TOXICOLOGICAL PROFILE FOR
 

CHLORINATED DIBENZO-p-DIOXINS
 

U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES
 

Public Health Service
 

Agency for Toxic Substances and Disease Registry
 

December 1998 
CDDs 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. 
CDDs iii 

UPDATE STATEMENT
 

A toxicological profile for chlorinated dibenzo-p-dioxins was released in February 1998. This edition 

supersedes any previously released draft or final profile. 

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
 
.
 
CDDs v 

FOREWORD
 

This toxicological profile is prepared in accordance with guidelines developed by ATSDR and 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 being described. Each profile identifies and reviews the key 

literature (that has been peer-reviewed) that describes a hazardous substance's toxicological properties. 

Other pertinent literature is also presented but 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. 

Each toxicological profile begins with a public health statement, which describes in nontechnical 

language a substance's relevant toxicological properties. Following the public health statement is 

information concerning levels of significant 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 protect public health will be identified by ATSDR 

and EPA. The focus of the profiles is on health and toxicological information; therefore, we have 

included this information in the beginning of the document. 

Each profile must include the following: 

(A) An examination, summary, and interpretation of available toxicological information and 

epidemiological evaluations on the hazardous substance in order to ascertain the levels of 

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

health effects. 

(B) 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. 

(C) Where appropriate, identification of toxicological testing needed to identify the types or 

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

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 the Agency 

for Toxic Substances and Disease Registry (ATSDR) to prepare toxicological profiles for hazardous 

substances which are 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 

Environmental Protection Agency (EPA). The availability of the revised priority list of the 275 

hazardous substances was announced in the Federal Register on February 28, 1994 (59 FR 9486). For 

prior versions of the list of substances, see Federal Register notices dated 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); and October 28, 1992 (57 FR 48801). 
CDDs vi 

Section 104 (i) (3) of CERCLA, as amended, directs the Administrator of ATSDR to prepare a 

toxicological profile for each substance on the list. 

This profile reflects our assessment of all relevant toxicological testing and information that has been peer 

reviewed. It has been reviewed by scientists from ATSDR, the Centers for Disease Control and 

Prevention (CDC), and other federal agencies. It has also been reviewed by a panel of nongovernment 

peer reviewers and is being made available for public review. Final responsibility for the contents and 

views expressed in this toxicological profile resides with ATSDR. 

Jeffrey P. Koplan, M.D., M.P.H.
 

Administrator
 

Agency for Toxic Substances and
 

Disease Registry
 
 

 

CDDs 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: Health Effects: Specific health effects of a given hazardous compound are reported by 

route of exposure, by type of health effect (death, systemic, immunologic, reproductive), 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 2.6 Children’s Susceptibility
 

Section 5.6 Exposures of Children
 

Other Sections of Interest: 

Section 2.7 Biomarkers of Exposure and Effect 

Section 2.10 Methods for Reducing Toxic Effects 

ATSDR Information Center 

Phone: 1-800-447-1544 (to be replaced by 1-888-42-ATSDR in 1999)

 or 404-639-6357 Fax: 404-639-6359 

E-mail: atsdric@cdc.gov Internet: http://atsdr1.atsdr.cdc.gov:8080 

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 
CDDs viii 

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. 

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@dgs.dgsys.com • AOEC Clinic Director: http://occ-envmed.mc.duke.edu/oem/aoec.htm. 

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: 847-228-6850 • FAX: 847-228-1856. 
CDDs	 ix 

CONTRIBUTORS 

CHEMICAL MANAGER(S)/AUTHORS(S): 

Hana Pohl, M.D., Ph.D.
 

ATSDR, Division of Toxicology, Atlanta, GA
 

Fernando Llados, Ph.D.
 

Research Triangle Institute, Research Triangle Park, NC
 

Lisa Ingerman, Ph.D.
 

Syracuse Research Corporation, Syracuse, NY
 

Patricia Cunningham, Ph.D.
 

Research Triangle Institute, Research Triangle Park, NC
 

James H. Raymer, Ph.D.
 

Research Triangle Institute, Research Triangle Park, NC
 

Constance Wall, B.S.
 

Research Triangle Institute, Research Triangle Park, NC
 

Thomas Gasiewicz, Ph.D.
 

