 

http://www.epa.gov/iris/subst/0147.htm. September 05, 2001. 

*Iwata H, Tanabe S, Sakai N, et al. 1993. Distribution of persistent organochlorines in the oceanic air 

and surface seawater and the role of ocean on their global transport and fate. Environ Sci Technol 

27:1080-1098. 

*Janak K, Becker G, Colmsjo A, et al. 1998. Methyl sulfonyl polychlorinated biphenyls and 2,2-bis(4­

chlorophenyl)-1,1-dichlorethene in gray seal tissues determined by gas chromatography with electron 

capture detection and atomic emission detection. Environ Toxicol Chem 17(6):1046-1055. 

*Jansen L, Jongen W. 1996. The use of initiated cell as a test system for the detection of inhibitors of 

gap junctional intercellular communication. Carcinogenesis 17(2):333-339. 

Jarman W, Springer A, Walker W. 1986. A metabolic derivation of DDE from kelthane. Environ 

Toxicol Chem 5:13-19. 

*Jefferies DJ. 1967. The delay in ovulation produced by pp'-DDT and its possible significance in the 

field. Ibis 109:266-272. 

*Jefferies DJ. 1969. Induction of apparent hyperthyroidism in birds fed DDT. Nature 222:578-579. 

*Jefferies DJ. 1972. Organochlorine insecticide residues in British bats and their significance. J Zool 

166:245-263. 

*Jefferies DJ, French MC. 1969. Avian thyroid: Effect of p,p'-DDT on size and activity. Science 

166:1278-1280. 

*Jefferies DJ, French MC. 1971. Hyper- and hypothyroidism in pigeons fed DDT: An explanation for 

the 'thin eggshell phenomenon'. Environ Pollut 1:235-242. 

*Jefferies DJ, French MC. 1972. Changes induced in the pigeon thyroid by p,p'-DDE and dieldrin. J 

Wildl Manage 36(1):24-30. 

*Jefferies DJ, Walker CH. 1966. Uptake of p,p'-DDT and its post-mortem breakdown in the avian liver. 

Nature 212:533-534. 
DDT, DDE, and DDD 348 

9. REFERENCES 

*Jefferies DJ, French MC, Osborne BE. 1971. The effect of pp'-DDT on the rate, amplitude and weight 

of the heart of the pigeon and bengalese finch. Br Poult Sci 12:387-399. 

*Jensen J, Cueto C, Dale W, et al. 1957. DDT metabolites in feces and bile of rats. J Agric Food Chem 

5(12):919-925. 

*Jensen S, Jansson B. 1976. Methyl sulfone metabolites of PCB and DDE. Ambio 5:257-260. 

Jeyaratnam J, Forshaw P. 1974. A study of the cardiac effects of DDT in laboratory animals. Bull 

World Health Organ 51:531-535. 

*Johanson CE. 1980. Permeability and vascularity of the developing brain: Cerebellum vs cerebral 

cortex. Brain Res 190:3-16. 

*Johansson M, Larsson C, Bergman A, et al. 1998. Structure-activity relationship for inhibition of 

CYP11B1-dependent glucocorticoid synthesis in Y1 cells by aryl methyl sulfones. Pharmacol Toxicol 

83:225-230. 

*Johansson U, Fredriksson A, Eriksson P. 1995. Bioallethrin causes permanent changes in behavioural 

and muscarinic acetylcholine receptor variables in adult mice exposed neonatally to DDT. Eur J 

Pharmacol 293:159-166. 

*Johansson U, Fredriksson A, Eriksson P. 1996. Low-dose effects of paraoxon in adult mice exposed 

neonatally to DDT: Changes in behavioural and cholinergic receptor variables. Environ Toxicol 

Pharmacol 2:307-314. 

*Johnsen R. 1976. DDT metabolism in microbial systems. Residue Rev 61:1-28. 

*Johnson G, Jalal S. 1973. DDT-induced chromosomal damage in mice. J Hered 64:7-8. 

*Johnson A, Norton D, Yake B. 1988b. Persistence of DDT in the Yakima River drainage, Washington. 

Arch Environ Contam Toxicol 17:289-297. 

Johnson DC. 1996. Estradiol-chlordecone (kepone) interactions: additive effect of combinations for 

utertropic and embryo implantation functions. Toxicol Lett 89:57-64. 

Johnson DC, Kogo H, Sam M, et al. 1988a. Multiple estrogenic action of o,p'-DDT: Initiation and 

maintenance of pregnancy in the rat. Toxicology 53:79-87. 

*Johnson DC, Sen M, Dey SK. 1992. Differential effects of dichlorodiphenyl-trichloroethane analogs, 

chlordecone, and 2,3,7,8-tetrachlorodibenzo-p-dioxin on establishment of pregnancy in the 

hypophysectomized rat. Proc Soc Exp Biol Med 199:42-48. 

*Jones FJS, Summers DDB. 1968. Relation between DDT in diets of laying birds and viability of their 

eggs. Nature 217:1162-1163. 

*Jones JL, Roberts LM. 1999. The relative merits of monitoring and domestic wells for ground water 

quality investigations. Ground Water Monit Remed 19(3):138-144. 

*Jonsson C-J. 1994. Decreased plasma corticosterone levels in suckling mice following injection of the 

adrenal toxicant, MeSO2-DDE, to the lactating dam. Pharmacol Toxicol 74:58-60. 
DDT, DDE, and DDD 349 

9. REFERENCES 

*Jonsson C-J, Lund B-O. 1994. In vitro bioactivation of the environmental pollutant 3-methylsulphonyl2,2-bis(4-chlorophenyl)-1,1-dichloroethene in the human adrenal gland. Toxicol Lett 71:169-175. 

*Jonsson C-J, Lund B-O, Bergman A, et al. 1992. Adrenocortical toxicity of 3-methylsulphonyl- DDE; 

3: Studies in fetal and suckling mice. Reprod Toxicol 6:233-240. 

*Jonsson C-J, Lund BO, Brunstrom B, et al. 1994. Toxicity and irreversible binding of two DDT 

metabolites-3-methylsulfonyl-DDE and o,p'-DDD- in adrenal interrenal cells in birds. Environ Toxicol 

Chem 13(8):1303-1310. 

*Jonsson C-J, Rodriguez-Martinez H, Lund B-O, et al. 1991. Adrenocortical toxicity of 

3-methylsulfonyl-DDE in mice: II. Mitochondrial changes following ecologically relevant doses. 

Fundam Appl Toxicol 16:365-374. 

*Jonsson H, Keil J, Gaddy R, et al. 1976. Prolonged ingestion of commercial DDT and PCB; effects on 

progesterone levels and reproduction in the mature female rat. Arch Environ Contam Toxicol 3:479-490. 

*Jonsson H, Walker E, Greene W, et al. 1981. Effects of prolonged exposure to dietary DDT and PCB 

on rat liver morphology. Arch Environ Contam Toxicol 10:171-183. 

*Jonsson J, Rodriguez-Martinez H, Brandt I. 1995. Transplacental toxicity of 3-methylsulphonyl-DDE 

in the developing adrenal cortex in mice. Reprod Toxicol 9(3):257-264. 

Joy RM. 1982. Chlorinated hydrocarbon insecticides. In: Ecobichon DJ, Joy RM, eds. Pesticides and 

neurological diseases. Boca Raton, FL: CRC Press, 91-150. 

Juberg D, Loch-Caruso R. 1991. Increased contraction frequency in rat uterine strips treated in vitro 

with o,p'-DDT. Bull Environ Contam Toxicol 46:751-755. 

*Juberg D, Loch-Caruso R. 1992. Investigation of the role of estrogenic action and prostaglandin E2 in 

DDT-stimulated rat uterine contractions ex vivo. Toxicology 74:161-172. 

*Juberg D, Stuenkel E, Loch-Caruso R. 1995. The chlorinated insecticide 1,1-dichloro-2,2-bis(4­

chlorophenyl)ethane (p,p'-DDD) increases intracellular calcium in rat myometrial smooth muscle cells. 

Toxicol Appl Pharmacol 135:147-155. 

*Kale SP, Murthy NBK, Raghu K, et al. 1999. Studies on degradation of 14C-DDT in the marine 

environment. Chemosphere 39:959-968 

*Kallenborn R, Oehme M, Wynn-Williams D, et al. 1998. Ambient air levels and atmospheric longrange transport of persistent organochlorines to Signy Island, Antarctica. Sci Total Environ 220:167-180. 

Kamel F, Boyes WK, Gladen BC, et al. 2000. Retinal degeneration in licensed pesticide applicators. 

Am J Ind Med 37:618-628. 

Kaminski N, Wells D, Dauterman W, et al. 1986. Macrophage uptake of a lipoprotein-sequestered 

toxicant: A potential route of immunotoxicity. Toxicol Appl Pharmacol 82:474-480. 

*Kan-Do Office and Pesticides Team. 1995. Accumulated pesticide and industrial chemical findings 

from a ten-year study of ready-to-eat foods. J Assoc Off Anal Chem 78(3):614-631. 
DDT, DDE, and DDD 350 

9. REFERENCES 

Kaneko T, Horiuchi J, Sato A. 2000. Development of a physiologically based pharmacokinetic model of 

organic solvent in rats. Pharm Res 42(5):465-470. 

*Kang K, Wilson M, Hayashi T, et al. 1996. Inhibition of gap junctional intercellular communication in 

normal human breast epithelial cells after treatment with pesticides, PCBs, and PBBs, alone or in 

mixtures. Environ Health Perspect 104(2):192-200. 

Kanja LW, Skaare JU, Ojwang SBO, et al. 1992. A comparison of organochlorine pesticide residues in 

maternal adipose tissue, maternal blood, cord blood, and human milk and mother/infant pairs. Arch 

Environ Contam Toxicol 22:21-24. 

Kapoor IP, Metcalf RL, Nystrom RF, et al. 1970. Comparative metabolism of methoxychlor, 

methiochlor, and DDT in mouse, insects, and in a model ecosystem. J Agric Food Chem 18(6):1145­

1152. 

*Kar P, Dikshith T. 1970. Dermal toxicity of DDT. Experientia 26:634-635. 

*Karlsson B, Persson B, Sodergren A, et al. 1974. Locomotory and dehydrogenase activities of redstarts 

Phoenicurus phoenicurus L. (aves) given PCB and DDT. Environ Pollut 7:53-63. 

*Karmaus W, Ajakevich S, Indurkhy A, et al. 2002. Childhood growth and exposure to 

dichlorodiphenyl dichloroethene and polychlorinated biphenyls. J Pediatr 140:33-39. 

Karmaus W, DeKoning EP, Kruse H, et al. 2001. Early childhood determinants of organochlorine 

concentrations in school-aged children. Pediatr Res 50(3):331-336. 

*Kashyap S, Nigam S, Karnik A, et al. 1977. Carcinogenicity of DDT (dichlorodiphenyl 

trichloroethane) in pure inbred Swiss mice. Int J Cancer 19:725-729. 

Kasperlik-Zaluska AA. 2000. Clinical results of the use of mitonane for adreocortical carcinoma. Braz J 

Med Biol Res 33:1191-1196. 

Kearney JP, Cole DC, Ferron LA, et al. 1999. Blood PCB, p,p'-DDE, and mirex levels in Great Lakes 

fish and waterfowl consumers in two Ontario communities. Environ Res A80:S138-S149. 

*Keith JO, Mitchell CA. 1993. Effects of DDE and food stress on reproduction and body condition of 

ringed turtle doves. Arch Environ Contam Toxicol 25:192-203. 

*Keith L, Garrison A, Allen R, et al. 1979. Identification of organic compounds in drinking water from 

thirteen U.S. cities. In: Keith L, ed. Identification and analysis of organic pollutants in water. Ann 

Arbor, MI: Ann Arbor Science, 329-373. 

Kelce WR, Gray LE, Wong C, et al. 1998. Developmental effects and molecular mechanisms of 

environmental antiandrogens. In: Eisenbrand G, ed. Hormonally active agents in food: Symposium. 

Weinheim, Germany: Wiley-VCH, 128-141. 

*Kelce W, Lambright C, Gray E, et al. 1997. Vinclozolin and p,p'-DDE alter androgen-dependent gene 

expression: In vivo confirmation of an androgen receptor-mediated mechanism. Toxicol Appl Pharmacol 

142:192-200. 
DDT, DDE, and DDD 351 

9. REFERENCES 

*Kelce W, Stone C, Laws S, et al. 1995. Persistent DDT metabolite p,p’-DDE is a potent androgen 

receptor antagonist. Nature 375:581-585. 

*Keller W, Yeary R. 1980. A comparison of the effects of mineral oil, vegetable oil, and sodium sulfate 

on the intestinal absorption of DDT in rodents. Clin Toxicol 16(2):223-231. 

*Kelly-Garvert F, Legator M. 1973. Cytogenetic and mutagenic effects of DDT and DDE in a Chinese 

hamster cell line. Mutat Res 17:223-229. 

*Kenney AD, Dacke CG, Wagstaff DJ, et al. 1972. Effects of DDT on calcium metabolism in the 

Japanese quail. Trace Subst Environ Health :247-255. 

*Keplinger M, Deichmann W, Sala F. 1970. Effects of combinations of pesticides on reproduction in 

mice. In: Deichmann W, ed. Pesticide symposia. 6th-7th International-American Conference of 

Toxicology and Occupational Medicine, University of Miami Medical School, Miami, FL, 125-138. 

Keplinger M, Dressler I, Deichmann W. 1966. Tissue storage in mice fed combinations of pesticides. 

Toxicol Appl Pharmacol 8:344. 

*Khairy M. 1959. Changes in behaviour associated with a nervous system poison (DDT). Q J Exp 

Psychol 11:94-91. 

*Khan IA, Thomas P. 1998. Estradiol-17β and o,p’-DDT stimulate gonadotropin release in Atlantic 

croaker. Marine Environmental Research 46:149-152. 

*Khanna N, Santerre C, Xu D, et al. 1997. Changes in dieldrin and p,p’-DDE residues following 

cooking of channel fish. J Food Prot 60(3):300-304. 

*Kidd KA, Bootsma HA, Hesslein RH, et al. 2001. Biomagnification of DDT through the benthic and 

pelagic food webs of Lake Malawi, East Africa: Importance of trophic level and carbon source. Environ 

Sci Technol 35:14-20 

*Kihlstrom J, Lundberg C, Orberg J, et al. 1975. Sexual functions of mice neonatally exposed to DDT or 

PCB. Environ Physiol Biochem 5:54-57. 

*Kimbrough R, Gaines T, Sherman J. 1964. Nutritional factors, long-term DDT intake, and 

chloroleukemia in rats. J Natl Cancer Inst 33:215-225. 

*Kiriluk R, Servos M, Whittle D, et al. 1995. Using ratios of stable nitrogen and carbon isotopes to 

characterize the biomagnification of DDE, mirex, and PCB in a Lake Ontario pelagic food web. Can J 

Fish Aquat Sci 52:2660-2674. 

*Kirk G, Jensen H. 1975. Toxic effects of o,p'-DDD in the normal dog. J Am Anim Hosp Assoc 

11:765-768. 

*Kirk G, Boyer S, Hutcheson D. 1974. Effects of o,p'-DDD on plasma cortisol levels and histology of 

the adrenal gland in the normal dog. J Am Anim Hosp Assoc 10:179-182. 

*Kitagawa T, Hino O, Nomura K, et al. 1984. Dose-response studies on promoting and anticarcinogenic 

effects of phenobarbital and DDT in the rat hepatocarcinogenesis. Carcinogenesis 5:1653-1656. 
DDT, DDE, and DDD 352 

9. REFERENCES 

*Kitchin K, Brown J. 1988. Biochemical effects of DDT and DDE in rat and mouse liver. Environ Res 

46:39-47. 

*Klaassen C. 1990. Principles of toxicology. In: Gilman A, Rall T, Nies A, et al., eds. The 

pharmacological basis of therapeutics. 8th ed. New York, NY: Pergamon Press, 49-61. 

Klauser S, Gantner D, Salgam P, et al. 1991. Structure-function studies of designed DDT-binding 

polypeptides. Biochem Biophys Res Commun 179(3):1212-1219. 

*Klein A, Laug E, Datta P, et al. 1964. Metabolites: Reductive dechlorination of DDT to DDD and 

isomeric transformation of o,p`-DDT to p,p'-DDT in vivo. J Assoc Off Agric Chem 47:1129-1145. 

*Kolmodin B, Azarnoff D, Sjoqvist F. 1969. Effect of environmental factors on drug metabolism: 

Decreased plasma half-life of antipyrine in workers exposed to chlorinated hydrocarbon insecticides. 

Clin Pharmacol Ther 10:638-642. 

*Komori M, Nishio K, Kitada M, et al. 1990. Fetus-specific expression of a form of cytochrome P-450 

in human livers. Biochemistry 29:4430-4433. 

*Koner B, Banergee B, Ray A. 1998. Organochlorine pesticide-induced oxidative stress and immune 

suppression in rats. Indian J Exp Biol 36:395-398. 

*Kornbrust D, Gillis B, Collins B, et al. 1986. Effects of 1-dichloro-2,2-bis[p-chlorophenyl]ethane 

(DDE) on lactation in rats. J Toxicol Environ Health 17:23-36. 

*Korrick SA, Altshul LM, Tolbert PE, et al. 2000a. Measurement of PCBs, DDE, and 

hexachlorobenzene in cord blood from infants born in towns adjacent to a PCB-contaminated waste site. 

J Expo Anal Environ Epidemiol 10:743-754. 

Korrick S, Damokosh A, Chen C, et al. 2000b. Serum DDT and DDE levels and spontaneous abortion 

[Abstract]. Am J Epidemiol 151(12):S24. 

Koschier F, Gigliotti P, Hong S. 1980. The effect of bis(p-chlorophenyl) acetic acid on the renal 

function of the rat. J Environ Pathol Toxicol 4:209-217. 

*Kostka G, Palut D, Kopec-Szlezak J, et al. 2000. Early hepatic changes in rats induced by permethrin 

in comparison with DDT. Toxicology 142:135-143. 

Kramer VJ, Giesy JP. 1999. Specific binding of hydroxylated polychlorinated biphenyl metabolites and 

other substances to bovine calf uterine estrogen receptor: structure-binding relationships. Sci Total 

Environ 233:141-161. 

Krantzberg G, Boyd D. 1991. The biological significance of contaminants in sediment from Hamilton 

Harbor Lake Ontario Canada. Seventeenth annual aquatic toxicity workshop, Vancouver, British 

Columbia, Canada, November 5-7, 1990. Can Tech Rep Fish Aquat Sci 1774:847-884. 

*Kratzer CR. 1999. Transport of sediment-bound organochlorine pesticides to the San Joaquin River, 

California. J Am Water Resources Assoc 35:957-981 

*Krause W. 1977. Influence of DDT, DDVP and malathion on FSH, LH and testosterone serum levels 

and testosterone concentration in testis. Bull Environ Contam Toxicol 18(2):231-242. 
DDT, DDE, and DDD 353 

9. REFERENCES 

*Krause W, Hamm K, Weissmuller J. 1975. The effect of DDT on spermatogenesis of the juvenile rat. 

Bull Environ Contam Toxicol 14(2):171-179. 

*Krauthacker B, Alebic-Kolbah T, Buntic A, et al. 1980. DDT residues in samples of human milk, and 

in mothers' and cord blood serum, in a continental town in Croatia (Yugoslavia). Int Arch Occup Environ 

Health 46:267-273. 

*Krauthaker B, Reiner E, Votava-Raic A, et al. 1998. Organochlorine pesticides and PCBs in human 

milk collected from mothers nursing hospitalized children. Chemosphere 37(1):27-32. 

Kraybill H. 1977. Global distribution of carcinogenic pollutants in water. Ann NY Acad Sci 298:80-89. 

*Kreiss K, Zack MM, Kimbrough RD, et al. 1981. Association of blood pressure and polychlorinated 

biphenyl levels. JAMA 245(24):2505-2509. 

*Krieger N, Wolff M, Hiatt R, et al. 1994. Breast cancer and serum organochlorines: A prospective 

study among white, black, and Asian women. J Natl Cancer Inst 86(8):589-599. 

*Krishnan K, Andersen ME. 1994. Physiologically-based pharmacokinetic modeling in toxicology. In: 

Hayes AW, ed. Principles and methods of toxicology. 3rd. Ed. New York, NY: Raven Press, Ltd.,399­

437. 

*Krishnan K, Andersen ME, Clewell H, et al. 1994. Physiologically based pharmacokinetic modeling of 

chemical mixtures. In: Yang R, ed. Toxicology of chemical mixtures: Case studies, mechanism, and 

novel approaches. New York, NY: Academic Press, 399-437. 

Krstevska-Konstantinova M, Charlier C, Craen M, et al. 2001. Sexual precocity after immigration from 

developing countries to Belgium: evidence of previous exposure to organochlorine pesticides. Hum 

Reprod 16(5):1020-1026. 

*Kubinski H, Nigam SK, Karnick AB, et al. 1981. DNA cell-binding (DCB) assay for suspected 

carcinogens and mutagens. Mutat Res 89:95-136. 

Kuiper GGJM, Carlsson B, Grandien K et al. 1997. Comparison of the ligand binding specificity and 

transcript tissue distribution of estrogen receptors α and β. Endocrinology 138:863-870. 

*Kuiper GGJM, Enmark E, Pelto-Huikko M et al. 1996. Cloning of a novel estrogen receptor expressed 

in rat prostate and ovary. Proc Natl Acad Sci USA 93:5925-5930 

Kuiper G, Lemmen J, Carlsson V, et al. 1998. Interaction of estrogenic chemicals and phytoestrogens 

with estrogen receptor β. Endocrinology 139(10):4252-4263. 

*Kuhnlein H, Receveur O, Muir D, et al. 1995. Arctic indigenous women consume greater than 

acceptable levels of organochlorines. Bethesda, MD: American Institute of Nutrition, 2501-2510. 

*Kurtz D. 1977. Adsorption of PCB’s and DDT’s on membrane filters- A new analysis method. Bull 

Environ Contam Toxicol 17(4):391-398. 

*Kutz F, Strassman S, Sperling J. 1979. Survey of selected organochlorine pesticides in the general 

population of the United States: Fiscal years 1970-75. Ann NY Acad Sci 320:60-68. 
DDT, DDE, and DDD 354 

9. REFERENCES 

*Kutz F, Wood P, Bottimore D. 1991. Organochlorine pesticides and polychlorinated biphenyls in 

human adipose tissue. Rev Environ Contam Toxicol 120:1-82. 

*Kutz F, Yobs A, Strassman S, et al. 1977. Pesticides in people: Effects of reducing DDT usage on total 

DDT storage in humans. Pestic Monit J 11(2):61-63. 

Lacorte S, Guiffard I, Fraisse D, et al. 2000. Broad spectrum analysis of 109 priority compounds listed 

in the 76/464/CEE council directive using solid-phase extraction and GC/EI/MS. Anal Chem 72:1430­

1440. 

*Laden F, Collman G, Iwamoto K, et al. 2001a. 1,1-Dichloro-2,2-bis(p-chlorophenyl)ethylene and 

polychlorinated biphenyls and breast cancer: Combined analysis of five U.S. studies. J Natl Cancer Inst 

93(10):768-776. 

*Laden F, Hankinson SE, Wolff MS, et al. 2001b. Plasma organochlorine levels and the risk of breast 

cancer: An extended follow-up in the nurses' health study. Indian J Cancer 91:568-574. 

Laden F, Hunter D, Cantor K, et al. 2000. Non-Hodgkin's lymphoma and plasma levels of DDE and 

PCBs [Abstract]. Am J Epidemiol 151(12):S23. 

*Laden F, Neas L, Spiegelman D, et al. 1999. Predictors of plasma concentrations of DDE and PCBs in 

a group of U.S. women. Environ Health Perspect 107(1):75-81. 

*LaKind JS, Berlin CM, Park CN, et al. 2000. Methodology for characterizing distributions of 

incremental body burdens of 2,3,7,8-TCDD and DDE from breast milk in North American nursing 

infants. J Toxicol Environ Health A 59:605-639. 

*Larsen K, Jalal S. 1974. DDT induced chromosome mutations in mice: Further testing. Can J Genet 

Cytol 16:491-497. 

Lascombe I, Beffa D, Rüegg U, et al. 2000. Estrogenic activity assessment of environmental chemicals 

using in vitro assays: Identification of two new estrogenic compounds. Environ Health Perspect 

108(7):621-629. 

*Lauenstein G. 1995. Comparison of organic contaminants found in mussels and oysters from a current 

mussel watch project with those from archived mollusc samples of the 1970s. Mar Pollut Bull 

30(12):826-833. 

*Laug E, Nelson A, Fitzhugh O, et al. 1950. Liver cell alternation and DDT storage in the fat of the rat 

induced by dietary levels of 1 to 50 ppm DDT. J Pharmacol Exp Ther 98:268. 

La Vecchia C, Negri E, D’Avanzo B, et al. 1990. Occupation and the risk of bladder cancer. Int J 

Epidemiol 19(2):264-268. 

*Laws E, Curley A, Biros F. 1967. Men with intensive occupational exposure to DDT: A clinical and 

chemistry study. Arch Environ Health 15:766-775. 

*Laws E, Maddrey W, Curley A, et al. 1973. Long-term occupational exposure to DDT. Arch Environ 

Health 27:318-321. 
DDT, DDE, and DDD 355 

9. REFERENCES 

*Lebel G, Dodin S, Ayotte.P, et al. 1998. Organochlorine exposure and the risk of endometriosis. Fertil 

Steril 69(2):221-228. 

*LeBlanc G. 1995. Trophic-level differences in the bioconcentration of chemicals: Implications in 

assessing environmental biomagnification. Environ Sci Technol 29:154-160. 

*Ledoux T, Lodge J, Touchberry R, et al. 1977. The effects of low dietary levels of DDT on breeding 

performance in hybrid mice. Arch Environ Contam Toxicol 6:435-446. 

Lee AG, East JM, Balgavy P. 1991. Interactions of insecticides with biological membranes. Pestic Sci 

32(3):317-327. 

*Leeder JS, Kearns GL. 1997. Pharmacogenetics in pediatrics: Implications for practice. Pediatr Clin 

North Am 44(1):55-77. 

*Legator M, Palmer K, Adler I-D. 1973. A collaborative study of in vitro cytogenetic analysis. I. 

Interpretation of slide preparations. Toxicol Appl Pharmacol 24:337. 

*Lehman A. 1965. Summaries of pesticide toxicity, Topeka, Kansas, Association of Food and Drug 

Officials of the United States. 

*Lehman JW, Peterle TJ, Mills CM. 1974. Effects of DDT on bobwhite quail (Colinus virginianus) 

adrenal gland. Bull Environ Contam Toxicol 11(5):407-413. 

Lembowicz K, Sitarska E, Gorski T, et al. 1991. The effect of organic chlorine compounds and their 

metabolites present in human milk on newborn mice. Toxicol Lett 57(2):215-226. 

*Leoni V, Fabiani L, Marinelli G, et al. 1989. PCB and other organochlorine compounds in blood of 

women with or without miscarriage: A hypothesis of correlation. Ecotoxicol Environ Safety 17:1-11. 

*Lessa J, Becak W, Rabello M, et al. 1976. Cytogenetic study of DDT on human lymphocytes in vitro. 

Mutat Res 40:131. 

*Letcher R, Norstrom R. 1995. An integrated analytical method for determination of polychlorinated 

aryl methyl sulfone metabolites and polychlorinated hydrocarbon contaminants in biological matrices. 

Anal Chem 67:4155-4163. 

*Letcher RJ, Norstrom RJ, Bergman A. 1995. Geographical distribution and identification of methyl 

sulphone PCB and DDE metabolites in pooled polar bear (Ursis maritimus) adipose tissue from western 

hemisphere arctic and subarctic regions. Sci Total Environ 160/161:409-420. 

*Letcher R, Norstrom R, Muir C. 1998. Biotransformation versus bioaccumulation: Sources of methyl 

Sulfones PCB and 4,4'-DDE metabolites in the polar bear food chain. Environ Sci Technol 32:1656­

1661. 

*Leung H-W. 1993. Physiologically-based pharmacokinetics modeling. In: Ballentine B, Marro T, 

Turner P, eds. General and applied toxicology. New York, NY: Stockton Press, 153-164. 

*Levy G. 1982. Gastrointestinal clearance of drugs with activated charcoal. New Engl J Med 

307(11):676-677. 
DDT, DDE, and DDD 356 

9. REFERENCES 

Lewin V, McBlain W, Wolfe F. 1972. Acute intraperitoneal toxicity of DDT and PCB's in mice using 

two solvents. Bull Environ Contam Toxicol 8:245-250. 

Lewis DFV. 1999. Frontier orbitals in chemical and biological activity: Quantitative relationships and 

mechanistic implications. Drug Metab Rev 31(3):755-816. 

Lewis R, Lee R. 1976. Air pollution from pesticides: Sources, occurrence, and dispersion. In: Lee R, 

ed. Air pollutions from pesticides and agricultural processes. Cleveland, OH: CRC Press, 5-50. 

*Lewis RG, Fortmann RC, Camann DE. 1994. Evaluation of methods for monitoring the potential 

exposure of small children to pesticides in the residential environment. Arch Environ Contam Toxicol 

26:37-46. 

Li JJ, Li SA. 1998. Natural and anthropogenic environmental oestrogens: the scientific basis for risk 

assessment. Breast cancer: evidence for xeno-oestrogen involvement in altering its incidence and risk. 

Pure Appl Chem 70(9):1713-1723. 

Liao W, Joe T, Cusick WG. 1991. Multiresidue screening method for fresh fruits and vegetables with 

gas chromatographic/mass spectrometric detection. J Assoc Off Anal Chem 74(3):554-565. 

*Lichtenberg J, Eichelberger J, Dreeman R, et al. 1970. Pesticides in surface waters of the United States: 

A 5-year summary, 1964-68. Pestic Monit J 4(2):71-86. 

*Lichtenstein E, Schulz K. 1959. Persistence of some chlorinated hydrocarbon insecticides as influenced 

by soil types, rate of application and temperature. J Econ Entomol 52(1):124-131. 

*Lichtenstein E, Schulz K. 1960. Translocation of some chlorinated hydrocarbon insecticides into the 

aerial parts of pea plants. Agric Food Chem 8(6):452-456. 

*Lide, DR, ed. 1998. Handbook of chemistry and physics. Boca Raton, FL: CRC Press, Inc., 3-65. 

*Lieberg-Clark P, Bacon C, Burns S, et al. 1995. DDT in California sea-lions: A follow-up study after 

twenty years. Mar Pollut Bull 30(11):744-745. 

*Ligocki M, Leuenberger C, Pankow J. 1985. Trace organic compounds in rain: II. Gas scavenging of 

neutral organic compounds. Atmos Environ 19(10):1609-1617. 

*Liljegren G, Hardell L, Lindstrom G, et al. 1998. Case-control study on breast cancer and adipose 

tissue concentrations of congener specific polychlorinated biphenyls, DDE and hexachlorobenzene. Eur J 

Cancer Prev 7:135-140. 

*Lindenau A, Fischer B, Seiler P, et al. 1994. Effects of persistent chlorinated hydrocarbons on 

reproductive tissues in female rabbits. Hum Reprod 9(5):772-780. 

*Linder RE, Strader LF, Slott VL, et al. 1992. Endpoints of spermatotoxicity in the rat after short 

duration exposures to fourteen reproductive toxicants. Reprod Toxicol 6:491-505. 

*Lindhe O, Lund B, Bergman A, et al. 2001. Irreversible binding and adrenocorticolytic activity of the 

DDT metabolite 3-methylsulfonyl-DDE examined in tissue-slice culture. Environ Health Perspect 

109(2):105-110. 
DDT, DDE, and DDD 357 

9. REFERENCES 

*Livingston, AL. 1978. Forage plant estrogens. J Toxicol Environ Health 4:301. 

*Locke LN, Chura NJ, Stewart PA. 1966. Spermatogenesis in bald eagles experimentally fed a diet 

containing DDT. The Condor 68:497-502. 

Lodi G, Betti A, Kahie YD, et al. 1991. Multiple-development high-performance thin-layer 

chromatography of organochlorine pesticides. J Chromatogr 545:214-218. 

*Loeffler IK, Peterson RE. 1999. Interactive effects of TCDD and p,p'-DDE on male reproductive tract 

development in in utero and lactationally exposed rats. Toxicol Appl Pharmacol 154:28-39. 

Long AR, Crouch MD, Barker SA. 1991a. Multiresidue matrix solid phase dispersion (MSPD) 

extraction and gas chromatographic screening of nine chlorinated pesticides in catfish (Ictalurus 

punctatus) muscle tissue. J Assoc Off Anal Chem 74(4):667-670. 

*Long AR, Soliman MM, Barker SA. 1991b. Matrix solid phase dispersion (MSPD) extraction and gas 

chromatographic screening of nine chlorinated pesticides in beef fat. J Assoc Off Anal Chem 

74(3):493-496. 

Long ER, MacDonald DD, Smith SL, et al. 1995. Incidence of adverse biological effects within ranges 

of chemical concentrations in marine and estuarine sediments. Environ Manag 19:81-97. 

*Longcore JR, Samson FB, Whittendale TW. 1971. DDE thins eggshells and lowers reproductive 

success of captive black ducks. Bull Environ Contam Toxicol 6(6):485-490. 

Longnecker MP, Rogan WJ. 2001. Persistent organic pollutants in children. Pediatr Res 50(3):322-323. 

Longnecker MP, Klebanoff MA, Brock JA, et al. 2000. DDE is associated with increased risk of preterm 

delivery and small-for-gestational-age birthweight in humans. Organohalogen Compounds 48:161-162. 

Longnecker MP, Klebanoff MA, Brock JWZ H, et al. 2001. Maternal serum level of DDE and risk of 

cryptorchidism, hypospadias, and polythelia among male offspring [Abstract]. Am J Epidemiol 

153(11):S267. 

Longnecker MP, Klebanoff MA, Gladen BC, et al. 1999. Serial levels of serum organochlorines during 

pregnancy and postpartum. Arch Environ Health 54(2):110-114. 

*Longnecker MP, Klebanoff MA, Zhou H, et al.. 2001. Association between maternal serum 

concentration of the DDT metabolite DDE and preterm and small-for-gestational-age babies at birth. 

Lancet 358:110-114. 

*Loomis AK, Thomas P. 1999. Binding characteristics of estrogen receptor (ER) in Atlantic croaker 

(Micropogonias undulatus) testis: different affinity for estrogens and xenobiotics from that of hepatic 

ER. Biol Reprod 61:51-60. 

*Looser R, Froescheis O, Cailliet GM, et al. 2000. The deep-sea as a final global sink of semivolatile 

persistent organic pollutants? Part II: organochlorine pesticides in surface and deep-sea dwelling fish of 

the North and South Atlantic and the Monterey Bay Canyon (California). Chemosphere 40:661-670. 

Lopez-Avila V, Benedicto J, Baldin E, et al. 1992. Analysis of classes of compounds of environmental 

concern: III. Organochlorine pesticides. J High Resolut Chromatogr 15:319-328. 
DDT, DDE, and DDD 358 

9. REFERENCES 

*Lopez-Carrillo L, Blair A, Lopez-Cervantes M, et al. 1997. Dichlorodiphenyltrichloroethane serum 

levels and breast cancer risk: A case-control study from Mexico. Cancer Res 57:3728-3732. 

López-Carrillo L, Torres-Sánchez L, López-Cervantes M, et al. 1999. The adipose tissue to serum 

dichlorodiphenyldichloroethane (DDE) ratio: Some methodological considerations. Environmental 

Research (Section A) 81:142-145. 

*Lordo RA, Dinh KT, Schwemberger JG. 1996. Semivolatile organic compounds in adipose tissue: 

Estimated averages for the US population and selected subpopulations. Am J Public Health 86(9):1253­

1259. 

*Lott HM, Barker SA. 1993. Matrix solid-phase dispersion extraction and gas chromatographic 

screening of 14 chlorinated pesticides in oysters (Crassostrea virginica). J Assoc Off Anal Chem Int 

76(1):67-72. 

*Lu F, Dourson M. 1992. Safety/risk assessment of pesticides: Principles, procedures and examples. 

Toxicol Lett 64/65:783-787. 

*Lu F, Jessup D, Lavallee A. 1965. Toxicity of pesticides in young versus adult rats. Food Cosmet 

Toxicol 3:591-596. 

*Luckens MM, Davis WH. 1964. Bats: Sensitivity to DDT. Science 145:948. 

*Lund B-O, Lund J. 1995. Novel involvement of a mitochondrial steroid hydroxylase (P450c11) in 

xenobiotic metabolism. J Biol Chem 270(36):20895-20897. 

*Lund B-O, Bergman A, Brandt I. 1988. Metabolic activation and toxicity of a DDT-metabolite, 

3-methylsulphonyl-DDE, in the adrenal zona fasciculata in mice. Chem Biol Interact 65:25-40. 

Lund B-O, Klasson-Wehler E, Brandt I. 1986. o,p'-DDT in the mouse lung: Selected uptake, covalent 

binding and effect on drug metabolism. Chem Biol Interact 60:129-141. 

*Lundberg C. 1973. Effects of long-term exposure to DDT on the oestrus cycle and the frequency of 

implanted ova in the mouse. Environ Physiol Biochem 3:127-131. 

*Lundberg C. 1974. Effect of DDT on cytochrome P-450 and oestrogen-dependent functions in mice. 

Environ Physiol Biochem 4:200-204. 

*Lundberg C, Kihlstrom J. 1973. DDT and the frequency of implanted ova in the mouse. Bull Environ 

Contam Toxicol 9(5):267-270. 

Lundholm CE. 1991. Influence of chlorinated hydrocarbons, mercury and methylmercury on steroid 

hormone receptors from eggshell gland mucosa of domestic fowls and ducks. Arch Toxicol 

65(3):220-227. 

*Lundholm CE. 1997. DDE-induced eggshell thinning in birds: Effects of p,p'-DDE on the calcium and 

prostaglandin metabolism of the eggshell gland. Comp Biochem Physiol 118C(2):113-128. 

Lundholm CE, Bartonek M. 1991. A study of the effects of p,p'-DDE and other related chlorinated 

hydrocarbons on inhibition of platelet aggregation. Arch Toxicol 65(7):570-574. 
DDT, DDE, and DDD 359 

9. REFERENCES 

*Lustick S, Voss T, Peterle TJ. 1972. Effects of DDT on steroid metabolism and energetics in bobwhite 

quail (Colinus virginianus). In: Morrison JA, Lewis JC, eds. proceedings of the first national bobwhite 

quail symposium. Stillwater, Oklahoma: Oklahoma State University, Research Foundation, 213-233. 

Machala M, Vondracek J, Ulrich R, et al. 1999. Evidence for estrogenic and TCDD-like activities in 

extracts of blood and semen of men. Organohalogen Compounds 42:83-86. 

*MacIntosh DL, Hammerstrom K, Ryan PB. 1999. Longitudinal exposure to selected pesticides in 

drinking water. Hum Ecol Risk Assess 5(3):575-588. 

Macintosh DL, Kabiru CW, Ryan PB. 2001. Longitudinal investigation of dietary exposure to selected 

pesticides. Environ Health Perspect 109(2):145-150. 

MacIntosh DL, Spengler JD, Ozkaynak H, et al. 1996. Dietary exposures to selected metals and 

pesticides. Environ Health Perspect 104(2):202-209. 

MacLusky NJ, Brown TJ, Schantz S, et al. 1998. Hormonal interactions in the effects of halogenated 

aromatic hydrocarbons on the developing brain. In: Colborn T, vom Saal F, Short P, eds. Advances in 

modern environmental toxicology. Princeton, NJ: Princeton Scientific Publishing Co., Inc., 219-246. 

Maestroni BM, Skerritt JH, Ferris IG, et al. 2001. Analysis of DDT residues in soil by ELISA: An 

international interlaboratory study. J AOAC Int 84(1):134-142. 

Magnussen I. 1985. p,p'-DDT-induced myoclonus in mice: The effect of enhanced 5-HT 

neurotransmission. Acta Pharmacol Toxicol 56:87-90. 

*Mahoney JJ. 1975. DDT and DDE effects on migratory condition in white-throated sparrows. J Wildl 

Manage 39(3):520-527. 

*Mahr U, Miltenburger H. 1976. The effect of insecticides on Chinese hamster cell cultures. Mutat Res 

40:107-118. 

Maliwal BP, Guthrie FE. 1981. Interactions of insecticides with human plasma lipoproteins. Chem Biol 

Interact 35:177-188. 

Maliwal BP, Guthrie FE. 1982. In vitro uptake and transfer of chlorinated hydrocarbons among human 

lipoproteins. J Lipid Res 23:474-479. 

*Maness S, McDonnell D, Gaido K. 1998. Inhibition of androgen receptor-dependent transcriptional 

activity by DDT isomers and methoxychlor in HepG2 human hepotema cells. Toxicol Appl Pharmacol 

151:135-142. 

Mangani F, Maione M, Palma P. 2000. Analysis of organochlorinated pesticides in water. In: Nollet 

LML, ed. Handbook of water analysis. New York, NY: Marcel Dekker, Inc., 517-536. 

Manigold D, Schulze J. 1969. Pesticides in selected western streams: A progress report. Pestic Monit J 

3(2):124-135. 

Marei A, Quirke J, Rinaldi G, et al. 1978. The environmental fate of DDT. Chemosphere 12:993-998. 
DDT, DDE, and DDD 360 

9. REFERENCES 

*Marien K, Laflamme D. 1995. Determination of a tolerable daily intake of DDT for consumers of DDT 

contaminated fish from the lower Yakima River, Washington. Risk Anal 15(6):709-717. 

*Martz F, Straw JA. 1980. Metabolism and covalent binding of 1-(o-chlorophenyl)-1-(p-chlorophenyl)­

2,2-dichloroethane (o,p'-DDD): Correlation between adrenocorticolytic activity and metabolic activation 

by adrenocortical mitochondria. Drug Metab Dispos 8(3):127-130. 

Maruyama S, Fujimoto N, Ito A. 1999. Growth stimulation of a rat pituitary cell line MtT/E-2 by 

environmental estrogens in vitro and in vivo. Endocr J (Tokyo) 46(4):513-520. 

*Maslansky C, Williams G. 1981. Evidence for an epigenetic mode of action in organochlorine pesticide 

hepatocarcinogenicity: A lack of genotoxicity in rat, mouse, and hamster hepatocytes. J Toxicol Environ 

Health 8:121-130. 

Masse R, Martineau D, Tremblay L, et al. 1986. Concentrations and chromatographic profile of DDT 

metabolites and polychlorobiphenyl (PCB) residues in stranded Beluga whales (Delphinapterus leucas) 

from the St. Lawrence Estuary, Canada. Arch Environ Contam Toxicol 15:567-579. 

*Masters PM, Inman DL. 2000. Transport and fate of organochlorines discharged to the salt marsh at 

upper Newport Bay, California, USA. Environ Toxicol Chem 19:2076-2084 

Matin M, Nigar F. 1979. A study of the effects of certain barbiturates on the level of striatal 

acetylcholine and convulsions in 1,1,1-trichloro2,2-bis(4'-chlorophenyl)ethane (p,p'-DDT) treated rats. 

Pharmacol Res Commun 11:371-377. 

*Matin M, Jaffery F, Siddiqui R. 1981. A possible neurochemical basis of the central stimulatory effects 

of p,p'-DDT. J Neurochem 36(3):1000-1005. 

*Matsumura F, ed. 1985. Toxicology of insecticides. New York, NY: Plenum Press. 

*Matsumura F, Patil KC. 1969. Adenosine triphosphatase sensitive to DDT in synapses in rat brain. 

Science 166:121-122. 

*Matter JM, McMurry CS, Anthony AB, et al. 1998. Development and implementation of endocrine 

biomarkers of exposure and effects in American alligators (Alligator mississippiensis). Chemosphere 

37:1905-1914. 

*Mattson A, Spillane J, Baker C, et al. 1953. Determination of DDT and related substances in human 

fat. Anal Chem 25(7):1065. 

*Matuo YK, Mamede MV, Clapis MJ, et al. 2000. Niveis de DDT no soro sanguineo de mulheres e 

risco de câncer de mama na regiao de Ribeirao Preto, Brasil. Rev Bras Toxicol 13(2):5-15. 

*Mayr U, Butsch A and Schneider S. 1992. Validation of two in vitro test systems for estrogenic 

activities with zearalenone, phytoestrogens and cereal extracts. Toxicology 74:135-149. 

Mazhitova Z, Jensen S, Ritzen M, et al. 1998. Chlorinated contaminants, growth and thyroid function in 

schoolchildren from the Aral Sea region in Kazakhstan. Acta Paediatr 87:991-995. 

*McCann J, Choi E, Yamasaki E, et al. 1975. Detection of carcinogens as mutagens in the Salmonella/ 

microsome test: Assay of 300 chemicals. Proc Natl Acad Sci 72(12):5135-5139. 
DDT, DDE, and DDD 361 

9. REFERENCES 

*McConnell L, Bidleman T, Cotham W, et al. 1998. Air concentrations of organochlorine insecticides 

and polychlorinated biphenyls over Green Bay, WI, and the four lower Great Lakes. Environ Pollut 

101:391-399. 

*McDougal A, Safe S. 1998. Induction of 16α -/2-hydroxyestrone metabolite ratios in MCM-7 cells by 

pesticides, carcinogens, and antiestrogens does not predict mammary carcinogens. Environ Health 

Perspect 104(4):203-206. 

*McKinney J, Moore L, Prokopetz A, et al. 1984. Validated extraction and cleanup procedures for 

polychlorinated biphenyls and DDE in human body fluids and infant formula. J Assoc Off Anal Chem 

67(1):122. 

McLachlan JA, Arnold SF. 1996. Environmental estrogens: Found internally, certain compounds are 

important biological signals; found in the environment they can become just so much noise. Am Sci 

84:452-461. 

McLachlan JA, Dixon RL. 1972. Gonadal function in mice exposed prenatally to p,p'-DDT. Toxicol 

Appl Pharmacol 22:327-331. 

McLachlan JA, Newbold RR, Teng CT, et al. 1998. Environmental estrogens: Orphan receptors and 

genetic imprinting. J Clean Technol Environ Toxicol Occup Med 7(2):221-226. 

*McMahon B, Burke J. 1978. Analytical behavior data for chemicals determined using AOAC 

multiresidue methodology for pesticide residues in foods. J Assoc Off Anal Chem 61(3):640-652. 

Mehendale H. 1978. Pesticide-induced modification of hepatobiliary function: Hexachlorobenzene, 

DDT and toxaphene. Food Cosmet Toxicol 16:19-25. 

*Meister R, Sine C. 1999. Farm Chemicals Handbook. Willoughby, OH: Meister Publication 

Company, C123-C124. 

Mendel J, Walton M. 1966. Conversion of p,p'-DDT to p,p'-DDD by intestinal flora of the rat. Science 

151:1527-1528. 

*Mendenhall VM, Klaas EE, McLane MAR. 1983. Breeding success of barn owls (Tyto alba) fed low 

levels of DDE and dieldrin. Arch Environ Contam Toxicol 12:235-240. 

*Mendonca GAS, Eluf-Neto J, Andrada-Serpa MJ, et al. 1999. Organochlorines and breast cancer: A 

case-control study in Brazil. Int J Cancer 83:596-600. 

Menzer RE. 1991. Water and soil pollutants. In: Amdur M, Doull J, Klaassen C, eds. Casarett and 

Doull’s toxicology: The basic science of poisons. 4th ed. New York, NY: Pergamon Press, 872-902. 

*Menzie C. 1980. Reaction types in the environment. In: Hutzinger O, ed. Reactions and processes. 

New York, NY: Springer-Verlag Berlin Heidelberg, 247-302. 

*Menzie CA, Burmaster DE, Freshman JS, et al. 1992. Assessment of methods for estimating ecological 

risk in the terrestrial component: A case study at the Baird and McGuire Superfund site in Holbrook, 

Massachusetts. Environ Toxicol Chem 11(2):245-260. 
DDT, DDE, and DDD 362 

9. REFERENCES 

*Mes J, Malcolm S. 1992. Comparison of chlorinated hydrocarbon residues in human populations from 

the Great Lakes and other regions of Canada. Chemosphere 25(3):417-424. 

*Mes J, Davies DJ, Turton D, et al. 1986. Levels and trends of chlorinated hydrocarbon contaminants in 

the breast milk of Canadian women. Food Addit Contam 3(4):313-322. 

*Mes J, Marchand L, Davies DJ. 1990. Specific polychlorinated biphenyl congener distribution in 

adipose tissue of Canadians. Environ Technol 11:747-756. 

*Metcalf RL. 1995. Insect Control Technology. In: Kroschwitz J, Howe-Grant M, eds. Kirk-Othmer 

encyclopedia of chemical technology. Volume 14. New York, NY: John Wiley and Sons, Inc., 524-602. 

*Metcalf R, Booth G, Schuth C, et al. 1973a. Uptake and fate of di-2-ethylhexyl phthalate in aquatic 

organisms and in a model ecosystem. Environ Health Perspect 4:27-34. 

*Metcalf RL, Kapoor IP, Lu P-O, et al. 1973b. Model ecosystem studies of the environmental fate of six 

organochlorine pesticides. Environ Health Perspect 4:35-44. 

*Metcalfe PJ. 1989. Organolead compounds in the atmosphere. Analyt Proc 26:134-136. 

*Meylan WM, Howard PH. 1993. Computer estimation of the atmospheric gas-phase reaction rate of 

organic compounds with hydroxyl radicals and ozone. Chemosphere 26(2):2293-2299. 

*Meylan W, Howard P, Boethling R. 1992. Molecular topology/fragment contribution method for 

predicting soil sorption coefficients. Environ Sci Technol 26(8):1560-1567. 

*Miller DS, Kinter WB, Peakall DB, et al. 1976. DDE feeding and plasma osmoregulation in ducks, 

guillemots, and puffins. Am J Physiol 231(1):370-376. 

*Miller G, Zepp R. 1979. Photoreactivity of aquatic pollutants sorbed on suspended sediments. Environ 

Sci Toxicol 13(7):860-863. 

*Miller M. 1994. Organochlorine concentration dynamics in Lake Michigan Chinook Salmon 

(Oncorhynchus tshawytscha). Arch Environ Contam Toxicol 27:367-374. 

Miller M, Alexander M. 1991. Kinetics of bacterial degradation of benzylamine in a montmorillonite 

suspension. Environ Sci Technol 25(2):240-245. 

*Millet M, Wortham H, Sanusi A, et al. 1997. Atmospheric contamination by pesticides: Determination 

in the liquid, gaseous and particulate phases. Environ Sci Pollut Res 4(3):172-180. 

*Minelli E, Ribeiro M. 1996. DDT and HCH residues in the blood serum of Malaria control sprayers. 

Bull Environ Contam Toxicol 57:691-696. 

*Minyard JP, Roberts WE. 1991. Chemical contaminants monitoring: State findings on pesticide 

residues in foods—1988 and 1989. J Assoc Off Anal Chem 74(3):438-452. 

Mitjavila S, Carrera G, Boigegrain R-A, et al. 1981a. I. Evaluation of the toxic risk of DDT in the rat: 

During accumulation. Arch Environ Contam Toxicol 10:459-469. 
DDT, DDE, and DDD 363 

9. REFERENCES 

Mitjavila S, Carrera G, Fernandez Y. 1981b. II. Evaluation of the toxic risk of accumulated DDT in the 

rat: During fat mobilization. Arch Environ Contam Toxicol 10:471-481. 

*Mohammed A, Eklund A, Ostlund-Lindqvist AM, et al. 1990. Distribution of toxaphene, DDT, and 

PCB among lipoprotien fractions in rat and human plasma. Arch Toxicol 64:567-571. 

*Monosmith CL, Hermanson MH. 1996. Spatial and temporal trends of atmospheric organochlorine 

vapors in the central and upper Great Lakes. Environ Sci Technol 30:3464-3472. 

*Moody RP, Chu I. 1995. Dermal exposure to environmental contaminants in the Great Lakes. Environ 

Health Perspect Suppl 103(9):103-114. 

Moody RP, Nadeau B, Chu I. 1994. In vitro dermal absorption of pesticides: VI. In vivo and in vitro 

comparison of the organochlorine insecticide DDT in rat, guinea pig, pig, human and tissue-cultured skin. 

Toxicol in Vitro 8(6):1225-1232. 

Moore ML, Butcher MK, Connor K, et al. 1999. Fingerprinting analysis of PCDD/PCDF sources in a 

surface water outfall near a petroleum refinery. Organohalogen Compounds 40:219-220. 

*Morello A. 1965. Induction of DDT-metabolizing enzymes in microsomes of rat liver after 

administration of DDT. Can J Biochem 43:1289-1293 

Moreno AJ, Madeira VM. 1991. Mitochondrial bioenergetics as affected by DDT. Biochim Biophys 

Acta 1060(2):166-174. 

*Morgan D, Lin L. 1978. Blood organochlorine pesticide concentrations, clinical hematology and 

biochemistry in workers occupationally exposed to pesticides. Arch Environ Contam Toxicol 7:423-447. 

*Morgan D, Roan C. 1971. Absorption, storage and metabolic conversion of ingested DDT and DDT 

metabolites in man. Arch Environ Health 22:301. 

*Morgan D, Roan C. 1974. The metabolism of DDT in man. In: Hayes W, Jr, ed. Essays in toxicology. 

New York, NY: Academic Press, 39-97. 

*Morgan D, Lin L, Saikaly H. 1980. Morbidity and mortality in workers occupationally exposed to 

pesticides. Arch Environ Contam Toxicol 9:349. 

*Morgan D, Roan C, Paschal E. 1972. Transport of DDT, DDE, and dieldrin in human blood. Bull 

Environ Contam Toxicol 8(6):321-326. 

*Morimoto S. 1996. Alteration of intercellular communication in a human urothelial carcinoma cell- line 

by tumor-promoting agents. Int J Urol 3:212-217. 

Morio Y, Imanishi T, Nishimura M, et al. 1987. Characteristics of neuronal effects of DDT and dieldrin 

in mice. Jpn J Vet Sci 49:1039-1044. 

Morozova O, Riboli E Turusov V. 1997. Estrogenic effect of DDT in CBA female mice. Exp Toxicol 

Pathol 49:483-485. 

*Morrison DE, Robertson BK, Alexander M. 1999. Written communication to Mario Citra, Syracuse 

Research Corporation. Possible use of a solid-phase extractant to predict availability of DDT, DDE, 
DDT, DDE, and DDD 364 

9. REFERENCES 

DDD, and dieldrin in soil to Eisenia foetida. Ithaca, NY: Institute for Comparative and Environmental 

Toxicology and Department of Soil, Crop, and Atmospheric Sciences, Cornell University. 

*Morselli PL, Franco-Morselli R, Bossi L. 1980. Clinical pharmacokinetics in newborns and infants: 

Age-related differences and the therapeutic implications. Clin Pharmacokin 5:485-527. 

*Mosselman S, Polman J, Dijkema R. 1996. ER-β: identification and characterization of a novel human 

estrogen receptor. FEBS Lett 392:49-53. 

*Moysich KB, Ambrosone CB, Mendola P, et al. 2002. Exposures associated with serum organochlorine 

levels among postmenopausal women from western New York State. Am J Ind Med 41:102-110. 

*Moysich K, Ambrosone C, Vena J, et al. 1998. Environmental organochlorine exposure and 

postmenopausal breast cancer risk. Cancer Epidemiol Biomarkers Prev 7:181-188. 

*Mühlebach S, Moor MJ, Wyss PA, et al. 1991. Kinetics of distribution and elimination of DDE in rats. 

Xenobiotica 21(1):111-120. 

*Mühlebach S, Wyss P, Bickel M. 1985. Comparative adipose tissue kinetics of thiopental, DDE and 

2,4,5,2',4',5'-hexachlorobiphenyl in the rat. Xenobiotica 15(6):485-491. 

*Muir D, Grift N, Lockhart W, et al. 1995. Special trends and historical profiles of organochlorine 

pesticides in arctic lake sediments. Sci Total Environ 160:447-457. 

*Muir DCG, Nordstrom RJ, Simon M. 1988. Organochlorine contaminants in arctic marine food chains: 

Accumulation of specific polychlorinated biphenyls and chlordane-related compounds. Environ Sci 

Technol 22:1071-1079. 

*Mukherjee I, Gopal M. 1996. Chromatographic techniques in the analysis of organochlorine pesticide 

residues. J Chromatogr A 754:33-42. 

*Mulhens K. 1946. [The importance of bis(chlorophenyl) (trischloromethyl)-methane preparations as 

insecticides in combating infectious diseases]. Dtsch Med Wochenschr 71:164-169 (German). 

*Murphy R, Harvey C. 1985. Residues and metabolites of selected persistent halogenated hydrocarbons 

in blood specimens from a general population survey. Environ Health Perspect 60:115-120. 

*Mussalo-Rauhamaa H. 1991. Partitioning and levels of neutral organochlorine compounds in human 

serum, blood cells, and adipose and liver tissue. Sci Total Environ 103:159-175. 

*Mussalo-Rauhamaa H, Hasanen E, Pyysalo H, et al. 1990. Occurrence of beta-hexachlorocyclohexane 

in breast cancer patients. Cancer 66(10):2124-2128. 

Myers JP. 2001. Endocrine disruptors: Lessons from new research [Abstract]. Neurotoxicology 22:131. 

Myers RA, Stella VJ. 1992. Factors affecting the lymphatic transport of penclomedine (NSC-338720), a 

lipophilic cytotoxic drug: Comparison to DDT and hexachlorobenzene. Int J Pharm 80:51-62. 

*Nachman G, Freal J, Barquet A, et al. 1972. A simplified method for analysis of DDT and DDE in 

blood for epidemiologic purposes. In: Davies J, Edmundson, W, eds. Epidemiology of DDT. Mount 

Kisco, NY: Futura Publishing Company, Inc., 133-139. 
DDT, DDE, and DDD 365 

9. REFERENCES 

Naevested R. 1947. [Poisoning with DDT powder and some other incidence of poisoning.] Tidsskr 

Norske Laegeforen 67:261-263. (Norwegian) 

Nagayama J, Fukushige J, Iida T, et al. 2000a. Developmental condition in 10-month-old Japanese 

infants perinatally exposed to organochlorine pesticides, PCBs and dioxins. Organohalogen Compounds 

48:244-247. 

Nagayama J, Fukushige J, Iida T, et al. 2000b. Effects of exposure to organochlorine pesticides, PCBs 

and dioxins through human milk on total development in 10-month-old Japanese infants. Organohalogen 

Compounds 48:240-243. 

Nagayama J, Iida T, Nakagawa R, et al. 2000c. Condition of thyroid hormone system in 10-month-old 

Japanese infants perinatally exposed to organochlorine pesticides, PCBs and dioxins. Organohalogen 

Compounds 48:236-239. 

Nagayama J, Tsuji H, Iida T, et al. 2000d. Condition of helper and suppressor T lymphocyte 

subpopulations in 10-month-old Japanese infants perinatally exposed to organochlorine pesticides, PCBs 

and dioxins. Organohalogen Compounds 49:87-90. 

*Nair A, Pillai MK. 1992. Trends in ambient levels of DDT and HCH residues in humans and the 

environment of Delhi, India. Sci Total Environ 121:145-157. 

Nair A, Samuel T, Pillai MKK. 1992. Behaviour of DDT in three soils exposed to solar radiations under 

different conditions. Pestic Sci 34:333-340. 

*Naishtein S, Leibovich D. 1971. Effect of small doses of DDT and lindane and their mixture on sexual 

function and embryogenesis in rats. Hygiene and Sanitation 36:190-194. 

*Nanni O, Falcini F, Buiatti E, et al. 1998. Multiple myeloma and work in agriculture: Results of a 

case-control study in Forli, Italy. Cancer Causes Control 9:277-283. 

Narahashi T. 1992. Nerve membrane Na+ channels as targets of insecticides. Trends Pharmacol Sci 

13:236-241. 

*Narahashi T, Haas H. 1967. DDT: Interaction with nerve membrane conductance changes. Science 

157:1438-1440. 

Narahashi T, Aistrup GL, Ikeda T, et al. 1999. Ion channels as targets for insecticides: Past, present, and 

future. In: Beadle DJ, ed. Progress in neuropharmacology and neurotoxicology of pesticides and drugs. 

Cambridge: Royal Society of Chemistry, 21-33. 

*Narayan S, Bajpai A, Chauhan S, et al. 1985a. Lipid peroxidation in lung and liver of rats given DDT 

and endosulfan intratracheally. Bull Environ Contam Toxicol 34:63-67. 

*Narayan S, Bajpai A, Tyagi S, et al. 1985b. Effect of intratracheal administration of DDT and 

endosulfan on cytochrome P-450 and glutathione-s-transferase in lung and liver of rats. Bull Environ 

Contam Toxicol 34:55-62. 

NAS. 1977. National Academy of Sciences. Drinking water and health. Washington, DC: National 

Academy Press. 
DDT, DDE, and DDD 366 

9. REFERENCES 

*NAS/NRC. 1989. Biologic markers in reproductive toxicology. National Academy of Sciences 

National Research Council. Washington, DC: National Academy Press, 15-35. 

*Nash R, Beall L. 1970. Chlorinated hydrocarbon insecticides: Root uptake versus vapor contamination 

of soybean foliage. Science 168:1109-1111. 

NATICH. 1992. Data base report on state, local, and EPA air toxics activities. National Air Toxics 

Information Clearinghouse. Washington, DC: U.S. Environmental Protection Agency, Office of Air 

Quality Planning and Standards. March 13, 1992. 

*NCI. 1978. National Cancer Institute. Bioassays of DDT, TDE, and p,p'-DDE for possible 

carcinogenicity. CAS No. 50-29-3, 72-54-8, 72-55-9 NCI-CG-TR-131. 

*Neal P, Von Oettingen W, Smith W, et al. 1944. Toxicity and potential dangers of aerosols, mists, and 

dusting powders containing DDT. Public Health Rep 117:1-32. 

*Neidert E and Saschenbrecker PW. 1996. Occurrence of pesticide residues in selected agricultural food 

commodities available in Canada. J Assoc Off Anal Chem Int 79(2):549-566. 

*Nelson A, Woodard G. 1949. Severe adrenal cortical atrophy (cytotoxic) and hepatic damage produced 

in dogs by feeding 2,2-bis(parachlorophenyl)-1,1-dichloroethane (DDD or TDE). Arch Pathol 48:387­

394. 

*Nelson JA. 1974. Effects of dichlorodiphenyltrichloroethane (DDT) analogues and polychlorinated 

biphenyl (PCB) mixture on 17β-[3

H] estradiol binding to rat uterine receptor. Biochem Pharm 

23:447-451. 

*Newsome WH, Davies D, Doucet J. 1995. PCB and organochlorine pesticides in Canadian human milk 

- 1992. Chemosphere 30(11):2143-2153. 

*Nhachi CFB, Kasilo O. 1990. Occupational exposure to DDT among mosquito control sprayers. Bull 

Environ Contam Toxicol 45:189-192. 

*Nhachi CFB, Loewenson R. 1989. Comparison study of the sensitivities of some indices of DDT 

exposure in human blood and urine. Bull Environ Contam Toxicol 43:493-498. 

Nichols WK, Covington MO, Seiders CD, et al. 1992. Bioactivation of halogenated hydrocarbons by 

rabbit pulmonary cells. Pharmacol Toxicol 71:335-339. 

*Nigam S, Kashyap S, Kumar A, et al. 1977. Effect of technical grade DDT on neonatal mice. Indian J 

Med Res 65(4):576-578. 

*Nigg H, Metcalf R, Kapoor I. 1974. DDT and selected analogues as microsomal oxidase inducers in the 

mouse. Arch Environ Contam 3:426-436. 

*Nims R, Lubet R, Fox S, et al. 1998. Comparative pharmacodynamics of CYP2B induction by DDT, 

DDE, and DDD in male rat liver and cultured rat hepatocytes. J Toxicol Environ Health Part A 53:455­

477. 
DDT, DDE, and DDD 367 

9. REFERENCES 

*NIOSH. 1977. NIOSH manual of analytical methods. 2nd ed. Vol 3. Washington, DC: U.S. 

Department of Health, Education, and Welfare, Public Health Service, Center for Disease Control, 

National Institute for Occupational Safety and Health, S274-1 - S274-7. 

NIOSH. 1992. NIOSH pocket guide to chemicals hazards. Washington, DC: U.S. Department of 

Health and Human Service, Center for Disease Control, National Institute for Occupational Safety and 

Health, Division of Standard Development and Technology Transfer. 

NIOSH. 1999. Pocket guide to chemical hazards. Washington DC: National Institute for Occupational 

Safety and Health, U.S. Department of Health and Human Services, 88-89 

*NIOSH. 2002. DDT. NIOSH pocket guide to chemical hazards. National Institute for Occupational 

Safety and Health. http://www.cdc.gov/niosh/npg/npgd0174.html. January 02, 2002. 

Nishihara T, Nishikawa J, Kanayama T, et al. 2000. Estrogenic activities of 517 chemicals by yeast twohybrid assay. J Health Sci 46(4):282-298. 

Nishimura H, Tanimura T, Mizutani T, et al. 1977. Levels of polychlorinated biphenyls and organochlorine insecticides in human embryos and fetuses. Pediatrician 6:45-57. 

*Nishizumi M. 1979. Effect of phenobarbital, dichlorodiphenyltrichloroethane, and polychlorinated 

biphenyls on diethylnitrosamine-induced hepatocarcinogenesis. Gann 70:835-837. 

*NOAA. 1988. PCB and chlorinated pesticide contamination in U.S. fish and shellfish: A historical 

assessment report. Rockville, MD: National Oceanic and Atmospheric Administration, Office of 

Oceanography and Marine Assessment. PB91172742. 

*Noguchi T, Charman W, Stella V. 1985. Lymphatic appearance of DDT in thoracic or mesenteric 

lymph duct cannulated rats. Int J Pharm 24:185-192. 

*Norén K. 1988. Changes in the levels of organochlorine pesticides, polychlorinated biphenyls, dibenzop-dioxins and dibenzofurans in human milk from Stockholm, 1972-1985. Chemosphere 17(1):39-49. 

*Norén K. 1993. Contemporary and retrospective investigations of human milk in the trend studies of 

organochlorine contaminants in Sweden. Sci Total Environ 139/140:347-355 

*Norén K, Meironyté D. 2000. Certain organochlorine and organobromine contaminants in Swedish 

human milk in perspective of past 20-30 years. Chemosphere 40:1111-1123. 

Norén K, Sjovall J. 1993. Liquid-gel partitioning and enrichment in the analysis of organochlorine 

contaminants. J Chromatogr 642:243-251. 

*Norén K, Lunden A, Pettersson E, et al. 1996. Methylsulfonyl metabolites of PCBs and DDE in human 

milk in Sweden, 1972-1992. Environ Health Perspect 104(7):766-772. 

*Norén K, Weistrand C, Karpe F. 1999. Distribution of PCB congeners, DDE, hexachlorobenzene, and 

methylsulfonyl metabolites of PCB and DDE among various fractions of human blood plasma. Arch 

Environ Contam Toxicol 37:408-414. 

Norpoth K, Heger M, Muller G, et al. 1986. Investigations on metabolism, genotoxic effects and 

carcinogenicity of 2,2'-dichlorodiethylether. J Cancer Res Clin Oncol 112:125-130. 
DDT, DDE, and DDD 368 

9. REFERENCES 

*Norstrom RJ, Belikov SE, Born EW, et al. 1998. Chlorinated hydrocarbon contaminants in polar bears 

from eastern Russia, North America, Greenland, and Svalbard: Biomonitoring of arctic pollution. Arch 

Environ Contam Toxicol 35:354-367. 

*NRC. 1993. Pesticides in the diets of infants and children. Washington, DC: National Research 

Council. National Academy Press. 

*NRC. 1991. Animals as Sentinels of Environmental Health Hazards. Washington, DC: National 

Research Council. National Academy Press. 

*NRC. 1999. Hormonally active agents in the environment. National Research Council. Washington, 

DC: National Academy Press. 

NREPC. 1986. Natural Resources and Environmental Protection Cabinet. Department for 

Environmental Protection, Division of Pollution. Kentucky. 401 KAR 63:022. 

*NTP. 1999. National Toxicology Program Report on Carcinogens. Substances reasonably anticipated 

to be human carcinogens: Names and synonyms of carcinogens. National Toxicology Program. 

http://ntpserver.niehs.nih.gov/NewHomeRoc/RAHC_list.html 

*NTP. 2002. DDT. Eighth report on carcinogens. National Toxicology Program. http://ntpserver.niehs.nih.gov/htdocs/8_RoC/RAC/DDT.html. January 03, 2002. 

*Numoto S, Tanaka T, Williams G. 1985. Morphologic and cytochemical properties of mouse liver 

neoplasms induced by diethylnitrosamine and promoted by 4,4'-dichlorodiphenyltrichloroethane, 

chlordane, or heptachlor. Toxicol Pathol 13(4):325-334. 

O'Connor R. 1984. Degradation of DDT in lake sediments [Abstract]. Diss Abstr Int 44(9):2772-B. 

*Odsjo T, Bignert A, Olsson M, et al. 1997. The Swedish Environmental Specimen Bank-application in 

trend monitoring of mercury and some organohalogenated compounds. Chemosphere 34(9/10):2059­

2066. 

Ohmiya Y, Nakai K. 1977. Effect of starvation on excretion, distribution and metabolism of DDT in 

mice. Tohoku J Exp Med 122:143-153. 

Ohta Y, Sahashi D, Sasaki E, et al. 1999. Alleviation of carbon tetrachloride-induced chronic liver injury 

and related dysfunction by L-tryptophan in rats. Ann Clin Biochem 36:504-510. 

*Okey A, Page D. 1974. Acute toxicity of o,p'-DDT to mice. Bull Environ Contam Toxicol 

11(4):359-363. 

*O'Leary J, Davies J, Edmundson W, et al. 1970a. Correlation of prematurity and DDE levels in fetal 

whole blood. Am J Obstet Gynecol 106(6):939. 

O'Leary J, Davies J, Feldman M. 1970b. Spontaneous abortion and human pesticide residues of DDT 

and DDE. Am J Obstet Gynecol 108:1291-1293. 

*Oliver B, Niimi A. 1985. Bioconcentration factors of some halogenated organics for rainbow trout: 

Limitations in their use for prediction of environmental residues. Environ Sci Technol 19:842-849. 
DDT, DDE, and DDD 369 

9. REFERENCES 

*Opelanio L, Rack E, Blotcky A, et al. 1983. Determination of chlorinated pesticides in urine by 

molecular neutron activation analysis. Anal Chem 55:677-681. 

*Orberg J, Lundberg C. 1974. Some effects of DDT and PCB on the hormonal system in the male 

mouse. Environ Physiol Biochem 4:116-120. 

*Ortega P. 1956. DDT in the diet of the rat. Public Health Monogr 43:1-27. 

*Ortelee M. 1958. Study of men with prolonged intensive occupational exposure to DDT. Am Med 

Assoc Arch Ind Health 18:433-440. 

OSHA. 1987. Access to employee exposure and medical records. Occupational Safety and Health 

Administration. Code of Federal Regulations. 29 CFR 1910.20. 

OSHA. 1988. Access to employee exposure and medical records. Occupational Safety and Health 

Administration. Federal Register 53(189):30163-30164. 

OSHA. 1989a. Toxic and hazardous substances. Occupational Safety and Health Administration. Code 

of Federal Regulations. 29 CFR 1910.1000. 

OSHA. 1989b. Toxic and hazardous substances. Occupational Safety and Health Administration. 

Federal Register 54(12):2920-2960. 

OSHA. 1992. Toxic and hazardous substances. Occupational Safety and Health Administration. Code 

of Federal Regulations. 29 CFR 1910.1000. 

OSHA. 1998a. U. S. Department of Labor. Occupational Safety and Health Administration. Code of 

Federal Regulations. 29 CFR 1910.1000. 

OSHA. 1998b. U. S. Department of Labor. Occupational Safety and Health Administration. Code of 

Federal Regulations. 29 CFR 1915.1000. 

OSHA. 1998c. U. S. Department of Labor. Occupational Safety and Health Administration. Code of 

Federal Regulations. 29 CFR 1926.55. 

*OSHA. 2001a. Air contaminants—shipyards. Occupational Safety and Health Administration. Code 

of Federal Regulations. 29 CFR 1915.1000. http://www.sha-slc.gov/OshStd_data/1915_1000.html. 

September 17, 2001. 

*OSHA. 2001b. Limits for air contaminants. Occupational Safety and Health Administration. Code of 

Federal Regulations. 29 CFR 1910.1000. http://www.sha-slc.gov/OshStd_data/1910_1000_ 

TABLE_Z.html. September 17, 2001. 

*OSHA. 2001c. Gases, vapors, fumes, dusts, and mists. Construction industry. Occupational Safety 

and Health Administration. Code of Federal Regulations. 29 CFR 1926.55. 

http://frwebgate.access.gpo.gov/cgi-bin/get-cfr.cgi?TITLE=29&PART=1926&SECTION=55& 

YEAR=2001&TYPE=TEXT. September 18, 2001. 

Ostrea EM, Tan E, Hernandez E, et al. 1998. Exposure to environmental pollutants adversely affects 

fetal outcome. Pediatr Res 43(4):224A. 
DDT, DDE, and DDD 370 

9. REFERENCES 

*OTA. 1990. Neurotoxicity, identifying and controlling poisons of the nervous system. Washington, 

DC: Office of Technology Assessment, U.S. Congress. OTA-BA-436. April 1990. 

*Ottar B. 1981. The transfer of airborne pollutants to the Arctic region. Atmos Environ 15:1439-1445 

*Ottoboni A. 1969. Effect of DDT on reproduction in the rat. Toxicol Appl Pharmacol 14:74-81. 

*Ottoboni A. 1972. Effect of DDT on the reproductive life-span in the female rat. Toxicol Appl 

Pharmacol 22:497-502. 

*Ottoboni A, Bissell G, Hexter A. 1977. Effects of DDT on reproduction in multiple generations of 

beagle dogs. Arch Environ Contam Toxicol 6:83-101. 

*Owen GM, Brozek J. 1966. Influence of age, sex, and nutrition on body composition during childhood 

and adolescence. In: Falkner F, ed. Human development. Philadelphia, PA: WB Saunders, 222-238. 

*Pace P, Taylor J, Suntharalingam S, et al. 1997. Human estrogen receptor β binds DNA in a manner 

similar to and dimerizes with estrogen receptor α. J Biol Chem 272(41):25832-25838. 

Palanza P, Morellini F, Parmigiani S, et al. 1999a. Prenatal exposure to endocrine disrupting chemicals: 

effects on behavioral development. Neurosci Biobehav Rev 23:1011-1027. 

*Palanza P, Parmigiani S, Liu H, et al. 1999b. Prenatal exposure to low doses of the estrogenic 

chemicals diethylstilbestrol and p,p'-DDT alters aggressive behavior of male and female house mice. 

Pharmacol Biochem Behav 64(4):665-672. 

*Palin KJ, Wilson CG, Davis SS, et al. 1982. The effects of oils on the lymphatic absorption of DDT. J 

Pharm Pharmacol 34:707-710. 

*Palmer BD, Palmer SK. 1995. Vitellogenin induction by xenobiotic estrogens in the red-eared turtle 

and African clawed frog. Environ Health Perspect Suppl 103(4):19-25. 

*Palmer K, Green S, Legator M. 1972. Cytogenetic effects of DDT and derivatives of DDT in a cultured 

mammalian cell line. Toxicol Appl Pharmacol 22:355-364. 

*Palmer K, Green S, Legator M. 1973. Dominant lethal study of p,p'-DDT in rats. Food Cosmet Toxicol 

11:53-62. 

*Pardio V, Waliszewski S, Aguirre A, et al. 1998. DDT and its metabolites in human milk collected in 

Veracruz City and suburban areas (Mexico). Bull Environ Contam Toxicol 60:852-857. 

*Parkki M, Marniemi J, Vainio H. 1977. Long-term effects of single and combined doses of DDT and 

PCB on drug-metabolizing enzymes in rat liver. J Toxicol Environ Health 3:903-911. 

*Paschal EH, Roan CC, Morgan DP. 1974. Evidence of excretion of chlorinated hydrocarbon pesticides 

by the human liver. Bull Environ Contam Toxicol 12(5):547-554. 

*Pasha S. 1981. Changes in hepatic microsomal enzymes due to DDT, DDE and DDD feeding in CF-1 

mice. Indian J Med Res 74:926-930. 
 

DDT, DDE, and DDD 371 

9. REFERENCES 

Paulsen K, Adesso V, Porter J. 1973. Report on pilot study of effects of DDT on maternal behavior. 

Bull Psychonomic Soc 1(3):20-206. 

Payne J, Rajapakse N, Wilkins M, et al. 2000. Prediction and assessment of the effects of mixtures of 

four xenoestrogens. Environ Health Perspect 108(10):983-987. 

*PDR. 1999. Physicians Desk Reference. 53rd ed. Montvale, NJ: Medical Economics Company. 

*Peakall DB. 1967. Pesticide-induced enzyme breakdown of steroids in birds. Nature 216:505-506. 

*Peakall DB. 1969. Effect of DDT on calcium uptake and vitamin D metabolism in birds. Nature 

224:1219-1220. 

*Peakall DB. 1970. p,p'-DDT: Effect on calcium metabolism and concentration of estriol in the blood. 

Science 168(931):592-594. 

*Peakall DB, Miller DS, Kinter WB. 1975. Blood calcium levels and the mechanism of DDE-induced 

eggshell thinning. Environ Pollut 9:289-294. 

*Pearce G, Mattson A, Hayes W. 1952. Examination of human fat for the presence of DDT. Science 

116:254. 

Peaslee M, Goldman M, Milburn S. 1972. Effects of acute and chronic administration of DDT upon 

pituitary melanocyte-stimulating hormone (MSH) activity. Comp Gen Pharmacol 3:191-194. 

*Peraino C, Fry R, Staffeldt E, et al. 1975. Comparative enhancing effects of phenobarbital, 

amobarbital, diphenylhydantoin, and dichlorodiphenyltrichloroethane on 2-acetylaminofluorene-induced 

hepatic tumorigenesis in the rat. Cancer Res 35:2884-2890. 

*Pereira W, Domagalski J, Hostettler F, et al. 1996. Occurrence and accumulation of pesticides and 

organic contaminants in river sediment, water and clam tissues from the San Joaquin River and 

tributaries, California. Environ Toxicol Chem 15(2):172-180. 

*Persson B, Fredriksson M, Olsen K, et al. 1993. Some occupational exposures as risk factors for 

malignant lymphomas. Cancer 72(5):1773-1778. 

*Peterle TJ, Lustick SI, Nauman LE, et al. 1973. Some physiological effects of dietary DDT on mallard, 

bobwhite quail and rabbits. In: Kjerner I, Bjurholm P, ed. XIth International Congress of Game 

Biologists: Stockholm, September 3-7, 1973. Stockholm: National Swedish Environment Protection 

Board, 457-478. 

*Peterson J, Robison W. 1964. Metabolic products of p,p'-DDT in the rat. Toxicol Appl Pharmacol 

6:321-327. 

*Peterson JR, Adams RS, Cutkomp LK. 1971. Soil properties influencing DDT bioactivity. Soil Sci Am 

Proc 35:73-59. 

Petersson DN, Tkalcevic GT, Koza-Taylor PH, et al. 1998. Identification of estrogen receptor β2, a 

functional variant of estrogen receptor β expressed in normal rat tissues. Endocrinology 139(3):1082­

1092. 
DDT, DDE, and DDD 372 

9. REFERENCES 

*Pettersson K, Grandien K, Kuiper G, et al. 1997. Mouse estrogen receptor β forms estrogen response 

element-binding heterodimers with estrogen receptor α. Mol Endocrinol 11:1486-1496. 

*Philips F, Gilman A. 1946. Studies on the pharmacology of DDT (2,2 bis-(parachlorphenyl)-1,1,1­

trichloroethane). J Pharmacol Exp Ther 86:213-221. 

*Phillips DL, Pirkle JL, Burse VW, et al. 1989. Chlorinated hydrocarbon levels in human serum: 

Effects of fasting and feeding. Arch Environ Contam Toxicol 18:495-500. 

Pinowski J, Niewiadöwska A, Juricová Z, et al. 1999. Chlorinated aromatic hydrocarbons in the brains 

ans lipids of sparrows (Passer domesticus and Passer montanus) from rural and suburban areas near 

warsaw. Bull Environ Contam Toxicol 63:736-743. 

*Pinto J, Camien M, Dunn M. 1965. Metabolic fate of p,p'-DDT [1,1,1-trichloro-2,2-bis(pchlorophenyl) ethane] in rats. J Biol Chem 240(5):2148-2154. 

*Pocock D, Vost A. 1974. DDT absorption and chylomicron transport in rat. Lipids 9(6):374-381. 

*Podreka S, Georges A, Maher B, et al. 1998. The environmental contaminant DDE fails to influence 

the outcome of sexual differentiation in the marine turtle Chelonia mydas. Environ Health Perspect 

106(4):185-188. 

*Poissant L, Koprivnjak J-F, Matthieu R. 1997. Some persistent organic pollutants and heavy metals in 

the atmosphere over a St. Lawrence River valley site (Villeroy) in 1992. Chemosphere 34(3):567-585. 

*Poland A, Smith D, Kuntzman R, et al. 1970. Effect of intensive occupational exposure to DDT on 

phenylbutazone and cortisol metabolism in human subjects. Clin Pharmacol Ther 11:724-732. 

Porta M. 2001. Role of organochlorine compounds in the etiology of pancreatic cancer: A proposal to 

develop methodological standards. Epidemiology 12:272-276. 

*Porta M, Jariod M, Malats N, et al. 2000. Prevalence of K-ras mutations at diagnosis and serum levels 

of DDT, DDE, PCBs and other organochlorine compounds in exocrine pancreatic cancer. In: Gress TM, 

ed. Molecular pathogenesis of pancreatic cancer. Amsterdam: IOS Press, 37-44. 

Porta M, Malats N, Jariod M, et al. 1999. Serum concentrations of organochlorine compounds and K-ras 

mutations in exocrine pancreatic cancer. Lancet 354:2125-2129. 

*Porter RD, Wiemeyer SN. 1969. Dieldrin and DDT: Effects on sparrow hawk eggshells and 

reproduction. Science 165:199-200. 

*Porter RD, Wiemeyer SN. 1972. DDE at low dietary levels kills captive American kestrels. Bull 

Environ Contam Toxicol 8(4):193-199. 

*Powers J, Hennigar G, Grooms G, et al. 1974. Adrenal cortical degeneration and regeneration 

following administration of DDD. Am J Pathol 75(1):181-194. 

*Pranzatelli MR, Tkach K. 1992. Regional glycine receptor binding in the p,p'-DDT myoclonic rat 

model. Arch Toxicol 66:73-76. 
DDT, DDE, and DDD 373 

9. REFERENCES 

*Pratt J, Rothwell J, Jenner P, et al. 1985. Myclonus in the rat induced by p,p-DDT and the role of 

altered monoamine function. Neuropharmacology 24(5):361-373. 

*Pratt J, Rothwell J, Jenner P, et al. 1986. p,p'-DDT-induced myoclonus in the rat and its application as 

an animal model of 5-HT-sensitive action myoclonus. In: Fahn S, et al., Advances in neurology. Vol. 

43: II. Myoclonus. New York, NY: Raven Press, 577-588. 

*Preat V, de Gerlache J, Lans M, et al. 1986. Comparative analysis of the effect of phenobarbital, 

dichlorodiphenyltrichloroethane, butylated hydroxytoluene and nafenopin on rat hepatocarcinogenesis. 

Carcinogenesis 7(6):1025-1028. 

Prest HF, Huckins JN, Petty JD, et al. 1995. A survey of recent results in passive sampling of water and 

air by semipermeable membrane devices. Mar Pollut Bull 31:306-312. 

*Preston BD, Miller JA, Miller EC. 1984. Reactions of 2,2',5,5'-tetrachlorobiphenyl 3,4-oxide with 

methionine, cysteine and glutathione in relation to the formation of methylthio-metabolites of 2,2',5,5'­

tetrachlorobiphenyl in the rat and mouse. Chem Biol Interact 50:289-312. 

*Probst G, Hill L. 1980. Chemically-induced DNA repair synthesis in primary rat hepatocytes: A 

correlation with bacterial mutagenicity. Ann NY Acad Sci 349:405-406. 

*Probst G, McMahon R, Hill L, et al. 1981. Chemically-induced unscheduled DNA synthesis in primary 

rat hepatocyte cultures: A comparison with bacterial mutagenicity using 218 compounds. Environ 

Mutagen 3:11-32. 

*Procianoy R, Schvartsman S. 1981. Blood pesticide concentration in mothers and their newborn 

infants. Acta Paediatr Scand 70:925-928. 

*Quensen J, Mueller S, Jain M, et al. 1998. Reductive dechlorination of DDE to DDMU in marine 

sediment microcosms. Science 280:722-724. 

*Quensen JF, Tiedje JM, Jain MK, et al. 2001. Factors controlling the rate of DDE dechlorination to 

DDMU in Palos Verdes margin sediments under anaerobic conditions. Environ Sci Technol 35:286-291. 

*Rabello M, Becak W, Dealmeida W, et al. 1975. Cytogenetic study on individuals occupationally 

exposed to DDT. Mutat Res 28:449. 

*Racke K, Skidmore M, Hamilton D, et al. 1997. Pesticide fate in tropical soils. Pure Appl Chem 

69(6):1349-1371. 

Radomski J, Deichmann W, MacDonald W, et al. 1965. Synergism among oral carcinogens. I. Results 

of the simultaneous feeding of four tumorigens to rats. Toxicol Appl Pharmacol 7:652-656. 

Rajukkannu K, Basha A, Habeebullah B, et al. 1985. Degradation and persistence of DDT, HCH, 

carbaryl and malathion in soil. Indian J Environ Health 27(3):237-243. 

*Ramachandran M, Zaidi S, Banerjee B, et al. 1984. Urinary excretion of DDA: 2,2-bis 

(4-chlorophenyl)acetic acid as an index of DDT exposure in men. Indian J Med Res 80:483-486. 
DDT, DDE, and DDD 374 

9. REFERENCES 

Randall DJ; Brauner CJ. 1991. Toxicant uptake across fish gills. Seventeenth annual aquatic toxicity 

workshop, Vancouver, British Columbia, Canada, November 5-7, 1990. Can Tech Rep Fish Aquat Sci 

1774:501-517. 

Rao V, Krishnaiah K, Sreeraman P, et al. 1978. Effect of oral DDT in male rats. Indian J Exp Biol 

16:1002-1003. 

*Rapaport R, Urban N, Capel P, et al. 1985. "New" DDT inputs to North America: Atmospheric 

deposition. Chemosphere 14(9):1167-1174. 

Rappolt R. 1973. Use of oral DDT in three human barbiturate intoxications: Hepatic enzyme induction 

by reciprocal detoxicants. Clin Toxicol 6(2):147-151. 

Rappolt R, Hale W. 1968. p,p'-DDE and p,p'-DDT residues in human placentas, cords, and adipose 

tissue. Clin Toxicol 1(1):57-61. 

Rattner BA, Pearson JL, Golden NH, et al. 2000. Contaminant exposure and effects - terrestrial 

vertabrates database: Trends and data gaps for Atlantic coast estuaries. Environ Monit Assess 63:131­

142. 

*Rehana T, Rao PR. 1992. Effect of DDT on the immune system in Swiss albino mice during adult and 

perinatal exposure: Humoral responses. Bull Environ Contam Toxicol 48:535-540. 

*Reich AR, Perkins JL, Cutter G. 1986. DDT contamination of a north Alabama aquatic ecosystem. 

Environ Toxicol Chem 5(8):725-736. 

*Reif V, Sinsheimer J. 1975. Metabolism of 2-(o-chlorophenyl)-1-(p-chlorophenyl)-2,2-dichloroethane 

(o,p'-DDD) in rats. Drug Metab Dispos 3(1):15-25. 

Reifenrath WG, Hawkins GS, Kurtz MS. 1991. Percutaneous penetration and skin retention of topically 

applied compounds: An in vitro-in vivo study. J Pharm Sci 80(6):526-532. 

*Reish DJ, Kauwling TJ, Mearns AJ, et al. 1978. Marine and estuarine pollution. J Water Pollut Control 

Fed :1424-1462. 

*Ren P, Mehta P, Ruch R. 1998. Inhibition of gap junctional intercellular communication by tumor 

promoters in connexin43 and connexin32-expressing liver cells: Cell specificity and role of protein 

kinase C. Carcinogenesis 19(1):169-175. 

Rhainds M, Levallois P, Dewailly E, et al. 1999. Lead, mercury, and organochlorine compound levels in 

cord blood in Quebec, Canada. Arch Environ Health 54(1):40-47. 

Rhouma KB, Tebourbi O, Sakly M. 2000. Acute hepatotoxicity of DDT: Effect on glucocorticoid 

receptors and serum transcortin. Indian J Exp Biol 38:452-456. 

Ribeiro ML, Minelli EV. 1991. On-line DDT determination in blood serum: Experimental parameters. 

Bull Environ Contam Toxicol 47(6):804-810. 

*Richie PJ, Peterle TJ. 1979. effect of DDE on circulating luteinizing hormone levels in ring doves 

during courtship and nesting. Bull Environ Contam Toxicol 23:220-226. 
DDT, DDE, and DDD 375 

9. REFERENCES 

*Risebrough RW, Anderson DW. 1975. Some effects of DDE and PCB on mallards and their eggs. J 

Wildl Manage 39(3):508-513. 

*Risebrough R, Jarman W, Springer A, et al. 1986. A metabolic derivation of DDE from kelthane. 

Environ Toxicol Chem 5:13-19. 

*Rivedal E, Mikalsen SO, Sanner T. 2000. Morphological transformatin and effect on gap junction 

intercellular communication in Syrian hamster embryo cells as screening tests for carcinogens devoid of 

mutagenic activity. Toxicol in Vitro 14:185-192. 

*Rivedal E, Yamasaki H, Sanner T. 1994. Inhibition of gap junctional intercellular communication in 

Syrian hamster embryo cells by TPA, retinoic acid and DDT. Carcinogenesis 15(4):689-694. 

*Rivero-Rodriquez L, Borja-Aburto V, Santos-Burgoa C, et al. 1997. Exposure assessment for workers 

applying DDT to control Malaria in Veracruz, Mexico. Environ Health Perspect 105(1):98-101. 