University of Rochester Medical Center, Rochester, NY
 

THE PROFILE HAS UNDERGONE THE FOLLOWING ATSDR INTERNAL REVIEWS: 

1.	 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. 

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

consistency across profiles and adherence to instructions in the Guidance. 
.
 
 

CDDs	 xi 

PEER REVIEW
 

A peer review panel was assembled for chlorinated dibenzo-p-dioxins. The panel consisted of the 

following members: 

1.	 James Olson, Department of Pharmacology and Toxicology, State University of New York at 

Buffalo, Buffalo, NY; 

2.	 John Ryan, Bureau of Chemical Safety, Health and Welfare Canada, Ottawa, Ontario, Canada; and 

3.	 Arnold Schecter, College of Medicine, State University of New York, Health Science Center, 

Binghamton, NY. 

These experts collectively have knowledge of chlorinated dibenzo-p-dioxins' 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. 
CDDs	 xii 

PAST PEER-REVIEWS
 

Several drafts of the Toxicological Profile for Chlorinated Dibenzo-p-Dioxins were submitted for a peer-review in 

the past. 

Draft 1997 

Ross Norstrom, National Wildlife Research Center, Environment Canada, Hull, Quebec, Canada 

James Olson, Department of Pharmacology and Toxicology, State University of New York at Buffalo, 

Buffalo, NY 

Dennis Paustenbach, McLaren/Hart ChemRisk, Alameda, CA 

John Ryan, Bureau of Chemical Safety, Health and Welfare Canada, Ottawa, Ontario, Canada 

Arnold Schecter, College of Medicine, State University of New York, Health Science Center, Binghamton, 

NY 

Draft 1994 (scientific panel meeting) 

Anthony DeCaprio,	 Wadsworth Center for Laboratories and Research, New York State Department of 

Health, Albany, NY 

Thomas Gasiewicz,	 University of Rochester Medical Center, Rochester, NY 

Ross Norstrom,	 National Wildlife Research Center, Environment Canada, Hull, Quebec, Canada 

James Olson,	 Department of Pharmacology and Toxicology, State University of New York at Buffalo, 

Buffalo, NY 

Dennis Paustenbach,	 McLaren/Hart ChemRisk, Alameda, CA 

John Ryan,	 Bureau of Chemical Safety, Health and Welfare Canada, Ottawa, Ontario, Canada 

Arnold Schecter,	 College of Medicine, State University of New York, Health Science Center, Binghamton, 

NY 

Draft 1992 

Thomas Gasiewicz,	 University of Rochester Medical Center, Rochester, NY 

James Olson,	 Department of Pharmacology and Toxicology, State University of New York at Buffalo, 

Buffalo, NY 

Arnold Schecter,	 College of Medicine, State University of New York, Health Science Center, Binghamton, 

NY

 Draft 1991 

Thomas Gasiewicz,	 University of Rochester Medical Center, Rochester, NY 

Arleen Rifkind,	 Cornell University of Medical College, New York, NY 

Arnold Schecter,	 College of Medicine, State University of New York, Health Science Center, Binghamton, 

NY 

Jay Silkworth,	 Wadsworth Center for Laboratories and Research, New York State Department of 

Health, Albany, NY 
CDDs	 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 CDDS? ......................................................... 1 
 

1.2 WHAT HAPPENS TO CDDS WHEN THEY ENTER THE ENVIRONMENT? .......... 3 
 

1.3 HOW MIGHT I BE EXPOSED TO CDDS? ....................................... 5 
 

1.4 HOW CAN CDDS ENTER AND LEAVE MY BODY? .............................. 8 
 

1.5 HOW CAN CDDS AFFECT MY HEALTH? ...................................... 9 
 

1.6	 HOW CAN CDDs AFFECT CHILDREN? ....................................... 13 
 

1.7 	 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO CDDs? ............. 14 
 