*Roan C, Morgan D, Paschal E. 1971. Urinary excretion of DDA following ingestion of DDT and DDT 

metabolites in man. Arch Environ Health 22:309-315. 

*Roberts DR, Laughlin LL, Hsheih P, et al. 1997. DDT, global strategies, and a malaria control crisis in 

South America. Emerging Infectious Diseases 3:295-302 

Roberts DR, Manguin S, Mouchet J. 2000. DDT house spraying and re-emerging malaria. Lancet 

356:330-332. 

*Robertson BK, Alexander M. 1998. Sequestration of DDT and dieldrin in soil: Disappearance of acute 

toxicity but not the compounds. Environ Toxicol Chem 17(6):1034-1038. 

Robinson J, Hunter C. 1966. Organochlorine insecticides: Concentrations in human blood and adipose 

tissue. Arch Environ Health 13:558-563. 

Robison A, Stancel G. 1982. The estrogenic activity of DDT: Correlation of estrogenic effect with 

nuclear level of estrogen receptor. Life Sci 31:2479-2484. 

*Rodriguez P, Permanyer J, Grases JM, et al. 1991. Confirmation method for the identification and 

determination of some organophosphorus and organochlorine pesticides in cocoa beans by gas 

chromatography-mass spectrometry. J Chromatogr 562:547-553. 

Rogan WJ, Ragan NB. 2000. Hazards to infants in the food chain: Exposure through breast milk. In: 

Aggett PJ, Kuiper HA, eds. Risk assessment in the food chain of children. Philadelphia, PA: Nestec 

Ltd., 163-178. 

*Rogan W, Gladen B, McKinney J. 1986. Polychlorinated biphenyls (PCBs) and dichlorodiphenyl 

dichloroethene (DDE) in human milk: Effects of maternal factors and previous lactation. Am J Public 

Health 76(2):172-177. 

*Rogan W, Gladen B, McKinney J, et al. 1987. Polychlorinated biphenyls (PCBs) and dichlorodiphenyl 

dichloroethene (DDE) in human milk: Effects on growth, morbidity, and duration of lactation. Am J 

Public Health 77(10):1294. 
DDT, DDE, and DDD 376 

9. REFERENCES 

*Rohrig L, Meisch H. 2000. Application of solid phase micro extraction for the rapid analysis of 

chlorinated organics in breast milk. Fresenius J Anal Chem 366:106-111. 

*Rojanapo W, Kupradinun P, Tepsuwan A, et al. 1993. Effect of varying the onset of exposure to DDT 

on its modulation of AFB1-induced hepatocarcinogenesis in the rat. Carcinogenesis 14(4):663-667. 

*Rojanapo W, Tepsuwan A, Kupradinun P, et al. 1988. Modulation of hepatocarcinogenicity of 

aflatoxin B1 by the chlorinated insecticide DDT. In: Nigam ed. Eicosanoids, Lipid Peroxidation and 

Cancer. Springer-Verlag Berlin. 

Romero Y. 1998. Effects of p,p’-DDE on male reproductive organs in peripubertal Wistar rats following 

a single intraperitoneal injection. Fukuoka Acta Med 88(2):64-77. 

*Romieu I, Hernandez-Avila M, Lazcano-Ponce E, et al. 2000. Breast cancer, lactation history, and 

serum organochlorines. Am J Epidemiol 152(4):363-370. 

*Ron M, Cucos B, Rosenn B, et al. 1988. Maternal and fetal serum levels of organochlorine compounds 

in cases of premature rupture of membranes. Acta Obstet Gynecol Scand 67:695-697. 

Roncevic N, Pavkov S, Galetin-Smith R, et al. 1987. Serum concentrations of organochlorine 

compounds during pregnancy and the newborn. Bull Environ Contam Toxicol 38:117-124. 

*Roseng L, Rivedal E, Skaare J, et al. 1994. Effect of 1,1'-(2,2,2-trichloroethylidene)-bis(4­

chlorobenzene) (DDT) on gap junctional intercellular communication and morphological transformation 

of Syrian hamster embryo cells. Chem Biol Interact 90:73-85. 

*Ross MH and Reith EJ. 1985. Histology: A Text and Atlas. New York, NY: Harper & Row. 

Rosselli M, Reinhart K, Imthurn B, et al. 2000. Cellular and biochemical mechanisms by which 

environmental oestrogens influence reproductive function. Hum Reprod 6(4):332-350. 

*Rossi L, Barbieri O, Sanguineti M, et al. 1983. Carcinogenicity study with technical-grade 

dichlorodiphenyltrichloroethane and 1,1-dichloro-2,2-bis(p-chlorophenyl)ethylene in hamsters. Cancer 

Res 43:776-781. 

*Rossi L, Ravera M, Repetti G, et al. 1977. Long-term administration of DDT or phenobarbital-Na in 

Wistar rats. Int J Cancer 19:179-185. 

*Rothe CF, Mattson AM, Nueslein RM, et al. 1957. Metabolism of chlorophenothane (DDT): Intestinal 

lymphatic absorption. Arch Ind Health 16:82-86. 

*Rothman N, Cantor K, Blair A, et al. 1997. A nested case-control study of non-Hodgkin lymphoma and 

serum organochlorine residues. Lancet 350:240-244. 

*Rowlands S, Hall A, McCormick P, et al. 1994. Destruction of toxic materials. Nature 367:223. 

*Rozman K, Ballhorn L, Rozman T. 1983. Mineral oil in the diet enhances fecal excretion of DDT in the 

rhesus monkey. Drug Chem Toxicol 6:311-316. 

RTECS. 1999a. o,p’-DDT - 789-02-6. Registry of Toxic Effects of Chemical Substances. National 

Institute for Occupational Safety and Health. April 19, 1999. 
DDT, DDE, and DDD 377 

9. REFERENCES 

RTECS. 1999b. o,p’-DDD - 53-19-0. Registry of Toxic Effects of Chemical Substances. National 

Institute for Occupational Safety and Health. April 19, 1999. 

RTECS. 1999c. p,p’-DDT - 50-29-3. Registry of Toxic Effects of Chemical Substances. National 

Institute for Occupational Safety and Health. April 19, 1999. 

RTECS. 1999d. p,p’-DDE - 72-55-9. Registry of Toxic Effects of Chemical Substances. National 

Institute for Occupational Safety and Health. April 19, 1999. 

RTECS. 1999e. p,p’-DDD - 72-54-8. Registry of Toxic Effects of Chemical Substances. National 

Institute for Occupational Safety and Health. April 19, 1999. 

*Rubin JG, Payne GT, Soderlund DM. 1993. Structure activity relationships for pyrethroids and DDT 

analogues as modifiers of [3

H] batrachotoxin A 20α-benzoate binding to mouse brain sodium channels. 

Pest Biochem Physiol 45:130-140. 

*Ruch R, Bonney W, Sigler K, et al. 1994. Loss of gap junctions from DDT-treated liver epithelial cells. 

Carcinogenesis 15(2):301-306. 

*Rudolph SG, Anderson DW, Risebrough.RW. 1983. Kestrel predatory behaviour under chronic lowlevel exposure to DDE. Environ Pollut Ser A 32:121-136. 

*Rupa DS, Reddy PP, Sreemannarayana K, et al. 1991. Frequency of sister chromatid exchange in 

peripheral lymphocytes of male pesticide applicators. Environ Mol Mutagen 18:136-138. 

*Rupa DS, Rita P, Reddy PP, et al. 1988. Screening of chromosomal aberrations and sister chromatid 

exchanges in peripheral lymphocytes of vegetable garden workers. Human Toxicol 7:333-336. 

*Rupa DS, Rita P, Reddy PP. 1989. Analysis of sister-chromatid exchanges, cell kinetics and mitotic 

index in lymphocytes of smoking pesticide sprayers. Mutat Res 223:253-258. 

*Russell RW, Hecnar SJ, Haffner GD. 1995. Organochlorine pesticide residues in Southern Ontario 

spring peepers. Environ Toxicol Chem 14(5):815-817. 

*Saady JJ, Fitzgerald RL, Poklis A. 1992. Measurement of DDT and DDE from serum, fat, and stool 

specimens. J Environ Sci Health A27(4):967-981. 

*Sabljic A. 1984. Predictions of the nature and strength of soil sorption of organic pollutants by 

molecular topology. J Agric Food Chem 32:243-246. 

*Safe S. 1995. Environmental and dietary estrogens and human health: Is there a problem? Environ 

Health Perspect 103(4):346-351. 

*Safe S. 1998a. Interactions between hormones and chemicals in breast cancer. Annu Rev Pharmacol 

Toxicol 38:121-158. 

Safe S. 1998b. Male reproductive capacity and exposure to environmental chemicals. In: Eisenbrand G, 

ed. Hormonally active agents in food: Symposium. Weinheim, Germany: Wiley-VCH, 253-260. 
DDT, DDE, and DDD 378 

9. REFERENCES 

*Safe S, Zackarewski T. 1997. Organochlorine exposure and risk for breast cancer. In: Aldaz CM, 

Gould MN, McLachlan J, et al, eds. Etiology of breast and gynecological cancers. New York, NY: 

Wiley-Liss, Inc., 133-145. 

*Safe S, Connor K, Ramamoorthy K, et al. 1997. Human exposure to endocrine-active chemicals: 

Hazard assessment problems. Reg Toxicol Pharmacol 26:52-58. 

Safe S, Connor K, Gaido K. 1998. Methods for xenoestrogen testing. Toxicol Lett 102-103:665-670. 

Sala M, Ribas-Fito N, de Muga ME, et al. 1999. Hexachlorobenzene and other organochlorine 

compounds incorporation to the new-borns and its effects on neonatal neurological development at 6-8 

weeks of life. Organohalogen Compounds 44:241-242. 

*Sala M, Ribas-Fitó N, Cardo E, et al. 2001. Levels of hexachlorobenzene and other organochlorine 

compounds in cord blood: exposure across placenta. Chemosphere 43:895-901. 

Salonen L, Vaajakorpi H. 1976. Bioaccumulation of 14C-DDT in a small pond. Environ Qual Safety 

5:130-140. 

*Samuel T, Pillai MKK. 1990. Effect of temperature and solar radiations on volatilisation, 

mineralisation and degradation of [14C] DDT in soil. Environ Pollut 57:63-77. 

Sandhu S, Warren W, Nelson P. 1978. Pesticidal residue in rural potable water. J Am Water Works 

Assoc 41-45. 

*Sanger DM, Holland AF, Scott GI. 1999. Tidal creek and salt marsh sediments in South Carolina 

coastal estuaries: II. Distribution of organic contaminants. Arch Environ Contam Toxicol 37:458-471. 

*Sanyal S, Agarwal N, Subrahmanyam D. 1986. Effect of acute sublethal and chronic administration of 

DDT (chlorophenothane) on brain lipid metabolism of Rhesus monkeys. Toxicol Lett 34:47-54. 

*Sasaki K, Kawasaki Y, Sekita K, et al. 1992. Disposition of β-hexachlorocyclohexane, p,p'-DDT, and 

trans-chlordane administered subcutaneously to monkeys (Macaca fascicularis). Arch Environ Contam 

Toxicol 22:25-29. 

Sasaki K, Ishizaka T, Suzuki T, et al. 1991a. Accumulation levels of organochlorine pesticides in human 

adipose tissue and blood. Bull Environ Contam Toxicol 46(5):662-669. 

*Sasaki K, Ishizaka T, Suzuki T, et al. 1991b. Organochlorine chemicals in skin lipids as an index of 

their accumulation in the human body. Arch Environ Contam Toxicol 21:190-194. 

Sathiakumar N, Delzell E, Austin H, et al. 1992. A follow-up study of agricultural chemical production 

workers. Am J Ind Med 21(3):321-330. 

*Sax E. 1979. Dangerous properties of industrial materials. 5th ed. New York, NY: Van Nostrand 

Reinhold, 534-535. 

*Saxena M, Siddiqui M, Bhargava A, et al. 1980. Role of chlorinated hydrocarbon pesticides in 

abortions and premature labor. Toxicology 17:323-331. 
DDT, DDE, and DDD 379 

9. REFERENCES 

*Saxena M, Siddiqui M, Agarwal V, et al. 1983. A comparison of organochlorine insecticide contents in 

specimens of maternal blood, placenta, and umbilical-cord blood from stillborn and live-born cases. J 

Toxicol Environ Health 11:71-79. 

*Saxena M, Siddiqui M, Seth T, et al. 1981. Organochlorine pesticides in specimens from women 

undergoing spontaneous abortion, premature or full-term delivery. J Anal Toxicol 5:6-9. 

*Saxena S, Khare C, Farooq A, et al. 1987. DDT and its metabolites in leiomyomatous and normal 

human uterine tissue. Arch Toxicol 59:453-455. 

Schantz SL, Jacobson JL, Humphrey HEB, et al. 1994. Determinants of polychlorinated biphenyls 

(PCBs) in the sera of mothers and children from Michigan farms with PCB-contaminated silos. Arch 

Environ Health 49(6):452-458. 

*Schattenberg H, Hsu J-P. 1992. Pesticide residue survey of produce from 1989 to 1991. J Assoc Off 

Anal Chem Int 75(5):925-933. 

*Schecter A, Toniolo P, Dai L, et al. 1997. Blood levels of DDT and breast cancer risk among women 

living in the north of Vietnam. Arch Environ Contam Toxicol 33:453-456. 

*Scheele J. 1998. A comparison of the concentrations of certain pesticides and polychlorinated 

hydrocarbons in bone marrow and fat tissue. J Environ Pathol Toxicol Oncol 17(1):65-68. 

*Scheele J, Teufel M, Niessen K-H. 1995. A comparison of the concentrations of certain chlorinated 

hydrocarbons and polychlorinated biphenyls in bone marrow and fat tissue of children and their 

concentrations in breast milk. J Environ Pathol Toxicol Oncol 14(1):11-14. 

*Schildkraut JM, Demark-Wahnefried W, DeVoto E, et al. 1999. Environmental contaminant and body 

fat distribution. Cancer Epidemiol Biomarkers Prev 8:179-183. 

Schlebusch H, Wagner U, Schneider C, et al. 1994. Placental transfer of chlorinated hydrocarbons in 

humans [Abstract]. Clin Chem 40(6):1084. 

*Schmid K, Lederer P, Goen T, et al. 1997. Internal exposure to hazardous substances of persons from 

various continents: Investigations on exposure to different organochlorine compounds. Int Arch Occup 

Environ Health 69:399-406. 

*Schmitt CJ, Zajicek JL, Peterman PH. 1990. National contaminant biomonitoring program: Residues 

of organochlorine chemicals in U.S. freshwater fish, 1976-1984. Arch Environ Contam Toxicol 19:748­

781. 

*Schneider J, Huh MM, Bradlow HL, et al. 1984. Antiestrogen action of 2-hydroxyestrone on MCF-7 

human breast cancer cells. J Biol Chem 259(8):4840-4845. 

*Schteingart DE. 2000. Conventional and novel strategies in the treatment of adrenocortical cancer. 

Braz J Med Biol Res 33:1197-1200. 

Schteingart D, Sinsheimer J, Counsell R, et al. 1993. Comparison of the adrenalytic activity of mitotane 

and a methylated homolog on normal adrenal cortex and adrenal cortical carcinoma. Cancer Chemother 

Pharmacol 31:459-466. 
DDT, DDE, and DDD 380 

9. REFERENCES 

*Schulte-Hermann R. 1974. Induction of liver growth by xenobiotic compounds and other stimuli. CRC 

Crit Rev Toxicol 3:97-158. 

*Schwack W. 1988. Photoinduced additions of pesticides to biomolecules. 2. Model reactions of DDT 

and methoxychlor with methyl oleate. J Agric Food Chem 36:645-648. 

Schulze J, Manigold D, Andrews F. 1973. Pesticides in selected western streams—1968-71. Pestic 

Monit J 7:73-84. 

*Scott ML, Zimmerman JR, Marinsky S, et al. 1975. Effects of PCBs, DDT, and mercury compounds 

upon egg production, hatchability and shell quality in chickens and Japanese quail. Poult Sci 54:350-368. 

Seiber JN. 1991. Analytical chemistry and pesticide regulation. In: Marco GJ, Hollingworth RM, 

Plimmer JR, eds. Regulation of agrochemicals: A driving force in their evolution. Washington, DC: 

American Chemical Society, 101-120. 

*Seidensticker JC. 1968. Response of juvenile raptors to DDT in the diet. Helena, Montana: University 

of Montana. 

Seifert M, Haindl S, Hock B. 1999. Development of an enzyme linked receptor assay (ELRA) for 

estrogens and xenoestrogens. Anal Chim Acta 386:191-199. 

*Seiler J. 1977. Inhibition of testicular DNA synthesis by chemical mutagens and carcinogens. 

Preliminary results in the validation of a novel short term test. Mutat Res 46:305-310. 

*Seiler P, Fischer B, Lindenau A, et al. 1994. Effects of persistent chlorinated hydrocarbons on fertility 

and embryonic development in the rabbit. Hum Reprod 9:1920-1926. 

Serat W, Lee M, VanLoon A, et al. 1977. DDT and DDE in the blood and diet of Eskimo children from 

Hooper Bay, Alaska. Pestic Monit J 11:1-4. 

*Setchell BP, Waites GMH. 1975. The blood testis barrier. In: Creep RO, Astwood EB, Geiger SR, 

eds. Handbook of physiology: Endocrinology V. Washington, DC: American Physiological Society, 

143-172. 

*Shabad L, Kolesnichenko T, Nikonova T. 1973. Transplacental and combined long-term effect of DDT 

in five generations of A-strain mice. Int J Cancer 11:688-693. 

Sharom M, Miles J. 1981. The degradation of parathion and DDT in aqueous systems containing organic 

additives. J Environ Sci Health B16(6):703-711. 

Sheeler CQ, Dudley MW, Khan SA. 2000. Environmental estrogens induce transcriptionally active 

estrogen receptor dimers in yeast: Activity potentiated by the coactivator RIP140. Environ Health 

Perspect 108(2):97-103. 

*Shekhar P, Werdell J, Basrur V. 1997. Environmental estrogen stimulation of growth and estrogen 

receptor function in preneoplastic and cancerous human breast cell lines. J Natl Cancer Inst 89(23):1774­

1782. 

*Shelby M, Newbold R, Tully D, et al. 1996. Assessing environmental chemicals for estrogenicity using 

a combination of in vitro and in vivo assays. Environ Health Perspect 104(12):1296-1300. 
DDT, DDE, and DDD 381 

9. REFERENCES 

*Shellenberger TE. 1978. A multi-generation toxicity evaluation of p,p'-DDT and dieldrin with Japanese 

quail: 1. Effects on growth and reproduction. Drug Chem Toxicol 1(2):137-146. 

*Shen K, Novak R. 1997a. DDT stimulates c-erbB2, c-met, and STATS tyrosine phosphorylation, Grb2­

SOS association, MAPK phosphorylation, and proliferation of human breast epithelial cells. Biochem 

Biophys Res Commun 231:17-21. 

*Shen K, Novak R. 1997b. Differential effects of Aroclors and DDT on growth factor gene expression 

and receptor tyrosine kinase activity in human breast epithelial cells. In: Honn ed. Eicosanoids and 

other bioactive lipids in cancer inflammation and radiation injury. New York, NY: Plenum Press, 295­

302. 

*Shiplov J, Graber C, Keil J, et al. 1972. Effect of DDT on antibody response to typhoid vaccine in 

rabbits and man. Immunol Commun 1(4):385-394. 

*Shirasu et al. 1976. Mutagenicity screening of pesticides in the microbial system. Mutat Res 40:19. 

Shivapurkar N, Hoover K, Poirier L. 1986. Effect of methionine and choline on liver tumor promotion 

by phenobarbital and DDT in diethylnitrosamine-initiated rats. Carcinogenesis 7:547-550. 

*Sieber S. 1976. The lymphatic absorption of p,p'-DDT and some structurally-related compounds in the 

rat. Pharmacology 14:443-454. 

*Sieber SM, Cohn VH, Wynn, WT. 1974. The entry of foreign compounds into the thoracic duct lymph 

of rat. Xenobiotica 4(5):265-284. 

*Sierra-Santoyo A, Hernandez M, Albores A, et al. 2000. Sex-dependent regulation of hepatic 

cytochrome P-450 by DDT. Toxicol Sci 54:81-87. 

*Silinskas K, Okey A. 1975. Protection by 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT) against 

mammary tumors and leukemia during prolonged feeding of 7,12-dimethylbenz [a]anthracene to female 

rats. J Natl Cancer Inst 55(3):653-657. 

*Singh BK, Kuhad RC, Singh A, et al. 1999. Biochemical and molecular basis of pesticide degradation 

by microorganisms. Crit Rev Biotechnol 19(3):197-225. 

*Singhal R, Valadares J, Schwark W. 1970. Metabolic control mechanisms in mammalian systems. 

Biochem Pharmacol 19:2145. 

Sjödin HL, Erfurth EM, Axmon A, et al. 1998. Dietary exposure to persistent organochlorines and 

pituitary thyroid and testosterone hormone levels in male adults. Organohalogen Compounds 37:229­

232. 

*Sjoeib F, Anwar E, Tungguldihardjo M. 1994. Laboratory studies of dissipation of 14C-DDT from soil 

and plywood surfaces. J Environ Sci Health B29(1):153-159. 

*Slebos RJC, Hoppin JA, Tolbert PE, et al. 2000. K-ras and p53 in pancreatic cancer: Association with 

medical history, histopathology, and environmental exposures in a population-based study. Cancer 

Epidemiol Biomarkers Prev 9:1223-1232. 
DDT, DDE, and DDD 382 

9. REFERENCES 

*Sleicher C, Hopcraft J. 1984. Persistence of pesticides in surface soil and relation to sublimation. 

Environ Sci Technol 18(7):514-518.
 

Smith AG. 2000. How toxic is DDT? Lancet 356:267-268.
 

*Smith D. 1999. Worldwide trends in DDT levels in human breast milk. Int J Epidemiol 28:179-188.
 

*Sobotka T. 1971. Behavioral effects of low doses of DDT. Proc Soc Exp Biol Med 137:952-955.
 

*Sohoni P, Sumpter J. 1998. Several environmental oestrogens are also anti-androgens. J Endocrinol
 

158:327-339.
 

*Soloman K. 1998. Endocrine-modulating substances in the environment: The wildlife connection. Int
 

J Toxicol Occup Environ Health 17:159-172.
 

Soto A, Sonnenschein C. 1985. The role of estrogens on the proliferation of human breast tumor cells
 

(MCF-7). Steroid Biochemistry 23(1):87-94.
 

*Soto A, Fernandez M, Luizzi M, et al. 1997. Developing a marker of exposure to xenoestrogen
 

mixtures in human serum. Environ Health Perspect Suppl 105(3):647-654.
 

*Soto AM, Lin FM, Justicia H, et al. 1998. An "in culture" bioassay to assess the estrogenicity of
 

xenobiotics (E-screen). J Clean Technol Environ Toxicol Occup Med 7(3):331-343.
 

*Spencer WF, Singh G, Taylor CD, et al. 1996. DDT persistence and volatility as affected by
 

management practices after 23 years. J Environ Qual 25:815-821.
 

*Sperry TS, Thomas P. 1999. Identification of two nuclear androgen receptors in kelp bass (Paralabrax
 

clathratus) and their binding affinities for xenobiotics: comparison with Atlantic croaker (Micropogonias
 

undulatus) androgen receptors. Biol Reprod 61:1152-1161. 

*Stanley C, Barney J, Helton M, et al. 1971. Measurement of atmospheric levels of pesticides. Environ
 

Sci Technol 5(5):430-435.
 

*Staples C, Werner A, Hoogheem T. 1985. Assessment of priority pollutant concentrations in the United
 

States using STORET database. Environ Toxicol Chem 4:131-142.
 

*Stehr-Green PA. 1989. Demographic and seasonal influences on human serum pesticide residues. J
 

Toxicol Environ Health 27:405-421.
 

Steinberg FM. 2001. It is time to dismiss calls to ban DDT. Br Med J 322(7287):676-677.
 

*Stellman SD, Djordjevic MV, Muscat JE, et al. 1998. Relative abundance of organochlorine pesticides
 

and polychlorinated biphenyls in adipose tissue and serum of women in Long Island, New York. Cancer
 

Epidemiol Biomarkers Prev 7:489-4986.
 

Steuerwald U, Weihe P, Jorgensen PJ, et al. 2000. Maternal seafood diet, methylmercury exposure, and
 

neonatal neurological function. J Pediatr 136:599-605.
 

Stevens MF, Ebell GF, Psaila-Savona P. 1993. Organochlorine pesticides in Western Australian nursing
 

mothers. Med J Aust 158:238-241.
 
DDT, DDE, and DDD 383 

9. REFERENCES 

*Stewart D, Chisholm D. 1971. Long-term persistence of BHC, DDT and chlordane in a sandy loam 

soil. Can J Soil Sci 51:379-383. 

Stewart P, Pagano J, Sargent D, et al. 2000. Effects of Great Lakes fish consumption on brain PCB 

pattern, concentration, and progressive-ratio performance. Environ Res 82:18-32. 

*Stickel LF, Stickel WH, Christensen R. 1966. Residues of DDT in brain and bodies of birds that died 

on dosage and in survivors. Science 151:1549-1551. 

Storelli MM, Marcotrigiano GO. 2000a. Chlorobiphenyls, HCB, and organochlorine pesticides in some 

tissues of Caretta caretta (Linnaeus)specimens beached along the Adriatic Sea, Italy. Bull Environ 

Contam Toxicol 64:481-488. 

Storelli MM, Marcotrigiano GO. 2000b. Environmental contamination in Bottlenose dolphin (Tursiops 

truncatus): Relationship between levels of metals, methylmercury, and organochlorine compounds in an 

adult female, her neonate, and a calf. Bull Environ Contam Toxicol 64:333-340. 

Strassman S, Kutz F. 1977. Insecticide residues in human milk from Arkansas and Mississippi. Pestic 

Monit J 10:130. 

*Streck G, Herrmann R. 2000. Distribution of endocrine disrupting semivolatile organic compounds in 

several compartments of a terrestrial ecosystem. Water Sci Technol 42:39-44 

*Street J, Chadwick R. 1967. Stimulation of dieldrin metabolism by DDT. Toxicol Appl Pharmacol 

11:68-71. 

*Street J, Sharma R. 1975. Alteration of induced cellular and humoral immune responses by pesticides 

and chemicals of environmental concern: Quantitative studies on immunosuppression by DDT, Aroclor 

1254, carbaryl, carbofuran, and methylparathion. Toxicol Appl Pharmacol 32:587-602. 

Stringer RL, Jacobs MN, Johnston PA, et al. 1996. Organochlorine residues in fish oil dietary 

supplements. Organohalogen compounds 28:551-556. 

*Strom P. 2000. Pesticides in yard waste compost. Compost Sci Util 8:54-60 

*Strompl C, Thiele J. 1997. Comparative fate of 1,1-diphenylethylene(DPE), 1,1-dichloro-2,2-bis(4­

chlorophenyl)-ethylene(DDE), and pentachlorophenol(PCP) under alternating aerobic and anaerobic 

conditions. Arch Environ Contam Toxicol 33:350-356. 

*Sturgeon SR, Brock JW, Potischman N, et al. 1998. Serum concentrations of organochlorine 

compounds and endometrial cancer risk (United States). Cancer Causes Control 9:417-424. 

Subba-Rao R, Alexander M. 1985. Bacterial and fungal cometabolism of 1,1,1-trichloro-2,2-bis­

(4-chlorophenyl)ethane (DDT) and its breakdown products. Appl Environ Microbiol 49:509-516. 

Suedel BC, Boraczek JA, Peddicord RK, et al. 1994. Trophic transfer and biomagnification potential of 

contaminants in aquatic ecosystems. Rev Environ Contam Toxicol 136:21-59. 

*Suggs J, Hawk R, Curley A, et al. 1970. DDT metabolism: Oxidation of the metabolite 2,2-bis­

(p-chlorophenyl) ethanol by alcohol dehydrogenase. Science 168:582. 
DDT, DDE, and DDD 384 

9. REFERENCES 

*Sunami O, Kunimatsu T, Yamada T, et al. 2000. Evaluation of a 5-day Hershberger assay using young 

mature male rats: methyltestosterone and p,p'-DDE, but not fenitrothion, exhibited androgenic or 

antiandrogenic activity in vivo. J Toxicol Sci 25(5):403-415. 

Sundstrom G. 1977. Metabolic hydroxylation of the aromatic rings of 1,1-dichloro-2,2-bis­

(4-chlorophenyl)ethylene (p,p'-DDE) by the rat. J Agric Food Chem 25(1):18-21. 

*Svensson B-G, Hallberg T, Nilsson A, et al. 1994. Parameters of immunological competence in 

subjects with high consumption of fish contaminated with persistent organochlorine compounds. Int 

Arch Occup Environ Health 65:351-358. 

*Swaneck GE, Fishman J. 1988. Covalent binding of the endogenous estrogen 16α-hydroxyestrone to 

estradiol receptor in human breast cancer cells: Characterization and intranuclear localization. Proc Natl 

Acad Sci U S A 85:7831-7835 

Swann RL, McCall PJ, Laskowski DA, et al. 1981. Estimation of soil sorption constants of organic 

chemicals by high-performance liquid chromatography. ASTM Spec Tech Pub 737:43-48. 

Syracuse Research Corporation. 1999. EPIWIN database. 

Tabak H, Quave S, Mashni C, et al. 1981. Biodegradability studies with organic priority pollutant 

compounds. J Water Pollut Control Fed 53:1503-1518. 

*Tadi-Uppala P, Sathiakumar N, Lewis M, et al. 1998. Cancer in residents of Triana, Alabama exposed 

to DDT [Abstract]. Proc Am Assoc Cancer Res 39:99. 

*Takayama S, Sieber SM, Dalgard DW, et al. 1999. Effects of long-term oral administration of DDT on 

nonhuman primates. Z Krebsforsch Klin Onkol 125:219-225. 

*Takei G, Kauahikaua S, Leong G. 1983. Analyses of human milk samples collected in Hawaii for 

residues of organochlorine pesticides and polychlorobiphenyls. Bull Environ Contam Toxicol 30:606­

613. 

*Talts U, Talts JF, Eriksson P. 1998. Differential expression of muscarinic subtype mRNAs after 

exposure to neurotoxic pesticides. Neurobiol Aging 19(6):553-559. 

*Tanaka T, Mori H, Williams G. 1987. Enhancement of dimethylnitrosamine-initiated 

hepatocarcinogenesis in hamsters by subsequent administration of carbon tetrachloride but not 

phenobarbital or p,p’-dichlorodiphenyltrichloroethane. Carcinogenesis 8(9):1171-1178. 

Tarjan R. 1970. Multigeneration studies on DDT. Food Cosmet Toxicol 8:479-481. 

*Tarjan R, Kemeny T. 1969. Multigeneration studies on DDT in mice. Food Cosmet Toxicol 

7:215-222. 

Tas JW, Keizer A, Opperhuizen A. 1996. Bioaccumulation and lethal body burden of four triorganotin 

compounds. Bull Environ Contam Toxicol 57:146-154. 

*Tateno C, Ito S, Tanaka M, et al. 1994. Effect of DDT on hepatic gap junctional intercellular 

communication in rats. Carcinogenesis 15(3):517-521. 
DDT, DDE, and DDD 385 

9. REFERENCES 

Telang N, Suto A, Wong G, et al. 1992. Induction by estrogen metabolite 16α-hydroxyestrone of 

genotoxic damage and aberrant proliferation in mouse mammary epithelial cells. J Nat Cancer Inst 

84(8):634-638. 

*Telang S, Tong C, Williams G. 1981. Induction of mutagenesis by carcinogenic polycyclic aromatic 

hydrocarbons but not by organochlorine pesticides in the ARL/HGPRT mutagenesis assay. Environ 

Mutagen 3:359. 

Terracini B. 1970. Multigeneration studies on DDT. Food Cosmet Toxicol 8:478-479. 

*Terracini B, Testa M, Cabral J, et al. 1973. The effects of long-term feeding of DDT to balb/c mice. Int 

J Cancer 11:747-764. 

Terzolo M, Pia A, Berruti A, et al. 2000. Low-dose monitored mitotane treatment achieves the 

therapeutic range and manageable side effects in patients with adrenocortical cancer. J Clin Endocrinol 

Metab 85(6):2234-2238. 

Tessari JD, Archibeque-Engle SL. 1999. Correspondence re: S.D. Stellman et al., Relative abundance of 

organochlorine pesticides and polychlorinated biphenyls in adipose tissue and serum of women in Long 

Island, New York. Cancer Epidemiol. Biomark. Prev., 7: 489-496, 1998. Cancer Epidemiol Biomarkers 

Prev 8:111-114. 

Teufel M, Niessen KH, Sartoris J, et al. 1990. Chlorinated hydrocarbons in fat tissue: Analyses of 

residues in healthy children, tumor patients, and malformed children. Arch Environ Contam Toxicol 

19:646-652. 

*Thomas RG. 1990. Volatilization from soil. In: Lyman WJ, Reehl WF, Rosenblatt DH, eds. 

Handbook of chemical property estimation methods: environmental behavior of organic compounds. 

Washington, DC: American Chemical Society, 16-25 to 16-28. 

*Thomas P. 1999. Nontraditional sites of endocrine disruption by chemicals on the hypothalamuspituitary-gonadal axis: interactions with steroid membrane receptors, monoaminergic pathways, and 

signal transduction systems. In: Naz RK, ed. Endocrine Disruptors: Effects on Male and Female 

Reproductive Systems. New York, NY: CRC Press. 

*Thorpe E, Walker A. 1973. The toxicology of dieldrin. II. Comparative long-term oral toxicity studies 

in mice with dieldrin, DDT, phenobarbitone, beta-BHC and alpha-BHC. Food Cosmet Toxicol 

11:433-442. 

*Tiemann U, Pohland R. 1999. Inhibitory effects of organochlorine pesticides on intercellular transfer of 

lucifer yellow in cultured bovine oviductal cells. Reprod Toxicol 13(2):123-130. 

*Tilson H, Hudson P, Hong J. 1986. 5,5-Diphenylhydantoin antagonizes neurochemical and behavioral 

effects of p,p’-DDT but not of chlordecone. J Neurochem 47(6): 1870-1878. 

*Tilson H, Shaw S, McLamb R. 1987. The effects of lindane, DDT, and chlordecone on avoidance 

responding and seizure activity. Toxicol Appl Pharmacol 88:57-65. 

To-Figueras J, Gomez-Catalan J, Rodamilans M, et al. 1988. Mobilization of stored hexachlorobenzene 

and p,p'-dichlorodiphenyldichloroethylene during partial starvation in rats. Toxicol Lett 42:79-86. 
DDT, DDE, and DDD 386 

9. REFERENCES 

Tomatis L, Huff J. 2000. Evidence of carcinogenicity of DDT in nonhuman primates. J Cancer Res Clin 

Oncol 126(4):246. 

*Tomatis L, Turusov V. 1975. Studies in the carcinogenicity of DDT. Gann 17:219-241. 

*Tomatis L, Turusov V, Charles R, et al. 1974a. Effect of long-term exposure to 1,1-dichloro-2,2-bis 

(p-chlorphenyl)ethylene, to 1,1-dichloro-2,2-bis(p-chlorophenyl)ethane and to the two chemicals 

combined on CF-1 mice. J Natl Cancer Inst 52(3):883-891. 

*Tomatis L, Turusov V, Charles R, et al. 1974b. Liver tumours in CF-1 mice exposed for limited periods 

to technical DDT. Cancer Res Clin Oncol 82:25-35. 

*Tomatis L, Turusov V, Day N, et al. 1972. The effect of long-term exposure to DDT on CF-1 mice. Int 

J Cancer 10:489-506. 

*Tong C, Fazio M, Williams G. 1981. Rat hepatocyte-mediated mutagenesis of human cells by 

carcinogenic polycyclic aromatic hydrocarbons but not organochlorine pesticides (41217). Proc Soc Exp 

Biol Med 167:572-575. 

*Torres-Arreola L, Lopez-Carrillo L, Torres-Sanchez L, et al. 1999. Levels of dichloro-dyphenyltrichloroethane (DDT) metabolites in maternal milk and their determinant factors. Arch Environ Health 

54(2):124-129. 

Touitou Y, Moolenaar A, Bogdan A, et al. 1985. o,p'-DDD (mitotane) treatment for Cushing's 

syndrome: Adrenal drug concentration and inhibition in vitro of steroid synthesis. Eur J Clin Pharmacol 

29:483-487. 

Trebicka-Kwiatkowska B, Radomanski T, Stec J. 1971. [A case of intrauterine fetal death in a woman 

exposed to pesticides during her work.] Polski Tydognik Lekarski 26:1862-1863. (Czechoslovakian) 

*Treon J, Cleveland F. 1955. Pesticide toxicity: Toxicity of certain chlorinated hydrocarbon insecticides 

for laboratory animals with special reference to aldrin and dieldrin. J Agric Food Chem 3(5):402-408 

*Treon J, Boyd J, Berryman G, et al. 1954. Final report on the effects on the reproductive capacity of 

three generations of rats being fed on diets containing aldrin, dieldrin, or DDT. Kettering Laboratory, 

University of Cincinnati, College of Medicine, Cincinnati, OH. 

TRI96. 1999. Toxic Chemical Release Inventory. National Library of Medicine, National Toxicology 

Information Program, Bethesda, MD. 

Trosko JE, Chang C-C, Upham B, et al. 1998. Epigenetic toxicology as toxicant-induced changes in 

intracellular signaling leading to altered gap junctional intercellular communication. Toxicol Lett 102­

103:71-78. 

*Tsushimoto G, Chang C, Trosko J, et al. 1983. Cytotoxic, mutagenic, and cell-cell communication 

inhibitory properties of DDT, lindane, and chlordane on Chinese hamster cells in vitro. Arch Environ 

Contam Toxicol 12:721-730. 

*Tully DB, Cox VT, Mumtaz MM, et al. 2000. Six high-priority organochlorine pesticides, either singly 

or in combination, are nonestrogenic in transfected HeLa cells. Reprod Toxicol 14:95-102. 
DDT, DDE, and DDD 387 

9. REFERENCES 

*Turner J, Shanks V. 1980. Absorption of some organochlorine compounds by the rat small intestine 

in vivo. Bull Environ Contam Toxicol 24:652-655. 

*Turusov V, Day N, Tomatis L, et al. 1973. Tumors in CF-1 mice exposed for six consecutive 

generations to DDT. J Natl Cancer Inst 51(3):983-997. 

Turusov V, Rakitsky V, Tomatis L. 2002. Dichlorodiphenyltrichloroethane (DDT): Ubiquity, 

persistence, and risks. Environ Health Perspect 110(2):125-128. 

*UATW. 1999a. Unified Air Toxics Website, Colorado. U.S. Environmental Protection Agency, Office 

of Air Quality Planning and Standards. wysiwyg://17/http://www.epa.gov/ttn/uatw/stprogs.html. May 26, 

1999. 

*UATW. 1999b. Unified Air Toxics Website, Alabama. U.S. Environmental Protection Agency, Office 

of Air Quality Planning and Standards. wysiwyg://17/http://www.epa.gov/ttn/uatw/stprogs.html. May 26, 

1999. 

*UATW. 1999c. Unified Air Toxics Website, Hawaii. U.S. Environmental Protection Agency, Office 

of Air Quality Planning and Standards. wysiwyg://17/http://www.epa.gov/ttn/uatw/stprogs.html. May 5, 

1999. 

*UATW. 1999d. Unified Air Toxics Website, Idaho. U.S. Environmental Protection Agency, Office of 

Air Quality Planning and Standards. wysiwyg://17/http://www.epa.gov/ttn/uatw/stprogs.html. May 5, 

1999. 

Ulrich EM, Caperell-Grant A, Jung S-H, et al. 2000. Environmentally relevant xenoestrogen tissue 

concentrations correlated to biological responses in mice. Environ Health Perspect 108(10):973-977. 

*Unger M, Kiaer H, Blichert-Toft M, et al. 1984. Organochlorine compounds in human breast fat from 

deceased with and without breast cancer and in a biopsy material from newly diagnosed patients 

undergoing breast surgery. Environ Res 34:24-28. 

*Unger M, Olsen J, Clausen J. 1982. Organochlorine compounds in the adipose tissue of deceased 

persons with and without cancer: A statistical survey of some potential confounders. Environ Res 

29:371-376. 

*Uppal R, Garg B, Ahmad A. 1983. Effect of malathion and DDT on the action of chlorpromazine and 

diazepam with reference to conditioned avoidance response in rats. Indian J Exp Biol 21:254-257. 

Uppal R, Garg B, Ahmad A. 1984. Neurologic effects of malathion and DDT. Indian Vet J 61:201-210. 

*U.S. Fish and Wildlife Service. 1963. Pesticide-wildlife studies: a review Fish and Wildlife Service 

investigations during 1961 and 1962. Washington, DC: U.S. Fish and Wildlife Service. Circular 167, 

74-96. 

*U.S. Fish and Wildlife Service. 1965. Effects of pesticides in fish and wildlife. Washington, DC: U.S. 

Fish and Wildlife Service. Circular 226. 

*U.S. Fish and Wildlife Service. 1975. Lethal dietary toxicities of environmental pollutants to birds. 

Special Scientific Report –Wildlife No. 191. Washington, DC: U.S. Fish and Wildlife Service, 

Department of the Interior. 
DDT, DDE, and DDD 388 

9. REFERENCES 

*U.S. Fish and Wildlife Service. 1984. Handbook of toxicity of pesticides to wildlife. 2nd ed. 

Washington, DC: U.S. Fish and Wildlife Service, Department of the Interior. Resource publication 153. 

*USC. 2001. Hazardous air pollutants. United States Code. 42 USC 7412. 

http://www4.law.cornell.edu/uscode/42/7412.text.html. November 29, 2001. 

*USDA. 2001. Labeling treated seed. United States Department of Agriculture. Code of Federal 

Regulations. 7 CFR 201.31a©)(2). http://frwebgate.access.gpo.gov/cgi-bin/...PART=201 

&SECTION=31a&YEAR=2001&TYPE=TEXT. September 24, 2001. 

*USGS. 1999. U.S. Geological Survey. The quality of our nation’s waters: Nutrients and pesticides. 

Reston, VA: U.S. Department of the Interior, U.S. Geological Survey Circular 1225. 

Vancheeswaran S, Hyman MR, Semprini L. 1999. Anaerobic biotransformation of trichlorofluoroethane 

in groundwater microsms. Environ Sci Technol 33:2040-2045. 

*Vangilder LD, Peterie TJ. 1980. South Louisiana crude oil and DDE in the diet of mallard hens: 

Effects on reproduction and duckling survival. Bull Environ Contam Toxicol 25:23-28. 

*Van Metre P, Callender E. 1997. Water quality trends in White Rock Creek Basin from 1912-1994 

identified using sediment cores from White Rock Lake Reservoir, Dallas, Texas. Journal of 

Paleolimnology 17:239-249. 

*Van Metre P, Callender E, Fuller C. 1997. Historical trends in organochlorine compounds in river 

basins identified using sediment cores from reservoirs. Environ Sci Technol 31:2339-2344. 

*van't Veer P, Lobbezoo I, Martin-Moreno J, et al. 1997. DDT (dicophane) and postmenopausal breast 

cancer in Europe: A case-control study. Br Med J 315:81-85. 

*Van Velzen A, Kreitzer. 1975. The toxicity of p,p'-DDT to the clapper rail. J Wildl Manage 39(2):305­

309. 

*Van Wendel de Joode B, Wesseling C, Kromhout H, et al. 2001. Chronic nervous-system effects of 

long-term occupational exposure to DDT. Lancet 357:1014-1016. 

*Veith G, DeFoe D, Bergstedt B. 1979. Measuring and estimating the bioconcentration factor of 

chemicals in fish. J Fish Res Board Can 36:1040-1048. 

*Velbinger H. 1947a. [On the question of DDT toxicity of man.] Dtsch Gesundheitwes 2:355-358. 

(German) 

*Velbinger H. 1947b. [Contribution to the toxicology of "DDT"-agent, dichlorodiphenyltrichloromethylmethane.] Pharmazie 2:268-274. (German) 

*Venkatesan, MI. 1996. DDTs and dumpsite in the Santa Monica Basin, California. Sci Total Environ 

179:61-71. 

*Verma A, Pillai MKK. 1991a. Bioavailability of soil-bound residues of DDT and HCH to certain 

plants. Soil Biol Biochem 23(4):347-351. 
DDT, DDE, and DDD 389 

9. REFERENCES 

*Verma A, Pillai MKK. 1991b. Bioavailability of soil-bound residues of DDT and HCH to earthworms. 

Curr Sci 61(12):840-843. 

*Verschueren K, ed. 1983. Handbook of environmental data on organic chemicals. 2nd ed. New York, 

NY: Van Nostrand Reinhold Company, 433-445. 

*Vetter W, Luckas B, Heidemann G, et al. 1996. Organochlorine residue in marine mammals from the 

Northern hemisphere-A consideration of the composition of organochlorine residues in the blubber of 

marine mammals. Sci Total Environ 186:29-39. 

Vieira EDR, Torres PM, Malm O. 2001. DDT environmental persistence from its use in a vector 

program: A case study. Environ Res A 86:174-182. 

*Vieira I, Sonnier M, Cresteil T. 1996. Developmental expression of CYP2E1 in the human liver: 

Hypermethylation control of gene expression during the neonatal period. Eur J Biochem 238:476-483. 