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

EXPOSED TO CDDS? ...................................................... 15 
 

1.9	 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO
 

PROTECT HUMAN HEALTH? ............................................... 16 
 

1.10 WHERE CAN I GET MORE INFORMATION? .................................. 17 
 

2. HEALTH EFFECTS .............................................................. 19 
 

2.1 HUMAN STUDIES ......................................................... 19 
 

2.1.1 Death ............................................................. 26 
 

2.1.2 Systemic Effects ..................................................... 27 
 

2.1.3 Immunological Effects ................................................ 42 
 

2.1.4 Neurological Effects .................................................. 46 
 

2.1.5 Reproductive Effects ................................................. 49 
 

2.1.6 Developmental Effects ................................................ 52 
 

2.1.7 Genotoxic Effects .................................................... 55 
 

2.1.8 Cancer ............................................................ 56 
 

2.2. ANIMAL STUDIES ......................................................... 68 
 

2.2.1 Inhalation Exposure .................................................. 69 
 

2.2.1.1 Death ..................................................... 69 
 

2.2.1.2 Systemic Effects ............................................ 69 
 

2.2.1.3 Immunological Effects ....................................... 69 
 

2.2.1.4 Neurological Effects ......................................... 69 
 

2.2.1.5 Reproductive Effects ......................................... 69 
 

2.2.1.6 Developmental Effects ....................................... 69 
 

2.2.1.7 Genotoxic Effects ........................................... 69 
 

2.2.1.8 Cancer .................................................... 69 
 

2.2.2 Oral Exposure ...................................................... 69 
 

2.2.2.1 Death ..................................................... 70 
 
 

CDDs xiv 

2.2.2.2 Systemic Effects ............................................ 73 
 

2.2.2.3 Immunological Effects ...................................... 161 
 

2.2.2.4 Neurological Effects ........................................ 165 
 

2.2.2.5 Reproductive Effects ........................................ 166 
 

2.2.2.6 Developmental Effects ...................................... 168 
 

2.2.2.7 Genotoxic Effects .......................................... 176 
 

2.2.2.8 Cancer ................................................... 176 
 

2.2.3 Dermal Exposure ................................................... 178 
 

2.2.3.1 Death .................................................... 178 
 

2.2.3.2 Systemic Effects ........................................... 178 
 

2.2.2.3 Immunological Effects ...................................... 186 
 

2.2.3.4 Neurological Effects ........................................ 186 
 

2.2.3.5 Reproductive Effects ........................................ 186 
 

2.2.3.6 Developmental Effects ...................................... 186 
 

2.2.3.7 Genotoxic Effects .......................................... 187 
 

2.2.3.8 Cancer ................................................... 187 
 

2.3 TOXICOKINETICS ........................................................ 187 
 

2.3.1 Absorption ........................................................ 188 
 

2.3.1.1 Inhalation Exposure ......................................... 188 
 

2.3.1.2 Oral Exposure ............................................. 189 
 

2.3.1.3 Dermal Exposure ........................................... 190 
 

2.3.2 Distribution ....................................................... 192 
 

2.3.2.1 Inhalation Exposure ......................................... 193 
 

2.3.2.2 Oral Exposure ............................................. 193 
 

2.3.2.3 Dermal Exposure ........................................... 196 
 

2.3.3 Metabolism ........................................................ 197 
 

2.3.4 Elimination and Excretion ............................................ 198 
 

2.3.4.1 Inhalation Exposure ......................................... 201 
 

2.3.4.2 Oral Exposure ............................................. 201 
 

2.3.4.3 Dermal Exposure ........................................... 202 
 

2.3.4.4 Transfer of CDDs Through the Placenta and Breast Milk ........... 203 
 

2.3.5 Physiologically Based Pharmacokinetic (PBPK)/Pharmacodynamic (PD) Models 210
 

2.3.5.1 Summary of PBPK Models. .................................. 213 
 

2.3.5.2 Comparison of PBPK Models for 2,3,7,8-TCDD. ................ 215 
 

2.3.5.3 TCDD Models ............................................. 217 
 

2.3.5.4 Risk Assessment. ........................................... 226 
 

2.4 MECHANISMS OF ACTION ................................................ 227 
 

2.4.1 Pharmacokinetic Mechanisms ......................................... 227 
 

2.4.2 Mechanisms of Toxicity .............................................. 230 
 

2.4.3 Animal-To-Human Extrapolations ...................................... 247 
 

2.5 RELEVANCE TO PUBLIC HEALTH ......................................... 252 
 

2.6 CHILDREN’S SUSCEPTIBILITY ............................................ 312 
 

2.7 BIOMARKERS OF EXPOSURE AND EFFECT ................................. 318 
 

2.7.1 Biomarkers Used to Identify or Quantify Exposure to CDDs ................ 319 
 

2.7.2 Biomarkers Used to Characterize Effects Caused by CDDs .................. 322 
 

2.8 INTERACTIONS WITH OTHER CHEMICALS ................................. 323 
 

2.9 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE ...................... 326 
 

2.10 METHODS FOR REDUCING TOXIC EFFECTS ................................ 327 
 

2.10.1 Reducing Peak Absorption Following Exposure ........................... 327 
 

2.10.2 Reducing Body Burden .............................................. 328 
 

2.10.3 Interfering with the Mechanism of Action for Toxic Effects .................. 329 
 
CDDs xv 

2.11 ADEQUACY OF THE DATABASE ........................................... 330 
 

2.11.1 Existing Information on Health Effects of CDDs .......................... 330 
 

2.11.2 Identification of Data Needs .......................................... 333 
 

2.11.4 Ongoing Studies .................................................... 345 
 

3. CHEMICAL AND PHYSICAL INFORMATION ...................................... 357 
 

3.1 CHEMICAL IDENTITY .................................................... 357 
 

3.2 PHYSICAL AND CHEMICAL PROPERTIES ................................... 357 
 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL ............................ 369 
 

4.1 PRODUCTION ........................................................... 369 
 

4.2 IMPORT/EXPORT ........................................................ 371 
 

4.3 USE ..................................................................... 371 
 

4.4 DISPOSAL ............................................................... 371 
 

5. POTENTIAL FOR HUMAN EXPOSURE ............................................ 377 
 

5.1 OVERVIEW .............................................................. 377 
 

5.2 RELEASES TO THE ENVIRONMENT ........................................ 382 
 

5.2.1 Air .............................................................. 390 
 

5.2.2 Water ............................................................ 396 
 

5.2.3 Soil .............................................................. 399 
 

5.3 ENVIRONMENTAL FATE .................................................. 400 
 

5.3.1 Transport and Partitioning ............................................ 401 
 

5.3.2 Transformation and Degradation ....................................... 419 
 

5.3.2.1 Air ...................................................... 419 
 

5.3.2.2 Water .................................................... 420 
 

5.3.2.3 Sediment and Soil .......................................... 423 
 

5.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT ............... 426 
 

5.4.1 Air .............................................................. 426 
 

5.4.2 Water ............................................................ 431 
 

5.4.3 Sediment and Soil .................................................. 434 
 

5.4.4 Other Environmental Media ........................................... 442 
 

5.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE .................. 461 
 

5.5.1 General Population .................................................. 461 
 

5.5.2 Occupational Exposure .............................................. 474 
 

5.6 EXPOSURES OF CHILDREN ............................................... 477 
 

5.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES ...................... 485 
 

5.8 ADEQUACY OF THE DATABASE ........................................... 497 
 

5.8.1 Identification of Data Needs .......................................... 498 
 

5.8.2 Ongoing Studies .................................................... 502 
 

6. ANALYTICAL METHODS ....................................................... 507 
 

6.1 BIOLOGICAL SAMPLES ................................................... 507 
 

6.2 ENVIRONMENTAL SAMPLES .............................................. 522 
 

6.3 ADEQUACY OF THE DATABASE ........................................... 525 
 

6.3.1 Identification of Data Needs .......................................... 526 
 

6.3.2 Ongoing Studies .................................................... 527 
 

7. REGULATIONS AND ADVISORIES .............................................. 529 
 
 

CDDs xvi 

8. REFERENCES ................................................................. 543 
 

9. GLOSSARY ................................................................... 673 
 

APPENDICES
 

A. ATSDR MINIMAL RISK LEVELS AND WORKSHEETS ........................ A-1 
 

B. ATSDR POLICY GUIDELINE (Updated September 2008).................................................... B-1
 

C. USER'S GUIDE.........................................................................................................................C-1
 

D. ACRONYMS, ABBREVIATIONS, AND SYMBOLS ............................. D-1 
 
CDDs xvii 

LIST OF FIGURES 

2-1 Levels of Significant Exposure to 2,3,7,8-TCDDs - Oral ............................... 118 
 

2-2 Levels of Significant Exposure to Dioxins (non-2,3,7,8-TCDD) - Oral ................... 141 
 

2-3 A Generalized Scheme of Pathways for the Biotransformation of CDDs Based on 

Information from In Vivo Mammalian Studies ....................................... 199 
 