*Vine MF, Stein L, Weigle K, et al. 2000. Effects on the immune system associated with living near a 

pesticide dump site. Environ Health Perspect 108(12):1113-1124. 

*Vine MF, Stein L, Weigle K, et al. 2001. Plasma 1,1-dichloro-2n2-bis(p-chlorophenyl)ethylene, (DDE) 

levels and immune response. Am J Epidemiol 153(1):53-63. 

*Vollner L, Klotz D. 1994. Behaviour of DDT under laboratory and outdoor conditions in Germany. J 

Environ Sci Health B29(1):161-167. 

*vom Saal FS, Nagel SC, Palanza P, et al. 1995. Estrogenic pesticides: Binding relative to estradiol in 

MCF-7 cells and effects of exposure during fetal life on subsequent territorial behaviour in male mice. 

Toxicol Lett 77:343-350. 

*Waliszewski S, Szymczynski G. 1983. Determination of selected chlorinated pesticides, bound and 

free, in human semen. Arch Environ Contam Toxicol 12:577-580. 

Waliszewski SM, Aguirre AA, Infanzon RM. 1999. Levels of organochlorine pesticides in blood serum 

and umbilical blood serum of mothers living in Veracruz, Mexico. Fresenius Environ Bull 8:171-178. 

*Waliszewski SM, Aguirre AA, Infanzon RM, et al. 2000. Partitioning coefficients of organochlorine 

pesticides between mother blood serum and umbilical blood serum. Bull Environ Contam Toxicol 

65:293-299. 

*Waliszewski SM, Aguirre AA, Infanzon RM, et al. 2001. Organochlorine pesticide levels in maternal 

adipose tissue, maternal blood serum, umbilical blood serum, and milk from inhabitants of Veracruz, 

Mexico. Arch Environ Contam Toxicol 40:432-438. 

*Walker A, Thorpe E, Stevenson D. 1972. The toxicology of dieldrin (HEOD). I. Long-term oral 

toxicity studies in mice. Food Cosmet Toxicol 11:415-432. 

Waller DP, Presperin C, Drum ML, et al. 1996. Great Lakes fish as a source of maternal and fetal 

exposure to chlorinated hydrocarbons. Toxicol Ind Health 12(3/4):335-345. 

*Wang J, Simpson K. 1996. Accumulation and depuration of DDTs in the food chain from Artemia to 

brook trout (Salvelinus fontinalis). Bull Environ Contam Toxicol 56:888-895. 
DDT, DDE, and DDD 390 

9. REFERENCES 

Wang T, Johnson R, Bricker J. 1979. Residues of polychlorinated biphenyls and DDT in water and 

sediment of the St. Lucie Estuary, Florida, 1977. Pestic Monit J 13(2):69-71. 

Wang T, Johnson R, Bricker J. 1980. Residues of polychlorinated biphenyls and DDT in water and 

sediment of the Indian River Lagoon, Florida, 1977-78. Pestic Monit J 13:141-144. 

*Wania F, MacKay D. 1993. Global fractionation and cold condensation of low volatility 

organochlorine compounds in polar regions. Ambio 22:10-18 

*Ward EM, Schulte P, Grajewski B, et al. 2000. Serum organochlorine levels and breast cancer: A 

nested case-control study of Norwegian women. Cancer Epidemiol Biomarkers Prev 9:1357-1367. 

Ware G. 1980. Effects of pesticides on nontarget organisms. Residue Reviews 76:173-201. 

*Ware G, Good E. 1967. Effects of insecticides on reproduction in the laboratory mouse: II. Mirex, 

telodrin, and DDT. Toxicol Appl Pharmacol 10:54-61. 

*Ware GW, Crosby DG, Giles JW. 1980. Photodecomposition of DDA. Arch Environ Contam Toxicol 

9:135-146. 

*Ware G, Estesen B, Buck N, et al. 1978. DDT moratorium in Arizona- agricultural residues after seven 

years. Pestic Monit J 12(1):1-3. 

Warnick S. 1972. Organochlorine pesticide levels in human serum and adipose tissue, Utah- fiscal years 

1967-71. Pestic Monit J 6(1):9-13. 

*Wassermann M, Noguiera DP, Tomatis L, et al. 1976. Organochlorine compounds in neoplastic and 

adjacent apparently normal breast tissue. Bull Environ Contam Toxicol 15(4):478-484. 

*Wassermann M, Rapolt M, Bercovici, et al. 1982. Premature delivery and organochlorine compounds: 

Polychlorinated biphenyls and some organochlorine insecticides. Environ Res 28:106-112. 

*Wassermann M, Wassermann D, Gershon Z, et al. 1969. Effects of organochlorine insecticides on body 

defense systems. Ann NY Acad Sci 160:393-401. 

Wassermann M, Wassermann D, Kedar E, et al. 1971. Immunological and detoxication interaction in 

p,p-DDT fed rabbits. Bull Environ Contam Toxicol 6(5):426-435. 

WDNR. 1987. Surface water quality criteria for toxic substances. Department of Natural Resources, 

Wisconsin. 1987. Order of the State of Wisconsin Natural Resources Board Repealing, Renumbering, 

Renumbering and Amending, Amending, Repealing and Recreating, and Creating Rules. Section 27, 

Chapter NR 105. 

*Webber EC, Bayne DR, Seesock WC. 1989. DDT contamination of benthic macroinvertebrates and 

sediments from tributaries of Wheeler Reservoir, Alabama. Arch Environ Contam Toxicol 18:728-733. 

*Weistrand C, Norén K. 1997. Methylsulfonyl metabolites of PCBs and DDE in human tissues. 

Environ Health Persp 105:644-649 

*Weistrand C, Norén K. 1998. Polychlorinated naphthalenes and other organochlorine contaminants in 

human adipose and liver tissue. J Toxicol Environ Health Part A 53:293-311. 
DDT, DDE, and DDD 391 

9. REFERENCES 

*Welch R, Levin W, Conney A. 1969. Estrogenic action of DDT and its analogs. Toxicol Appl 

Pharmacol 14:358-367. 

*Wessel JR, Yess NJ. 1991. Pesticide residues in foods imported into the United States. Rev Environ 

Contam Toxicology 120:83-104. 

*West JR, Smith HW, Chasis H. 1948. Glomerular filtration rate, effective renal blood flow, and 

maximal tubular excretory capacity in infancy. J Pediatr 32:10-18. 

*Wester RC, Maibach HI, Bucks DA, et al. 1990. Percutaneous absorption of [14C]DDT and 

[

14C]benzo[a]pyrene from soil. Fundam Appl Toxicol 15:510-516. 

*Westin JB, Richter E. 1990. The Israeli breast-cancer anomaly. Ann N Y Acad Sci :269-279. 

*Whitmore RW, Immerman FW, Camann DE, et al. 1994. Non-occupational exposures to pesticides for 

residents of two U.S. cities. Arch Environ Contam Toxicol 26:47-59. 

*WHO. 1979. DDT and its derivatives. Environmental Health Criteria 9. Geneva, Switzerland: World 

Health Organization. 

*WHO. 1989. International programme on chemical safety: Environmental health criteria 83: DDT and 

its derivatives - environmental aspects. Lyon, France: World Health Organization. 

*WHO. 2002. Guidelines for drinking water. DDT. World Health Organization. 

http://www.who.int/water_sanitation_health/GDWQ/Chemicals/ddsum..htm. January 02, 2002. 

*Wickstrom K, Pyysalo H, Siimes M. 1983. Levels of chlordane, hexachlorobenzene, PCB and DDT 

compounds in Finnish human milk in 1982. Bull Environ Contam Toxicol 31:251-256. 

*Widdowson EM, Dickerson JWT. 1964. Chapter 17: Chemical composition of the body. In: Comar 

CL, Bronner F, eds. Mineral metabolism: An advanced treatise. Volume II: The elements Part A. New 

York, NY: Academic Press, 2-247 

Wilcox A. 1967. USPH investigation, DDT health effects. Inter-Office Correspondence to M.V. 

Anthony, Stauffer Chemical Co. (As cited in WHO, 1979). 

*Williams G, Weisburger J. 1991. Chemical carcinogens. In: Amdur M, Doull J, Klaassen C, eds. 

Casarett and Doull's toxicology: The basic science of poisons. 4th edition. New York, NY: Pergamon 

Press, 127-200. 

*Williams S. 1984. Official methods of analysis. Association of Official Analytical Chemists, Inc., 

Arlington, VA. 

Willis G, Parr J, Smith S. 1971. Volatilization of soil-applied DDT and DDD from flooded and 

nonflooded plots. Pestic Monit J 4(4):240-248. 

Wilson D, Locker D, Ritzen C, et al. 1973. DDT concentrations in human milk. Am J Dis Child 

125:814-817. 

*Wilson HR, Thompson NP, Roland DA. 1973. The effect of oral doses of DDT on various 

physiological parameters in the bobwhite laying hen. Poult Sci 52(5):2103. 
DDT, DDE, and DDD 392 

9. REFERENCES 

*Wilson N, Shear N, Paustenbach D, et al. 1998. The effect of cooking practices on the concentration of
 

DDT and PCB compounds in the edible tissue of fish. J Expo Anal Environ Epidemiol 8(3):423-440.
 

Witorsch RJ. 2000. Endocrine disruption: A critical review of environmental estrogens from a
 

mechanistic perspective. Toxic Subst Mech 19:53-78.
 

*Wolfe H, Armstrong J. 1971. Exposure of formulating plant workers to DDT. Arch Environ Health
 

23:169-176.
 

*Wolfe J, Esher R, Robinson K, et al. 1979. Lethal and reproductive effects of dietary mirex and DDT
 

on old-field mice, Peromyscus polionotus. Bull Environ Contam Toxicol 21:397-402.
 

*Wolfe MF, Seiber JN. 1993. Environmental activation of pesticides. Occup Med 8:561-573 

Wolff MS. 1999. Letter to the editor. Arch Environ Contam Toxicol 36:504. 

Wolff MS, Anderson HA. 1999. Correspondence re: J.M. Schildkraut et al., environmental 

contaminants and body fat distribution. Cancer Epidemiol Biomark. Prev., 8:179-183, 1999. Cancer 

Epidemiol Biomarkers Prev 8:951-952. 

Wolff MS, Landrigan PJ. 2002. Organochlorine chemicals and children's health. J Pediatr 140:10-13.
 

Wolff M, Anderson H, Selikoff I. 1982. Human tissue burdens of halogenated aromatic chemicals in
 

Michigan. JAMA 247(15):2112-2116.
 

Wolff M, Aubrey B, Camper F, et al. 1978a. Relation of DDE and PBB serum levels in farm residents,
 

consumers, and Michigan Chemical Corporation employees. Environ Health Perspect 23:177-181.
 

*Wolff MS, Berkowitz GS, Brower S, et al. 2000a. Organochlorine exposures and breast cancer risk in
 

New York City women. Environ Res A 84:151-161.
 

Wolff M, Haymes N, Anderson H, et al. 1978b. Family clustering of PBB and DDE values among
 

Michigan dairy farmers. Environ Health Perspect 23:315-319.
 

*Wolff MS, Toniolo P, Lee EW, et al. 1993. Blood levels of organochlorine residues and risk of breast
 

cancer. J Natl Cancer Inst 85(8):648-652.
 

*Wolff MS, Zeleniuch-Jacquotte A, Dubin N, et al. 2000b. Risk of breast cancer and organochlorine
 

exposure. Cancer Epidemiol Biomarkers Prev 9:271-277.
 

*Wong O, Brocker W, Davis H, et al. 1984. Mortality of workers potentially exposed to organic and
 

inorganic brominated chemicals, DBCP, TRIS, PBB, and DDT. Br J Ind Med 41:15-24.
 

Woodruff T, Wolff MS, Davis DL, et al. 1994. Organochlorine exposure estimation in the study of
 

cancer etiology. Environ Res 65:132-144.
 

*Woodward AR, Percival HF, Jennings ML, et al. 1993. Low clutch viability of American alligators on
 

Lake Apopka. Fla Sci 56(1):52-62.
 

*Woodwell G, Craig, P, Johnson H. 1971. DDT in the biosphere: Where does it go? Science 

174:1101-1107. 
DDT, DDE, and DDD 393 

9. REFERENCES 

*Woolcott CG, Aronson KJ, Hanna WM, et al. 2001. Organochlorines and breast cancer risk by receptor 

status, tumor size, and grade (Canada). Cancer Causes Control 12:395-404. 

Woolley D. 1970. Effects of DDT and of drug-DDT interactions on electroshock seizures in the rat. 

Toxicol Appl Pharmacol 16:521-532. 

Woolley D. 1973. Studies on 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT)-induced 

hyperthermia: Effects of cold exposure and aminopyrine injections. J Pharmacol Exp Ther 184(1):261­

268. 

*Woolley D, Talens G. 1971. Distribution of DDT, DDD, and DDE in tissues of neonatal rats and in 

milk and other tissues of mother rats chronically exposed to DDT. Toxicol Appl Pharmacol 18:907-916. 

*Wooten M, King D. 1993. Adrenal cortical carcinoma: Epidemiology and treatment with mitotane and 

a review of the literature. Cancer 72(11):3145-3155. 

*Wrenn T, Weyant J, Fries G, et al. 1971. Effect of several dietary levels of o,p'-DDT on reproduction 

and lactation in the rat. Bull Environ Contam Toxicol 6(5):471-480. 

*Wright S, Zhong J, Larrick J. 1994. Inhibition of apoptosis as a mechanism of tumor promotion. 

FASEB J 8:654-660. 

Xu-Qing W, Peng-Yuan G, Yuan-Zhen L, et al. 1988. Studies on hexachlorocyclohexane and DDT 

contents in human cerumen and their relationships to cancer mortality. Biomed Environ Sci 1:138-151. 

*Yagüe C, Bayarri S, Lázaro R, et al. 2001. Multiresidue determination of organochlorine pesticides and 

polychlorinated biphenyls in milk by gas chromatography with electron-capture detection after extraction 

by matrix solid-phase dispersion. J AOAC Int 84:1561-1568. 

Yau E, Mennear J. 1977. The inhibitory effect of DDT on insulin secretion in mice. Toxicol Appl 

Pharmacol 39:81-88. 

*Yess N, Gunderson E, Roy R. 1993. U.S. Food and Drug Administration monitoring of pesticide 

residues in infant foods and adult foods eaten by infants/children. J Assoc Off Anal Chem 76(3):492-507. 

*Yoder J, Watson M, Benson W. 1973. Lymphocyte chromosome analysis of agricultural workers 

during extensive occupational exposure to pesticides. Mutat Res 21:335-340. 

Yoon HR, Cho SY, Kim JM, et al. 1999. Analysis of multi-component pesticide residues in herbal 

medicines by GC-MS with electron impact ionization and with positive- and negative-ion chemical 

ionization. Chromatographia 49(9/10):525-534. 

You G, Sayles GD, Kupferle MJ, et al. 1996. Anaerobic DDT biotransformation: Enhancement by 

application of surfactants and low oxidation reduction potential. Chemosphere 32(11):2269-2284. 

*You L, Brenneman KA, Heck Hd'A. 1999a. In utero exposure to antiandrogens alters the 

responsiveness of the prostate to p,p'-DDE in adult rats and may induce prostatic inflammation. Toxicol 

Appl Pharmacol 161:258-266. 
DDT, DDE, and DDD 394 

9. REFERENCES 

*You L, Casanova M, Archibeque-Engle S, et al. 1998. Impaired male sexual development in perinatal 

Sprague-Dawley and Long-Evans hooded rats exposed in utero and lactationally to p,p'-DDE. Toxicol 

Sci 45:162-173. 

You L, Chan SK, Bruce JM, et al. 1999b. Modulation of testosterone-metabolizing hepatic cytochrome 

P-450 enzymes in developing Sprague-Dawley rats following in utero exposure t p,p'-DDE. Toxicol 

Appl Pharmacol 158:197-205. 

*You L, Gazi E, Archibeque-Engle S, et al. 1999c. Transplacental and lactational transfer of p,p'-DDE 

in Sprague-Dawley rats. Toxicol Appl Pharmacol 157:134-144. 

*You L, Sar M, Bartolucci E, et al. 2001. Induction of hepatic aromatase by p,p'-DDE in adult male rats. 

Mol Cell Endocrinol 178:207-214. 

Zabik ME, Zabik MJ, Booren AM, et al. 1995. Pesticides and total polychlorinated biphenyls in 

Chinook salmon and carp harvested from the Great Lakes: Effects of skin-on and skin-off processing and 

selected cooking methods. J Agric Food Chem 43:993-1001. 

*Zacharewski T. 1998. Identification and assessment of endocrine disrupters: Limitations of in vivo and 

in vitro assays. Environ Health Perspect Suppl 106(2):577-582. 

Zaidi S, Banerjee B. 1987. Enzymatic detoxication of DDT to DDD by rat liver: Effects of some 

inducers and inhibitors of cytochrome P-450 enzyme system. Bull Environ Contam Toxicol 38:449-455. 

Zaidi SA, Anthwal S, Sharma N, et al. 1994a. Interaction of DDT and malathion on the level of 

5-hydroxytryptamine [5-HT] in rat brain. Orient J Chem 10(1):75-76. 

Zaidi SA, Sharma N, Singh VS. 1994b. Effect of DDT on gamma aminobutyric acid (GABA) in rat 

brain. Orient J Chem 10(1):71-72. 

*Zava D, Blen M, Duwe G. 1997. Estrogenic activity of natural and synthetic estrogens in human breast 

cells in culture. Environ Health Perspect 105(Suppl 3):637-645. 

*Zeng E, Yu C, Tran K. 1999. In situ measurements of chlorinated hydrocarbons in the water column 

off the Palos Verdes Peninsula, California. Environ Sci Technol 33:392-398. 

*Zepp RG, Wolfe NL, Azarraga LV, et al. 1977. Photochemical transformation of DDT and 

methoxychlor degradation products, DDE and DMDE, by sunlight. Arch Environ Contam Toxicol 

6:305-314. 

*Zheng T, Holford TR, Mayne ST, et al. 1999. DDE and DDT in breast adipose tissue and risk of 

female breast cancer. Am J Epidemiol 150(5):453-458. 

*Zheng T, Holford TR, Mayne ST, et al. 2000. Risk of female breast cancer associated with serum 

polychlorinated biphenyls and 1,1-dichloro-2,2'-bis(p-chlorophenyl)ethylene. Cancer Epidemiol 

Biomarkers Prev 9:167-174. 

*Ziegler EE, Edwards BB, Jensen RL et al. 1978. Absorption and retention of lead by infants. Pediatr 

Res 12:29-34. 
DDT, DDE, and DDD 395 

9. REFERENCES 

Zimmerman LR, Thurman EM, Bastian KC. 2000. Detection of persistent organic pollutants in the 

Mississippi Delta using semipermeable membrane devices. Sci Total Environ 248:169-179. 

*Zook C, Feng J. 1999. 1,1,1-trichloro-2,2-bis-(4'-chlorophenyl)ethane (DDT) pathway map. 

http://www.labmed.umn.edu/umbbd/ddt/ddt_map.html. June 23, 1999. 

DDT, DDE, and DDD 397 

10. GLOSSARY 

Absorption—The taking up of liquids by solids, or of gases by solids or liquids. 

Acute Exposure—Exposure to a chemical for a duration of 14 days or less, as specified in the 

Toxicological Profiles. 

Adsorption—The adhesion in an extremely thin layer of molecules (as of gases, solutes, or liquids) to the 

surfaces of solid bodies or liquids with which they are in contact. 

Adsorption Coefficient (Koc)—The ratio of the amount of a chemical adsorbed per unit weight of 

organic carbon in the soil or sediment to the concentration of the chemical in solution at equilibrium. 

Adsorption Ratio (Kd)—The amount of a chemical adsorbed by a sediment or soil (i.e., the solid phase) 

divided by the amount of chemical in the solution phase, which is in equilibrium with the solid phase, at a 

fixed solid/solution ratio. It is generally expressed in micrograms of chemical sorbed per gram of soil or 

sediment. 

Benchmark Dose (BMD)—Usually defined as the lower confidence limit on the dose that produces a 

specified magnitude of changes in a specified adverse response. For example, a BMD10 would be the 

dose at the 95% lower confidence limit on a 10% response, and the benchmark response (BMR) would be 

10%. The BMD is determined by modeling the dose response curve in the region of the dose response 

relationship where biologically observable data are feasible. 

Benchmark Dose Model—A statistical dose-response model applied to either experimental toxicological 

or epidemiological data to calculate a BMD. 

Bioconcentration Factor (BCF)—The quotient of the concentration of a chemical in aquatic organisms 

at a specific time or during a discrete time period of exposure divided by the concentration in the 

surrounding water at the same time or during the same period. 

Biomarkers—Broadly defined as indicators signaling events in biologic systems or samples. They have 

been classified as markers of exposure, markers of effect, and markers of susceptibility. 

Cancer Effect Level (CEL)—The lowest dose of chemical in a study, or group of studies, that produces 

significant increases in the incidence of cancer (or tumors) between the exposed population and its 

appropriate control. 

Carcinogen—A chemical capable of inducing cancer. 

Case-Control Study—A type of epidemiological study which examines the relationship between a 

particular outcome (disease or condition) and a variety of potential causative agents (such as toxic 

chemicals). In a case-controlled study, a group of people with a specified and well-defined outcome is 

identified and compared to a similar group of people without outcome. 

Case Report—Describes a single individual with a particular disease or exposure. These may suggest 

some potential topics for scientific research but are not actual research studies. 
DDT, DDE, and DDD 398 

10. GLOSSARY 

Case Series—Describes the experience of a small number of individuals with the same disease or 

exposure. These may suggest potential topics for scientific research but are not actual research studies. 

Ceiling Value—A concentration of a substance that should not be exceeded, even instantaneously. 

Chronic Exposure—Exposure to a chemical for 365 days or more, as specified in the Toxicological 

Profiles. 

Cohort Study—A type of epidemiological study of a specific group or groups of people who have had a 

common insult (e.g., exposure to an agent suspected of causing disease or a common disease) and are 

followed forward from exposure to outcome. At least one exposed group is compared to one unexposed 

group. 

Cross-sectional Study—A type of epidemiological study of a group or groups which examines the 

relationship between exposure and outcome to a chemical or to chemicals at one point in time. 

Data Needs—Substance-specific informational needs that if met would reduce the uncertainties of human 

health assessment. 

Developmental Toxicity—The occurrence of adverse effects on the developing organism that may result 

from exposure to a chemical prior to conception (either parent), during prenatal development, or 

postnatally to the time of sexual maturation. Adverse developmental effects may be detected at any point 

in the life span of the organism. 

Dose-Response Relationship—The quantitative relationship between the amount of exposure to a 

toxicant and the incidence of the adverse effects. 

Embryotoxicity and Fetotoxicity—Any toxic effect on the conceptus as a result of prenatal exposure to 

a chemical; the distinguishing feature between the two terms is the stage of development during which the 

insult occurs. The terms, as used here, include malformations and variations, altered growth, and in utero 

death. 

Environmental Protection Agency (EPA) Health Advisory—An estimate of acceptable drinking water 

levels for a chemical substance based on health effects information. A health advisory is not a legally 

enforceable federal standard, but serves as technical guidance to assist federal, state, and local officials. 

Epidemiology—Refers to the investigation of factors that determine the frequency and distribution of 

disease or other health-related conditions within a defined human population during a specified period. 

Genotoxicity—A specific adverse effect on the genome of living cells that, upon the duplication of 

affected cells, can be expressed as a mutagenic, clastogenic or carcinogenic event because of specific 

alteration of the molecular structure of the genome. 

Half-life—A measure of rate for the time required to eliminate one half of a quantity of a chemical from 

the body or environmental media. 

Immediately Dangerous to Life or Health (IDLH)—The maximum environmental concentration of a 

contaminant from which one could escape within 30 minutes without any escape-impairing symptoms or 

irreversible health effects. 
DDT, DDE, and DDD 399 

10. GLOSSARY 

Incidence—The ratio of individuals in a population who develop a specified condition to the total 

number of individuals in that population who could have developed that condition in a specified time 

period. 

Intermediate Exposure—Exposure to a chemical for a duration of 15–364 days, as specified in the 

Toxicological Profiles. 

Immunologic Toxicity—The occurrence of adverse effects on the immune system that may result from 

exposure to environmental agents such as chemicals. 

Immunological Effects—Functional changes in the immune response. 

In Vitro—Isolated from the living organism and artificially maintained, as in a test tube. 

In Vivo—Occurring within the living organism. 

Lethal Concentration(LO) (LCLO)—The lowest concentration of a chemical in air which has been 

reported to have caused death in humans or animals. 

Lethal Concentration(50) (LC50)—A calculated concentration of a chemical in air to which exposure for a 

specific length of time is expected to cause death in 50% of a defined experimental animal population. 

Lethal Dose(LO) (LDLO)—The lowest dose of a chemical introduced by a route other than inhalation that 

has been reported to have caused death in humans or animals. 

Lethal Dose(50) (LD50)—The dose of a chemical which has been calculated to cause death in 50% of a 

defined experimental animal population. 

Lethal Time(50) (LT50)—A calculated period of time within which a specific concentration of a chemical 

is expected to cause death in 50% of a defined experimental animal population. 

Lowest-Observed-Adverse-Effect Level (LOAEL)—The lowest exposure level of chemical in a study, 

or group of studies, that produces statistically or biologically significant increases in frequency or severity 

of adverse effects between the exposed population and its appropriate control. 

Lymphoreticular Effects—Represent morphological effects involving lymphatic tissues such as the 

lymph nodes, spleen, and thymus. 

Malformations—Permanent structural changes that may adversely affect survival, development, or 

function. 

Minimal Risk Level (MRL)—An estimate of daily human exposure to a hazardous substance that is 

likely to be without an appreciable risk of adverse noncancer health effects over a specified route and 

duration of exposure. 

Modifying Factor (MF)—A value (greater than zero) that is applied to the derivation of a minimal risk 

level (MRL) to reflect additional concerns about the database that are not covered by the uncertainty 

factors. The default value for a MF is 1. 
DDT, DDE, and DDD 400 

10. GLOSSARY 

Morbidity—State of being diseased; morbidity rate is the incidence or prevalence of disease in a specific 

population. 

Mortality—Death; mortality rate is a measure of the number of deaths in a population during a specified 

interval of time. 

Mutagen—A substance that causes mutations. A mutation is a change in the DNA sequence of a cell’s 

DNA. Mutations can lead to birth defects, miscarriages, or cancer. 

Necropsy—The gross examination of the organs and tissues of a dead body to determine the cause of 

death or pathological conditions. 

Neurotoxicity—The occurrence of adverse effects on the nervous system following exposure to a 

chemical. 

No-Observed-Adverse-Effect Level (NOAEL)—The dose of a chemical at which there were no 

statistically or biologically significant increases in frequency or severity of adverse effects seen between 

the exposed population and its appropriate control. Effects may be produced at this dose, but they are not 

considered to be adverse. 

Octanol-Water Partition Coefficient (Kow)—The equilibrium ratio of the concentrations of a chemical 

in n-octanol and water, in dilute solution. 

Odds Ratio (OR)—A means of measuring the association between an exposure (such as toxic substances 

and a disease or condition) which represents the best estimate of relative risk (risk as a ratio of the 

incidence among subjects exposed to a particular risk factor divided by the incidence among subjects who 

were not exposed to the risk factor). An odds ratio of greater than 1 is considered to indicate greater risk 

of disease in the exposed group compared to the unexposed. 

Organophosphate or Organophosphorus Compound—A phosphorus containing organic compound 

and especially a pesticide that acts by inhibiting cholinesterase. 

Permissible Exposure Limit (PEL)—An Occupational Safety and Health Administration (OSHA) 

allowable exposure level in workplace air averaged over an 8-hour shift of a 40-hour workweek. 

Pesticide—General classification of chemicals specifically developed and produced for use in the control 

of agricultural and public health pests. 

Pharmacokinetics—The science of quantitatively predicting the fate (disposition) of an exogenous 

substance in an organism. Utilizing computational techniques, it provides the means of studying the 

absorption, distribution, metabolism and excretion of chemicals by the body. 

Pharmacokinetic Model—A set of equations that can be used to describe the time course of a parent 

chemical or metabolite in an animal system. There are two types of pharmacokinetic models: data-based 

and physiologically-based. A data-based model divides the animal system into a series of compartments 

which, in general, do not represent real, identifiable anatomic regions of the body whereby the 

physiologically-based model compartments represent real anatomic regions of the body. 

Physiologically Based Pharmacodynamic (PBPD) Model—A type of physiologically-based doseresponse model which quantitatively describes the relationship between target tissue dose and toxic end 
DDT, DDE, and DDD 401 

10. GLOSSARY 

points. These models advance the importance of physiologically based models in that they clearly 

describe the biological effect (response) produced by the system following exposure to an exogenous 

substance. 

Physiologically Based Pharmacokinetic (PBPK) Model—Comprised of a series of compartments 

representing organs or tissue groups with realistic weights and blood flows. These models require a 

variety of physiological information: tissue volumes, blood flow rates to tissues, cardiac output, alveolar 

ventilation rates and, possibly membrane permeabilities. The models also utilize biochemical information 

such as air/blood partition coefficients, and metabolic parameters. PBPK models are also called 

biologically based tissue dosimetry models. 

Prevalence—The number of cases of a disease or condition in a population at one point in time. 

Prospective Study—A type of cohort study in which the pertinent observations are made on events 

occurring after the start of the study. A group is followed over time. 

q1*—The upper-bound estimate of the low-dose slope of the dose-response curve as determined by the 

multistage procedure. The q1* can be used to calculate an estimate of carcinogenic potency, the 

incremental excess cancer risk per unit of exposure (usually μg/L for water, mg/kg/day for food, and 

μg/m3

 for air). 

Recommended Exposure Limit (REL)—A National Institute for Occupational Safety and Health 

(NIOSH) time-weighted average (TWA) concentrations for up to a 10-hour workday during a 40-hour 

workweek. 

Reference Concentration (RfC)—An estimate (with uncertainty spanning perhaps an order of 

magnitude) of a continuous inhalation exposure to the human population (including sensitive subgroups) 

that is likely to be without an appreciable risk of deleterious noncancer health effects during a lifetime. 

The inhalation reference concentration is for continuous inhalation exposures and is appropriately 

expressed in units of mg/m3 or ppm. 

Reference Dose (RfD)—An estimate (with uncertainty spanning perhaps an order of magnitude) of the 

daily exposure of the human population to a potential hazard that is likely to be without risk of deleterious 

effects during a lifetime. The RfD is operationally derived from the no-observed-adverse-effect level 

(NOAEL-from animal and human studies) by a consistent application of uncertainty factors that reflect 

various types of data used to estimate RfDs and an additional modifying factor, which is based on a 

professional judgment of the entire database on the chemical. The RfDs are not applicable to 

nonthreshold effects such as cancer. 

Reportable Quantity (RQ)—The quantity of a hazardous substance that is considered reportable under 

the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Reportable 

quantities are (1) 1 pound or greater or (2) for selected substances, an amount established by regulation 

either under CERCLA or under Section 311 of the Clean Water Act. Quantities are measured over a 

24-hour period. 

Reproductive Toxicity—The occurrence of adverse effects on the reproductive system that may result 

from exposure to a chemical. The toxicity may be directed to the reproductive organs and/or the related 

endocrine system. The manifestation of such toxicity may be noted as alterations in sexual behavior, 

fertility, pregnancy outcomes, or modifications in other functions that are dependent on the integrity of 

this system. 
DDT, DDE, and DDD 402 

10. GLOSSARY 

Retrospective Study—A type of cohort study based on a group of persons known to have been exposed 

at some time in the past. Data are collected from routinely recorded events, up to the time the study is 

undertaken. Retrospective studies are limited to causal factors that can be ascertained from existing 

records and/or examining survivors of the cohort. 

Risk—The possibility or chance that some adverse effect will result from a given exposure to a chemical. 

Risk Factor—An aspect of personal behavior or lifestyle, an environmental exposure, or an inborn or 

inherited characteristic, that is associated with an increased occurrence of disease or other health-related 

event or condition. 

Risk Ratio—The ratio of the risk among persons with specific risk factors compared to the risk among 

persons without risk factors. A risk ratio greater than 1 indicates greater risk of disease in the exposed 

group compared to the unexposed. 

Short-Term Exposure Limit (STEL)—The American Conference of Governmental Industrial 

Hygienists (ACGIH) maximum concentration to which workers can be exposed for up to 15 min 

continually. No more than four excursions are allowed per day, and there must be at least 60 min 

between exposure periods. The daily Threshold Limit Value - Time Weighted Average (TLV-TWA) may 

not be exceeded. 

Standardized Mortality Ratio (SMR)—A ratio of the observed number of deaths and the expected 

number of deaths in a specific standard population. 

Target Organ Toxicity—This term covers a broad range of adverse effects on target organs or 

physiological systems (e.g., renal, cardiovascular) extending from those arising through a single limited 

exposure to those assumed over a lifetime of exposure to a chemical. 

Teratogen—A chemical that causes structural defects that affect the development of an organism. 

Threshold Limit Value (TLV)—An American Conference of Governmental Industrial Hygienists 

(ACGIH) concentration of a substance to which most workers can be exposed without adverse effect. 

The TLV may be expressed as a Time Weighted Average (TWA), as a Short-Term Exposure Limit 

(STEL), or as a ceiling limit (CL). 

Time-Weighted Average (TWA)—An allowable exposure concentration averaged over a normal 8-hour 

workday or 40-hour workweek. 

Toxic Dose(50) (TD50)—A calculated dose of a chemical, introduced by a route other than inhalation, 

which is expected to cause a specific toxic effect in 50% of a defined experimental animal population. 

Toxicokinetic—The study of the absorption, distribution and elimination of toxic compounds in the 

living organism. 

Uncertainty Factor (UF)—A factor used in operationally deriving the Minimal Risk Level (MRL) or 

Reference Dose (RfD) or Reference Concentration (RfC) from experimental data. UFs are intended to 

account for (1) the variation in sensitivity among the members of the human population, (2) the 

uncertainty in extrapolating animal data to the case of human, (3) the uncertainty in extrapolating from 

data obtained in a study that is of less than lifetime exposure, and (4) the uncertainty in using lowestobserved-adverse-effect level (LOAEL) data rather than no-observed-adverse-effect level (NOAEL) data. 
DDT, DDE, and DDD 403 

10. GLOSSARY 

A default for each individual UF is 10; if complete certainty in data exists, a value of one can be used; 

however a reduced UF of three may be used on a case-by-case basis, three being the approximate 

logarithmic average of 10 and 1. 

Xenobiotic—Any chemical that is foreign to the biological system. 

DDT, DDE, and DDD A-1 

APPENDIX A
 

ATSDR MINIMAL RISK LEVEL AND WORKSHEETS
 

The Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) [42 U.S.C. 

9601 et seq.], as amended by the Superfund Amendments and Reauthorization Act (SARA) [Pub. L. 

99–499], requires that the Agency for Toxic Substances and Disease Registry (ATSDR) develop jointly 

with the U.S. Environmental Protection Agency (EPA), in order of priority, a list of hazardous substances 

most commonly found at facilities on the CERCLA National Priorities List (NPL); prepare toxicological 

profiles for each substance included on the priority list of hazardous substances; and assure the initiation 

of a research program to fill identified data needs associated with the substances. 

The toxicological profiles include an examination, summary, and interpretation of available toxicological 

information and epidemiologic evaluations of a hazardous substance. During the development of 

toxicological profiles, Minimal Risk Levels (MRLs) are derived when reliable and sufficient data exist to 

identify the target organ(s) of effect or the most sensitive health effect(s) for a specific duration for a 

given route of exposure. An MRL is an estimate of the daily human exposure to a hazardous substance 

that is likely to be without appreciable risk of adverse noncancer health effects over a specified duration 

of exposure. MRLs are based on noncancer health effects only and are not based on a consideration of 

cancer effects. These substance-specific estimates, which are intended to serve as screening levels, are 

used by ATSDR health assessors to identify contaminants and potential health effects that may be of 

concern at hazardous waste sites. It is important to note that MRLs are not intended to define clean-up or 

action levels. 

MRLs are derived for hazardous substances using the no-observed-adverse-effect level/uncertainty factor 

approach. They are below levels that might cause adverse health effects in the people most sensitive to 

such chemical-induced effects. MRLs are derived for acute (1–14 days), intermediate (15–364 days), and 

chronic (365 days and longer) durations and for the oral and inhalation routes of exposure. Currently, 

MRLs for the dermal route of exposure are not derived because ATSDR has not yet identified a method 

suitable for this route of exposure. MRLs are generally based on the most sensitive chemical-induced end 

point considered to be of relevance to humans. Serious health effects (such as irreparable damage to the 

liver or kidneys, or birth defects) are not used as a basis for establishing MRLs. Exposure to a level 

above the MRL does not mean that adverse health effects will occur. 
DDT, DDE, and DDD A-2 

APPENDIX A 

MRLs are intended only to serve as a screening tool to help public health professionals decide where to 

look more closely. They may also be viewed as a mechanism to identify those hazardous waste sites that 

are not expected to cause adverse health effects. Most MRLs contain a degree of uncertainty because of 

the lack of precise toxicological information on the people who might be most sensitive (e.g., infants, 

elderly, nutritionally or immunologically compromised) to the effects of hazardous substances. ATSDR 

uses a conservative (i.e., protective) approach to address this uncertainty consistent with the public health 

principle of prevention. Although human data are preferred, MRLs often must be based on animal studies 

because relevant human studies are lacking. In the absence of evidence to the contrary, ATSDR assumes 

that humans are more sensitive to the effects of hazardous substance than animals and that certain persons 

may be particularly sensitive. Thus, the resulting MRL may be as much as a hundredfold below levels 

that have been shown to be nontoxic in laboratory animals. 

Proposed MRLs undergo a rigorous review process: Health Effects/MRL Workgroup reviews within the 

Division of Toxicology, expert panel peer reviews, and agencywide MRL Workgroup reviews, with 

participation from other federal agencies and comments from the public. They are subject to change as 

new information becomes available concomitant with updating the toxicological profiles. Thus, MRLs in 

the most recent toxicological profiles supersede previously published levels. For additional information 

regarding MRLs, please contact the Division of Toxicology, Agency for Toxic Substances and Disease 

Registry, 1600 Clifton Road, Mailstop E-29, Atlanta, Georgia 30333. 
DDT, DDE, and DDD A-3 

APPENDIX A 

MINIMAL RISK LEVEL WORKSHEET 

Chemical Name: DDT 

CAS Number: 50-29-3 

Date: September 2002 

Profile Status: Third Draft Post Public 

Route: [ ] Inhalation [X] Oral 

Duration: [X] Acute [ ] Intermediate [ ] Chronic 

Graph Key: 47m 

Species: Mice 

Minimal Risk Level: 0.0005 [X] mg/kg/day [ ] ppm 

Reference: Eriksson and Nordberg 1986; Eriksson et al. 1990a, 1990b, 1992, 1993; Johansson et al. 1995, 

1996; Talts et al. 1998. 

Experimental design and effects noted: The acute oral MRL is based on results from a group of studies 

conducted by the same group of investigators in which the most significant finding was the presence of 

altered motor behavior in adult mice treated with DDT perinatally. Groups of 10-day-old male NMRI 

mice were treated by gavage with a single dose of 0 (vehicle control) or 0.5 mg DDT/kg in a fat emulsion 

vehicle by gavage (Eriksson et al. 1990a). At the age of 4 months, the mice were subjected to behavioral 

tests of spontaneous activity (locomotion, rearing, and total activity). Tests were conducted for 1 hour, 

and scores were summed for three 20-minute periods. During the last 40 minutes of testing, the treated 

mice showed significantly more activity than untreated controls. This was interpreted as disruption of a 

simple, non-associative learning process, (i.e., habituation), or a retardation in adjustment to a new 

environment. These same results were reported in a later paper (Eriksson et al. 1990b) in which the 

authors also reported results of neurochemical evaluations conducted 2–3 weeks after behavioral testing. 

They measured muscarinic acetylcholine (MACh) receptor density and choline acetyltransferase (ChAT) 

activity in the cerebral cortex and hippocampus (MACh also in striatum), and also measured 

K+

-stimulated ACh release from cerebral cortex slices. In addition, five 10-day-old mice were 

administered 0.5 mg 14C-DDT and the radioactivity in the brain was assayed 24 hours, 7 days, or 1 month 

after dosing. The results showed that K+

-evoked ACh release in treated mice was significantly increased 

relative to controls, ChAT activity was not changed in the cerebral cortex or hippocampus, and the 

density of MACh was not significantly changed in the hippocampus or striatum, but a decreasing trend 

was seen in the cerebral cortex. DDT-derived radioactivity could be detected until day 7 after dosing, but 

none could be detected 1 month after dosing. 

Previous studies have shown a significant increase in density of MACh in the cerebral cortex of 10-dayold mice 7 days after dosing, but not at 1 day post-exposure compared to controls (Eriksson and Nordberg 

1986). No increased binding was noted in the hippocampus either 1 or 7 days post-treatment. This was 

further investigated by evaluating the proportion of high- and low-affinity binding sites and the affinity 

constants of the muscarinic receptors. A significant increase in the percentage of low-affinity binding 

sites accompanied by a significant decrease in high-affinity binding sites was measured in the cerebral 

cortex 7 days post-exposure. No significant changes in affinity constants were noted. According to the 

authors, these low-affinity binding sites correspond to the M1 receptor in the cerebral cortex, which are 

thought to be associated with neuronal excitation. No changes were observed in the sodium-dependent 

choline uptake system in the cerebral cortex 7 days post-exposure. 

In a follow-up study, Eriksson et al. (1992) treated 3-, 10-, and 19-day-old mice, and conducted 

behavioral testing and neurochemical evaluations at 4 months of age. As previously published, mice 
DDT, DDE, and DDD A-4 

APPENDIX A 

treated at 10 days exhibited hyperactivity relative to controls and also a significant decrease in the density 

of MACh in the cerebral cortex. No such changes were seen in mice treated at 3 or 19 days old. The 

authors suggested that the changes in MACh density and behavior might be the consequence of early 

interference with muscarinic cholinergic transmission specifically around the age of 10 days. In 

subsequent studies by the same group, 5- and 7-month-old mice were tested (Eriksson et al. 1993; 

Johansson et al. 1995). At both time points, mice treated with DDT perinatally showed increased 

spontaneous motor activity relative to controls, and decreased density of MACh in the cerebral cortex. 

No changes were seen regarding percentages of high- or low-affinity muscarinic binding sites in the 

cerebral cortex. Mice in these studies were also treated orally at the age of 5 months with the type I 

pyrethroid insecticide, bioallethrin, and tested for motor activity at this age (Eriksson et al. 1993) and at 

7 months (Johansson et al. 1995). In general, mice treated with DDT at the age of 10 days and later with 

bioallethrin showed increased motor behavior relative to those treated with bioallethrin alone, suggesting 

a DDT-induced increased susceptibility to bioallethrin. Mice treated first with DDT and later on with 

bioallethrin also showed increased difficulties in learning a skill, such as the swim maze test, compared 

with untreated mice, mice treated with DDT alone, or mice treated with bioallethrin alone (Johansson et 

al. 1995). Also, treatment with DDT followed by bioallethrin significantly increased the density of 

muscarinic receptors in the cerebral cortex relative to DDT alone. This increase was later attributed to 

increased expression of muscarinic receptor m4 mRNA (Talts et al. 1998). In yet another study from this 

group, paraoxon replaced bioallethrin, and the mice were tested at 5 and 7 months (Johansson et al. 1996). 

In addition, acetylcholinesterase activity was measured in cerebral cortex of 5-month-old mice and MACh 

and nicotinic cholinergic receptors in cortex of 7-month-old mice. Relevant new findings include that: 

(1) DDT did not significantly alter acetylcholinesterase activity; (2) DDT did not alter the effects of 

paraoxon on acetylcholinesterase activity (decreased); (3) DDT altered (increased or decreased) some of 

motor responses due to paraoxon alone at 7 months but not at 5 months; (4) none of the treatments altered 

performance in the swim maze test; and (5) none of the treatments altered the density of nicotinic 

cholinergic receptors in the cortex. 

In this series of studies, two responses seem to be consistent from study to study in mice treated with 

DDT perinatally and tested as adults, a decrease in the density of muscarinic cholinergic receptors in the 

cerebral cortex and increased spontaneous motor activity. It is not clear whether there is a causality 

relationship. DDT also altered some motor responses induced by other pesticides, but a pattern was not 

always clear. The investigators interpreted the latter findings as DDT inducing changes early in the brain 

that translated into increased susceptibility to other pesticides later in life. The DDT-induced increase in 

spontaneous motor activity at the dose of 0.5 mg/kg is considered a less serious LOAEL. 

Dose and end point used for MRL derivation: 0.5 mg/kg; neurodevelopmental effects. 

[ ] NOAEL [X] LOAEL 

Uncertainty Factors used in MRL derivation: 

[X] 10 for use of a LOAEL 

[X] 10 for extrapolation from animals to humans 

[X] 10 for human variability 

The DDT in this experiment was administered in a fat emulsion vehicle via gavage. There was no 

adjustment made for effects of the fat vehicle on absorption because neonatal children are likely to be 

exposed to DDT via high fat breast milk. 
 

DDT, DDE, and DDD A-5 

APPENDIX A 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? 

No 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: NA 

Was a conversion used from intermittent to continuous exposure? 

No 

Agency Contact (Chemical Manager): Obaid Faroon 
DDT, DDE, and DDD A-6 

APPENDIX A 

MINIMAL RISK LEVEL WORKSHEET 

Chemical Name: DDT 

CAS Number: 50-29-3 

Date: September 2002 

Profile Status: Third Draft Post Public 

Route: [ ] Inhalation [X] Oral 

Duration: [ ] Acute [X] Intermediate [ ] Chronic 

Graph Key: 60r 

Species: Rat 

Minimal Risk Level: 0.0005 [X] mg/kg/day [ ] ppm 

In this dietary study the amount of DDT added to the food was measured, but the actual food 

consumption and body weights of the rats were not measured. The calculated value for the NOAEL on 

which the MRL is based ranges from 0.05 to 0.09 mg/kg/day depending on the food consumption values 

used to calculate the actual dose of DDT consumed. More details are provided below in the discussion of 

conversion factors. 

In protecting public health ATSDR recommends using the conservative lower end of this calculated 

NOAEL range, 0.05 mg/kg/day, to derive an intermediate-duration MRL of 0.0005 mg/kg/day. This 

MRL value is consistent with and supported by the acute-duration MRL of 0.0005 mg/kg/day, 

particularly if intermediate exposure were to occur during the sensitive critical window of development 

(postnatal day 10 in the mouse model) identified in the acute duration MRL exposure studies (Eriksson 

and Nordberg 1986; Eriksson et al. 1990a, 1990b, 1992, 1993; Johansson et al. 1995, 1996). 

Reference: Fitzhugh and Nelson 1947; Laug et al. 1950 

Experimental design: Groups of male and female Osborne-Mendel rats (15/sex/group) were administered 

technical DDT(dissolved in corn oil) added to the diet at dosage levels of 0, 1, 5, 10, or 50 ppm for 15–27 

weeks. This study was essentially designed to examine whether DDT accumulates in adipose tissue and to 

what extent, how age and dose level affect accumulation, and how rapidly it is eliminated. Seventy-seven 

rats were used for microscopic evaluation of only the liver and kidney. This was based on findings from 

a previous study from the same group (Fitzhugh and Nelson 1947, see below) in which higher dietary 

levels of DDT had been used. Based on the previous findings, only the liver was expected to show 

microscopic changes. Although not explicitly stated, it is assumed that morphologic evaluations were 

conducted at the times when DDT levels in fat were determined (after 15, 19, 23, and 27 weeks of 

treatment). 

Effects noted in study and corresponding doses: These dose ranges were calculated as shown below in 

the discussion of conversion factors. There were no morphologic alterations in the kidneys. Liver 

alterations were noticed at the 5 ppm (0.25–0.5 mg/kg/day) dietary level of DDT and higher, but not at 

1 ppm (0.05–0.09 mg/kg/day). Liver changes consisted of hepatic cell enlargement, especially in central 

lobules, increased cytoplasmic oxyphilia with sometimes a semihyaline appearance, and more peripheral 

location of the basophilic cytoplasmic granules. Necrosis was not observed. The severity of the effects 

was dose-related, and males tended to show more hepatic cell changes than females. Changes seen at the 

5 ppm level (0.25–0.5 mg/kg/day) were considered by the authors as ”minimal”; changes seen at the 

50 ppm level (2.5–4.6 mg/kg/day) were slight, sometimes moderate; the authors do not comment about 

what they saw in the 10 ppm (0.5–0.9 mg/kg/day) group, presumably the results were intermediate to the 

doses above and below. The results from the kinetic studies revealed that accumulation of DDT in fat 

occurred at all dietary levels tested and that females stored more DDT than males; storage reached a 
DDT, DDE, and DDD A-7 

APPENDIX A 

maximum at 19–23 weeks; age did not affect the rate of DDT-accumulation; about 50–75% of DDT 

stored in fat remained after a 1-month DDT-free diet, and 25% remained after 3 months. 

Dose and end point used for MRL derivation: 0.05 mg/kg/day; liver effects. 

[X] NOAEL [ ] LOAEL 

Uncertainty Factors used in MRL derivation: 

[ ] 10 for use of a LOAEL 

[X] 10 for extrapolation from animals to humans 

[X] 10 for human variability 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? 

Yes 

In this dietary study, the amount of DDT added to the food was measured, but the actual food 

consumption and body weights of the rats were not measured. The calculated value for the NOAEL on 

which the MRL is based ranges from 0.05 to 0.09 mg/kg/day depending on the food consumption values 

used to calculate the actual dose of DDT consumed. Using the most conservative estimates of food 

consumption, the NOAEL for 1 ppm DDT in food = 0.05 mg DDT/kg/day; this was the value used for 

calculating the MRL. 

Calculation of Food Consumption Values (Three Models) 

(1) Using EPA 1986d Reference Values 

Using the EPA 1986d Reference Values for Risk Assessment, as was done in the 1994 edition of this 

toxicological profile, yields a NOAEL of 0.05 mg/kg/day equivalent to feeding 1 ppm DDT in the food. 

This reference considers the food consumption of an average rat to be 0.05 kg food/kg body weight/day 

(averaged over a lifetime) and the average weight of a rat to be 0.35 kg. No allometric equation is used 

for calculating the food factor. However, this reference recommends using a food consumption value of 

0.09 mg/kg/day for a subchronic 90-day study. 

NOAEL= 1 ppm in food = 1 mg/kg food 

1 mg DDT/kg food x 0.05 kg food/kg body weight/day=0.05 mg/kg/day 

Equivalents for other food concentrations used in this study, as calculated by using the EPA 1986d 

Reference Values for Risk Assessment: 

5 ppm= 0.25 mg/kg/day 

10 ppm= 0.5 mg/kg/day 

50 ppm= 2.5 mg/kg/day 

(2) EPA 1988g Method, Chronic Duration Average Body Weights 

The NOAEL is 1 ppm of DDT in dietary study feeding Osborne-Mendel rats for 15–27 weeks

(105–109 days). EPA 1988g has time weighted average body weights for male or female OsborneMendel rats for either of two study durations: subchronic (weaning to 90 days) and chronic (weaning to

730 days or 2 years). This study does not exactly fall into either of the intervals for which the time

weighted average weights were calculated, but the convention is to pick the chronic category for studies 
 

DDT, DDE, and DDD A-8 

APPENDIX A 

longer than 90 days. The risk assessment calculations recommended in EPA 1988g were used for

calculating doses for all the dietary studies discussed in this toxicological profile. 

F= 0.056(W)0.6611 where F = kg food/day and W = body weight in kilograms 

For chronic duration, the average body weight of male and female Osborne-Mendel rats is 0.452 kg,

yielding an F=0.033 kg/food/day
 

NOAEL= 1 ppm in food = 1 mg/kg food

1 mg DDT/kg food x 0.033 kg food/day/0.452 kg body weight = 0.07 mg DDT/kg/day.
 

Equivalents for other food concentrations used in this study, as calculated by using the EPA 1986d

Reference Values for Risk Assessment:
 

5 ppm= 0.4 mg/kg/day

10 ppm= 0.7 mg/kg/day

50 ppm= 3.7 mg/kg/day
 

(3) EPA 1988g Method, Subchronic Duration Average Body Weights 

For subchronic duration (actually less than the duration of this study), the average body weight of male

and female Osborne-Mendel rat 0.232 kg, yielding an F=0.0213 kg/food/day 

NOAEL= 1 ppm in food = 1 mg/kg food

1 mg DDT/kg food x 0.0213 kg food/day/0.232 kg body weight = 0.09 mg DDT/kg/day 

Equivalents for other food concentrations used in this study, as calculated by using the EPA 1986d

Reference Values for Risk Assessment: 

5 ppm= 0.5 mg/kg/day

10 ppm= 0.9 mg/kg/day

50 ppm= 4.6 mg/kg/day 

If an inhalation study in animals, list conversion factors used in determining human equivalent dose: NA 

Was a conversion used from intermittent to continuous exposure?

No 

Other additional studies or pertinent information that lend support to this MRL: In the Fitzhugh and

Nelson (1947) study, 16 female Osborne-Mendel rats were fed a diet containing 1,000 ppm technical

DDT for 12 weeks. Using EPA 1988g reference values for female body weight and subchronic food

consumption, the diet provided approximately 96 mg DDT/kg/day. Sacrifices were conducted at 

cessation of dosing and at various intervals after a DDT-free period. Liver changes were similar to those

seen in the Laug et al. (1950) study, although of increased severity, and were still present in rats killed

after 2 weeks in a DDT-free diet. Minimal liver changes were apparent after 4–6 weeks of recovery, and

complete recovery was seen after 8 weeks. Hepatic effects ranging from increased liver weights to

cellular necrosis have been reported in animals after chronic exposure in the diet. 

The Laug et al. (1950) study serves also as the basis for an oral RfD derived by EPA for DDT (IRIS

2001a). The Fitzhugh and Nelson (1947) study is considered supportive for the RfD. 

Agency Contact (Chemical Manager): Obaid Faroon 
DDT, DDE, and DDD A-9 

APPENDIX A 

MINIMAL RISK LEVEL WORKSHEET 

Chemical Name: DDT 

CAS Number: 50-29-3 

Date: September 2002 

Profile Status: Third Draft Post Public 

Route: [ ] Inhalation [X] Oral 

Duration: [ ] Acute [ ] Intermediate [X] Chronic 

Graph Key: N/A 

Species: N/A 

An oral MRL for chronic-duration exposure to DDT was not derived because of the inadequacy of the 

available data on liver effects in animals to describe the dose-response relationship at low-dose levels. In 

a brief communication, Fitzhugh (1948) stated that histopathological lesions occurred in the liver of rats 

fed 10 ppm DDT in the diet for 2 years, but no experimental details were given, so the quality of the 

study cannot be evaluated. Using reference values for body weight and food consumption from EPA 

(1988), it can estimated that the 10 ppm dietary level provided DDT doses of approximately 

0.7 mg/kg/day. This dietary level was the lowest level tested in the study, but was still higher than the 

lowest level resulting in hepatic effects in the Laug et al. 1950 study used for derivation of the 

intermediate-duration MRL. 

Agency Contact (Chemical Manager): Obaid Faroon 

DDT, DDE, and DDD B-1 

APPENDIX B 

USER'S GUIDE 

Chapter 1 

Public Health Statement 

This chapter of the profile is a health effects summary written in non-technical language. Its intended 

audience is the general public especially people living in the vicinity of a hazardous waste site or 

chemical release. If the Public Health Statement were removed from the rest of the document, it would 

still communicate to the lay public essential information about the chemical. 

The major headings in the Public Health Statement are useful to find specific topics of concern. The 

topics are written in a question and answer format. The answer to each question includes a sentence that 

will direct the reader to chapters in the profile that will provide more information on the given topic. 

Chapter 2 

Relevance to Public Health 

This chapter provides a health effects summary based on evaluations of existing toxicologic, 

epidemiologic, and toxicokinetic information. This summary is designed to present interpretive, 

weight-of-evidence discussions for human health end points by addressing the following questions. 

1.	 What effects are known to occur in humans? 

2.	 What effects observed in animals are likely to be of concern to humans? 

3.	 What exposure conditions are likely to be of concern to humans, especially around hazardous 

waste sites? 

The chapter covers end points in the same order they appear within the Discussion of Health Effects by 

Route of Exposure section, by route (inhalation, oral, dermal) and within route by effect. Human data are 

presented first, then animal data. Both are organized by duration (acute, intermediate, chronic). In vitro 

data and data from parenteral routes (intramuscular, intravenous, subcutaneous, etc.) are also considered 

in this chapter. If data are located in the scientific literature, a table of genotoxicity information is 

included. 

The carcinogenic potential of the profiled substance is qualitatively evaluated, when appropriate, using 

existing toxicokinetic, genotoxic, and carcinogenic data. ATSDR does not currently assess cancer 

potency or perform cancer risk assessments. Minimal risk levels (MRLs) for noncancer end points (if 

derived) and the end points from which they were derived are indicated and discussed. 

Limitations to existing scientific literature that prevent a satisfactory evaluation of the relevance to public 

health are identified in the Chapter 3 Data Needs section. 
DDT, DDE, and DDD B-2 

APPENDIX B 

Interpretation of Minimal Risk Levels 

Where sufficient toxicologic information is available, we have derived minimal risk levels (MRLs) for 

inhalation and oral routes of entry at each duration of exposure (acute, intermediate, and chronic). These 

MRLs are not meant to support regulatory action; but to acquaint health professionals with exposure 

levels at which adverse health effects are not expected to occur in humans. They should help physicians 

and public health officials determine the safety of a community living near a chemical emission, given the 

concentration of a contaminant in air or the estimated daily dose in water. MRLs are based largely on 

toxicological studies in animals and on reports of human occupational exposure. 

MRL users should be familiar with the toxicologic information on which the number is based. Chapter 2, 

"Relevance to Public Health," contains basic information known about the substance. Other sections such 

as Chapter 3, Section 3.10, "Interactions with Other Substances,” and Section 3.11, "Populations that are 

Unusually Susceptible" provide important supplemental information. 

MRL users should also understand the MRL derivation methodology. MRLs are derived using a 

modified version of the risk assessment methodology the Environmental Protection Agency (EPA) 

provides (Barnes and Dourson 1988) to determine reference doses for lifetime exposure (RfDs). 

To derive an MRL, ATSDR generally selects the most sensitive end point which, in its best judgement, 

represents the most sensitive human health effect for a given exposure route and duration. ATSDR 

cannot make this judgement or derive an MRL unless information (quantitative or qualitative) is available 

for all potential systemic, neurological, and developmental effects. If this information and reliable 

quantitative data on the chosen end point are available, ATSDR derives an MRL using the most sensitive 

species (when information from multiple species is available) with the highest NOAEL that does not 

exceed any adverse effect levels. When a NOAEL is not available, a lowest-observed-adverse-effect 

level (LOAEL) can be used to derive an MRL, and an uncertainty factor (UF) of 10 must be employed. 

Additional uncertainty factors of 10 must be used both for human variability to protect sensitive 

subpopulations (people who are most susceptible to the health effects caused by the substance) and for 

interspecies variability (extrapolation from animals to humans). In deriving an MRL, these individual 

uncertainty factors are multiplied together. The product is then divided into the inhalation concentration 

or oral dosage selected from the study. Uncertainty factors used in developing a substance-specific MRL 

are provided in the footnotes of the LSE Tables. 

Chapter 3 

Health Effects 

Tables and Figures for Levels of Significant Exposure (LSE) 

Tables (3-1, 3-2, and 3-3) and figures (3-1 and 3-2) are used to summarize health effects and illustrate 

graphically levels of exposure associated with those effects. These levels cover health effects observed at 

increasing dose concentrations and durations, differences in response by species, minimal risk levels 

(MRLs) to humans for noncancer end points, and EPA's estimated range associated with an upper- bound 

individual lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. Use the LSE tables and figures for a 

quick review of the health effects and to locate data for a specific exposure scenario. The LSE tables and 

figures should always be used in conjunction with the text. All entries in these tables and figures 

represent studies that provide reliable, quantitative estimates of No-Observed-Adverse-Effect Levels 

(NOAELs), Lowest-Observed-Adverse-Effect Levels (LOAELs), or Cancer Effect Levels (CELs). 
DDT, DDE, and DDD B-3 

APPENDIX B 

The legends presented below demonstrate the application of these tables and figures. Representative 

examples of LSE Table 3-1 and Figure 3-1 are shown. The numbers in the left column of the legends 

correspond to the numbers in the example table and figure. 

LEGEND 

See LSE Table 3-1 

(1)	 Route of Exposure One of the first considerations when reviewing the toxicity of a substance using 

these tables and figures should be the relevant and appropriate route of exposure. When sufficient 

data exists, three LSE tables and two LSE figures are presented in the document. The three LSE 

tables present data on the three principal routes of exposure, i.e., inhalation, oral, and dermal (LSE 

Table 3-1, 3-2, and 3-3, respectively). LSE figures are limited to the inhalation (LSE Figure 3-1) 

and oral (LSE Figure 3-2) routes. Not all substances will have data on each route of exposure and 

will not therefore have all five of the tables and figures. 

(2)	 Exposure Period Three exposure periods - acute (less than 15 days), intermediate (15–364 days), 

and chronic (365 days or more) are presented within each relevant route of exposure. In this 

example, an inhalation study of intermediate exposure duration is reported. For quick reference to 

health effects occurring from a known length of exposure, locate the applicable exposure period 

within the LSE table and figure. 

(3)	 Health Effect The major categories of health effects included in LSE tables and figures are death, 

systemic, immunological, neurological, developmental, reproductive, and cancer. NOAELs and 

LOAELs can be reported in the tables and figures for all effects but cancer. Systemic effects are 

further defined in the "System" column of the LSE table (see key number 18). 

(4)	 Key to Figure Each key number in the LSE table links study information to one or more data points 

using the same key number in the corresponding LSE figure. In this example, the study represented 

by key number 18 has been used to derive a NOAEL and a Less Serious LOAEL (also see the 2 

"18r" data points in Figure 3-1). 

(5)	 Species The test species, whether animal or human, are identified in this column. Chapter 2, 

"Relevance to Public Health," covers the relevance of animal data to human toxicity and 

Section 3.5, "Toxicokinetics," contains any available information on comparative toxicokinetics. 

Although NOAELs and LOAELs are species specific, the levels are extrapolated to equivalent 

human doses to derive an MRL. 

(6)	 Exposure Frequency/Duration The duration of the study and the weekly and daily exposure 

regimen are provided in this column. This permits comparison of NOAELs and LOAELs from 

different studies. In this case (key number 18), rats were exposed to 1,1,2,2-tetrachloroethane via 

inhalation for 6 hours per day, 5 days per week, for 3 weeks. For a more complete review of the 

dosing regimen refer to the appropriate sections of the text or the original reference paper, i.e., 

Nitschke et al. 1981. 

(7)	 System This column further defines the systemic effects. These systems include: respiratory, 

cardiovascular, gastrointestinal, hematological, musculoskeletal, hepatic, renal, and dermal/ocular. 

"Other" refers to any systemic effect (e.g., a decrease in body weight) not covered in these systems. 

In the example of key number 18, 1 systemic effect (respiratory) was investigated. 
DDT, DDE, and DDD B-4 

APPENDIX B 

(8)	 NOAEL A No-Observed-Adverse-Effect Level (NOAEL) is the highest exposure level at which no 

harmful effects were seen in the organ system studied. Key number 18 reports a NOAEL of 3 ppm 

for the respiratory system which was used to derive an intermediate exposure, inhalation MRL of 

0.005 ppm (see footnote "b"). 

(9)	 LOAEL A Lowest-Observed-Adverse-Effect Level (LOAEL) is the lowest dose used in the study 

that caused a harmful health effect. LOAELs have been classified into "Less Serious" and 

"Serious" effects. These distinctions help readers identify the levels of exposure at which adverse 

health effects first appear and the gradation of effects with increasing dose. A brief description of 

the specific end point used to quantify the adverse effect accompanies the LOAEL. The respiratory 

effect reported in key number 18 (hyperplasia) is a Less serious LOAEL of 10 ppm. MRLs are not 

derived from Serious LOAELs. 

(10)	 Reference The complete reference citation is given in Chapter 9 of the profile. 

(11)	 CEL A Cancer Effect Level (CEL) is the lowest exposure level associated with the onset of 

carcinogenesis in experimental or epidemiologic studies. CELs are always considered serious 

effects. The LSE tables and figures do not contain NOAELs for cancer, but the text may report 

doses not causing measurable cancer increases. 

(12)	 Footnotes Explanations of abbreviations or reference notes for data in the LSE tables are found in 

the footnotes. Footnote "b" indicates the NOAEL of 3 ppm in key number 18 was used to derive an 

MRL of 0.005 ppm. 

LEGEND 

See Figure 3-1 

LSE figures graphically illustrate the data presented in the corresponding LSE tables. Figures help the 

reader quickly compare health effects according to exposure concentrations for particular exposure 

periods. 

(13)	 Exposure Period The same exposure periods appear as in the LSE table. In this example, health 

effects observed within the intermediate and chronic exposure periods are illustrated. 

(14)	 Health Effect These are the categories of health effects for which reliable quantitative data exists. 

The same health effects appear in the LSE table. 

(15)	 Levels of Exposure concentrations or doses for each health effect in the LSE tables are graphically 

displayed in the LSE figures. Exposure concentration or dose is measured on the log scale "y" axis. 

Inhalation exposure is reported in mg/m3

 or ppm and oral exposure is reported in mg/kg/day. 

(16)	 NOAEL In this example, the open circle designated 18r identifies a NOAEL critical end point in 

the rat upon which an intermediate inhalation exposure MRL is based. The key number 18 

corresponds to the entry in the LSE table. The dashed descending arrow indicates the extrapolation 

from the exposure level of 3 ppm (see entry 18 in the Table) to the MRL of 0.005 ppm (see 

footnote "b" in the LSE table). 

(17)	 CEL Key number 38r is 1 of 3 studies for which Cancer Effect Levels were derived. The diamond 

symbol refers to a Cancer Effect Level for the test species-mouse. The number 38 corresponds to 

the entry in the LSE table. 
DDT, DDE, and DDD B-5 

APPENDIX B 

(18)	 Estimated Upper-Bound Human Cancer Risk Levels This is the range associated with the 

upper-bound for lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. These risk levels are 

derived from the EPA's Human Health Assessment Group's upper-bound estimates of the slope of 

the cancer dose response curve at low dose levels (q1*). 

(19)	 Key to LSE Figure The Key explains the abbreviations and symbols used in the figure. 
2 

SAMPLE 

6 Table 3-1. Levels of Significant Exposure to [Chemical x] – Inhalation 1 

3 4 

6

Key to figurea Species 

Exposure frequency/ duration System 

NOAEL (ppm) 

INTERMEDIATE EXPOSURE 

5 6 7 8 

6 Systemic 9 9 9 96 18 Rat 13 wk 

5 d/wk 6 hr/d 

Resp 3b 

CHRONIC EXPOSURE Cancer 38 Rat 18 mo 

5 d/wk 7 hr/d 

39 Rat 89–104 wk 

5 d/wk 6 hr/d 

40 Mouse 79–103 wk 

5 d/wk 6 hr/d 

a The number corresponds to entries in Figure 3-1. 

Less serious (ppm) 

LOAEL (effect) 

Serious (ppm) 

9 

9

10 (hyperplasia) 

11 

9

20 (CEL, multiple 

organs) 

10 (CEL, lung tumors

nasal tumors) 

10 (CEL, lung tumors

hemangiosarcom

b Used to derive an intermediate inhalation Minimal Risk Level (MRL) of 5 x 10-3 ppm; dose adjusted for intermittent ex

an uncertainty factor of 100 (10 for extrapolation from animal to humans, 10 for human variability). 12 6


DDT, DDE, and DDD C-1 

APPENDIX C 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

ACOEM American College of Occupational and Environmental Medicine 

ACGIH American Conference of Governmental Industrial Hygienists 

ADI acceptable daily intake 

ADME absorption, distribution, metabolism, and excretion 

AED atomic emission detection 

AOEC Association of Occupational and Environmental Clinics 

AFID alkali flame ionization detector 

AFOSH Air Force Office of Safety and Health 

ALT alanine aminotransferase 

AML acute myeloid leukemia 

AOAC Association of Official Analytical Chemists 

AP alkaline phosphatase 

APHA American Public Health Association 

AST aspartate aminotranferase 

atm atmosphere 

ATSDR Agency for Toxic Substances and Disease Registry 

AWQC Ambient Water Quality Criteria 

BAT best available technology 

BCF bioconcentration factor 

BEI Biological Exposure Index 

BSC Board of Scientific Counselors 

C centigrade 

CAA Clean Air Act 

CAG Cancer Assessment Group of the U.S. Environmental Protection Agency 

CAS Chemical Abstract Services 

CDC Centers for Disease Control and Prevention 

CEL cancer effect level 

CELDS Computer-Environmental Legislative Data System 

CERCLA Comprehensive Environmental Response, Compensation, and Liability Act 

CFR Code of Federal Regulations 

Ci curie 

CI confidence interval 

CL ceiling limit value 

CLP Contract Laboratory Program 

cm centimeter 

CML chronic myeloid leukemia 

CPSC Consumer Products Safety Commission 

CWA Clean Water Act 

DHEW Department of Health, Education, and Welfare 

DHHS Department of Health and Human Services 

DNA deoxyribonucleic acid 

DOD Department of Defense 

DOE Department of Energy 

DOL Department of Labor 

DOT Department of Transportation 
 

DDT, DDE, and DDD C-2 

APPENDIX C 

DOT/UN/ Department of Transportation/United Nations/

 NA/IMCO North America/International Maritime Dangerous Goods Code 

DWEL drinking water exposure level 

ECD electron capture detection 

ECG/EKG electrocardiogram 

EEG electroencephalogram 

EEGL Emergency Exposure Guidance Level 

EPA Environmental Protection Agency 

F Fahrenheit 

F1 first-filial generation 

FAO Food and Agricultural Organization of the United Nations 

FDA Food and Drug Administration 

FEMA Federal Emergency Management Agency 

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act 

FPD flame photometric detection 

fpm feet per minute 

FR Federal Register 

FSH follicle stimulating hormone 

g gram 

GC gas chromatography 

gd gestational day 

GLC gas liquid chromatography 

GPC gel permeation chromatography 

HPLC high-performance liquid chromatography 

HRGC high resolution gas chromatography 

HSDB Hazardous Substance Data Bank 

IARC International Agency for Research on Cancer 

IDLH immediately dangerous to life and health 

ILO International Labor Organization 

IRIS Integrated Risk Information System 

Kd adsorption ratio 

kg kilogram 

Koc organic carbon partition coefficient 

Kow octanol-water partition coefficient 

L liter 

LC liquid chromatography 

LCLo lethal concentration, low 

LC50 lethal concentration, 50% kill 

LDLo lethal dose, low 

LD50 lethal dose, 50% kill 

LDH lactic dehydrogenase 

LH luteinizing hormone 

LT50 lethal time, 50% kill 

LOAEL lowest-observed-adverse-effect level 

LSE Levels of Significant Exposure 

m meter 

MA trans,trans-muconic acid 

MAL maximum allowable level 

mCi millicurie 

MCL maximum contaminant level 

MCLG maximum contaminant level goal 
DDT, DDE, and DDD C-3 

APPENDIX C 

MFO mixed function oxidase 

mg milligram 

mL milliliter 

mm millimeter 

mmHg millimeters of mercury 

mmol millimole 

mppcf millions of particles per cubic foot 

MRL Minimal Risk Level 

MS mass spectrometry 

NAAQS National Ambient Air Quality Standard 

NAS National Academy of Science 

NATICH National Air Toxics Information Clearinghouse 

NATO North Atlantic Treaty Organization 

NCE normochromatic erythrocytes 

NCEH National Center for Environmental Health 

NCI National Cancer Institute 

ND not detected 

NFPA National Fire Protection Association 

ng nanogram 

NIEHS National Institute of Environmental Health Sciences 

NIOSH National Institute for Occupational Safety and Health 

NIOSHTIC NIOSH's Computerized Information Retrieval System 

NLM National Library of Medicine 

nm nanometer 

NHANES National Health and Nutrition Examination Survey 

nmol nanomole 

NOAEL no-observed-adverse-effect level 

NOES National Occupational Exposure Survey 

NOHS National Occupational Hazard Survey 

NPD nitrogen phosphorus detection 

NPDES National Pollutant Discharge Elimination System 

NPL National Priorities List 

NR not reported 

NRC National Research Council 

NS not specified 

NSPS New Source Performance Standards 

NTIS National Technical Information Service 

NTP National Toxicology Program 

ODW Office of Drinking Water, EPA 

OERR Office of Emergency and Remedial Response, EPA 

OHM/TADS Oil and Hazardous Materials/Technical Assistance Data System 

OPP Office of Pesticide Programs, EPA 

OPPTS Office of Prevention, Pesticides and Toxic Substances, EPA 

OPPT Office of Pollution Prevention and Toxics, EPA 

OR odds ratio 

OSHA Occupational Safety and Health Administration 

OSW Office of Solid Waste, EPA 

OW Office of Water 

OWRS Office of Water Regulations and Standards, EPA 

PAH polycyclic aromatic hydrocarbon 

PBPD physiologically based pharmacodynamic 
DDT, DDE, and DDD C-4 

APPENDIX C 

PBPK physiologically based pharmacokinetic 

PCE polychromatic erythrocytes 

PEL permissible exposure limit 

PID photo ionization detector 

pg picogram 

pmol picomole 

PHS Public Health Service 

PMR proportionate mortality ratio 

ppb parts per billion 

ppm parts per million 

ppt parts per trillion 

PSNS pretreatment standards for new sources 

RBC red blood cell 

REL recommended exposure level/limit 

RfC reference concentration 

RfD reference dose 

RNA ribonucleic acid 

RTECS Registry of Toxic Effects of Chemical Substances 

RQ reportable quantity 

SARA Superfund Amendments and Reauthorization Act 

SCE sister chromatid exchange 

SGOT serum glutamic oxaloacetic transaminase 

SGPT serum glutamic pyruvic transaminase 

SIC standard industrial classification 

SIM selected ion monitoring 

SMCL secondary maximum contaminant level 

SMR standardized mortality ratio 

SNARL suggested no adverse response level 

SPEGL Short-Term Public Emergency Guidance Level 

STEL short term exposure limit 

STORET Storage and Retrieval 

TD50 toxic dose, 50% specific toxic effect 

TLV threshold limit value 

TOC total organic carbon 

TPQ threshold planning quantity 

TRI Toxics Release Inventory 

TSCA Toxic Substances Control Act 

TWA time-weighted average 

UF uncertainty factor 

U.S. United States 

USDA United States Department of Agriculture 

USGS United States Geological Survey 

VOC volatile organic compound 

WBC white blood cell 

WHO World Health Organization 

> greater than $ greater than or equal to 

= equal to 

< less than 
DDT, DDE, and DDD C-5 

APPENDIX C 

# less than or equal to 

% percent α alpha β beta γ gamma δ delta 

μm micrometer 

μg microgram q1

* cancer slope factor 

– negative 

+ positive 

(+) weakly positive result 

(–) weakly negative result 

 

DDT, DDE, and DDD D-1 

APPENDIX D
 

D. HEALTH EFFECTS IN WILDLIFE POTENTIALLY RELEVANT TO HUMAN HEALTH 

Overview. The 1972 EPA decision to ban DDT for most crop uses in the United States was 

significantly influenced by a large body of scientific information indicating adverse health effects in 

wildlife (EPA 1975). The wildlife health effects that were considered by EPA in banning DDT were 

severe, including lethality of DDT in birds and fish and DDE-induced reproductive effects in birds, 

particularly eggshell thinning (EPA 1975). Wildlife may be regarded as sentinels for human health (NRC 

1991). Although it is difficult to draw firm conclusions about human health from the toxicity observed in 

sentinel wildlife species in the environment, these observations have motivated more precise experiments 

with known doses of DDT, DDD, and DDE in both wildlife species and more traditional laboratory 

animal models, as well as identifying key health end points for epidemiological investigation. Wildlife 

field observations have also stimulated investigation of reproductive effects in mammalian models more 

directly relevant to humans and in vitro and mechanism of action studies that have resulted in the 

identification of some DDT isomers and metabolites as androgen antagonists and estrogen agonists. 

There have been a number of intriguing mechanistic studies of DDT isomers and metabolites in fish that 

relate to reproductive and developmental effects (Das and Thomas 1999; Faulk et al. 1999; Khan and 

Thomas 1998; Loomis and Thomas 1999; Sperry and Thomas 1999; Thomas 1999); these are discussed 

in Section 3.6.2, Mechanisms of Toxicity. Environmental monitoring studies have shown that DDT, 

DDE, and DDD are also highly persistent in the environment (see Section 6.3, Environmental Fate), and 

therefore, continue to present a potential health hazard both to humans and wildlife. It should not be 

forgotten that wildlife is part of the same broad ecological web as humans, and thus, toxic effects on 

wildlife ultimately affect the quality of human life as well. 

Field observations of health effects in wildlife have strongly influenced the design of experimental 

studies. A high degree of causal uncertainty usually exists in field studies because wildlife species are 

frequently co-exposed to many other toxicants, such as other organochlorine pesticides and heavy metals, 

and are subjected to a variety of other uncontrolled and unknown stresses that may affect their health and 

confound an analysis of causation of a particular effect. 

The purpose of this section is to provide a qualitative synopsis of health effects in terrestrial wildlife to 

address the potential concern that effects observed in wildlife that are attributable to DDT/DDE/DDD 

exposure may also occur in humans. The organization of Appendix D, Health Effects in Wildlife 

Potentially Relevant to Human Health, closely parallels that of Section 3.2, Discussion of Health Effects 

by Route of Exposure, to facilitate weight-of-evidence evaluations that may use both human health and 

wildlife toxicological effects data. The primary focus of this section is on experimental studies with 

known doses of DDT isomers and metabolites, but key observations from ecological field studies of 

wildlife have also been highlighted. These ecological field studies include observations of reproductive 

developmental effects in alligators living in contaminated Lake Apopka, Florida. No case reports of 

lethality in wildlife populations from exposure during DDT application were included because the 

exposure scenario is no longer relevant since the banning of DDT in 1972. 

Where statistical significance of an observed effect was evaluated in a wildlife experimental study, 

statistical significance (p#0.05) was indicated in the text of this section using the term “significant”; if the 

term “significant” was not used in describing an observed effect, then no statistical evaluation is implied. 

This convention for this section is in contrast to the convention used in Section 3.2, Discussion of Health 

Effects by Route of Exposure, in which observed effects that are discussed in the text are usually assumed 

to be significant unless otherwise specified. Unfortunately, many experimental studies in wildlife used a 

small number of animals, making meaningful statistical tests of significance difficult. Study designs for 
 

DDT, DDE, and DDD D-2 

APPENDIX D 

testing toxicity in more traditional laboratory animals have generally included larger samples sizes, 

facilitating tests of significance. 

Experimental wildlife health effects information concerning DDT/DDE/DDD exposures was located for 

approximately 40 terrestrial species and on numerous end points. A large proportion of the information, 

however, is represented by a relatively small number of species and end points. The most heavily studied 

taxonomic group was birds, and among birds, the mallard/white Pekin duck (Anas platyrhyncus), 

Japanese quail (Coturnix coturnix japonica), domestic fowl, and ringed turtle dove (Streptopelia risoria) 

had the greatest representation. The most commonly reported end points were lethality, and neurological 

and reproductive end points. Of particular interest are those effects that were observed consistently across 

species and in spite of variability in exposure scenarios. The significant health effects most consistently 

reported were lethality (several taxa), hepatic (liver enzyme induction and liver damage in birds), 

endocrine (estrogenic effects in several taxa, and reduced thyroid weight and altered thyroid activity in 

birds), neurological (tremors in several taxa), reproductive (oviposition delay and eggshell thinning in 

birds), and developmental (reduced chick survival in birds, testicular feminization). 

A hazard identification table (Table D-1) is provided so the reader may quickly scan the wildlife database 

for species or toxicological end points that are of particular interest. The table is divided into the 

following four sections based on species taxonomy: Wild mammals, reptiles, and amphibians; 

Birds—raptors, wading birds, water birds; Birds—gallinaceous birds; and Birds—passerines and 

nonpasserine ground birds. The organization of effect categories within each section of the table parallels 

that of the text so that more detailed information for particular table entries may be readily located. The 

specific toxicological effects reflected in the table under each effect category may vary between sections 

of the table, since specific end points may have been evaluated in certain species but not in others. Not 

every study mentioned in the text was reported in the table. Some studies that reported ambiguous results 

with respect to certain toxicological end points are included in the text, but no corresponding entry was 

made in Table D-1 for the ambiguous results. Not every significant effect reported in the text received a 

unique entry in Table D-1. Effects reported in different studies that would receive the same coding for a 

given species were all represented by a single entry in the table. Effects that were observed at high 

exposure levels but not at low exposure levels were entered once in the table as an observed effect. The 

individual isomers of DDT, DDE, and DDD, as well as technical-grades and unspecified mixtures, 

received unique codes in the table so that the reader may readily focus on compounds of particular 

interest. Since Table D-1 is intended for hazard identification, parenteral exposures were included, as 

well as oral, inhalation, and dermal routes that are more directly relevant to human environmental 

exposures. In contrast, the LSE tables in the previous section included only inhalation, oral, and dermal 

exposures. 

Table D-1 is intended to be a visual aid to illustrate the wildlife database; it is not intended to be a standalone document. The reader should refer to the text for more detailed experimental design information 

and for information concerning the statistical significance of the observed effects. 

D.1 Death 

Based on the following analysis, the order of susceptibility to DDT-induced mortality appears to be 

amphibians>mammals>birds. Animals on restricted diets, such as migrating or nesting animals, are 

generally more sensitive to DDT-induced lethality than animals fed on nutritionally sufficient diets. 

Mammals. Studies of DDT lethality in mammals are limited to bats and shrews, and generally attribute 

the cause of death to accumulation of DDT/DDE/DDD in the brain. In pipistrelle bats (pipistrellus), no 

mortality occurred after single oral doses of p,p’-DDT below 45 mg/kg body weight for up to 31 days 

postdosing, but 100% mortality occurred within 28 days in groups administered single doses of 
DDT, DDE, and DDD D-3 

APPENDIX D 

$95 mg/kg body weight; the estimated LD50 was 63 mg/kg body weight (Jefferies 1972). A single oral 

dose of technical-grade DDT at 20 mg/kg body weight caused some mortality in big brown bats 

(Eptesicus fuscus), while $40 mg/kg body weight was 100% lethal (Luckens and Davis 1964). For 

40 days, free-tailed bats (Tadarida brasiliensis) were fed mealworms that were raised in wheat bran 

containing 100 ppm p,p’-DDE; the treated bats lost body weight quicker and died sooner than an 

untreated control group during a postexposure starvation period (Clark and Kroll 1977). Among the 

17 treated bats, a strong negative relationship was seen between p,p’-DDE residue in the brain and 

percent lipid in the carcass, suggesting that p,p’-DDE mobilized from fat will accumulate in the brain 

(Clark and Kroll 1977). Based on a review of the experimental literature and DDT/DDE/DDD residues in 

dead wild bats, Clark (1981) estimated that the minimum lethal concentrations are 12 ppm (w/w) DDT in 

the brain of the little brown bat (Myotis liucifugus), and 460 and 540 ppm DDE in the free-tailed bat and 

the little brown bat, respectively. The combined effect of starvation and exercise on the disposition of 

existing DDE body burden was evaluated in wild-captured free-tailed bats (Geluso et al. 1976). Median 

brain DDE in the starved-sedentary group was increased by a factor of 12.7 in young bats to 53.8 in older 

bats relative to the unstarved group, and median brain DDE in the starved-exercised group was increased 

by a factor of 43.2 in young bats to 123.1 in older bats over levels in the unstarved group, suggesting that 

bats may be particularly vulnerable to DDE-induced mortality during migration. 

Dietary LC50 values in short-tailed shrews (Blarina brevicauda) for 14-day p,p’-DDT exposures ranged 

from 651 to 1,160 mg/kg diet when DDT was dissolved in oil prior to mixing in the diet, and ranged from 

839 to 2,550 mg/kg diet when DDT was added as the powder (Blus 1978). The concentration of DDT in 

the brains of shrews that died was highly variable, and DDE residues were relatively low, suggesting that 

accumulation in the brain had little effect on mortality in shrews. 

Amphibians. Lethality information in adult amphibians is limited to studies in the common frog (Rana 

temporaria) and the bullfrog (Rana catesbeiana). No mortality was seen in adult common frogs dosed 

twice weekly for 8 weeks with DDT (isomeric composition not specified) at 0.6 mg/kg, but in treated 

frogs that were not fed, 50% mortality was seen by the end of the exposure period (Harri et al. 1979). 

The LD50 in the adult common frog 20 days after a single oral administration of DDT (unspecified 

isomer) in gelatin capsules at unreported dose levels was estimated to be 7.6 mg/kg body weight; LD50 

values at 3 and 4 days after the single oral administration were approximately 85 and 25 mg/kg, 

respectively (Harri et al. 1979). Mortality was seen in common frog tadpoles immersed for 1 hour in 1 or 

10 ppm p,p’-DDT, but not in #0.1 ppm (Cooke 1970a). In the adult bullfrog, a 14-day oral LD50 of 

>2,000 mg/kg body weight was reported following a single oral administration of p,p’-DDT in gelatin 

capsules at unreported dose levels (U.S. Fish and Wildlife Service 1984). 

Birds. Historically, observations of high mortality in local wild bird populations occurred coincidentally 

with application of DDT for pest control (EPA 1975). At the site of DDT application, local bird 

populations were acutely exposed by inhalation of airborne DDT, by ingestion of DDT residues on 

insects and other invertebrates such as earthworms, and by direct ingestion of DDT while preening. 

Several authors have postulated that high mortality may occur during times of stress, such as during 

nesting or during migration, when energy from fat stores is mobilized (EPA 1975). As fat stores are 

depleted, fat-stored and newly absorbed DDT could distribute to the brain; as in mammals, accumulation 

of high levels in the brain of birds is hypothesized to be lethal (EPA 1975). Since DDT was banned, the 

primary route of exposure to DDT compounds in wild bird populations has been in the diet through the 

food chain. Available experimental data on bird lethality indicate that DDT/DDE/DDD have moderate to 

low toxicity in birds after ingestion in the diet or from gavage administration (WHO 1989). 

Acute LD50 values of orally administered DDT (unspecified isomeric composition) in 2-month-old 

Japanese quail (Coturnix coturnix japonica), p,p’-DDT in 4-month-old pheasant (Phasianus colchicus), 

technical-grade DDT in 6-month-old California quail (Callipepia californica), DDT (unspecified isomeric 
DDT, DDE, and DDD D-4 

APPENDIX D 

composition) in 3-month-old Mallard ducks (Anas platyrhyncus), DDT (unspecified isomeric 

composition) in the rock dove (Columba livia), and p,p’-DDT in the adult sandhill crane (Grus 

canadensis) ranged from 595 mg/kg body weight in 6-month-old male California quail to >4,000 mg/kg 

body weight in male and female rock doves (U.S. Fish and Wildlife Service 1984). Dietary LC50 values 

for DDT (unspecified isomeric composition) ingestion ranged from 311 to 1,869 mg/kg diet after 5-day 

exposures in immature bobwhite quail (Colinus virginianus), Japanese quail, Mallard duck, and pheasant 

(U.S. Fish and Wildlife Service 1965, 1975). A dietary LC50 of 1,612 mg/kg diet for p,p’-DDT after a 

5-day exposure was reported in 10-week-old clapper rails (Rallus longirostris) (Van Veltzen and Kreitzer 

1975). Dietary LC50 values for technical-grade DDT after 5-day exposures in immature bobwhite quail, 

cardinal (Richmondena cardinalis), house sparrow (Passer domesticus), and blue jay (Cyanocitta 

cristata) ranged from 415 to 1,610 mg/kg diet (Hill et al. 1971). Dietary LC50 values for p,p’-DDT after 

10-day exposures in immature Mallard ducks ranged from 1,202 to 1,622 mg/kg, and in adult Mallard 

ducks was 1,419 mg/kg (Friend and Trainer 1974a). The dietary LC50 values for DDT (unspecified 

isomeric composition) after acute (<10 days) exposure in immature bobwhite quail, pheasants, and 

Mallard ducks ranged from 500 to 1,000 mg/kg diet and in adult bobwhite and pheasants, ranged from 

1,000 to 2,500 mg/kg diet; after intermediate (<100 days) exposure, LC50 values in immature birds ranged 

from 100 to 400 mg/kg diet and in adult birds (including Mallards) from >100 to 1,000 mg/kg diet (U.S. 

Fish and Wildlife Service 1963). In the red-winged blackbird (Agelaius phoeniceus), dietary LC50 values 

after acute (<10 days) and intermediate (<30 days) exposures to DDT (unspecified isomeric composition) 

were 1,000 and 500 mg/kg diet, respectively (U.S. Fish and Wildlife Service 1963). Dietary LC50 values 

for ingestion of p,p’-DDE (isomeric composition not specified) ranged from 825 to 3,570 mg/kg diet after 

5-day exposures in immature bobwhite quail, Japanese quail, Mallard duck, and pheasant (U.S. Fish and 

Wildlife Service 1975). Acute (exposure duration not specified) oral LD50 values for DDD (unspecified 

isomeric composition) were 386, >700, and >2,000 mg/kg body weight in 3- to 4-month-old pheasants, 

6-month-old California quail, and 3-month-old Mallard ducks (U.S. Fish and Wildlife Service 1984). 

Dietary LC50 values for technical-grade DDD ranged from 445 to 4,810 mg/kg diet after 5-day exposures 

in immature bobwhite quail, Japanese quail, pheasant, and Mallard ducks (U.S. Fish and Wildlife Service 

1975). 