2-4 Conceptual Representation of a Physiologically Based Pharmacokinetic (PBPK) 

Model for a Hypothetical Chemical Substance ...................................... 212 
 

2-5 Existing Information on Health Effects of 2,3,7,8-TCDD .............................. 331 
 

2-6 Existing Information on Health Effects of Other CDDs ................................ 332 
 

5-1 Frequency of NPL Sites with 2,3,7,8-TCDD Contamination ............................ 383 
 

5-2 Frequency of NPL Sites with Total Dioxin Contamination ............................. 384 
 

5-3 Frequency of NPL Sites with Tetra Dioxins Contamination ............................ 385 
 

5-4 Frequency of NPL Sites with Penta Dioxins Contamination ............................ 386 
 

5-5 Frequency of NPL Sites with Hexa Dioxins Contamination ............................ 387 
 

5-6 Frequency of NPL Sites with Hepta Dioxins Contamination ............................ 388 
 

5-7 Frequency of NPL Sites with Octa Dioxins Contamination ............................. 389 
 

5-8 National Listing of Fish and Wildlife Consumption Advisories for Dioxins ................ 486 
 
.
 
CDDs xix 

LIST OF TABLES 

2-1 Health Effects in Humans Associated with Estimated 2,3,7,8-TCDD Body Burdens .......... 21 
 

2-2 Levels of Significant Exposure to 2,3,7,8-TCDD - Oral ................................ 74 
 

2-3 Levels of Significant Exposure to Dioxins (non-2,3,7,8-TCDD) - Oral ................... 127 
 

2-4 Levels of Significant Exposure to 2,3,7,8-TCDD - Dermal ............................. 179 
 

2-5 Levels of Significant Exposure to Dioxins (non-2,3,7,8-TCDD) - Dermal ................. 184 
 

2-6 Mean Levels of CDDs in Breast Milk ............................................. 204 
 

2-7 Pharmacokinetic Parameters of 2,3,7,8-TCDD Used in PBPK Models .................... 216 
 

2-8 Comparison of Health Effects Among Species Exposed to CDDs ........................ 248 
 

2-9 Comparison of LOAELs Among Animal Species Following a Single Oral Dose 

of 2,3,7,8-TCDD .............................................................. 249 
 

2-10 Comparison of Body Burden Effects Levels Among Humans and Animals ................ 251 
 

2-11 Toxicity Equivalency Factors (TEFs) for Halogenated Aromatic Hydrocarbons ............ 257 
 

2-12 World Health Organization TEFs for Humans, Mammals, Fish, and Birds ............... 258 
 

2-13 Estimated Body Burdens of 2,3,7,8-TCDD That Correspond to MRLs .................... 267 
 

2-14 Health Effects in Animals Following Lactation-Only Exposure to 2,3,7,8-TCDD ........... 301 
 

2-15 Health Effects in Humans Associated with CDD and CDF Levels in Breast Milk ........... 302 
 

2-16 Genotoxicity of 2,3,7,8-TCDD In Vivo ............................................ 305 
 

2-17 Genotoxicity of 2,3,7,8-TCDD In Vitro ............................................ 308 
 

2-18 Ongoing Studies on CDDs ...................................................... 346 
 

3-1 Chemical Identity of CDDs ..................................................... 358 
 

3-2 Physical and Chemical Properties of CDDs ......................................... 362 
 

5-1 Number of NPL Sites Where CDDs Have Been Detected in One or More 

Environmental Media .......................................................... 397 
 

5-2 Rain Scavenging Ratios (RS) and Percent Washout Due to Particulates (%W) for CDDs 

and CDFs in Ambient Air in Two Midwest Cities .................................... 402 
 

5-3 Bioconcentration Factors (BCFs) for Aquatic Organisms .............................. 408 
 
CDDs xx 

5-4 2,3,7,8-TCDD Levels Measured in Soil Samples Collected in 1984 from Industrialized 

Areas of U.S. Cities ........................................................... 436 
 

5-5 CDD/CDF Concentrations in Food Samples of Animal Origin .......................... 445 
 

5-6 Dioxins, Dibenzofurans, and Dioxin Toxicity Equivalencies (TEQs) in U.S. Foods .......... 446 
 

5-7 Overall National Averages of the Concentrations of Dioxin and Furan Congeners in Fat of 

Meat and Milk on a Lipid Basis .................................................. 447 
 

5-8 Summary of CDDs/CDFs Detected in Fish Tissue as Part of the EPA National 

Study of Chemical Residues in Fish ............................................... 451 
 

5-9 Estimated Average Daily Intake of 2,3,7,8-TCDD by the General U.S. Population. .......... 463 
 

5-10 Estimated Daily Background Exposure to CDDs/CDFs in the General U.S. Population ....... 464 
 

5-11 Estimated Average Daily Intake of TEQs Associated with Exposure to a Typical 

State-of-the-Art Municipal Waste Incinerator ....................................... 468 
 

5-12 Dioxins and Dibenzofurans in Adipose Tissue of the General U.S. Population ............. 471 
 

5-13 Reference Range Levels and TEQS of CDDs, CDFs, and Coplanar PCBs in Whole Blood 

(Lipid Basis for the General Population ............................................ 473 
 

5-14 Mean and Range (ppt) of Serum CDD (Lipid Adjusted) ............................... 492 
 

5-15 Ongoing Studies on Environmental Fate and Treatment of CDDs ........................ 504 
 

6-1 Analytical Methods for Determining CDDs in Biological Samples ....................... 509 
 

6-2 Analytical Methods for Determining CDDs in Environmental Samples ................... 513 
 

7-1 Regulations and Guidelines Applicable to CDDs ..................................... 533 
 
CDDs 1 

1. PUBLIC HEALTH STATEMENT
 

This public health statement tells you about chlorinated dibenzo-p-dioxins (CDDs) 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 clean-up. CDDs (all types) have been found in at least 126 of the 1,467 current 

or former NPL sites. However, it's unknown how many NPL sites have been evaluated for these 

substances. As more sites are evaluated, the number of sites with CDDs may increase. This 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 can be 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 CDDs, 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 it. You 

must also consider the other chemicals you're exposed to and your age, sex, diet, family traits, 

lifestyle, and state of health. 

1.1 WHAT ARE CDDs? 

CDDs are a family of 75 different compounds commonly referred to as polychlorinated dioxins. 

These compounds have varying harmful effects. The CDD family is divided into eight groups of 

chemicals based on the number of chlorine atoms in the compound. The group with one chlorine 

atom is called the mono-chlorinated dioxin(s). The groups with two through eight chlorine atoms 

are called di-chlorinated dioxin (DCDD), tri-chlorinated dioxin (TrCDD), tetra-chlorinated dioxin 

(TCDD), penta-chlorinated dioxin (PeCDD), hexa-chlorinated dioxin (HxCDD), hepta-chlorinated 

dioxin (HpCDD), and octa-chlorinated dioxin (OCDD). The chlorine atoms can be attached to the 

dioxin molecule at any one of eight positions. The name of each CDD indicates both the number 
CDDs 2 

1. PUBLIC HEALTH STATEMENT 

and the positions of the chlorine atoms. For example, the CDD with four chlorine atoms at 

positions 2, 3, 7, and 8 on the dioxin molecule is called 2,3,7,8-tetrachlorodibenzo-p-dioxin or 

2,3,7,8-TCDD. 2,3,7,8-TCDD is one of the most toxic of the CDDs to mammals and has received 

the most attention. Thus, 2,3,7,8-TCDD serves as a prototype for the CDDs. CDDs with toxic 

properties similar to 2,3,7,8-TCDD are called “dioxin-like” compounds. 

In the pure form, CDDs are colorless solids or crystals. CDDs enter the environment as mixtures 

containing a variety of individual components and impurities. In the environment they tend to be 

associated with ash, soil, or any surface with a high organic content, such as plant leaves. In air 

and water, a portion of the CDDs may be found in the vapor or dissolved state, depending on the 

amount of particulate matter, temperature, and other environmental factors. 2,3,7,8-TCDD is 

odorless. The odors of the other CDDs are not known. CDDs are known to occur naturally, and 

are also produced by human activities. They are naturally produced from the incomplete 

combustion of organic material by forest fires or volcanic activity. CDDs are not intentionally 

manufactured by industry, except in small amounts for research purposes. They are 

unintentionally produced by industrial, municipal, and domestic incineration and combustion 

processes. Currently, it is believed that CDD emissions associated with human incineration and 

combustion activities are the predominant environmental source. 