Time to death in adult and immature bald eagles (Haliaetus leucocephalus) was inversely related to 

dietary technical-grade DDT level in the feed; birds fed 4,000 ppm died within 23 days, while birds at 

lower exposure levels survived for up to 112 days (Chura and Stewart 1967). All three bald eagles (adult 

and immature) fed 4,000 ppm technical-grade DDT in the diet died after at least 15 days of exposure, and 

one of two eagles fed 160 ppm technical DDT in the diet died after at least 76 days of exposure (Chura 

and Stewart 1967; Locke et al. 1966). Groups of 10–30 Mallard ducks (5-day-old, 30-day-old, and adult 

ducks) were fed p,p’-DDT at 250–2,000 ppm for 10 days in a diet specially formulated to provide 

adequate nutrition, but minimize fat formation (Friend and Trainer 1974a); onset and mean time to 

mortality occurred earlier in younger ducks compared to older ducks, although mortality was seen in all 

age groups. Survival times up to 49 and 29 days were seen in house sparrows fed 200 and 300 ppm DDT 

(unspecified isomer), respectively, in chick starter mash (Bernard 1963). DDT (isomeric composition not 

specified) provided in drinking water (12% solution) or as residue in earthworms sprayed twice daily with 

a 12% solution (DDT residue concentration in earthworms was not reported) caused 100% mortality in 

wild-captured adult house sparrows within 17 days of the initial exposure (Boykins 1967). 

Liver DDE was 128 and 253 ppm in two male kestrels that died compared to a mean level of 24 ppm in 

surviving males, and brain DDE residues were 212 and 301 ppm in the two birds that died compared to a 

mean level of 15 ppm in surviving males (Porter and Wiemeyer 1972). DDT and DDE residue levels in 

the brains of ducks that died were between 6 and 17 times the levels seen in ducks that survived to 

10 days postexposure in groups of 10–30 Mallard ducks (5-day-old, 30-day-old, and adult ducks) 

provided with p,p’-DDT at 250–2,000 ppm for 10 days in a diet specially formulated to provide adequate 

nutrition, but minimize fat formation (Friend and Trainer 1974a); DDT residue levels were generally 
 

 

DDT, DDE, and DDD D-5 

APPENDIX D 

higher in the younger ducks than in older ducks. In contrast, DDE levels in the brains of cowbirds 

(Molothrus ater) that survived 8–12 days of dietary exposure to 500 ppm p,p’-DDT were higher than 

brain DDE levels in birds that died during exposure, suggesting that brain DDE was not the principle 

factor inducing mortality in cowbirds (Stickel et al. 1966). 

Observations of mortality were reported in several other studies. Mortality was observed in 2/12 male 

kestrels (Falco sparverius), one adult and one yearling, exposed for at least 14 months on diets containing 

2.8 ppm p,p’-DDE (Porter and Wiemeyer 1972). There was no significant increase in mortality rate in 

barn owls (Tyto alba) fed 3 ppm DDE (isomeric composition not reported) in the diet for 2 years 

compared with controls (Mendenhall et al. 1983); however, two DDE-treated females died when they 

were unable to lay eggs that were “virtually shell-less.” Mortality occurred in 6–29% of female pheasants 

(Phasiannus colchicus) and 33–100% of male pheasants fed $100 ppm technical-grade DDT for at least 

22 days; no deaths occurred in birds fed 10 ppm for up to 101 days (Azevedo et al. 1965). Survival of 

Japanese quail administered up to 50 ppm DDT (unspecified isomeric composition) was reportedly 

comparable to control group survival throughout a 3-generation reproductive toxicity assay 

(Shellenberger 1978). o,p’- and p,p’-DDT increased pentobarbital-induced mortality rates in Japanese 

quail when administered in the diet at 100 ppm for 2–14 days prior to a single intramuscular injection of 

pentobarbital (Bitman et al. 1971). Mortality in Japanese quail (Coturnix coturnix japonica) was 

significantly exposure-related in birds fed diets containing between 700 and 1,600 ppm p,p’-DDT for 

20 days; factors that affected weight, such as restricted diet, sex, strain, and breeding condition, also 

affected susceptibility to DDT intoxication—heavier birds were less susceptible than lighter, more 

stressed birds (Gish and Chura 1970). No effect on hen mortality was seen in laying bobwhite quail 

(Colinus virginianus) administered up to 20 mg/bird of DDT (unspecified isomeric composition) every 

other day during a 4-week exposure period (Wilson et al. 1973). p,p’-DDT provided in the diet caused 

mortality in Bengalese finches (Lonchura striata) at 84 ppm (within 46 days) and 168 ppm (within 

35 days), but not at #42 ppm (Jefferies and Walker 1966). 

D.2 Systemic Effects 

Respiratory Effects. No experimental studies were located regarding respiratory effects in wild 

mammals, reptiles, or amphibians from exposure to DDT/DDE/DDD. 

Birds. No gross lung lesions were observed in 6-week-old pheasant chicks (Phasianus colchicus) fed 

100 ppm technical-grade DDT for up to 101 days or 500 ppm for up to 23 days (Azevedo et al. 1965). 

Cardiovascular Effects. No experimental studies were located regarding cardiovascular effects in 

wild mammals, reptiles, or amphibians from exposure to DDT/DDE/DDD. 

No cardiovascular effects were observed in birds after dietary exposures to DDT/DDE/DDD, but capsular 

bolus administration of DDT or DDE resulted in changes in heart morphology and function that may be 

interpreted to be secondary to thyroid effects. Cardiovascular lesions were limited to a single observation 

of poor heart muscular tone at relatively high intermediate duration exposures. 

Birds. Heart weights were not significantly affected in white pelicans (Pelecanus erythrorhynchos) 

exposed by daily oral administrations of a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDE, and 15 mg 

p,p’-DDD for 10 weeks (Greichus et al. 1975); doses were injected into the first fish fed to the birds each 

day. Heart weights were not significantly decreased in double-crested cormorants (Phalacrocorax 

auritus) fed diets containing up to 25 ppm of total DDT/DDE/DDD for 9 weeks (Greichus and Hannon 

1973). Heart weight was increased in homing pigeons (Columba livia) administered 3 mg/kg/day of 

p,p’-DDT by capsule for 8 weeks, but heart weights showed exposure-related decreases at doses of 

6–54 mg/kg/day, with an overall significant decreasing trend in heart weight with dose (Jefferies et al. 
 

 

DDT, DDE, and DDD D-6 

APPENDIX D 

1971). A significant dose-related decreasing trend in heart weight was also seen in pigeons administered 

daily doses of 18–72 mg/kg/day of p,p’-DDE by capsule for 8 weeks (Jefferies and French 1972). A 

dose-related increase in relative heart weight was observed in Bengalese finches (Lonchura striata) fed up 

to 0.3 mg/kg/day of p,p’-DDT for at least 15 weeks (Jefferies 1969). Jefferies (1969) suggested that the 

significant increasing trend in heart weight with dose observed in Bengalese finches at low doses was due 

to DDT-induced hyperthyroidism. In a later study in pigeons, Jefferies et al. (1971) refined the 

hypothesis, suggesting a biphasic heart response to changes in thyroid activity with increasing dose. 

They postulated that increased heart weight in homing pigeons at low, intermediate-duration oral 

exposures was due to a hyperthyroid condition, while a dose-related decrease in heart weights at higher 

oral doses in pigeons occurred due to hypothyroidism. The hypothesis of a hyper-hypothyroidism 

continuum with increasing dose was supported by observations (discussed below) that pulse rate and 

ventricular S-wave amplitude peaked at low doses in homing pigeons, with a dose-related decline of both 

parameters at higher dose levels. Further data supporting the hypothesis of biphasic thyroid activity in 

pigeons is provided in an observation of biphasic metabolic rate (Jefferies and French 1971), increased 

thyroid weights with decreased thyroidal colloid (Jefferies and French 1969, 1971), and increased adrenal 

weights (Jefferies et al. 1971). These effects are discussed further below under Endocrine Effects. 

Pulse rate was increased above controls at all dose levels tested in the homing pigeon fed p,p’-DDT by 

capsule for 3 or 6 weeks; pulse rate peaked at 18% above controls at 3 mg/kg/day, and showed a 

decreasing trend with dose level at 6–36 mg/kg/day (Jefferies et al. 1971). Bengalese finches showed an 

exposure-related increase in pulse rate at oral doses of p,p’-DDT up to 11.7 mg/kg/day (Jefferies et al. 

1971). 

The amplitude of the ventricular beat in homing pigeons peaked at 3 mg/kg/day after 3- or 6-week oral 

exposures and showed a significant exposure-related decrease at exposure levels up to 36 mg/kg/day 

(Jefferies et al. 1971). In Bengalese finches, there was no clear effect on the amplitude of the ventricular 

beat at oral dose levels up to 11.7 mg/kg/day for up to 6 weeks of exposure (Jefferies et al. 1971). 

No gross heart lesions were observed in 6-week-old pheasant chicks (Phasianus colchicus) fed 100 ppm 

technical-grade DDT for up to 101 days or 500 ppm for up to 23 days (Azevedo et al. 1965). Hearts in 

two homing pigeons receiving daily doses of $36 mg/kg/day p,p’-DDE (but not at #18 mg/kg/day) for 

8 weeks were “flaccid, with thin musculature” (Jefferies and French 1972). 

Gastrointestinal Effects. No experimental studies were located regarding gastrointestinal effects in 

wildlife from exposure to DDT/DDE/DDD. 

Hematological Effects. No experimental studies were located regarding hematological effects in 

wild mammals, reptiles, or amphibians from exposure to DDT/DDE/DDD. In experimental studies in 

birds, hematological effects were observed inconsistently. 

Hematological effects were inconsistently observed in birds after oral and subcutaneous exposures to 

DDT compounds. Bird hematology appeared to be more sensitive to gavage exposure than to dietary 

exposure. Relatively high, acute gavage exposure in crowned guinea hens resulted in decreased 

hemoglobin, erythrocyte count, and hematocrit, while no effect on hemoglobin content or hematocrit 

were observed in Japanese quail and cormorants exposed in the diet for an intermediate duration. 

Alternative hypotheses explaining the decreases in red blood cell indices after bolus doses include a direct 

disruption of erythrocyte membranes by lipophilic DDT, a reduced erythrocyte viability from nuclear 

metabolic interference, an inhibition of erythrocyte proliferation in hematopoetic tissue, or an estrogenic 

inhibitory effect on red blood cells (Fourie and Hattingh 1979). Intermediate-duration dietary exposures 

in bobwhite quail and ringed turtle doves resulted in increased and decreased hematocrits, respectively. 
 

DDT, DDE, and DDD D-7 

APPENDIX D 

Except for decreased red blood cell count in cockerels (young male chickens; young roosters; domestic 

fowl), parenteral exposures in domestic fowl showed no effects in hematological parameters. 

Birds. Blood hemoglobin was not significantly affected in double-crested cormorants (Phalacrocorax 

auritus) fed diets containing up to 25 ppm of total DDT/DDE/DDD for 9 weeks (Greichus and Hannon 

1973). Hemoglobin was significantly decreased in crowned guinea fowl (Numida meleagris) after an 

assumed daily gavage dose of 75 mg/kg body weight (exposure units were ambiguously reported) of 

technical-grade DDT on 5 consecutive days (Fourie and Hattingh 1979), while a 12-week dietary 

exposure to up to 100 ppm p,p’-DDE did not significantly affect hemoglobin concentration in Japanese 

quail (Coturnix coturnix japonica) (Dieter 1974). No effect on hemoglobin level was seen in chickens 

administered daily subcutaneous injections of DDT (unspecified isomeric composition) for up to 81 days 

for a cumulative dose of 2–3 g/chicken (Burlington and Lindeman 1952). 

Erythrocyte count was significantly decreased in crowned guinea fowl after an assumed daily gavage 

dose of 75 mg/kg body weight (exposure units were ambiguously reported) of technical-grade DDT on 

5 consecutive days (Fourie and Hattingh 1979). No effect on erythrocyte count was seen in chickens 

administered daily subcutaneous injections of DDT (unspecified isomeric composition) for up to 81 days 

for a cumulative dose of 2–3 g/chicken (Burlington and Lindeman 1952). Erythrocyte count was reduced 

compared to controls by an average of 17.8% in White Leghorn cockerels injected with DDT (unspecified 

isomer) subcutaneously for between 60 and 89 days at dose levels that increased from 15 to 

300 mg/kg/day during the exposure period (Burlington and Lindeman 1950). 

Hematocrit was not significantly affected in double-crested cormorants fed diets containing up to 25 ppm 

of total DDT/DDE/DDD for 9 weeks (Greichus and Hannon 1973). Hematocrit was significantly 

decreased in crowned guinea fowl (Numida meleagris) after an assumed daily gavage dose of 75 mg/kg 

body weight (exposure units were ambiguously reported) of technical-grade DDT on 5 consecutive days 

(Fourie and Hattingh 1979). A 12-week dietary exposure to up to 100 ppm p,p’-DDE did not 

significantly affect hematocrit in Japanese quail (Dieter 1974). Hematocrit was unaffected compared to 

controls in White Leghorn cockerels injected with DDT (unspecified isomer) subcutaneously for between 

60 and 89 days at dose levels that increased from 15 to 300 mg/kg/day during the exposure period 

(Burlington and Lindeman 1950). Hematocrit was significantly increased in bobwhite quail (Colinus 

virginianus) fed diets containing from 10 (lowest level tested) to 150 ppm technical-grade DDT for 

several months (Lustick et al. 1972; Peterle et al. 1973). The increased hematocrit in bobwhite quail was 

hypothesized to be a compensatory increase in relative red blood cell volume related to increased 

metabolic oxygen demand caused by a DDT-induced increase in thyroxin secretion by the thyroid; the 

study confirmed increased uptake of I131 by the thyroid in DDT-exposed birds, indicating increased 

thyroid activity (Lustick et al. 1972; Peterle et al. 1973). Significantly decreased hematocrit was seen in 

ringed turtle doves (Streptopelia risoria) fed a diet containing 200 ppm (but not at #20 ppm) of DDE 

(unspecified isomeric composition) for 8 weeks (Heinz et al. 1980). The decrease in hematocrit observed 

in ringed turtle doves occurred only at a dietary exposure level of DDE greater than those used to 

administer DDT in the bobwhite quail study. 

Musculoskeletal Effects. No experimental studies were located regarding musculoskeletal effects in 

wild mammals, reptiles, or amphibians from exposure to DDT/DDE/DDD. 

Birds. No gross skeletal muscle lesions were observed in 6-week-old pheasant chicks (Phasianus 

colchicus) fed 100 ppm technical-grade DDT for up to 101 days or 500 ppm for up to 23 days (Azevedo 

et al. 1965). 

Calcium uptake in bone (which is mediated by estradiol) was significantly reduced (compared to controls) 

at 8 days postmating in ringed turtle doves (Streptopelia risoria) fed diets containing 10 ppm p,p’-DDT 
 

 

DDT, DDE, and DDD D-8 

APPENDIX D 

for 3 weeks prior to mating, but not in birds allowed to complete their clutch (Peakall 1970). “Extremely 

soft skulls” were observed in homing pigeons (Columba livia) that died within 35 days of oral exposure 

to 72 mg/kg/day of p,p’-DDE in capsules (Jefferies and French 1972). 

Hepatic Effects. No experimental studies were located regarding hepatic effects in reptiles or 

amphibians from exposure to DDT/DDE/DDD. 

DDT and DDE consistently induced hepatic microsomal enzyme activity in six species of birds after oral 

exposures; mixed results were obtained in five studies in Japanese quail. Five of the studies showing 

enzyme induction indicated that the increased hepatic enzyme activity significantly accelerated the 

breakdown of steroid hormones, including progesterone and estrogen in females and testosterone in 

males, potentially affecting hormone balance. No effect on liver weights was observed in birds orally 

administered DDT alone, but significantly increased liver weights were seen in three species fed DDE 

alone, probably due to hepatic enzymatic induction. Decreased liver weights were observed in two 

piscivorous species after oral exposure to a mixture of DDT/DDE/DDD; the authors did not speculate on 

the cause of the decrease in liver weights. Increased activity of liver enzymes in the blood was 

consistently observed in three species, indicating liver cell damage (which causes the enzymes to leak out 

from the cells), and microscopic lesions of the liver were observed in two species. With few exceptions, 

liver effects are a consistent, albeit nonspecific, indicator of DDT/DDE/DDD toxicity in birds, and 

induction of liver-mediated metabolism may affect steroid hormone balance. 

Mammals. Short-tailed shrews showed liver weights that were not significantly different from controls 

after consuming an earthworm diet containing an average of 16.6 ppm DDT for 3 weeks (unspecified 

isomeric composition) (Braham and Neal 1974). 

Birds. The hepatic microsomal mixed function oxidase system was significantly induced in puffins 

administered approximately 6 mg DDE (unspecified isomer)/bird for 16–21 days before sacrifice by 

decapitation (Bend et al. 1977). Hepatic microsomal ethylmorphine N-demethylase activity was 

significantly increased in kestrels (Falco sparverious) fed diets containing 10 ppm DDE (unspecified 

isomeric composition) for 5 years (Gillett et al. 1970). No effect on hepatic microsomal protein, but 

significantly increased hepatic P-450 activity, were seen in domestic chicken hens orally administered 

40 mg/hen of technical-grade DDT for 5 days (Chen et al. 1994). 

Japanese quail (Coturnix coturnix japonica) showed increased pentobarbital-induced sleeping time when 

fed each of o,p’- and p,p’-isomers of DDT, DDE, and DDD separately at 100 ppm in the diet for up to 

14 days (Bitman et al. 1971), indirectly suggesting that the DDT compounds inhibited the hepatic 

microsomal enzymes that metabolize pentobarbital. Gillett et al. (1970) reported significantly decreased 

hepatic microsomal epoxidase activity in Japanese quail fed up to 100 ppm DDT (unspecified isomeric 

composition) for 27 weeks. Another study provided some evidence that 28-day dietary exposure to 

50 ppm of p,p’-DDT, DDE, and DDD altered (causing both increases and decreases) hepatic enzymatic 

activity in Japanese quail (Bunyan et al. 1970), while yet another study indicated increased hepatic 

microsomal P-450 levels without a significant change in microsomal protein content in Japanese quail fed 

diets of up to 100 ppm p,p’-DDE for 21 days (Bunyan and Page 1973; Bunyan et al. 1972). Liver 

oxidative enzyme activity was not affected in female Japanese quail fed diets containing up to 30 ppm 

technical-grade DDT for 107 days (Kenney et al. 1972). 

Metabolism of 17β-estradiol was significantly increased in the hepatic microsomal fraction obtained from 

domestic fowl hens fed diets containing 300–1,200 ppm technical-grade DDT for 7 to 21 days (Britton 

1975). Hepatic microsomal extracts from female bobwhite quail (Colinus virginianus) fed a diet 

containing 5 ppm technical-grade DDT for 30–70 days produced a significantly greater conversion of 14C-labelled progesterone to metabolites in vitro than microsomal extracts from untreated female quail; 
DDT, DDE, and DDD D-9 

APPENDIX D 

microsomal enzyme extracts from DDT-treated male quail did not produce a significant increase in mean 

testosterone conversion in vitro (Lustick et al. 1972; Peakall 1967; Peterle et al. 1973). Metabolism of 

both testosterone and progesterone was significantly increased over control levels in liver microsomal 

fractions obtained from male and female White King pigeons, respectively, after a 1-week dietary 

exposure to 10 ppm p,p’-DDT (Peakall 1967). Similarly, estradiol metabolism was significantly 

increased in liver microsomal fractions obtained from ringed turtle doves (Streptopelia risoria) fed 

10 ppm of technical-grade DDT in the diet for 3 weeks (Peakall 1969). Hepatic enzyme metabolism of 

estradiol was significantly increased at 8 days postmating in ringed turtle doves fed diets containing 

10 ppm p,p’-DDT for 3 weeks prior to mating, but not in birds allowed to complete their clutch (Peakall 

1969, 1970). 

Relative liver weights were significantly depressed in white pelicans (Pelecanus erythrorhynchos) 

administered a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDD, and 15 mg p,p’-DDE for 10 weeks 

(Greichus et al. 1975). Liver weights were significantly decreased in double-crested cormorants 

(Phalacrocorax auritus) fed diets containing $5 ppm of total DDT/DDE/DDD for 9 weeks (Greichus and 

Hannon 1973). Liver weights and yield of microsomal protein were not significantly affected in puffins 

administered approximately 6 mg DDE (unspecified isomer)/bird for 16–21 days before sacrifice by 

decapitation (Bend et al. 1977). In Japanese quail, relative liver weights were significantly increased after 

12 weeks on a diet containing $25 ppm p,p’-DDE (Dieter 1974). In other studies with Japanese quail, 

liver weights were not significantly affected by nine oral administrations of 10 mg o,p’-DDT over 

3 weeks (Cooke 1970b) or by dietary exposure to up to 100 ppm of the p,p’- isomers of DDT, DDE, and 

DDD for 28 days (Bunyan et al. 1970, 1972). Liver weight and microsomal protein were not significantly 

different from controls in Japanese quail fed diets containing up to 100 ppm DDT (unspecified isomeric 

composition) for 27 weeks (Gillett et al. 1970). Significantly increased relative liver weights were seen 

compared to controls in bobwhite quail fed diets containing 50–150 ppm technical-grade DDT for 

30–70 days (Lustick et al. 1972). No consistent effect on chicken liver weights was seen in birds fed 

between 400 and 800 ppm p,p’-DDT in the feed for 2–6 weeks (Glick 1974). No significant effects on 

liver weights or hepatic microsomal protein levels were seen in domestic fowl hens orally administered 

40 mg/hen of technical-grade DDT for 5 days (Chen et al. 1994). No effect on liver weights was seen in 

redstarts (Phoenicurus phoenicurus) administered a cumulative oral dose of 126 μg p,p’-DDT 

administered in equal daily doses over a 12-day period (Karlsson et al. 1974). Significantly increased 

liver weights were seen in ringed turtle doves fed diets containing $20 ppm DDE (unspecified isomeric 

composition) for 8 weeks (Heinz et al. 1980). A significant dose-related increase in liver weight was seen 

in homing pigeons (Columba livia) administered daily doses of p,p’-DDE for 8 weeks (Jefferies and 

French 1972) or p,p’-DDT for 42 days (Jefferies and French 1969) by capsule at between 18 and 

72 mg/kg/day. 

Plasma liver enzyme activities (creatinine kinase, aspartate aminotransferase, cholinesterase, fructosediphosphate aldose, and lactate dehydrogenase) were significantly increased in Japanese quail, suggesting 

liver damage at exposure levels of $5 ppm p,p’-DDE in the diet for 12 weeks (Dieter 1974). Crowned 

guinea fowl (Numida meleagris) showed significantly increased hepatic enzyme activities in blood 

collected after 5 days of oral exposure to approximately 75 mg/kg technical-grade DDT (dose was 

ambiguously reported) (Fourie and Hattingh 1979). Significantly increased plasma aspartate 

aminotransferase was seen in ringed turtle doves fed diets containing 200 ppm (but not #20 ppm) of DDE 

(unspecified isomeric composition) for 8 weeks (Heinz et al. 1980). 

Centrilobular liver degeneration was seen in 6-day-old chickens within 2 days of a single intraperitoneal 

injection with 0.25 mmol o,p’-DDD, and increased in severity over the next several days (Jönsson et al. 

1994). No gross liver lesions were observed in 6-week-old pheasant chicks (Phasianus colchicus) fed 

100 ppm technical-grade DDT for up to 101 days or 500 ppm for up to 23 days (Azevedo et al. 1965). 

Dose-related “marked” liver hypertrophy was observed in homing pigeons orally administered capsules 
 

DDT, DDE, and DDD D-10 

APPENDIX D 

containing p,p’-DDT at doses ranging from 3 to 54 mg/kg/day over a 17-week period (Jefferies and 

French 1971). 

Daily oral exposure of homing pigeons to p,p’-DDT in capsules ($36 mg/kg/day) for up to 8 weeks 

resulted in a significant decrease in hepatic vitamin A storage (Jefferies and French 1971) (vitamin A is a 

fat-soluble compound that is essential for normal night vision, health of epithelial cells, and normal 

growth of bones and teeth, and is stored in the liver). The authors suggested that the reduced liver 

vitamin A was related to DDT-induced hypothyroidism, as evidenced also by reduced body temperature 

and oxygen consumption. In contrast, a lower dose of DDT (3 mg/kg/day), which caused 

hyperthyroidism, significantly increased hepatic vitamin A storage. Liver vitamin A was significantly 

increased in white pelicans administered a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDD, and 15 mg 

p,p’-DDE injected into the first fish fed to the birds each day over a 10-week period (Greichus et al. 

1975). Liver carotene (a vitamin A precursor converted to vitamin A in the liver) was not significantly 

different from controls (Greichus et al. 1975). Liver vitamin A content was significantly decreased in 

double-crested cormorants fed diets containing $5 ppm of total DDT/DDE/DDD for 9 weeks (Greichus 

and Hannon 1973). In this case, liver carotene levels were also not significantly affected. Thyroid status 

was not evaluated in the studies of Greichus and coworkers. 

Liver glycogen was marginally (not significantly) decreased and liver lipid levels were significantly 

increased in bobwhite quail fed diets containing 100 ppm technical-grade DDT for 10 weeks (Haynes 

1972). 

Serum lipid was not significantly affected in Japanese quail either by four consecutive daily intramuscular 

injections with 5 mg o,p’-DDT or by nine oral administrations of 10 mg/bird of o,p’-DDT over 3 weeks 

(Cooke 1970b). Crowned guinea fowl showed significantly increased serum cholesterol compared to preexposure levels after a 5-day oral exposure to approximately 75 mg/kg/day technical-grade DDT (Fourie 

and Hattingh 1979). 

Significantly depressed serum protein was observed in white pelicans (Pelecanus erythrorhynchos) 

administered daily oral doses of a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDD, and 15 mg 

p,p’-DDE for 10 weeks (Greichus et al. 1975). Blood total protein was not significantly affected in 

double-crested cormorants fed diets containing up to 25 ppm of total DDT/DDE/DDD for 9 weeks 

(Greichus and Hannon 1973). Plasma protein was not significantly affected in crowned guinea fowl after 

an assumed daily gavage dose of 75 mg/kg body weight (exposure units were ambiguously reported) of 

technical-grade DDT on 5 consecutive days (Fourie and Hattingh 1979). No effects on plasma protein 

level was seen in cockerels administered daily subcutaneous injections of DDT (unspecified isomeric 

composition) for up to 81 days starting on the 8th day posthatch; daily dose was unreported, but 

cumulative dose was 2–3 g/cockerel (Burlington and Lindeman 1952). No effects on fibrinogen level or 

prothrombin time were seen in blood of cockerels administered daily subcutaneous injections of DDT 

(unspecified isomeric composition) for up to 81 days for a cumulative dose of 2–3 g/cockerel (Burlington 

and Lindeman 1952). Prothrombin is formed and stored in the liver; in the presence of thromboplastin 

and calcium, it is converted into thrombin, which, in turn, converts fibrinogen to fibrin for blood 

coagulation. 

Renal Effects. No experimental studies were located regarding renal effects in wild mammals, 

reptiles, or amphibians from exposure to DDT/DDE/DDD. 

Birds. No renal effects were observed in experimental studies in birds. Kidney weights were not 

significantly affected in white pelicans (Pelecanus erythrorhynchos) exposed by daily oral 

administrations of a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDD, and 15 mg p,p’-DDE for 

10 weeks (Greichus et al. 1975). No effect on kidney weights was seen in redstarts (Phoenicurus 
 

DDT, DDE, and DDD D-11 

APPENDIX D 

phoenicurus) administered a cumulative oral dose of 126 μg p,p’-DDT administered in equal daily doses 

over a 12-day period (Karlsson et al. 1974). 

No gross kidney lesions were observed in 6-week-old pheasant chicks (Phasianus colchicus) fed 100 ppm 

technical-grade DDT for up to 101 days, or 500 ppm for up to 23 days (Azevedo et al. 1965). 

Endocrine Effects. No experimental studies were located regarding endocrine effects in wild 

mammals from exposure to DDT/DDE/DDD. 

Thyroid weights in birds were consistently elevated in six studies of dietary or capsular DDT exposures. 

Thyroid function in birds appeared to be biphasic after dietary exposure, showing decreased I131 uptake or 

no change in two studies at #25 ppm DDT, and increased I131 uptake in four studies at $100 ppm. The 

opposite pattern appears to have occurred in birds administered bolus doses of DDT in capsules, with 

hyperthyroidism occurring at relatively low bolus doses and hypothyroidism at high doses. The incidence 

of lesions in thyroid tissue increased with increasing dosage; thus, the hypothyroidism observed at high 

bolus oral doses may be due to degenerative changes in the thyroid. Similar degeneration may not occur 

at high dietary concentrations because dietary exposure occurs relatively gradually over time and may not 

result in blood DDT levels high enough to damage thyroid tissue. Gross adrenal morphology was 

unaffected by oral exposure to DDT, but histological examination showed changes in the corticomedullary ratio in birds fed DDT in the diet and degenerative changes in adrenals of birds exposed 

parenterally. Adrenal function has not been directly evaluated in wildlife species. Estrogenic changes 

have been reported in birds, reptiles, and amphibians after parenteral exposure to o,p’-DDT, and reduced 

phallus size and altered sex ratio at hatching have been reported in reptiles exposed in ovo to DDE. See 

Section 3.6.2, Mechanisms of Toxicity, for further discussion of estrogenicity of DDT/DDE/DDD in 

laboratory animals and in vitro. 

Reptiles. Significantly elevated plasma vitellogenin, suggesting estrogenic activity, was induced in adult 

male red-eared turtles (Trachemys scripta) intraperitoneally injected with o,p’-DDT at either 1 or 

250 μg/g/day for 7 days. The amounts of vitellogenin produced were significantly lower than those seen 

in turtles injected with 1 μg/g/day 17β-estradiol (Palmer and Palmer 1995). Vitellogenin is a protein 

precursor of egg proteins that is produced in the liver of oviparous and ovoviviparous species and is 

found in measurable quantities in blood; production is stimulated by estrogens in the blood that are 

produced in the ovary; thus, vitellogenin is normally absent in the blood of males (Hadley 1984). 

Amphibians. Plasma vitellogenin was significantly increased in adult male African clawed frogs 

(Xenopus laevis) intraperitoneally injected with o,p’-DDT at either 1 or 250 μg/g/day for 7 days (Palmer 

and Palmer 1995), suggesting estrogenic activity. However, the amounts of plasma vitellogenin induced 

were significantly lower than those seen in frogs injected with 1 μg/g/day 17β-estradiol (Palmer and 

Palmer 1995). 

Birds. Thyroid weights were significantly increased in an exposure-related manner in juvenile male 

Mallards (Anas platyrhyncus) fed technical-grade DDT at 2.5–250 ppm in the diet for 30 days (Peterle et 

al. 1973). Thyroid glands were significantly enlarged in bobwhite quail (Colinus virginianus) fed 

technical-grade DDT in the diet at 500 ppm (but not at #50 ppm) for at least 3 months (Hurst et al. 1974). 

Bobwhite quail fed on diets containing 100 ppm (but not 10 ppm) technical-grade DDT for at least 

13 weeks showed significantly increased thyroid weights relative to controls (Lustick et al. 1972). 

Significantly increased relative thyroid weight was seen in bobwhite quail held for 1 week at -18 EC after 

being fed 100 ppm technical-grade DDT for several months, but not in controls or birds fed 10 ppm 

(Peterle et al. 1973). A significant treatment-related increase in relative thyroid weight was seen in 

homing pigeons (Columba livia) fed p,p’-DDT or p,p’-DDE in capsules every other day over a 42- to 

56-day period at 18–72 mg/kg/day (Jefferies and French 1969, 1972). Another experiment by Jefferies 
DDT, DDE, and DDD D-12 

APPENDIX D 

and French (1971) confirmed the finding of increased thyroid weight in homing pigeons in response to 

daily oral exposure to p,p’-DDT (3–36 mg/kg/day) for up to 8 weeks. 

Uptake of I131 into the thyroid was decreased (not significantly) in juvenile male mallards (Anas 

platyrhyncus) fed 2.5 or 25 ppm technical-grade DDT in the diet for 30 days, and a significantly increased 

uptake of I131 into the thyroid was seen in birds at 250 ppm DDT (Peterle et al. 1973). Thyroid uptake of 

I

131 was significantly increased in bobwhite quail fed 500 ppm technical-grade DDT in the diet for 

1–3 months, but not after 4 months of treatment (Hurst et al. 1974). Bobwhite quail fed diets containing 

100 ppm (but not 10 ppm) technical-grade DDT for at least 13 weeks showed significantly increased 

thyroid activity as measured by I131 uptake relative to controls (Lustick et al. 1972). In another 

experiment, significantly increased I131 uptake into the thyroid was seen in bobwhite quail held for 1 week 

at -18 EC after being fed 100 ppm technical-grade DDT for several months, but not in controls or birds 

fed 10 ppm (Peterle et al. 1973). Jefferies and French (1971) found that body temperature and oxygen 

consumption were decreased in a dose-related manner in homing pigeons (Columba livia) in response to 

daily oral exposure to p,p’-DDT in capsules (3–36 mg/kg/day) for up to 8 weeks, and that significantly 

increased hepatic vitamin A storage (suggesting hyperthyroidism) was seen in the low-dose group, while 

significantly decreased hepatic vitamin A storage (suggesting hypothyroidism) was seen in the high-dose 

group; the authors noted that hypothyroidism has been associated with eggshell thinning in birds. 

The incidence of hyperplasia of the thyroid follicular epithelium was increased in treated birds, and 

follicular colloid was decreased, in homing pigeons fed 18–72 mg/kg/day of p,p’-DDT or p,p’-DDE in 

capsules every other day over a 42- to 56-day period (Jefferies and French 1969, 1972). Another 

experiment by Jefferies and French (1971) confirmed the finding of hyperplasia and decreased follicular 

colloid in homing pigeons in response to daily oral exposure to p,p’-DDT (3–36 mg/kg/day) for up to 

8 weeks. 

Adrenal weight was unaffected in juvenile male Mallards fed technical-grade DDT at 2.5–250 ppm in the 

diet for 30 days (Peterle et al. 1973). Adrenal weights were not significantly affected in domestic 

chickens by a 6-week dietary exposure to p,p’-DDT at 400 or 800 ppm in the feed (Glick 1974). Adrenal 

gland weight was not significantly affected in bobwhite quail fed technical-grade DDT in the diet at 

500 ppm (but not at #50 ppm) for at least 3 months (Hurst et al. 1974). Adrenal weights were not 

significantly affected in bobwhite quail fed diets containing up to 150 ppm technical-grade DDT for up to 

242 days (Lehman et al. 1974). No significant effect was seen on adrenal weight in bobwhite quail at 

exposure levels up to 150 ppm of technical-grade DDT in the diet for several months (Peterle et al. 1973). 

A significant dose-related increase in adrenal weights was observed in homing pigeons fed p,p’-DDT in 

the diet at estimated dose rates of up to 54 mg/kg/day for 8 weeks (Jefferies et al. 1971) and in homing 

pigeons administered p,p’-DDE in capsules at 18–72 mg/kg/day for 8 weeks (Jefferies and French 1972). 

Adrenal cortico-medullary ratio was unaffected in juvenile male mallards fed technical-grade DDT at 

2.5–250 ppm in the diet for 30 days (Peterle et al. 1973). An exposure-related increase in adrenal 

cortical-medullary ratio was observed in bobwhite quail at exposure levels up to 150 ppm of technicalgrade DDT in the diet for several months (Peterle et al. 1973); the authors speculated that this alteration 

of the cortico-medullary ratio may induce a reduced production and release of adrenaline compared to 

levels of corticosteroids. In bobwhite quail fed diets containing up to 150 ppm technical-grade DDT for 

up to 242 days, there was a significant exposure-related increase in the cortico-medullary ratio of adrenal 

gland (Lehman et al. 1974). Focal vacuolation and pycnotic nuclei were seen in adrenal interrenal cells of 

6-day-old chickens within 1 week of a single intraperitoneal injection with 0.25 mmol o,p’-DDD (Jönsson 

et al. 1994). 

Plasma corticosterone levels were unaffected in juvenile male mallards (Anas platyrhyncus) fed technicalgrade DDT at 2.5–250 ppm in the diet for 30 days (Peterle et al. 1973). 
 

 

 

 

DDT, DDE, and DDD D-13 

APPENDIX D 

No studies were located that directly evaluated for effects to the pancreas; however, a few studies 

measured blood sugar levels. Crowned guinea fowl showed significantly decreased serum blood sugar 

compared to pre-exposure levels after a 5-day oral exposure to approximately 75 mg/kg/day technicalgrade DDT (Fourie and Hattingh 1979). No effect on blood sugar level was seen in cockerels 

administered daily subcutaneous injections of DDT (unspecified isomeric composition) for up to 81 days 

starting on the 8th day posthatch; daily dose was unreported, but cumulative dose was 2–3 g/chicken 

(Burlington and Lindeman 1952). 

Dermal Effects. No experimental studies were located regarding dermal effects in wildlife from 

exposures to DDT/DDE/DDD. 

Ocular Effects. No experimental studies were located regarding ocular effects in wildlife from 

exposures to DDT/DDE/DDD. 

Body Weight Effects. No experimental studies were located regarding body weight effects in 

reptiles from exposure to DDT/DDE/DDD. 

Body weight loss or reduced body weight gain were observed in mammalian, avian, and amphibian 

species in response to DDT/DDE/DDD exposures. Birds were most intensively studied, and among birds, 

raptors, passerines, and nonpasserine ground birds were more sensitive to DDT/DDE/DDD exposures 

with respect to body weight changes than gallinaceous birds. Body weight loss has been associated with 

hyperthyroidism, and weight gain has been associated with hypothyroidism (Jefferies 1969); see further 

discussion of DDT-induced thyroid effects under Endocrine, above. No clear patterns in body weight 

changes were evident regarding relative sensitivities based on particular DDT compound or route of 

exposure. 

Mammals. Free-tailed bats (Tadarida brasiliensis) fed diets containing 107 ppm DDE (unspecified 

isomeric composition) showed significantly decreased body weight gain compared to controls after a 

40-day exposure period (Clark and Kroll 1977). Body weights were generally not significantly affected 

in short-tailed shrews (Blarina brevicauda) fed DDT (unspecified isomeric composition) in the diet for 7, 

14, or 17 days (Blus 1978); exposure levels were not reported, but were sufficiently high to calculate LC50 

levels of $651 mg/kg diet. Another study in short-tailed shrews also reported body weights comparable 

to controls after consuming an earthworm diet containing an average of 16.6 ppm DDT for 3 weeks 

(unspecified isomeric composition) (Braham and Neal 1974). 

Amphibians. Body weight loss was significantly exposure-related in tadpoles of the common frog (Rana 

temporaria) exposed for 1 hour to 0.01–10 ppm p,p’-DDT in the water column (Cooke 1970a). 

Birds. Body weights were not significantly affected in white pelicans (Pelecanus erythrorhynchos) by 

daily oral administrations of a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDE, and 15 mg p,p’-DDD 

for 10 weeks (Greichus et al. 1975); doses were injected into the first fish fed to the birds each day. 

American kestrels (Falco sparverius) that died after exposure to 2.8 ppm p,p’-DDE in the diet for 

14–18 months lost between 30 and 35% of their body weight; kestrels that survived the chronic dietary 

exposure lost an average of 16% of their body weight (Porter and Wiemeyer 1972). Bald eagles 

(Haliaetus leucocephalus) experienced weight loss of up to 49% after 10–16 weeks on diets containing 

10–4,000 ppm technical-grade DDT (Chura and Stewart 1967). 

Body weights of mallard ducklings (Anas platyrhyncus) were not significantly affected by a 10-day oral 

exposure to p,p’-DDT at up to 2,000 ppm in the diet (Friend and Trainer 1974a), but in apparent 
 

DDT, DDE, and DDD D-14 

APPENDIX D 

contradiction, body weight gain was “suppressed” in Mallard ducklings fed DDT (unspecified 

composition) at 500–900 ppm in the diet for 10 days (Friend and Trainer 1974b). 

Body weights of adult female Japanese quail (Coturnix coturnix japonica), but not males, were 

significantly decreased after daily intramuscular injections of 5 mg o,p’-DDT for 4 days (Cooke 1970b); 

male body weights were also unaffected by nine oral administrations of 10 mg DDT over a 3-week 

period. In another study in Japanese quail, body weights were not significantly affected by dietary 

exposure to up to 100 ppm p,p’-DDE and p,p’-DDT for 21 days (Bunyan and Page 1973; Bunyan et al. 

1972). Growth of Japanese quail as measured by body weight was comparable to control group growth 

throughout a 3-generation reproductive toxicity assay in which birds were administered up to 50 ppm 

DDT (unspecified isomeric composition) in the diet (Shellenberger 1978). 

Bobwhite quail (Colinus virginianus) lost between 2 and 24% of body weight after 5 days of feeding on 

diets containing $400 ppm technical-grade DDT (Hill et al. 1971); wild quail lost appreciably more 

weight than farm-raised birds at the same exposure levels. No effect on hen body weight was seen in 

laying bobwhite quail orally administered up to 20 mg/bird of DDT (unspecified isomeric composition) 

by capsule every other day during a 4-week exposure period (Wilson et al. 1973). 

Chicken body weights were not significantly affected by 6-week dietary exposure to 400 or 800 ppm 

(Glick 1974) or by #81 days of daily subcutaneous injections of DDT (isomeric composition 

unspecified) for a cumulative dose of 2–3 grams (Burlington and Lindeman 1952). However, a more 

recent study found a slight but significant loss of body weight, accompanied by a significant reduction in 

food consumption, in hens orally administered 40 mg/hen of technical-grade DDT for 5 days (Chen et al. 

1994). No effect on body weight was observed in chicken hens fed p,p’-DDT at #200 ppm in the diet for 

12 weeks (Davison and Sell 1972). 

Pheasant (Phasianus colchicus) adult females fed diets containing 400 ppm DDT (unspecified isomeric 

composition) for at least 10 weeks lost weight, but females fed 100 ppm DDT gained weight (Genelly and 

Rudd 1956). 

Nonsignificant body weight loss was seen in ringed turtle doves (Streptopelia risoria) fed p,p’-DDE at 

10 or 50 ppm in the diet for 63 days (Haegele and Hudson 1977). No significant change in body weight 

was seen in ringed turtle doves fed diets containing up to 200 ppm DDE (unspecified isomeric 

composition) for 8 weeks (Heinz et al. 1980). Homing pigeons (Columba livia) that died after several 

weeks of oral exposure to p,p’-DDE at $36 mg/kg/day by capsule showed an average weight loss of 

approximately 33% of their original weights, and birds that survived 8 weeks of exposure had lost an 

average of 2% of body weight; controls showed no change in body weight (Jefferies and French 1972). 

House sparrows (Passer domesticus) fed diets containing 320–700 ppm technical-grade DDT for 5 days 

lost between 10 and 12% of their body weight on the average (Hill et al. 1971). Body fat content and 

body weight gain were significantly decreased in an exposure-related manner in white-throated sparrows 

(Zonotrichia albicollis) fed diets containing 5–25 ppm technical-grade DDT for 31 days prior to the 

beginning of the migratory season (Mahoney 1975); birds fed p,p’-DDE at 5 or 25 ppm in the diet for 

6 weeks showed no significant alteration in body weight gain or body fat indices. No effect on body 

weight was seen in redstarts (Phoenicurus phoenicurus) administered a cumulative oral dose of 126 μg 

p,p’-DDT administered in equal daily doses over a 12-day period (Karlsson et al. 1974). 
 

 

DDT, DDE, and DDD D-15 

APPENDIX D 

Metabolic Effects. No experimental studies were located regarding metabolic effects in reptiles or 

amphibians from exposure to DDT/DDE/DDD. Metabolic rate was increased in mammals and birds 

receiving relatively low oral bolus (capsular) doses of DDT, and decreased at relatively high bolus doses. 

Blood calcium was consistently unaffected by DDT/DDE/DDD treatment in birds after dietary or 

parenteral exposures; no effect on blood calcium was seen in seven studies using six bird species after 

acute or intermediate durations. Altered blood calcium was seen only after relatively high, intermediate 

duration oral bolus (capsular) exposure in one bird species. Two studies in birds reported increased blood 

pCO2 after acute oral exposures, although the significance of this is uncertain. 

Mammals. Metabolic rate as measured by O2 consumption rate was significantly increased compared to 

controls in short-tailed shrews (Blarina brevicauda) fed an earthworm diet containing an average of 

16.6 ppm DDT (unspecified isomeric concentration) for 1 week, but not after 2 or 3 weeks of exposure; 

treated shrews that were starved after exposure also showed an increase in metabolic rate compared to a 

control group that showed reduced metabolic rate in response to starvation (Braham and Neal 1974). 

Birds. Bobwhite quail (Colinus virginianus) fed on diets containing 10–150 ppm technical-grade DDT 

for at least 6 weeks showed increased (not statistically significant) metabolic rate compared to controls as 

measured by oxygen consumption at all ambient temperature levels tested between 5 and 40 EC, except at 

thermal neutrality at 30 EC, at which oxygen consumption was comparable to controls (Lustick et al. 