CDDs (mainly 2,3,7,8-TCDD) may be formed during the chlorine bleaching process used by pulp 

and paper mills. CDDs occur as a contaminant in the manufacturing process of certain chlorinated 

organic chemicals, such as chlorinated phenols. 2,3,7,8-TCDD is a by-product formed during the 

manufacture of 2,4,5-trichlorophenol (2,4,5-TCP). 2,4,5-TCP was used to produce 

hexachlorophene (used to kill bacteria) and the herbicide, 2,4,5-trichlorophenoxyacetic acid 

(2,4,5-T). Various formulations of 2,4,5-T have been used extensively for weed control on crops 

and range lands, and along roadways throughout the world. 2,4,5-T was a component of Agent 

Orange, which was used extensively by the U.S. military in the Vietnam War. In most 

industrialized countries the use of products contaminated with CDDs has been greatly reduced. 

Use of hexachlorophene and the herbicide 2,4,5-T is currently restricted in the United States. 

Other chlorinated chemicals, like pentachlorophenol (PCP), used to preserve wood, do contain 

some of the more highly chlorinated CDDs (those with more chlorine atoms), but 2,3,7,8-TCDD is 

not usually found. The use of PCP has been restricted to certain manufacturing applications. 
CDDs 3 

1. PUBLIC HEALTH STATEMENT 

Currently, CDDs are primarily released to the environment during combustion of fossil fuels 

(coal, oil, and natural gas) and wood, and during incineration processes (municipal and medical 

solid waste and hazardous waste incineration). While incineration may be the primary current 

source of release of CDDs into the environment, the levels of CDDs produced by incineration are 

extremely low. CDDs are associated with ash generated in combustion and incineration 

processes. Emissions from incinerator sources vary greatly and depend on management practices 

and applied technologies. CDDs also have been detected at low concentrations in cigarette smoke, 

home-heating systems, and exhaust from cars running on leaded gasoline or unleaded gasoline, 

and diesel fuel. Burning of many materials that may contain chlorine, such as plastics, wood 

treated with pentachlorophenol (PCP), pesticide-treated wastes, other polychlorinated chemicals 

(polychlorinated biphenyls or PCBs), and even bleached paper can produce CDDs. 

Although this public health statement will focus on CDDs, it is important to note that CDDs are 

found in the environment together with other structurally related chlorinated chemicals, such as 

chlorinated dibenzofurans (CDFs) and polychlorinated biphenyls (PCBs). Therefore, people are 

generally exposed to mixtures of CDDs and other classes of toxicologically and structurally 

similar compounds. 2,3,7,8-TCDD is one of the most toxic and extensively studied of the CDDs 

and serves as a prototype for the toxicologically relevant or “dioxin-like CDDs. Based on results 

from animal studies, scientists have learned that they can express the toxicity of dioxin-like CDDs 

as a fraction of the toxicity attributed to 2,3,7,8-TCDD. For example, the toxicity of dioxin-like 

CDDs can be half or one tenth or any fraction of that of 2,3,7,8-TCDD. Scientists call that 

fraction a Toxic Equivalent Factor (TEF). More information on TEFs can be found in Section 2.5. 

For more information on CDDs, please refer to Chapters 3, 4, and 5. 

1.2 WHAT HAPPENS TO CDDs WHEN THEY ENTER THE ENVIRONMENT? 

CDDs are released into the air in emissions from municipal solid waste and industrial incinerators. 

Exhaust from vehicles powered with leaded and unleaded gasoline and diesel fuel also release 

CDDs to the air. Other sources of CDDs in air include: emissions from oil- or coal-fired power 

plants, burning of chlorinated compounds such as PCBs, and cigarette smoke. CDDs formed 

during combustion processes are associated with small particles in the air, such as ash. The larger 

particles will be deposited close to the emission source, while very small particles may be 
CDDs 4 

1. PUBLIC HEALTH STATEMENT 

transported longer distances. Some of the lower chlorinated CDDs (DCDD, TrCDD, and some of 

the TCDDs) may vaporize from the particles (and soil or water surfaces) and be transported long 

distances in the atmosphere, even around the globe. It has been estimated that 20 to 60% of 

2,3,7,8-TCDD in the air is in the vapor phase. Sunlight and atmospheric chemicals will break 

down a very small portion of the CDDs, but most CDDs will be deposited on land or water. 