1972; Peterle et al. 1973). In low-temperature stressed quail, shivering occurred in many birds but was 

most pronounced in birds that died after being fed DDT, while at the highest temperatures tested, several 

DDT-treated birds collapsed while none of the controls collapsed; the authors surmised that dietary DDT 

may have reduced the upper critical temperature of the thermoneutral zone from 40 to 35 EC (Lustick et 

al. 1972; Peterle et al. 1973). It is possible that the increased severity of “shivering” observed in stressed 

quail that eventually died may actually have been DDT-induced neurotoxicity (i.e., tremors), since similar 

responses that have been attributed to neurotoxicity have been observed in several other studies of 

energy-stressed birds (Appendix D.4, Neurological/Behavioral effects). After 11 weeks consuming 

p,p’-DDT in capsules at 3 mg/kg/day, homing pigeons (Columba livia) showed significantly increased 

oxygen consumption (with a significant increasing time-trend), but at 36 mg/kg/day for 11 weeks, there 

was a significant decrease in oxygen consumption (with a significant decreasing time trend) (Jefferies and 

French 1971). Jefferies and French (1971) found that body temperature significantly decreased over time 

in homing pigeons in response to daily oral exposure to p,p’-DDT in capsules at 36 mg/kg/day or higher 

for up to 8 weeks. 

Blood calcium was not significantly affected in double-crested cormorants (Phalacrocorax auritus) fed 

diets containing #25 ppm total DDT/DDE/DDD for 9 weeks (Greichus and Hannon 1973). Serum 

calcium was not significantly affected in white pelicans (Pelecanus erythrorhynchos) exposed by daily 

oral administrations of a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDE, and 15 mg p,p’-DDD for 

10 weeks (Greichus et al. 1975); doses were injected into the first fish fed to the birds each day. Blood 

calcium levels during egg laying was not significantly affected in Pekin ducks (domesticated Mallards; 

Anas platyrhyncus) fed diets containing 250 ppm DDE (unspecified isomeric composition) for 10 days; 

no other exposure levels were tested (Peakall et al. 1975). Decreased serum calcium was seen in laying 

bobwhite quail (Colinus virginianus) administered 20 mg/bird of DDT (unspecified isomeric 

composition) by capsule every other day during a 4-week exposure period, but not at 10 mg DDT/bird 

(Wilson et al. 1973). Serum calcium was not significantly affected in Japanese quail (Coturnix coturnix 

japonica) exposed by four consecutive daily intramuscular injections with 5 mg o,p’-DDT or by nine oral 

administrations of 10 mg/bird of o,p’-DDT over 3 weeks (Cooke 1970b). No effect on plasma calcium 

levels was seen in cockerels administered daily subcutaneous injections of DDT (unspecified isomeric 

composition) for up to 81 days starting on the 8th day posthatch; daily dose was unreported, but 

cumulative dose was 2–3 g/cockerel (Burlington and Lindeman 1952). No significant effect on plasma 

calcium was seen in hens fed 40 mg/hen of technical DDT for 5 days (Chen et al. 1994). Blood calcium 
 

DDT, DDE, and DDD D-16 

APPENDIX D 

during egg laying was not significantly affected in ringed turtle doves (Streptopelia risoria) by a 3-week 

dietary exposure to 100 ppm DDE (unspecified isomeric composition); no other exposure levels were 

tested (Peakall et al. 1975). 

Blood pCO2 (partial pressure of CO2 in blood in units of mm Hg) was significantly increased in adult 

male Japanese quail (Coturnix coturnix japonica) fed diets containing $10 ppm technical-grade DDT (but 

not at 3 ppm) for 2 weeks, but at 6 weeks, there was a significant increase only at 10 ppm and not at 

higher exposure levels; there was no consistent effect in females (Kenney et al. 1972). Significantly 

increased blood pCO2, but no significant effect on pO2 and blood pH, was also seen in crowned guinea 

fowl (Numida meleagris) compared to pre-exposure levels after a 5-day oral gavage exposure to 

approximately 75 mg/kg/day technical-grade DDT (Fourie and Hattingh 1979). 

Significantly depressed serum potassium levels were observed in white pelicans (Pelecanus 

erythrorhynchos) administered daily oral doses of a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDD, 

and 15 mg p,p’-DDE for 10 weeks (Greichus et al. 1975); no significant effects were seen in serum 

inorganic phosphorus, uric acid, magnesium, or sodium. Blood sodium, urea nitrogen, and phosphorus 

were not significantly affected in double-crested cormorants fed diets containing up to 25 ppm of total 

DDT/DDE/DDD for 9 weeks (Greichus and Hannon 1973). Plasma osmolality and plasma sodium were 

significantly increased in white Pekin ducks fed 100 ppm DDE in the diet for a total of 15 days; plasma 

potassium was not significantly affected (Miller et al. 1976). Crowned guinea fowl showed significantly 

decreased serum potassium and urea nitrogen compared to pre-exposure levels after a 5-day gavage 

exposure to approximately 75 mg/kg/day technical-grade DDT (Fourie and Hattingh 1979); no significant 

effect was seen on blood sodium or osmolarity. No effect on plasma phosphorous level was seen in 

cockerels administered daily subcutaneous injections of DDT (unspecified isomeric composition) for up 

to 81 days starting on the 8th day posthatch; daily dose was unreported, but cumulative dose was 

2–3 g/chicken (Burlington and Lindeman 1952). Significantly decreased plasma sodium was seen in 

ringed turtle doves fed diets containing 200 ppm (but not at #20 ppm) of DDE (unspecified isomeric 

composition) for 8 weeks; no significant effect on plasma potassium was seen (Heinz et al. 1980). 

Other Systemic Effects. No experimental studies were located regarding other systemic effects in 

wild mammals, reptiles, or amphibians from exposure to DDT/DDE/DDD. 

Birds. Nasal gland (“salt gland”) secretion rate was not significantly affected in white Pekin ducks 

(domesticated Mallards; Anas platyrhyncus) fed 50 ppm DDE (unspecified isomeric composition) in the 

diet for a total of 7 days (Miller et al. 1976). In Mallards maintained on fresh water (but not in those 

maintained on saltwater), nasal gland secretion rate was significantly depressed when challenged with 

intravenous saltwater injections after up to 9 days of feeding on diets containing $10 ppm DDE 

(unspecified isomeric composition) (Friend et al. 1973). 

No significant changes compared to controls were seen in lactic, malic, α-glycero-phosphate, or glucose6-phosphate dehydrogenase activities in extracts of homogenated liver, brain, breast muscle, and kidney 

tissue from redstarts (Phoenicurus phoenicurus) administered a cumulative oral dose of 126 μg p,p’-DDT 

administered in equal daily doses over a 12-day period (Karlsson et al. 1974). 

D.3 Immunological and Lymphoreticular Effects 

No experimental studies were located regarding immunological or lymphoreticular effects in wild 

mammals, reptiles, or amphibians from exposure to DDT/DDE/DDD. Spleen weights were unaffected in 

several bird species. Immunological function in chickens was impaired after dietary exposure to DDT, 

but not in Mallards after similar exposures. 
 

DDT, DDE, and DDD D-17 

APPENDIX D 

Birds. Spleen weight was not significantly affected in white pelicans (Pelecanus erythrorhynchos) 

administered a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDD, and 15 mg p,p’-DDE per day for 

10 weeks (Greichus et al. 1975). Double-crested cormorants (Phalacrocorax auritus) fed diets containing 

up to 25 ppm of total DDT/DDE/DDD for 9 weeks showed no significant change in spleen weight 

(Greichus and Hannon 1973). Spleen and bursa weights were not significantly affected in chickens fed 

diets containing 500 ppm of p,p’-DDT for 5 weeks (Glick 1974). 

Serum immunoglobulin levels were not significantly affected in white pelicans exposed by daily oral 

administrations of a combination of 20 mg p,p’-DDT, 15 mg p,p’-DDD, and 15 mg p,p’-DDE for 

10 weeks (Greichus et al. 1975). The antibody titre in response to bovine serum albumin (BSA) 

injections was significantly decreased in domestic fowl chicks fed diets containing 400 ppm of p,p’-DDT 

(but not at 200 ppm) for 5 weeks and then had food withheld for 4 days prior to injection with BSA (but 

not in birds that did not have food withheld); serum IgG and IgM were depressed at exposure levels of 

$200 ppm for 4 weeks (Glick 1974). Plaque formation in response to sheep red blood cell challenge and 

an index of phagocytic activity were not significantly changed relative to controls in domestic fowl chicks 

fed diets containing 500 ppm of p,p’-DDT for 5 weeks (Glick 1974). A 10-day dietary exposure to up to 

900 ppm of p,p’-DDT in the feed did not significantly affect hepatitis-induced mortality or incidence liver 

lesions in Mallard ducks (Anas platyrhyncus) (Friend and Trainer 1974b). 

“Severe dissolution” of bursal follicles and vacuolation with loss of medullary cells was observed in the 

bursae of domestic fowl chicks fed diets containing 500 ppm of p,p’-DDT for 5 weeks (Glick 1974). 

D.4 Neurological/Behavioral Effects 

No experimental studies were located regarding neurological or behavioral effects in reptiles from 

exposure to DDT/DDE/DDD. 

Neurological effects (tremors, convulsions, hyperactivity, and behavioral changes) were observed in 

mammalian wildlife, amphibians, and birds experimentally exposed to DDT or DDE, particularly after 

administration of lethal doses or after administration of lower doses when food intake was restricted. The 

most commonly reported neurological effect was tremors. Studies generally did not offer explanations as 

to the possible mechanisms that caused tremors, although it is not unreasonable to assume a mechanism 

similar to that seen in laboratory animals (see Section 3.6.2, Mechanisms of Toxicity). Diets were 

experimentally restricted in several studies to simulate the health effects of DDT/DDE/DDD mobilized 

from fat during periods of energetic stress in the wild such as may occur, for example, during periods of 

nesting, migration, or thermal or other stress. Reviews (EPA 1975; WHO 1989) have postulated that 

during periods of energy stress, DDT mobilized from fat is redistributed to the brain (presumably because 

of the high lipid content in brain tissue) where it induces neurological effects and death. A study in bats 

(Clark and Kroll 1977) demonstrated that DDT residues in the brain increase substantially when the diet 

was restricted. Although a direct action on the central nervous system in wildlife has not been confirmed 

by observations of brain lesions, one study in birds did show significant decreases in brain 

neurotransmitter levels that were associated with increased brain DDE residue levels after sublethal 

dietary exposures (Heinz et al. 1980). Alterations in neurotransmitter levels may explain changes in bird 

behavior that were observed in several species. However, most available data suggest that wildlife 

species may not be sensitive sentinels of neurological effects in humans because the most prominent 

neurological effects in wildlife occurred primarily at lethal exposure levels or in energy-stressed animals 

at lower exposure levels. 

Mammals. Tremors, convulsions, and hyperactivity (“running fits”) were observed in 16 of 40 shorttailed shrews (Blarina brevicauda) that died after consuming p,p’-DDT in the diet ad libitum for 14 days 

at unreported concentrations (the range of concentrations was sufficient to identify an LC50 of 1,784 ppm) 
DDT, DDE, and DDD D-18 

APPENDIX D 

(Blus 1978). Prolonged tremors were seen in free-tailed bats (Tadarida brasiliensis) fed diets containing 

107 ppm DDE (unspecified isomeric composition) for 40 days and subsequently starved; tremors were 

observed only during starvation in treated bats but not in control bats (Clark and Kroll 1977). Tremors 

were seen in big brown bats (Eptesicus fuscus) within 2.5 hours of a single oral dose of technical-grade 

DDT at 800 mg/kg, and were seen up to 14 days postdosing in virtually all bats administered DDT at 

doses ranging from 20 to 800 mg/kg (Luckens and Davis 1964). Pipistrelle bats also showed tremors the 

day before dying from a single oral exposure to >45 mg/kg of p,p’-DDT (Jefferies 1972). 

Amphibians. Common frog (Rana temporaria) tadpoles showed uncoordinated movement and signs of 

hyperactivity after 1 hour of exposure to as low as 0.1 ppm p,p’-DDT (but not at 0.01 ppm) (Cooke 

1970b), and transient uncoordinated hyperactivity when exposed to 0.001 ppm p,p’-DDT for as few as 

5 days (Cooke 1973a). Similar neurological symptoms were seen in common frog tadpoles exposed as 

eggs to DDT for only 24–48 hours (Cooke 1972). Tadpoles of common frogs, common toads (Bufo 

bufo), and smooth newts (Triturus vulgaris) showed hyperactivity and abnormal movement when exposed 

to DDT (unspecified isomeric composition) for 24–48 hours during metamorphosis (Cooke 1972). Adult 

common frogs showed hyperactivity, tremors, lack of muscular coordination, and weakness several days 

after a single lethal oral dose of DDT (isomer not reported) at an unreported dose level; adult frogs who 

did not die within 10 days of the exposure recovered fully, and did not show further signs of neurotoxicity 

(Harri et al. 1979). Adult common frogs administered DDT (unspecified isomeric composition) in 

16 twice-weekly oral doses of 0.6 mg/kg/dose in gelatin capsules showed similar signs of neurotoxicity 

and eventually died, but only when food was withheld at the time of exposure; frogs that were fed at the 

time of dosing did not show any signs of toxicity (Harri et al. 1979). 

Birds. Brain weight was not significantly affected in white pelicans (Pelecanus erythrorhynchos) that 

consumed a diet of fish containing an unreported dose of a mixture of p,p’-DDT (40%), p,p’-DDE (30%), 

and p,p’-DDD (30%) for 10 weeks, and then subjected to a restricted diet for 2 weeks (Greichus et al. 

1975). Double-crested cormorants (Phalacrocorax auritus) fed diets containing up to 25 ppm of total 

DDT/DDE/DDD for 9 weeks showed no significant change in brain weight (Greichus and Hannon 1973). 

No effect on brain weights was seen in redstarts (Phoenicurus phoenicurus) administered a cumulative 

oral dose of 126 μg p,p’-DDT administered in equal daily doses over a 12-day period (Karlsson et al. 

1974). 

No gross brain lesions were observed in 6-week-old pheasant chicks (Phasianus colchicus) fed 100 ppm 

technical-grade DDT for up to 101 days or 500 ppm for up to 23 days (Azevedo et al. 1965). 

Brain dopamine levels were significantly decreased in ringed turtle doves after 8-week dietary exposures 

of 20 and 200 ppm DDE (unspecified isomeric composition), but not when fed 2 ppm, and brain 

norepinephrine was significantly decreased at 200 ppm, but not at 2 or 20 ppm (Heinz et al. 1980). A 

significant negative correlation was seen between neurotransmitter levels and DDE residues in the brain 

(Heinz et al. 1980). 

Bald eagles (Haliaetus leucocephalus) showed tremors before dying from consuming diets containing 

technical-grade DDT at 160 ppm or greater for 15 to 112 days, but tremors were not observed in eagles 

that did not die after consuming 10 ppm diets for up 120 days (Chura and Stewart 1967). Tremors were 

also seen in two out of three bald eagles before dying after administration of 4,000 ppm technical-grade 

DDT in the diet for at least 15 days (Locke et al. 1966). An adult and an immature male American kestrel 

(Falco sparverius) fed diets containing 2.8 ppm p,p’-DDE exhibited tremors prior to death during 

exposures that ranged from 14 to 16 months in duration (Porter and Wiemeyer 1972). Double-crested 

cormorants (Phalacrocorax auritus) fed diets containing #25 ppm total DDT/DDE/DDD for 9 weeks 

showed tremors and convulsions prior to death (Greichus and Hannon 1973). 
DDT, DDE, and DDD D-19 

APPENDIX D 

Pheasants exhibited trembling before death from consuming 100 ppm technical-grade DDT in the diet for 

at least 22 days (Azevedo et al. 1965). Bobwhite quail (Colinus virginianus) showed “slight” tremors 

during a 5-day dietary exposure to 800 ppm technical-grade DDT in the feed; more severe neurotoxic 

signs, including convulsions, were seen in quail fed 1,600 ppm (Hill et al. 1971). Premortality tremors 

were seen in Japanese quail fed diets containing $700 ppm (lowest level tested) of p,p’-DDT for up to 

20 days (Gish and Chura 1970). Japanese quail (Coturnix coturnix japonica) provided with an earthworm 

diet containing 298 ppm DDT (unspecified isomeric composition) for up to 5 days exhibited tremors 

before dying (Boykins 1967), but showed no tremors or convulsions after consuming #100 ppm 

p,p’-DDE in a dry diet of mash for 12 weeks (Dieter 1974). Pharol D-1 quail fed 40 ppm p,p’-DDT for 

12 weeks showed tremors, but did not die (Davison et al. 1976). 

House sparrows, cardinals (Richmondena cardinalis), and blue jays (Cyanocitta cristata) each showed 

tremors after consuming technical-grade DDT at concentrations in the diet ranging from 320 to 910 ppm 

for 5 days (Hill et al. 1971). Tremors were seen in house sparrows (Passer domesticus) provided with 

drinking water containing 12% DDT (unspecified isomeric composition) or earthworm diets containing at 

least 86 ppm DDT for 6 or fewer days (Boykins 1967). “Severe” tremors were observed prior to the 

death of cowbirds (Molothrus ater) fed diets containing 500 ppm of p,p’-DDT for at most 12 days 

(Stickel et al. 1966). Trembling accompanied by significantly decreased body temperature and oxygen 

consumption was seen in homing pigeons (Columba livia) orally administered p,p’-DDT at 36 mg/kg/day 

for 12 weeks and then 54 mg/kg/day for 6 weeks (Jefferies and French 1971). Trembling was observed in 

pigeons that died after oral exposure to p,p’-DDE at $36 mg/kg/day by capsule for up to 56 days 

(Jefferies and French 1972). 

Ataxia was observed in 8 chickens fed diets containing 1,600 ppm of p,p’-DDT for <2 to 4 weeks before 

dying, but no effects were seen at 800 ppm (Glick 1974). Uncoordinated movement was seen in 

pheasants that died after consuming 100 ppm technical-grade DDT in the diet for at least 22 days 

(Azevedo et al. 1965). Bobwhite quail (Colinus virginianus) showed “slight” balance disturbances during 

a 5-day dietary exposure to $800 ppm technical-grade DDT (Hill et al. 1971). Balance disturbances were 

seen in house sparrows, cardinals (Richmondena cardinalis), and blue jays (Cyanocitta cristata) after 

consuming technical-grade DDT at concentrations in the diet ranging from 320 to 910 ppm for 5 days 

(Hill et al. 1971). 

Predatory response time and attack rate were not significantly affected in American kestrels fed diets 

containing 6 ppm p,p’-DDE for 4 months, both in comparison to pretreatment rates and compared to 

controls (Rudolph et al. 1983). Mallard peck order was apparently not affected by consuming 

2.5–250 ppm technical-grade DDT in the diet for at least 29 days (Peterle et al. 1973). The onset of 

nocturnal restlessness indicative of normal migratory behavior (Zugunruhe) appeared to be delayed for 

1 week in white-throated sparrows (Zonotrichia albicollis) fed 5–25 ppm technical-grade DDT in the diet 

for 31 days prior to the migratory season and for approximately 2 weeks during the initial phases of the 

migratory season (followed by a 14-week observation period on untreated diet); the behavior was 

abnormally increased toward the end of the migratory season after exposure had ceased (Mahoney 1975). 

A similar late migratory season increase in nocturnal restlessness was seen in white-throated sparrows fed 

5–25 ppm DDE (unspecified isomeric composition) in the diet for 31 days prior to the migratory season 

and for approximately 2 weeks during the initial phases of the migratory season (Mahony 1975). 

Significantly decreased courting behavior was seen in ringed turtle doves fed p,p’-DDE at 10 or 50 ppm 

in the diet for 63 days (Haegele and Hudson 1977). Courtship behavior in ringed turtle doves 

(Streptopelia risoria) was not significantly affected by a 3-week pre-pairing dietary exposure to 100 ppm 

p,p’-DDE, but when birds had also been on a 10% food reduction, courtship behavior was “practically 

eliminated”; a significant interaction between DDE exposure and restricted diet was apparent, because the 

reduction in courting behavior was not as pronounced in birds on the restricted diet that had not been 

exposed previously to DDE (Keith and Mitchell 1993). Decreased nest attendance by parental birds was 
 

DDT, DDE, and DDD D-20 

APPENDIX D 

reported in ringed turtle doves fed diets ad libitum containing 100 ppm p,p’-DDE for a 3-week period 

prior to pairing for mating (Keith and Mitchell 1993); decrease in chick survival appeared to be related to 

nest attendance, and the authors hypothesized that losses in young birds were related to decreased food 

intake by the chicks. No significant effect on locomotor activity pattern was reported in redstarts 

(Phoenicurus phoenicurus) administered a cumulative oral dose of 126 μg p,p’-DDT administered in 

equal daily doses over a 12-day period (Karlsson et al. 1974). 

D.5 Reproductive Effects 

Little information is available concerning reproductive effects in reptiles, amphibians, and wild mammals 

from exposure to DDT/DDE/DDD. In birds, the well-publicized decline in wild raptor populations, 

including the bald eagle, during the 1950s and 1960s was attributed partly to reproductive impairment, 

particularly eggshell thinning (see Appendix D.6, Eggshell Thinning in Birds). Egg production, fertility, 

and hatchability were largely unaffected in numerous studies in a variety of bird species. However, 

increased embryolethality, decreased egg size, delayed oviposition after mating, and increased testicular 

effects were observed with some regularity among experimental studies in birds. Several authors 

speculated that the effects were due to DDT-induced hormonal imbalances, and in fact, blood hormone 

levels (estrogen, luteinizing hormone) were altered in three of four studies in birds consuming either DDT 

or DDE in the diet. While the mechanisms of toxicity for these effects have not been thoroughly 

investigated in wildlife, and thus the direct relevance to human health is uncertain, the consistency of 

certain reproductive effects suggests that wildlife species may be appropriate sentinels for reproductive 

toxicity of DDT/DDE/DDD in humans. 

Mammals. Short-tailed shrews showed testis weights that were not significantly different from controls 

after consuming an earthworm diet containing an average of 16.6 ppm DDT for 3 weeks (unspecified 

isomeric composition) (Braham and Neal 1974). Similar to the associations made between DDT and 

preterm deliveries in humans (Saxena et al. 1980, 1981; Wassermann et al. 1982), premature births in 

California sea lions (Zalophus californianus californianus) are associated with elevated DDT 

concentrations (DeLong et al. 1973). Mean blubber DDT in sea lions delivering preterm was 824.4 ppm, 

whereas blubber DDT in full term individuals averaged 103.2 ppm. However, the effect of preterm 

delivery could not be causally isolated to DDT, as PCBs were also significantly elevated in individuals 

having preterm deliveries. 

Reptiles. Organochlorine contaminants, in general, and p,p’-DDE specifically, are thought to influence 

sexual dimorphism in the common snapping turtle (Chelydra serpentina) (de Solla et al. 1998). In normal 

snapping turtles, the cloaca of the female is located very close to the plastron when compared to males. 

However, snapping turtles in Ontario, Canada, with high p,p’-DDE concentrations (mean concentrations 

of 10.1 ppm at one site and 21.7 ppm at another site) lacked such sexual dimorphism in the distance 

between cloaca and plastron. De Solla et al. (1998) attributed this lack of sexual dimorphism to the 

antiandrogenic effects of p,p’-DDE. 

Birds. Ovary weight and oviduct weight were not significantly affected in adult female Japanese quail 

(Coturnix coturnix japonica) administered o,p’-DDT by intramuscular injection at 5 mg/bird on each of 

4 consecutive days (Cooke 1970b). Ovary weights were significantly reduced in an exposure-related 

manner in Japanese quail fed diets containing between 700 and 1,600 ppm p,p’-DDT for 20 days (Gish 

and Chura 1970); the effect was more pronounced in treatment groups with relatively lower body 

weights. Increases in oviduct weight and uterine glycogen content were observed in chickens and 

Japanese quail administered 50 mg/bird of o,p’-DDT (but not in birds administered p,p’-DDT) by 

intraperitoneal injection on 3 consecutive days; the increases were comparable to those seen in birds 

injected 3 times with 0.5 mg 17β-estradiol (Bitman et al. 1968). Ovary and oviduct weights were 
DDT, DDE, and DDD D-21 

APPENDIX D 

significantly decreased in ringed turtle doves fed diets containing 100 ppm p,p’-DDE for a 3-week period 

prior to pairing for mating (Keith and Mitchell 1993). 

Latency time between first pairing and first oviposition was significantly increased compared to controls 

in Mallard hens fed diets containing 10 ppm DDE (isomeric composition not reported) for 2 months prior 

to pairing and through the breeding season (Vangilder and Peterle 1980). Delayed oviposition after initial 

pairing (compared to controls), attributed by the authors to delayed ovulation, was observed in Japanese 

quail fed diets containing either 100 ppm p,p’-DDT or 100 ppm p,p’-DDE (the only exposure levels 

tested) for up to 74 days (Cecil et al. 1971). The incidence of delayed oviposition was significantly 

increased in a dose-related manner in Bengalese finches (Lonchura striata) after dietary exposure to 

p,p’-DDT at up to 0.3 mg/bird/day (dose levels were not reported) for at least 15 weeks (Jefferies 1967, 

1969). Jefferies (1967) hypothesized that the proximal cause of the delay in oviposition in Bengalese 

finches was delayed ovulation because of underdeveloped ovaries; he further postulated that DDT may 

have induced an estrogen-like inhibition of FSH and LH secretion by the pituitary thus, FSH and LH 

stimulation of ovary development may have been inhibited. An exposure-related significant increase 

(compared to controls) in median latency period between pairing for mating and the first oviposition was 

seen in ringed turtle doves (Streptopelia risoria) fed diets containing 10 and 40 ppm p,p’-DDE for 

approximately 90 days (Richie and Peterle 1979); in doves fed 40 ppm, there was also a delay and a 

nonsignificant suppression in peak serum LH levels after pairing, compared to controls (Richie and 

Peterle 1979), although the authors did not believe the findings established a clear causal relationship. 

Significantly delayed oviposition, associated with significantly reduced blood estradiol (compared to 

controls) and significantly increased hepatic enzyme metabolism of estradiol at 8 days postmating (but 

not in birds after completing their clutch), was observed in ringed turtle doves fed diets containing 

10 ppm p,p’-DDT for 3 weeks prior to mating (Peakall 1970). 

Egg production was not significantly affected in American kestrels (Falco sparverius) fed diets 

containing unreported levels (reportedly calculated to be “just short” of the lethal level) of DDT 

(unspecified isomeric composition) for at least a full year that encompassed two breeding seasons, nor 

was an effect seen in first-year young that were fed DDT at the parental dietary levels and bred in the 

second breeding season of the study (Porter and Wiemeyer 1969). Interestingly, mean egg production 

was significantly increased in two consecutive breeding seasons in barn owls fed diets containing 3 ppm 

DDE (unspecified isomer) for several months prior to the first breeding season and for almost 2 years 

before the second (Mendenhall et al. 1983). The reasons for this were compensatory egg production that 

occurred in the DDE-fed owls, which replaced eggs that were lost due to breakage, and re-nesting that 

occurred among owls that lost their entire clutches. 

Egg production per hen was not significantly affected in mallards fed #40 ppm p,p’-DDT, p,p’-DDE, or 

technical-grade DDD in the diet either for several weeks or over 1 year (Heath et al. 1969). Egg 

production was not significantly affected in mallards (Anas platyrhyncus) exposed to 40 ppm p,p’-DDE in 

the diet for 1–4 months (Risebrough and Anderson 1975) or 10 ppm DDE (isomeric composition not 

reported) for 2 months prior to pairing and through the breeding season (Vangilder and Peterle 1980). 

Egg production and embryonation (the production of an embryo within the egg) were not significantly 

affected in black duck (Anas rubripes) clutches that were laid after approximately 5 months of exposure 

to 10 or 30 ppm (the only levels tested) p,p’-DDE in the diet (Longcore et al. 1971). 

Egg production was significantly decreased in Japanese quail fed technical-grade DDT in the diet at 

10–100 ppm (but not at 3 ppm) for up to 95 days (Kenney et al. 1972). However, egg production was not 

significantly affected in Japanese quail fed up to 200 ppm p,p’-DDE for 13 weeks or up to 40 ppm 

p,p’-DDT for 16 weeks, or in Pharol D-1 quail fed diets containing up to 40 ppm p,p’-DDT for 12 weeks 

(Davison et al. 1976). Egg production was significantly decreased in Japanese quail that were on 

restricted diets for approximately 10 days prior to dietary exposure to $700 ppm (lowest level tested) of 
DDT, DDE, and DDD D-22 

APPENDIX D 

p,p’-DDT for 20 days, but not consistently in birds fed normal quantities of food prior to exposure (Gish 

and Chura 1970). Egg production was not significantly affected in chickens or in Japanese quail by 

dietary exposure to up to 100 ppm of DDT/DDE/DDD (a commercial mix of DDT compounds) for up to 

10 weeks (Scott et al. 1975). There was no consistent effect on egg production in three consecutive 

generations of Japanese quail fed diets containing up to 50 ppm DDT (isomeric composition not 

specified) for their lifetime (Shellenberger 1978). Significantly reduced first-month egg production was 

observed in Japanese quail fed diets containing 100 ppm p,p’-DDE for up to 74 days, but not in those fed 

100 ppm p,p’-DDT (the only exposure levels tested); egg production during 3 subsequent months was not 

significantly different from control levels in both treatment groups (Cecil et al. 1971). Chickens fed 

p,p’-DDT at up to 200 ppm in the diet for 12 weeks showed no significant changes relative to controls in 

egg production (eggs/bird; eggs/clutch) (Davison and Sell 1972). No significant effect on egg production 

was observed in chickens orally administered 40 mg technical-grade DDT/chicken on 5 days (Chen et al. 

1994). No effect on egg production was seen in laying bobwhite quail (Colinus virginianus) orally 

administered up to 20 mg/bird of DDT (unspecified isomeric composition) by capsule every other day 

during a 4-week exposure period (Wilson et al. 1973). Pheasant (Phasiannus colchicus) egg production 

was not significantly affected by exposure to up to 500 ppm technical-grade DDT in the diet for at least 

21 days prior to egg-laying until either the beginning of egg-laying or through egg-laying (Azevedo et al. 

1965). 

Egg production was significantly reduced by an average of 13.5% in ringed turtled doves fed diets 

containing 40 ppm of p,p’-DDE for 126 days (Haegele and Hudson 1973). Clutch size was unaffected in 

Bengalese finches by dietary exposure to p,p’-DDT at estimated dose levels of up to 0.3 mg/bird/day for 

at least 15 weeks (Jefferies 1969). 

Egg weight and lipid content were reportedly not significantly affected compared with controls in barn 

owls (Tyto alba) fed 3 ppm DDE (isomeric composition not reported) in the diet for up to 2 years that 

encompassed two breeding seasons (Mendenhall et al. 1983). Egg weights were not significantly affected 

in chickens fed p,p’-DDT at up to 200 ppm in the diet for 12 weeks relative to controls (Davison and Sell 

1972). Decreased mean egg weight was observed in eggs of bobwhite quail hens orally administered 

20 mg/bird of DDT (unspecified isomeric composition) by capsule every other day during a 4-week 

exposure period, but not in birds administered 10 mg/bird (Wilson et al. 1973). Egg size was 

significantly decreased in a dose-related manner in Bengalese finches after dietary exposure to p,p’-DDT 

at up to 0.3 mg/bird/day (dose levels were not reported) for at least 15 weeks (Jefferies 1969). 

Significantly reduced mean egg weight was observed in ringed turtle doves (Streptopelia risoria) fed 

diets containing 10 ppm p,p’-DDT for 3 weeks prior to mating (Peakall 1970). 

Egg fertility was not significantly affected in American kestrels fed diets containing unreported levels 

(reportedly calculated to be “just short” of the lethal level) of DDT (unspecified isomeric composition) 

for at least a full year that encompassed two breeding seasons, nor was an effect seen in first-year young 

that were fed DDT at the parental dietary levels and bred in the second breeding season of the study 

(Porter and Wiemeyer 1969). However, among parental groups of kestrels, significantly decreased egg 

hatchability was observed in at least one of the two breeding seasons (Porter and Wiemeyer 1969), and 

among yearling breeders, significantly decreased egg hatchability was observed in DDT-treated groups 

compared to untreated controls (Porter and Wiemeyer 1969). Egg hatchability was significantly 

decreased compared to controls in Mallard hens fed diets containing 10 ppm DDE (isomeric composition 

not reported) for 2 months prior to pairing and through the breeding season (Vangilder and Peterle 1980). 

Hatchability and fertility of pheasant eggs were not significantly affected by exposure to up to 500 ppm 

technical-grade DDT in the diet for at least 2 days prior to egg-laying until either the beginning of egglaying or through egg-laying (Azevedo et al. 1965). Egg hatchability was not significantly affected in 

chickens or in Japanese quail exposed to #100 ppm of DDT/DDE/DDD (a commercial mix of DDT 

compounds) in the diet for up to 10 weeks (Scott et al. 1975). There was no consistent effect on egg 
DDT, DDE, and DDD D-23 

APPENDIX D 

fertility or hatchability in three consecutive generations of Japanese quail fed diets containing up to 

50 ppm DDT (isomeric composition not specified) for their lifetime (Shellenberger 1978). Egg 

hatchability was not significantly affected in Japanese quail fed 200 ppm technical-grade DDT in the diet 

for either the first or second week of the laying period (Jones and Summers 1968). Decreased fertility 

and hatchability were observed in eggs of bobwhite quail hens orally administered 20 mg/bird of DDT 

(unspecified isomeric composition) by capsule every other day during a 4-week exposure period, but not 

in birds administered 10 mg/bird (Wilson et al. 1973). Egg hatchability was reduced, but not 

significantly, in ringed turtled doves fed diets containing 40 ppm of p,p’-DDE for 126 days (Haegele and 

Hudson 1973). 

Compared to controls, significantly increased embryolethality was observed in the eggs of yearling 

kestrels fed diets, since fledging, containing unreported levels of DDT (unspecified isomer) (Porter and 

Wiemeyer 1969). Embryonic survival was significantly decreased late in the incubation period in 

Mallard ducks fed $10 ppm p,p’-DDE for 2–3 weeks or for over 1 year; no significant effect on 

embryolethality was seen in ducks fed p,p’-DDT or technical-grade DDD for over 1 year (Heath et al. 

1969). Embryolethality in intact eggs was significantly increased in black duck (Anas rubripes) clutches 

that were laid after approximately 5 months of exposure to 10 or 30 ppm (the only levels tested) 

p,p’-DDE in the diet (Longcore et al. 1971). 

Overall reproductive performance score (a composite index accounting for egg production, egg 

hatchability, and fledging rate) was significantly decreased (primarily from a decrease in fledging rate, 

i.e., chick survival) compared with untreated controls in ringed turtle doves fed diets ad libitum 

containing 100 ppm p,p’-DDE for a 3-week period prior to pairing for mating; egg production was 

eliminated in birds that were fed DDE and were provided 10–30% less food than controls that were fed 

ad libitum (Keith and Mitchell 1993). See Appendix D.7, Developmental Effects, for further discussion 

of fledgling survival. In another study, although egg production, egg fertility, and egg hatchability in 

pheasants fed diets containing to up to 400 ppm DDT (unspecified isomeric composition) for at least 

3 weeks were not significantly different from controls, the cumulative effect of the nonsignificant changes 

(reported as relative reproductive success rate-recruitment rate of chicks to 13 days of age) was markedly 

decreased in an exposure-related manner (Genelly and Rudd 1956). 

Testis weight was not significantly affected in adult male Japanese quail orally administered 9 doses of 

10 mg o,p’-DDT/bird over 3 weeks (Cooke 1970b). Testis weights were significantly reduced in an 

exposure-related manner in male Japanese quail fed diets containing between 700 and 1,600 ppm 

p,p’-DDT for 20 days (Gish and Chura 1970); the effect was more pronounced in treatment groups with 

relatively lower body weights. Testis size and secondary sex characteristics were markedly reduced 

compared to controls in White Leghorn cockerels injected subcutaneously for 60–89 days at DDT 

(unspecified isomer) dose levels that increased from 15 to 300 mg/kg/day during the exposure period 

(Burlington and Lindeman 1950). Testis weights and secondary sex characteristics were also “strikingly” 

decreased in cockerels orally administered (unspecified administration technique) DDT (unspecified 

isomer) at 6.25–50 mg/kg/day for 47 weeks (George and Sunararaj 1995); relative testis weight was 

decreased up to 20% of control testis weights. Testis weights were not significantly affected in Bengalese 

finches exposed to #0.3 mg/bird/day p,p’-DDT for at least 15 weeks (Jefferies 1969). Testis weights 

were unaffected in ringed turtle doves fed diets containing 100 ppm p,p’-DDE for a 3-week period prior 

to pairing for mating (Keith and Mitchell 1993). 

No effect on spermatogenesis was observed in testes of bald eagles (Haliaetus leucocephalus) fed a diet 

containing 10 ppm technical-grade DDT for 60–120 days (Locke et al. 1966). Semen volume, percentage 

of live sperm, sperm motility, and sperm concentration were decreased and semen cholesterol 

concentration was increased in a dose-related manner in White Leghorn cockerels administered oral doses 

(unspecified administration technique) of DDT (unspecified isomeric composition) ranging from 6.25 to 
DDT, DDE, and DDD D-24 

APPENDIX D 

50 mg/kg/day for 47 weeks (George and Sundararaj 1995); effects persisted for 21 weeks following 

cessation of exposure. 

“Marked” testicular degeneration was seen in one of three bald eagles fed 4,000 ppm technical-grade 

DDT in the diet for 15 days; the other two birds died after 15 and 23 days of exposure (Locke et al. 1966). 

Testicular degeneration was seen in one eagle fed 160 ppm technical-grade DDT for 112 days, but not in 

another eagle that died after 76 days at the same exposure level (Locke et al. 1966). No gross gonad 

lesions were observed in 6-week-old chicks of pheasants fed 100 ppm technical-grade DDT for up to 

101 days, or 500 ppm for up to 23 days (Azevedo et al. 1965). Testes of cockerels orally administered 

50 mg/kg/day (but not at lower doses) of DDT (unspecified isomeric composition) for 47 weeks showed a 

variety of microscopic pathologies that persisted after cessation of exposure, most notably testicular 

atrophy, and including “markedly” irregular seminiferous tubule size and shape, desquamation of 

spermatogonia, and pyknosis of germ cells (George and Sundararaj 1995). 

Plasma 17β-estradiol concentration was significantly decreased in chickens after a 5-day oral exposure to 

technical-grade DDT at 40 mg/hen/day (Chen et al. 1994). Blood estradiol was significantly reduced 

(compared to controls) at 8 days postmating in ringed turtle doves (Streptopelia risoria) fed diets 

containing 10 ppm p,p’-DDT for 3 weeks prior to mating, but not in birds allowed to complete their 

clutch (Peakall 1970). In ringed turtle doves fed 40 ppm p,p’-DDE for 90 days (but not in those fed 

10 ppm), there was a delay and a nonsignificant suppression in peak serum LH levels after pairing for 

mating, compared to controls (Richie and Peterle 1979); a concurrent delay in oviposition was observed. 

Since LH is a glycoprotein gonadotropin that stimulates corpora lutea formation and ovulation in females 

(Hadley 1984), the delay and transient suppression of peak LH levels after mating could have caused the 

delay in oviposition by delaying ovulation, although Richie and Peterle (1979) did not believe that the 

evidence indicated a clear cause and effect association. 

D.6 Eggshell Thinning in Birds 

Eggshell thinning in birds reached widespread public awareness in the 1960s and 1970s largely because 

of field observations of eggshell thinning in high-profile raptors like the bald eagle, peregrine falcon, and 

osprey, and the association of these observations with abrupt population declines. Field observations and 

experimental studies established a scientific link between DDT/DDE/DDD exposure, particularly DDE, 

and avian eggshell thinning, which weighed significantly in the decision to ban most domestic crop uses 

of DDT in the 1970s (EPA 1975). A large body of literature was developed during the 1960s and 1970s 

regarding the effects of DDT/DDE/DDD on eggshell integrity in birds; experimental findings on eggshell 

thinning are provided below. In general, raptors, waterfowl, passerines, and nonpasserine ground birds 

were more susceptible to eggshell thinning than domestic fowl and other gallinaceous birds, and DDE 

appears to have been a more potent inducer of eggshell thinning than DDT (Cooke 1973b; EPA 1975; 

Lundholm 1997; WHO 1989). Further, reproductive disturbances associated with DDT/DDE/DDD 

exposure continue to be reported in North American populations of predatory birds and/or birds that 

migrate to regions such as South America where DDT is still used (Lundholm 1997). 

There is some question, however, as to the relevance of avian eggshell thinning to human health. Birds 

possess a shell gland—a specialized segment of the oviduct—that has no anatomical or physiological 

counterpart in humans. The function of the shell gland is to lay down calcite (CaCO3 – calcium 

carbonate) onto the developing avian egg to form the eggshell (EPA 1975). Reduced total calcium 

deposition as calcite, rather than reduced percent calcium in the eggshell, has been hypothesized to be the 

proximal cause of avian eggshell thinning (Davison 1978). Mechanisms of action that involve a direct 

action of DDT/DDE/DDD on the shell gland itself probably have no human relevance, but mechanisms of 

action that involve intermediate effects, such as reduced blood calcium, may have relevance to human 

health. Possible mechanisms of eggshell thinning in birds have been extensively studied and reviewed 
DDT, DDE, and DDD D-25 

APPENDIX D 

(Cooke 1973b; EPA 1975; Lundholm 1997; Peakall et al. 1975; WHO 1989). Early experimental work 

focused the eggshell thinning mechanism debate on calcium metabolism as opposed to carbonate 

availability (EPA 1975). 

An early review (Peakall et al. 1975) summarized two sets of hypotheses based on a discussion by Cooke 

(1973c) concerning possible mechanisms of reduced calcium deposition by the shell gland on the 

developing egg. One set of hypotheses (including effects to the thyroid, parathyroid, and adrenals, as 

well as hypotheses involving estrogen mimicry) suggested that the calcium supply to the eggshell gland 

was reduced, and involved an expected intermediate effect of decreased blood calcium. However, 

available data indicate that blood calcium level in birds was generally not sensitive to DDT/DDE/DDD 

exposures even in species in which eggshell thinning has been observed (Lundholm 1997; Peakall et al. 

1975; see Appendix D.2, Systemic Effects, Metabolic, above). The alternative set of hypotheses 

suggested that the functionality of the shell gland itself in laying down the calcium-based shell was 

impaired in spite of an adequate supply of calcium. Shortly after the DDT ban, a leading hypothesis for 

DDE-induced eggshell thinning was an inhibition of calcium ATP-ase in the shell gland; calcium 

ATP-ase is believed to act as a pump to produce active transport of Ca+2 from blood in shell gland 

mucosal capillaries into the lumen of the shell gland (EPA 1975; Lundholm 1997). More recent work in 

ducks suggests that DDE does not directly inhibit calcium ATP-ase, but rather inhibits a signal for 

activating the calcium pump (Lundholm 1997). An estrogenic regulation of eggshell gland calcium 

secretion by DDT/DDE/DDD does not appear to be a tenable hypothesis (Lundholm 1997). The leading 

hypothesis for DDE-induced eggshell thinning involves an inhibition by p,p’-DDE (but not by o,p’-DDE 

or DDT or DDD isomers) of prostaglandin synthesis in the shell gland mucosa (Lundholm 1997). 

Lundholm (1997) postulated that shell gland prostaglandin E2 acts as a signal for calcium ATP-ase 

transport of Ca+2 (coupled to bicarbonate transport) during eggshell formation, and summarized 

experimental work that demonstrated a direct inhibition of prostaglandin synthetase by p,p’-DDE. 

There is still some question as to the primary mechanism of eggshell thinning, and reviewers have 

suggested that mechanisms of eggshell thinning may differ between bird species or differ with 

environmental conditions or physiological state for a given species. The following experimental results 

illustrate the potentially wide-spread occurrence of eggshell thinning in wild bird populations. 

Barn owls (Tyto alba) showed a substantial increase in egg breakage during incubation associated with 

significantly thinned eggshells during two breeding seasons (60–82% breakage in 3 ppm DDE-fed owls 

vs. 3–5% in controls) encompassed within a 2-year exposure period; mean number of eggs hatched and 

mean number of young fledged per mated pair were significantly decreased in DDE-treated owls 

(isomeric composition of DDE not specified), but this was likely due to the increased egg fragility 

(Mendenhall et al. 1983). Among parental and yearling groups of breeding American kestrels, 

significantly increased egg disappearance (presumably from breakage or egg-eating by parental birds) and 

significantly decreased eggshell thickness were seen in at least one of two consecutive breeding seasons 

after kestrels were fed unreported levels of DDT (unspecified isomer) in the diet for at least 1 year that 

encompassed the two breeding seasons (Porter and Wiemeyer 1969). 

Mallard ducks (Anas platyrhynchus) fed 25 ppm p,p’-DDT in the diet or $10 ppm p,p’-DDE for either 

2–3 weeks or for over 1 year showed a significant increase in percent eggs cracked and a decrease in 

eggshell thickness; no eggshell effects were seen in ducks fed technical-grade DDD for over 1 year or in 

ducks fed #10 ppm DDT (Heath et al. 1969). Egg breakage, due to significantly decreased eggshell 

thickness, was increased compared to untreated controls in Mallards (Anas platyrhyncus) after dietary 

exposure to 40 ppm p,p’-DDE for 1–4 months (Risebrough and Anderson 1975). Eggshell thickness was 

significantly decreased compared to controls in Mallard hens fed diets containing 10 ppm DDE (isomeric 

composition not reported) for 2 months prior to pairing and through the breeding season (Vangilder and 

Peterle 1980). Percent of eggs that cracked during incubation and eggshell thinning were significantly 
 

DDT, DDE, and DDD D-26 

APPENDIX D 

increased in black duck (Anas rubripes) clutches that were laid after approximately 5 months of exposure 

to 10 or 30 ppm (the only levels tested) p,p’-DDE in the diet (Longcore et al. 1971). 

Japanese quail (Coturnix coturnix japonica) showed significantly decreased eggshell weight after being 

fed technical-grade DDT in the diet at $10 ppm for up to 95 days (Kenney et al. 1972). However, neither 

eggshell weight nor eggshell thickness were significantly affected in Japanese quail fed up to 200 ppm 

p,p’-DDE for 13 weeks or up to 40 ppm p,p’-DDT for 16 weeks, or in Pharol D-1 quail fed diets 

containing up to 40 ppm p,p’-DDT for 12 weeks (Davison et al. 1976). Breaking strengths of eggshells 

was not significantly affected in chickens or in Japanese quail by dietary exposure to up to 100 ppm of 

DDT/DDE/DDD (a commercial mix of DDT compounds) for up to 10 weeks (Scott et al. 1975). A 

nonsignificant increase in egg breakage and significantly decreased eggshell calcium were observed in 

Japanese quail fed 100 ppm p,p’-DDT for up to 74 days, but not in those fed 100 ppm p,p’-DDE; eggshell 

thickness and eggshell weight were not significantly affected in both exposure groups (Cecil et al. 1971). 

Chickens fed p,p’-DDT at up to 200 ppm in the diet for 12 weeks showed no significant changes relative 

to controls in eggshell weights, eggshell calcium, and eggshell thickness (Davison and Sell 1972). 

Decreased eggshell thickness and shell membrane thickness, along with decreased serum calcium, were 

observed in eggs of bobwhite quail hens (Colinus virginianus) orally administered 20 mg/bird of DDT 

(unspecified isomer) by capsule every other day during a 4-week exposure period; no changes were seen 

in birds administered 10 mg/bird every other day for 4 weeks (Wilson et al. 1973). No significant effect 

on eggshell thickness was seen in chickens orally administered daily doses of 40 mg technical-grade 

DDT/bird for 5 days (Chen et al. 1994). 

Ringed turtle doves (Streptopelia risoria) fed diets containing 40 ppm p,p’-DDE for 126 days showed 

significantly reduced eggshell thickness by an average of 10% compared to controls (Haegele and 

Hudson 1973). Significantly reduced deposition of calcium in eggs (as measured by 45Ca incorporation) 

was observed in ringed turtle doves fed diets containing 10 ppm p,p’-DDT for 3 weeks prior to mating; 

breeding females were orally administered 45Ca on the day before pairing for mating (Peakall 1970). 

Peakall (1970) observed a significant decrease in eggshell weights associated with significantly reduced 

oviduct carbonic anhydrase (which is involved in the secretion of calcareous eggshell) activity in ringed 

turtle doves injected intraperitoneally with 150 mg/kg p,p’-DDE within 1 day of laying the first egg of the 

clutch; since exposure occurred at the end of egg development, the authors concluded that DDE interfered 

directly with eggshell formation at the oviduct, rather than indirectly by affecting blood estradiol. 

Eggshells were respectively 35 and 20% thinner than controls in ringed turtle doves after a 3-week dietary 

exposure to 100 ppm DDE (unspecified isomeric composition) and in Pekin ducks (domesticated 

Mallards: Anas platyrhyncus) fed diets containing 250 ppm DDE for 10 days; no other exposure levels 

were tested (Peakall et al. 1975). 

D.7 Developmental Effects 

No experimental studies were located regarding developmental effects in wild mammals from exposure to 

DDT/DDE/DDD. 

Little information is available concerning developmental effects in amphibians, but effects in reptile and 

bird populations have received considerable attention. Studies of alligator populations at Lake Apopka in 

Florida, where a pesticide spill occurred in 1980, have reported various effects that may ultimately affect 

reproduction in the population, including reduced clutch viability (Woodward et al. 1993), altered 

steroidogenesis (Crain et al. 1997; Guillette et al. 1995), abnormal ovarian morphology and plasma 

17β-estradiol levels in female alligators (Guillette et al. 1994), and reductions of phallus size and serum 

testosterone in male alligators (Guillette et al. 1994, 1995, 1996, 1997, 1999), compared to alligators at 

control lakes where environmental concentrations of pesticides including DDT/DDE/DDD were relatively 

low. The purported estrogenicity of DDT and other contaminants was hypothesized to have induced 
DDT, DDE, and DDD D-27 

APPENDIX D 

hormonal imbalance in Lake Apopka alligators, causing the observed effects (Guillette and Crain 1996). 

Since the alligators were exposed to a complex mixture of environmental contaminants at Lake Apopka, 

the contribution of DDT/DDE/DDD to the observed effects is uncertain. However, findings in an 

experimental study of in ovo DDE exposures in alligators (Matter et al. 1998) support the hypothesis that 

certain DDT-related compounds induce estrogenic effects in reptiles, potentially ultimately affecting 

reproduction in the population. In general, reptiles may be particularly susceptible to the endocrinealtering effects of DDT/DDE/DDD, as many reptiles have environmental sex-determining mechanisms 

compared to the genetic sex-determining mechanisms in birds and mammals (Crain and Guillette 1998). 

In birds, the most consistently reported developmental effect is a reduction in early posthatch survival in 

chicks after oral exposures to DDT or DDE in maternal birds. Reduced chick survival was observed in 

six bird species after acute to chronic experimental exposures, regardless of whether all of the maternal 

exposure occurred only prior to pairing for mating, only after mating (e.g., only during the hatching phase 

of reproduction), or both. While the mechanism of DDT-induced reduced chick survival has not been 

thoroughly studied, investigators have hypothesized that increased body burden of DDT in chicks may 

cause direct toxicity, or that reduction in parental care-giving among treated birds may result in chick 

malnutrition and poor survival. Other developmental effects in birds include a decreased ability to 

thermoregulate and behavioral alterations in chicks of treated parental birds, a reduced testicular 

development in chicks directly administered DDT, and development of ovarian tissue and oviducts in 

genetically male sea gulls exposed to DDT in the egg. Wildlife species may be appropriate sentinels of 

developmental effects in humans because certain effects, particularly reduced early survival in young, 

occurred consistently across several species under various exposure conditions. 

Amphibians. Common frog (Rana temporaria) tadpole metamorphosis was significantly delayed 

compared to controls in tadpoles exposed for 28 days to 0.001 ppm p,p’-DDT in the surrounding water, 

but not in tadpoles exposed to 0.0001 ppm; abnormal spinal development was seen in 3/8 and 

3/10 tadpoles in the low and high exposure groups, respectively, but not in controls (Cooke 1973a). 

Similar significantly delayed development was seen in common frog tadpoles exposed as eggs to 0.5 ppm 

DDT (unspecified isomeric composition) in the water column for only 24–48 hours (Cooke 1972). 

Altered facial features were observed in tadpoles of common frogs within 48 hours of a 1-hour exposure 

to $1 ppm p,p’-DDT; metamorphosis also appeared to be delayed (Cooke 1970a). 

Reptiles. Recently, vitellogenin has been used as a biomarker for exposure to estrogenic contaminants 

(Folmar et al. 1996). Vitellogenin is a protein precursor of egg proteins that is produced in the liver of 

oviparous and ovoviviparous species and is normally found in the blood of these animals in measurable 

quantities; production of vitellogenin is stimulated by estrogens in the blood that are produced in the 

ovary; thus, it is normally absent in the blood of males (Hadley 1984). Production of vitellogenin 

(suggestive of estrogenic activity) was not induced in 21-day-old American alligators (Alligator 

mississippiensis) exposed in ovo during gonadal differentiation (at male-producing incubation 

temperatures) with p,p’-DDE and o,p’-DDE painted onto the outer surface of the eggshell at exposure 

levels ranging from 0.1 to 10 μg DDE/kg of egg (Matter et al. 1998). However, a significant decrease in 

percent male hatchlings and significantly decreased clitoral and penis sizes compared to controls were 

seen in 21-day-old alligators pooled from groups exposed in ovo to $1 μg p,p’-DDE/kg of egg painted 

onto the eggshell; no significant effects were seen in alligators pooled from groups exposed to #0.3 μg 

p,p’-DDE/kg of egg in ovo (Matter et al. 1998). Treatment of alligator eggs with o,p’-DDE during 

gonadal differentiation led to a significant decrease in percent male hatchlings in the low exposure group 

(#0.3 ppm), but not in the high exposure group ($1 ppm); nonsignificant decreases in clitoral/penis size 

were also seen in low exposure (but not in high-exposure animals) that were exposed in ovo to o,p’-DDE 

(Matter et al. 1998). The lack of a dose-response for these effects after o,p’-DDE exposures in this study 

warrants further investigation into the dose-response of gender alterations. Treatment (topical 

application) of eggs of the green sea turtle (Chelonia mydas) with p,p’-DDE in doses that resulted in up to 
DDT, DDE, and DDD D-28 

APPENDIX D 

543 ng DDE/g egg (ppb) did not alter the sex ratio from what was expected based on temperature alone 

(Podreka et al. 1998). Furthermore, incubation time, hatching success, incidence of body deformities, 

hatching size, and weight were not significantly affected. 

Birds. Significantly decreased early chick survival was observed compared to untreated controls among 

chicks of yearling kestrels fed unreported levels of DDT (unspecified isomer) in the diet since they were 

fledged the previous year (Porter and Wiemeyer 1969). Survival to 14 days posthatch was significantly 

decreased (approximately 35% decrease from control survival rate) in Mallard ducklings (Anas 

platyrhynchus) hatched from hens fed 25 ppm p,p’-DDT in the diet (but not at #10 ppm) for either 

2–3 weeks or for over 1 year; no effect on posthatch survival was seen in ducklings of hens fed p,p’-DDE 

or technical-grade DDD for over 1 year (Heath et al. 1969). Duckling survival was significantly 

decreased in black duck (Anas rubripes) clutches that were laid after approximately 5 months of exposure 

to 10 or 30 ppm (the only levels tested) p,p’-DDE in the diet (Longcore et al. 1971). Pheasant chick 

(Phasiannus colchicus) survival was significantly decreased in an exposure-related manner in groups fed 

100 (lowest level tested) and 400 ppm DDT (isomeric composition not reported) for at least 3 weeks 

(Genelly and Rudd 1956). Chick survival was significantly depressed during the first 7 days posthatch in 

pheasants fed diets containing 500 ppm technical-grade DDT (but not at 10–100 ppm) from at least 

21 days prior to egg-laying until either the beginning of egg-laying or through egg-laying (Azevedo et 

al.1965). Survival of Japanese quail (Coturnix coturnix japonica) chicks was significantly reduced 

compared to controls in groups fed 200 ppm technical-grade DDT in the diet for either the first or second 

week of the laying period; egg hatchability was reportedly not significantly affected by DDT treatment, 

but 79% of chick mortality occurred within 3 days of hatching (Jones and Summers 1968); the authors 

concluded that lethal levels of DDT were transferred to the egg and accumulated in the developing chicks 

during incubation. Survival of Japanese quail, including chicks, was comparable to control group 

survival throughout a 3-generation reproductive toxicity assay in which birds were administered up to 

50 ppm DDT (unspecified isomeric composition) in the diet (Shellenberger 1978). Fledging rate (i.e., 

chick survival) was decreased to 85% (compared to 100% in controls) in ringed turtle doves (Streptopelia 

risoria) fed diets ad libitum containing 100 ppm p,p’-DDE for a 3-week period prior to pairing for mating 

(Keith and Mitchell 1993); the decrease in chick survival was related to decreased nest attendance by 

parental birds, and the authors hypothesized that losses in young birds were related to decreased food 

intake by the chicks. No association between posthatch parental nest attendance and decreased survival in 

chicks was observed during the first 21 days posthatch in another study in ringed turtle doves at a lower 

exposure level. Survival during the first 21 days posthatch was significantly reduced compared to 

controls (by a factor of 2) in chicks of ringed turtle doves fed diets containing 40 ppm of p,p’-DDE for 

126 days (Haegele and Hudson 1973). 

Growth, as measured by body weight and foot pad length was reportedly not significantly affected in 

nestling red-tailed hawks (Buteo jamaicensis) fed technical-grade DDT at 20 mg/kg body weight every 

4th day during an exposure period lasting up to 80 days (Seidensticker 1968). Growth of Japanese quail, 

as measured by body weight, was comparable to control group growth during a 3-generation reproductive 

toxicity assay in which birds were administered up to 50 ppm DDT (unspecified isomeric composition) in 

the diet throughout the study (Shellenberger 1978). No effect on chick weight was seen in bobwhite quail 

(Colinus virginianus) administered up to 20 mg/bird of DDT (unspecified isomeric composition) every 

other day during a 4-week exposure period (Wilson et al. 1973). 

Failure to thermoregulate under cooling thermal stress (measured by cloacal temperature) was seen in a 

greater proportion (50%) of Mallard ducklings (Anas platyrhyncus) of hens fed diets containing 10 ppm 

DDE (unspecified isomeric composition) for 2 months prebreeding and through the breeding season than 

in ducklings of control hens (30%) (Vangilder and Peterle 1980). Among ducklings that did 

thermoregulate during cooling thermal stress, there was significantly greater body weight loss and 
DDT, DDE, and DDD D-29 

APPENDIX D 

significantly decreased survival time compared to control ducklings that did thermoregulate (Vangilder 

and Peterle 1980). 

Testis development was “markedly retarded” and development of secondary sex characteristics was 

inhibited in cockerels administered daily subcutaneous injections of DDT (unspecified isomeric 

composition) for up to 81 days starting on the 8th day posthatch; daily dose was unreported, but 

cumulative dose was 2–3 g/cockerel (Burlington and Lindeman 1952). Similar results were seen in 

another subcutaneous administration study in cockerels administered DDT (unspecified isomer) at dose 

levels that increased from 15 to 300 mg/kg/day over a 60- to 89-day exposure period (Burlington and 

Lindeman 1950). 

Fry and Toone (1981) exposed sea gulls (Larus californicus), which are much more susceptible to the 

effects of estrogenic contaminants than domesticated chickens and quail, to different congeners of DDT 

and DDE. o,p’-DDT induced feminization (development of ovaries and oviduct) at all dosages tested 

(2–100 ppm) in genetically male gulls. p,p’-DDE induced such feminization only at high concentrations 

(20–100 ppm), whereas p,p’-DDT caused no abnormalities. Thus, the estrogenic action of o,p’-DDT has 

a gender-reversing effect in birds. 

No effect on plumage development was seen in nestling red-tailed hawks (Buteo jamaicensis) fed 

technical-grade DDT at 20 mg/kg body weight every 4th day during an exposure period lasting up to 

80 days (Seidensticker 1968). 

Behavioral changes were observed in Mallard ducklings of hens that were fed 3 ppm p,p’-DDE in the diet 

for at least 3 months compared to controls; effects included significant hyper-responsiveness in a maternal 

approach stimulus test and significant under-responsiveness in a noise avoidance test (Heinz 1976). 

D.8 Genotoxic Effects 

No experimental studies were located regarding genotoxic effects in wildlife from exposures to 

DDT/DDE/DDD. 

D.9 Cancer 

No experimental studies were located regarding carcinogenic effects in wildlife from exposures to 

DDT/DDE/DDD. 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife 

Wild mammals, reptiles, and amphibians 

Adverse biological effect 

Shorttailed 

shrew 

Freetailed 

bats 

Big 

brown 

bat 

Pipistrelle 

bat 

Common 

frog 

Bull 

frog 

Common 

toad 

Smooth 

newt 

African 

clawed 

frog 

Mortality 

immature mature OA 1 OI 5 OA 3 OA 1 

OA 1 

OA,I 4 OA 1 

Systemic Effects 

Respiratory Cardiovascular Gastrointestinal Hematological Musculoskeletal Hepatic 

liver weight NOI 4 

Renal Endocrine 

induction of vitellogenin in males OA 2 

Dermal/ocular Body weight 

loss or reduced gain NOA,I 4 OI 7 OA 1 

Metabolic 

O2 consumption OA 4 

NOI 4 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Wild mammals, reptiles, and amphibians (continued) 

Adverse biological effect 

Shorttailed 

shrew 

Freetailed 

bats 

Big 

brown 

bat 

Pipistrelle 

bat 

Common 

frog 

Bull 

frog 

Common 

toad 

Smooth 

newt 

African 

clawed 

frog 

Immunological/ Lymphoreticular Neurological/Behavioral 

tremors convulsions hyperactivity uncoordinated movement 

OA 1 OA 1 OA 1 

OI 7 OA 3 OA 1 OA,I 4 

OA 1,4 

OI 4 

OA,I 1,4 

OI 4 

OA 4 OA 4 

OA 4 OA 4 

Reproductive 

testis weight NOI 4 

Developmental 

facial features spinal abnormalities hatchling sex ratio penis size clitoris size induction of vitellogenin in males delayed development 

OA1

OI1OI1

OA4

Genotoxic Cancer 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: raptors, wading birds, water birds 

Double- White 

Adverse Bald White American crested Mallard Clapper Sandhill Pekin Black B

biological effect eagle pelican kestrel cormorant Puffin duck Rail crane duck duck o

Mortality 

3 5 4 1 immature OI OC OA,I OA 

1,5,10,11 OA 

3 5 4 1 

mature OI OC OA,I OA N

1 OA 

Systemic Effects 

Respiratory Cardiovascular 

12 12 heart weight NOI NOI 

Gastrointestinal Hematological 

12 hematocrit NOI

12 hemoglobin NOI

Musculoskeletal Hepatic 

12 12 7 liver weight OI OI NOI 

12 12 vitamin A level OI OI 

microsomal 

MFO 

7 7 induction OC OI 

12 12 liver carotene NOI NOI 

level 

12 12 serum protein OI NOI 

Renal 

12 kidney weight NOI

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: raptors, wading birds, water birds (continued) 

Double- White 

Adverse Bald White American crested Mallard Clapper Sandhill Pekin Black Bbiological effect eagle pelican kestrel cormorant Puffin duck Rail crane duck duck o

Endocrine 

3 adrenal weight NOI 

adrenal 

corticomedullary 

3 ratio NOI 

3 thyroid weight OI 

thyroid iodine 

3 uptake OI 

plasma 

corticosterone 

3 level NOI 

Dermal/ocular Body weight 

3 12 5 1 loss or OI NOI OC NOA 

4 reduced gain OA

Metabolic 

12 12 blood urea NOI NOI 

12 7 blood K+ OI NOI 

12 12 7 blood Na+ NOI NOI OI 

blood 

12 12 phosphorus NOI NOI 

12 blood Mg+2 NOI 

12 12 7 blood Ca+2 NOI NOI NOA 

plasma 

7 osmolality OI 

Other 

nasal gland 

secretion 

7 7 rate OA NOA 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: raptors, wading birds, water birds (continued) 

Adverse biological effect 

Bald 

eagle 

White 

pelican 

American 

kestrel 

Doublecrested 

cormorant Puffin 

Mallard 

duck 

Clapper 

Rail 

Sandhill 

crane 

White 

Pekin 

duck 

Black 

duck 

B

o

Immunological/ Lymphoreticular 

hepatitis 

susceptibility 

spleen weight serum 

immunoglobulins 

NOI 12 NOI 12 

NOI 12 

NOA 1 

Neurological/ 

Behavioral 

tremors brain weight convulsions predatory behavior social peck 

dominance 

OI 3 

NOI 12 

OC 5 

NOI 5 

OI 12 

NOI 12 

OI 12 

NOI 3 

Reproductive 

egg production egg size/weight embryo survival spermatogenesis testicular lesions egg lipid content egg fertility timing of 

oviposition 

egg hatchability 

NOI 3 

OI 3 

NOI,C 4 

OI,C 4 

NOI,C 4 

OI,C 4 

NOC 1,3,5,10 

NOI 1,5,7,10 

NOI,C 1,10 

OI,C 5 

OI 7 OI 7 

NOI 5 

OI 5 

O

NN

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: raptors, wading birds, water birds (continued) 

Adverse biological effect 

Bald 

eagle 

White 

pelican 

American 

kestrel 

Doublecrested 

cormorant Puffin 

Mallard 

duck 

Clapper 

Rail 

Sandhill 

crane 

White 

Pekin 

duck 

Black 

duck 

B

o

Eggshell Thinning 

egg breakage eggshell 

thickness 

OI,C 4 OI,C 4 

OI,C 1,5 

NOC 10 

OC 1,5 

OI 1,5,7 

NOC 10 

OA 7 

OI 5 OI 5 

OO

Developmental 

chick behavior chick early survival nestling weight nestling foot pad 

growth 

nestling plumage 

development 

chick thermo regulation testicular feminization abnormal oviducts 

OI5

4 5,10 5 OI,C NOI,C OI 

1 OI,C 

OI7 OI7 OI7 

Genotoxic Cancer 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Adverse Biological Effect 

Bobwhite 

quail 

Japanese 

quail 

Birds: gallinaceous birds 

California Pharol 

quail D-1 quail 

Crowned 

Guinea fowl 

Domest

fowl 

Mortality 

immature mature 

4,5 OI 

3,4,5,10 OA 

3,4 OA 

4,5 OI 

4 NOI 

4 NOC 

4,5,10 OA 

1,2 OA 

1 OI 

4 NOC 

3,11 OA 

Systemic Effects 

Respiratory 

gross lung lesions 

Cardiovascular 

gross heart lesions 

Gastrointestinal Hematological 

hemoglobin RBC count hematocrit 

Musculoskeletal 

gross muscle lesions 

TABLE KEY: 

NO = evaluated, but not observed; O = observed 

5 NOI 

3 5 OI NOI 

Subscript (Exposure Duration): 

A = acute; I = intermediate; C = chronic 

Superscript (Chemical Identity): 

1 = p,p’-DDT; 2 = o,p’-DDT; 3 = technical grade DDT; 4 = unspecified DDT; 5 = p,p’-DDE; 6 = o,p’-DDE; 

3 4 OA NOI 

3 4 OA OI

4 NOI

3 4 OA NOI 

7 = unspecified DDE; 8 = p,p’-DDD; 9 = o,p’-DDD; 10 = technical grade DDD; 11 = unspecified DDD; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Adverse Biological Effect 

Bobwhite 

quail 

Birds: gallinaceous birds (continued) 

Japanese California Pharol 

quail quail D-1 quail 

Crowned 

Guinea fowl 

Domest

fowl 

Hepatic 

liver weight plasma hepatic enzyme 

activity 

liver microsomal 

enzyme activity 

microsomal protein 

level

liver lipidliver microscopic

lesions 

blood cholesterol serum lipid serum protein fibrinogen level prothrombin time microsomal hormone 

metabolism 

Renal 

gross kidney lesions 

Endocrine 

adrenal weight adrenal microscopic 

lesions 

thyroid weight thyroid iodine uptake adrenal histology blood sugar level 

Dermal/ocular 

TABLE KEY:

NO = evaluated, but not observed;O = observed

3 1,2,4,5,8 OI NOI 

5 OI

5 OI

OA 1,2,5,6,8,9

OI 1,4,5,8

3 NOI

4 NOI

3 OI

2 NOA,I 

3 OI

3 NOI 

3 OI 

3 OI 3 OI 3 OI 

Subscript (Exposure Duration): 

A = acute; I = intermediate; C = chronic 

Superscript (Chemical Identity): 

1 = p,p’-DDT; 2 = o,p’-DDT; 3 = technical grade DDT; 4 = unspecified DDT; 5 = p,p’-DDE; 6 = o,p’-DDE; 

1 NOI 

3 NOA 

3 OA 

3 OA,I 

3 NOA 

3 9 OA OA 

3 4 NOA NOI 

4 NOI 4 NOI 

3 OA,I 

1 NOI 

9 OA 

3 4 OA NOI 

7 = unspecified DDE;8 = p,p’-DDD;9 = o,p’-DDD;10 = technical grade DDD;11 = unspecified DDD;12 = DDT/DDE/DDD mixture
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Adverse Biological Effect 

Bobwhite 

quail 

Birds: gallinaceous birds (

Japanese 

quail 

California 

quail 

continued) 

Pharol 

D-1 quail 

Crowned 

Guinea fowl 

Domest

fowl 

Body weight 

loss or reduced gain NOI 4 

OA 3 

NOI 1,2,5 

OA 2 

NOI 1,4 

OA 3 

NOC 4 

decreased food 

consumption OA 3 

Metabolic 

oxygen consumption blood urea 

NOI 3 

O A 3

blood K+ O A 3 

blood calcium OI 4 NOA,I 2 NOI 4 

NOA 3 

blood Na+ NOA 3 blood lactic acid blood osmolarity blood phosphorous blood pCO2 blood pO2 blood pH 

OI 3

NOA 3 

OA 3 

NOA 3 NOA 3 

NOI 4 

Immunological/ Lymphoreticular 

spleen weight bursa weight bursal lesions 

NOI 1 NOI 1 

OI 1 

BSA antibody formation OI 1 SRBC plaque-forming 

cells NOI 1 

serum immunoglobulins phagocytic index 

OI 1 

NOI 1 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: gallinaceous birds (continued) 

Adverse Biological Effect 

Bobwhite 

quail 

Japanese 

quail 

California 

quail 

Pharol 

D-1 quail 

Crowned 

Guinea fowl 

Domest

fowl 

Neurological/Behavioral 

tremors convulsions ataxia loss of balance gross brain lesions uncoordinated movement 

OA 3 OA 3 OA 3 

NOI 5 

OI 1 

OA 4 

NOI 5 

OI 1 

OA,I 1 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: gallinaceous birds (continued) 

Bobwhite Japanese California Pharol Crowned Domest

Adverse Biological Effect quail quail quail D-1 quail Guinea fowl fowl 

Reproductive 

egg hatchability egg production egg size/weightegg fertilitytiming of ovipositionovary weightoviduct weight testis weight testicular atrophy semen volume percentage viable sperm sperm motility sperm concentration semen cholesterol testicular microscopic 

lesions 

gross gonad lesions decreased plasma 

estrogen 

uterine glycogen 

OI4 

NOI4 

OI4OI4

NOA3 

NOI12

NOC4

1,3,5 OI 

1,5,12 NOI 

NOC4 NOC4 

1,3,5 OI 

OI1 

NOA2NOA2

OA2

NOI2

OI1

OA2

NOA1

NOI1 

NOI12 

NOI1 

NOA3 

NOI1 

OA2 

OI4 OI4OI4OI4OI4OI4OI4OI4 

OA3 OA2 

NOA1 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Adverse Biological Effect Eggshell Thinning 

eggshell calcium eggshell thickness eggshell weight egg breakage egg membrane thickness 

Developmental 

chick survival testis development secondary sex 

characteristics 

chick weight 

Genotoxic 

Bobwhite 

quail 

OI 4 OI 4 

NOI 4 

Birds: gallinaceous birds (continued) 

Japanese 

quail 

California 

quail 

Pharol 

D-1 quail 

OI 1 

NOI 5 

NOI 1,5 NOI 1,5 

OI 3 

NOI 1,5,12 

NOI 1 NOI 1 

OA 3 

NOC 4 NOC 4 

Crowned 

Guinea fowl 

Domest

fowl 

NOI 1 NOI 1 

NOA 3 

NOI 1 

NOI 12 

OI 4 OI 4 

Cancer 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: passerines and non-passerine ground birds 

Adverse biological effect 

House 

sparrow Cowbird 

Redwinged 

blackbird 

Blue 

jay Cardinal 

Ringed 

turtle 

dove 

Rock 

dove 

Pigeons (homing 

and white 

king) 

Bengalese 

finch 

Mortality 

3 3 3 immature OA OA OA 4 1 4 4 1 mature OA OA OA,I OA OI 

Systemic Effects 

Respiratory Cardiovascular 

pulse rate OI 1 OI 1 

ventricular beat 

amplitude 

heart weight heart muscle tone 

OI 1 

OI 1,5 

OI 5 

NOI 1 

OI 1 

Gastrointestinal Hematological 

hematocrit OI 7 

Musculoskeletal 

soft skull OI 5 

calcium uptake in 

bone OI 1 

Hepatic 

liver weight OI 7 OI 1,5 

plasma aspartate 

aminotransferase OI 7 

vitamin A storage hypertrophy 

OI 1 OI 1 

microsomal hormone 

metabolism OI 1,3 OA 1 

Renal 

kidney weight 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: passerines and non-passerine ground birds (continued) 

Pigeons 

Red- Ringed (homing 

House winged Blue turtle Rock and white Bengalese 

Adverse biological effect sparrow Cowbird blackbird jay Cardinal dove dove king) finch 

Endocrine 

1,5 adrenal weights OI

1 thyroid weight OI

hyperplasia of thyroid 

1 follicular epithelium OI

decreased follicular 

1 colloid OI

Dermal/ocular Body weight 

3 7 5 

loss or reduced gain OA NOI OI decreased body fat 

Metabolic rate 

1 body temperature OI

oxygen consumption 

1 rate OI

7 blood Na+ OI

7 blood K+ NOI

7 blood Ca+2 NOI

Other 

activity of various 

dehydrogenases in various tissues 

Immunological/ Lymphoreticular 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Birds: passerines and non-passerine ground birds (continued) 

Adverse biological effect 

House 

sparrow Cowbird 

Redwinged 

blackbird 

Blue 

jay Cardinal 

Ringed 

turtle 

dove 

Rock 

dove 

Pigeons (homing 

and white 

king) 

Bengalese 

finch 

Neurological/Behavioral 

tremors balance disturbance affected courtship 

behavior 

brain dopamine brain norepinephrine brain weight migratory behavior nesting behavior activity pattern 

OA 3,4 

OA 3 

OA 1 OA 3 

OA 3 

OA 3 OA 3 

OI 5 

NOI 5 

OI 7 OI 7 OI 5 

OI 1,5 

Reproductive 

egg hatchability egg production testis weight clutch size ovary weight oviduct weight timing of oviposition egg size/weight serum LH levels blood estradiol levels 

NOI5 

OI5 

NOI5NOI5

OI5OI5

OI1,5

OI1OI5OI1

NOI1NOI1

OI1OI1

Eggshell Thinning 

eggshell weight OA5 

eggshell thickness OI5,7 45Ca deposition in egg OI1 

oviduct carbonic 

anhydrase activity OA5 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
Table D-1. DDT/DDE/DDD Hazard Identification in Wildlife: Health Effects Observ

Experimental Toxicological Studies in Wildlife (continued)

Adverse biological effect Developmental 

chick early survival 

Genotoxic 

House 

sparrow

Birds: passerines and non-passerine ground birds (continued) 

 Cowbird 

Redwinged 

blackbird 

Blue 

jay Cardinal 

Ringed 

turtle 

dove 

Rock 

dove 

Pigeons (homing 

and white 

king) 

OI 5 

Bengalese 

finch 

Cancer 

TABLE KEY: Subscript (Exposure Duration): Superscript (Chemical Identity): 

NO = evaluated, but not observed; A = acute; 1 = p,p’-DDT; 7 = unspecified DDE; O = observed I = intermediate; 2 = o,p’-DDT; 8 = p,p’-DDD; 

C = chronic 3 = technical grade DDT; 9 = o,p’-DDD; 

4 = unspecified DDT; 10 = technical grade DDD; 5 = p,p’-DDE; 11 = unspecified DDD; 6 = o,p’-DDE; 12 = DDT/DDE/DDD mixture 
DDT, DDE, and DDD E-1 

APPENDIX E
 

INDEX 

acetylcholine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28, 92, 105, 209
 

acetylcholinesterase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29 
 

adenocarcinoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26, 37, 127 
 

adipose tissue . . . . . . . . . . . . . . . . . . . . 4, 18, 24, 25, 30, 92, 99, 109, 111, 117, 118, 120, 141-143, 146, 149, 150, 192, 193, 198,
 

199, 210-212, 230, 258, 271, 272, 279-281, 284, 288, 291, 293-295, 301
 

adrenal gland . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2, 7, 20, 84, 88, 127, 150, 172, 179, 183, 225, 226
 

adrenal glands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 163, 174
 

adrenals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84, 173, 174, 179, 190
 

adrenocorticolytic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18, 172 
 

adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 230, 236, 245, 286, 294
 

aerobic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 230, 243, 245
 

air . . . . . . . . . . . . . . . . . . . . . . . 2-6, 10-12, 15, 36, 37, 184, 202, 216, 222, 223, 227, 229-232, 234, 238, 241, 245-248, 268, 270,
 

282, 286, 290, 296, 300, 302-304, 308, 309, 311
 

alanine aminotransferase (see ALT) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78 
 

ALT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 80, 82, 172 
 

anaerobic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 230, 243, 244, 255
 

androgen receptor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23, 24, 98, 103, 165, 168-171, 182, 183, 187, 196
 

Antarctic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3, 15, 229, 246
 

antiandrogenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20, 21, 23, 98, 102, 139, 168-170, 181-183, 186, 206, 207, 210
 

antiestrogenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23, 24, 181 
 

Arctic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3, 15, 18, 229, 231, 232, 238, 246, 256-258, 260, 261, 268, 282, 284, 287
 

aspartate aminotransferase (see AST) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78 
 

AST . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79, 80, 82, 172 
 

average daily intake . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 265 
 

bass . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 170 
 

BCF . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239, 240, 287, 311
 

bioaccumulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16, 17, 236, 258, 286, 287, 301
 

bioavailability ........................................................................... 210, 237, 287, 288
 

bioconcentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239, 240, 287, 311
 

bioconcentration factor (see BCF) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239, 287, 311
 

biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 231, 237, 238, 242-245, 253, 286
 

biomagnification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16, 239, 240, 287
 

biomarker . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109, 190, 191, 193
 

birds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17, 135-138, 140, 183, 216, 239, 259, 261
 

birth weight . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 101, 186
 

blood . . . . . . . . . . . . . . . . . . . . 6, 8, 9, 11, 16, 20-22, 24, 25, 75-79, 85, 87, 88, 90, 92, 94, 95, 100, 101, 109, 110, 112-115, 117,
 

121, 123, 124, 131, 137, 141-146, 150, 155-157, 161-163, 180, 181, 184, 186, 188, 191-193, 198, 

204, 205, 210, 212, 215, 272-274, 276, 280, 281, 288, 290-292, 295, 301 

body weight effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 85 

breast cancer . . . . . . . . . . . . . . . . . . . . . . . . . . 20, 22, 24, 25, 109-118, 125, 166, 176-178, 181, 204, 215-217, 275, 278, 280, 290 

breast milk . . . . . . . . . . . . . . . . . . . . . . . 4, 6, 8, 11, 18, 19, 100-103, 141, 143, 144, 151, 186, 187, 192, 198, 205, 210, 213, 259, 

282-284, 288, 290, 295 

cancer . . . . . . . . . . . . . . . . . . . 2, 7, 12, 20, 22, 24-27, 30, 35, 37, 72, 109-126, 128, 131, 163, 166, 174, 176-178, 181, 184, 203, 

204, 210, 211, 215-217, 221, 225, 226, 270, 275, 278-280, 290, 311
 

carcinogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7, 25, 194, 195, 217, 303, 311
 

carcinogenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7, 26, 34, 35, 126, 127, 131, 174, 194, 204, 303, 311
 

carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26, 126, 127, 174, 175, 195, 204, 211, 304, 307, 311
 

carcinoma . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20, 26, 84, 120, 126, 172, 176
 

cardiovascular effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74, 130 
 

cholinesterase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 76, 82, 85 
 

Clean Water Act . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 295, 308
 

CYP1A1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 150 
 

CYP2B1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189 
 

CYP3A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 150 
 

dechlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 146, 149, 240, 243
 
DDT, DDE, and DDD E-2 

APPENDIX E 

deoxyribonucleic acid (see DNA) ............................................................... 132, 133, 217
 

Department of Health and Human Services (see DHHS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7, 26, 303 
 

dermal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36, 85, 130 
 

DHHS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26, 303 
 

DNA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 132-135, 173-175, 191, 194, 211, 215, 217
 

dog . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172, 173, 306
 

dopamine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 140 
 

dye . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 176 
 

endocrine effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30, 83, 84, 180 
 

eradication . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 122, 125
 

estrogen receptor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24, 111, 114, 116, 165-169, 171, 176, 177
 

estrogenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21-24, 96, 104, 108, 118, 138, 165-167, 170, 175, 177, 180-183, 186, 210
 

FDA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10-12, 17, 257, 264, 265, 267-269, 282, 305-307, 311
 

FEDRIP . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 214 
 

fetus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8, 9, 145, 155, 156, 161, 174, 183, 188-190, 212
 

fish . . . . . . . . . . . . . . . . . . . 4, 10-12, 16-18, 87, 135, 138, 170, 171, 196, 208, 217, 230, 232, 239, 240, 251, 256, 257, 262-268,
 

270, 275, 277, 279-284, 287, 289, 290, 295, 299, 302, 305, 310, 311
 

fish consumption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87, 275, 281, 284, 310
 

follicle stimulating hormone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97 
 

Food and Drug Administration (see FDA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10, 11, 257, 311
 

fruits . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10, 264, 265, 268
 

FSH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97, 137 
 

gastrointestinal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75 
 

general population . . . . . . . . . . . . . . . . . . . . . . . . 6, 8, 16, 19, 22, 76, 109, 144, 153, 190, 196, 202, 210, 267, 271, 284, 285, 288
 

groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3, 230, 235, 236, 249, 250, 297, 305
 

half-life . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3, 79, 114, 144, 153, 191, 192, 230, 231, 238, 239, 241, 242, 244, 245, 249, 271
 

hematological effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75, 130 
 

Henry’s law . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238 
 

hepatic effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20, 27, 31, 78-82, 130, 172, 203
 

hepatocellular carcinomas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 195 
 

hydrolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 150, 242
 

hydroxyl radical . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 241 
 

hyperthermia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73, 164, 185
 

hyporeflexia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 96 
 

IgG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89 
 

immune system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 88, 89, 208 
 

immunological effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20, 87, 90, 130, 208
 

insects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2, 3, 226, 235
 

Integrated Risk Information System . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 311 
 

kidney . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7, 30, 83, 102, 146, 155, 156, 163, 168, 261, 292, 301
 

lake . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17, 138, 183, 216, 230, 231, 234, 235, 238, 239, 252, 261-263, 277
 

LD50 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 73, 74, 129, 142, 185
 

leukemia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 125, 127
 

LH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97, 103, 137
 

lice . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 130, 226, 303, 307
 

liver . . . . . . . . . . . . . . . . . . . . 6, 7, 16, 18, 20, 23, 24, 26, 27, 30, 31, 78-83, 89, 125-128, 130, 132-134, 145, 146, 155-158, 161,
 

163, 169-172, 175, 176, 179, 185, 188, 191, 193-196, 198, 203, 204, 209, 211, 258, 261, 268, 278, 

291, 292, 294, 295, 301, 303
 

lung . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18, 35, 37, 126, 127, 141, 163, 191, 197, 203, 258, 294
 

luteinizing hormone (see LH) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97 
 

lymph . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25, 90, 114, 115, 141-143, 162, 294, 301
 

lymphatic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 141, 142, 156, 160-162
 

lymphoreticular effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37, 87, 130 
 

marine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4, 17, 231, 238, 243, 253, 256, 257, 260, 284, 286, 287, 310
 

marine mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4, 231, 238, 260
 

Mast cells . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 88 
 

menstrual . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 110, 217
 

methylsulfone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 152, 258, 280
 

methylsulfones . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 150 
 

milk . . . . . . . . . . . . . . . . . . . 4, 6, 8, 11, 17-19, 21, 96, 97, 100-103, 141, 143-145, 151, 155, 157, 161, 174, 186, 187, 189, 190,
 

192, 198, 205, 210, 213, 230, 259, 267, 272-274, 280, 282-284, 288, 290-295, 298, 300, 301 
DDT, DDE, and DDD E-3 

APPENDIX E 

Minimal Risk Levels (see MRL) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27, 34, 35, 188 
 

miscarriages . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21, 94, 205 
 

mosquito . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 226, 239
 

MRL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24, 27, 28, 30, 31, 35, 81, 108, 188, 202, 203, 206, 207, 303
 

MRLs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27, 34, 35 
 

Muscarinic receptor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29, 106 
 

Muscarinic receptors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28, 29, 105, 106
 

musculoskeletal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77, 78 
 

National Human Adipose Tissue Survey (see NHATS) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 272 
 

National Priorities List (see NPL) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1, 16, 209, 232
 

Native American . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10 
 

neoplastic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 303 
 

neurobehavioral . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9, 100, 105, 108, 180, 202
 

neurochemical . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24, 28, 29, 104, 105
 

neurodevelopmental . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28 
 

neurophysiological . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26 
 

neurotransmitter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24, 92, 105, 139, 140, 164, 196
 

New Bedford . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 276 
 

NHATS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 272 
 

Nicotinic cholinergic receptor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106 
 

Nicotinic cholinergic receptors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29 
 

NIOSH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 220, 221, 291, 296, 304, 311
 

NOAEL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30, 34, 74, 81, 86, 90, 93, 100, 108
 

NOAELs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34 
 

no-observed-adverse-effect level . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30 
 

NPL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1, 2, 5, 16, 209, 232-235, 285
 

ocean . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17, 246, 256, 257
 

ocular effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36, 85, 129, 202
 

odds ratio ............................................................................... 110, 111, 116-123
 

Organophosphate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121 
 

paresthesia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26 
 

particulate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 202, 223, 229-231, 236, 237, 241, 247, 257
 

particulates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 17 
 

partition coefficients .................................................................. 156, 222, 223, 236, 239
 

PBPD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 154 
 

PBPK . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145, 153-155, 157-160, 189, 213
 

pharmacodynamic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 153, 154, 213
 

pharmacokinetic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145, 153, 154, 156, 158, 162, 203, 210, 213
 

photolysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 230, 241, 242
 

physiologically based pharmacodynamic (see PBPD) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 154 
 

physiologically based pharmacokinetic (see PBPK) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145, 153, 158
 

plankton . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239, 256, 257
 

porpoise . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 260 
 

precipitation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239, 248
 

public health . . . . . . . . . . . . . . . . . . . . . 1, 2, 7, 11, 12, 15, 20, 33, 34, 179, 200, 215-217, 225, 226, 284-286, 290, 291, 300, 303
 

RCRA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 311 
 

reductive dechlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 146, 149
 

reference dose (see RfD) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 303, 311
 

regulations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11, 12, 303-305, 309, 311
 

renal effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 83, 130 
 

reportable quantity ............................................................................... 308, 311
 

reproductive effect . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 135 
 

reproductive system . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9, 83, 98, 102, 164, 171, 193, 199, 207, 213
 

RfD . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 303, 307, 311
 

sea . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87, 138, 150, 232, 235, 246, 250, 256, 257, 260, 280, 281, 289
 

seal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18, 150, 257, 258, 260
 

sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4, 5, 16, 229-231, 235-239, 242, 243, 248, 251-253, 255, 257, 270, 287
 

sedimentation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 257 
 

selenium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191 
 

serotonin . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 92, 164, 199
 
DDT, DDE, and DDD E-4 

APPENDIX E 

serum . . . . . . . . . . . . . . . . . . . . . 21-24, 26, 27, 72, 75, 77-80, 82, 87, 89, 95, 97, 98, 101-103, 109, 110, 112-117, 119-121, 123­

125, 138, 141, 142, 149, 165, 166, 171, 180, 181, 186, 187, 192, 193, 205, 207, 208, 210, 271, 

272, 274-277, 280, 281, 284, 290-293, 295
 

shrimp . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 240 
 

smelt . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 263 
 

SMR . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72, 73, 125 
 

soil . . . . . . . . . . . . . . . . . . . . . . 2-5, 8, 11, 15, 17, 143, 210, 229-231, 235-246, 250, 251, 253-255, 259, 264, 283, 284, 286-290,
 

295, 297, 300, 302
 

solubility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16, 222, 223, 240, 247, 270
 

Standardized mortality ratio (see SMR) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72 
 

Standardized mortality ratios . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 125 
 

Superfund . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 250, 255, 290
 

surface water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3-5, 229, 234, 235, 242, 248, 250, 252, 286, 295
 

thyroid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84, 127 
 

time-weighted average (see TWA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 311 
 

tomato . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 266, 267, 306
 

total diet studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17, 264, 265, 268, 269
 

toxic equivalents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23 
 

toxicokinetic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33 
 

Toxics Release Inventory (see TRI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 232 
 

transport . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 17, 151, 162-164, 180, 202, 230, 231, 234, 236, 239-242, 245, 247, 256, 283, 286
 

tremors . . . . . . . . . . . . . . . . . . . . . . . . . . 6, 8, 19, 72, 91-93, 99, 104, 131, 139, 140, 164, 185, 193, 196, 199, 202, 209, 210, 212
 

TRI . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 232 
 

triglycerides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112 
 

tumors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25, 26, 111, 114, 120, 126-128, 142, 172, 175, 194, 195, 204, 303
 

TWA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 304, 311
 

Type I . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29, 106 
 

U.S. Department of Agriculture (see USDA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 303, 311
 

USDA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 215-217, 307, 311
 

vapor phase . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3, 202, 223, 229, 241
 

vapor pressure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 222, 223
 

volatility . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 246 
 

volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15, 229-231, 234, 236-239, 242, 244, 245, 247, 266, 286
 

water . . . . . . . . . . . . . . . . . . . . 2-5, 10-12, 15-17, 140, 161, 173, 184, 222, 223, 229-231, 234-240, 242, 243, 245, 248-253, 255­

257, 259, 270, 286, 290, 295-297, 300, 302-305, 308, 310, 311 

whale . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 260, 261, 268, 289