CDDs occur as a contaminant in the manufacture of various chlorinated pesticides and herbicides, 

and releases to the environment have occurred during the use of these chemicals. Because CDDs 

remain in the environment for a long time, contamination from past pesticide and herbicide use 

may still be of concern. In addition, improper storage or disposal of these pesticides and waste 

generated during their production can lead to CDD contamination of soil and water. 

CDDs are released in waste waters from pulp and paper mills that use chlorine or chlorinecontaining chemicals in the bleaching process. Some of the CDDs deposited on or near the water 

surface will be broken down by sunlight. A very small portion of the total CDDs in water will 

evaporate to air. Because CDDs do not dissolve easily in water, most of the CDDs in water will 

attach strongly to small particles of soil or organic matter and eventually settle to the bottom. 

CDDs may also attach to microscopic plants and animals (plankton) which are eaten by larger 

animals, that are in turn eaten by even larger animals. This is called a food chain. Concentrations 

of chemicals such as the most toxic, 2,3,7,8-chlorine substituted CDDs, which are difficult for the 

animals to break down, usually increase at each step in the food chain. This process, called 

biomagnification, is the reason why undetectable levels of CDDs in water can result in measurable 

concentrations in aquatic animals. The food chain is the main route by which CDD concentrations 

build up in larger fish, although some fish may accumulate CDDs by eating particles containing 

CDDs directly off the bottom. 

CDDs deposited on land from combustion sources or from herbicide or pesticide applications bind 

strongly to the soil, and therefore are not likely to contaminate groundwater by moving deeper into 

the soil. However, the presence of other chemical pollutants in contaminated soils, such as those 

found at hazardous waste sites or associated with chemical spills (for example, oil spills), may 

dissolve CDDs, making it easier for CDDs to move through the soil. The movement of chemical 

waste containing CDDs through soil has resulted in contamination of groundwater. Soil erosion 
CDDs 5 

1. PUBLIC HEALTH STATEMENT 

and surface runoff can also transport CDDs into surface waters. A very small amount of CDDs at 

the soil surface will evaporate into air. Certain types of soil bacteria and fungus can break CDDs 

down, but the process is very slow. In fact, CDDs can exist in soil for many years. Plants take up 

only very small amounts of CDDs by their roots. Most of the CDDs found on the parts of plants 

above the ground probably come from air and dust and/or previous use of CDD-containing 

pesticides or herbicides. Animals (such as cattle) feeding on the plants may accumulate CDDs in 

their body tissues (meat) and milk. 

For more information on what happens to CDDs in the environment, see Chapter 5. 

1.3 HOW MIGHT I BE EXPOSED TO CDDs? 

CDDs are found at very low levels in the environment. These levels are measured in nanograms 

and picograms. One nanogram (ng) is one billionth of a gram, and one picogram (pg) is one 

trillionth of a gram. In some contaminated soils, concentrations of CDDs are reported as parts per 

billion. One part per billion (ppb) is one part CDD per billion parts of soil. The concentration of 

CDDs is often reported as parts per trillion, in samples of air, water, or soil. One part per trillion 

(ppt) is one part CDD per trillion parts of air, water, or soil. In some rural areas where CDD 

concentrations are very low in air or water, measurements are given in parts per quadrillion (ppq), 

which means one part CDD per quadrillion parts of air or water. 

CDDs are found everywhere in the environment, and most people are exposed to very small 

background levels of CDDs when they breath air, consume food or milk, or have skin contact 

with materials contaminated with CDDs. For the general population, more than 90% of the daily 

intake of CDDs, CDFs, and other dioxin-like compounds comes from food, primarily meat, dairy 

products, and fish. CDDs may be present at much lower levels in fruits and vegetables. The 

actual intake of CDDs from food for any one person will depend on the amount and type of food 

consumed and the level of contamination. Higher levels may be found in foods from areas 

contaminated with chemicals, such as pesticides or herbicides, containing CDDs as impurities. 

CDDs have been measured in human milk, cow's milk, and infant formula, so infants are known to 

be exposed to CDDs. 
CDDs 6 

1. PUBLIC HEALTH STATEMENT 

Most surface water in the United States typically does not contain 2,3,7,8-TCDD and other CDDs 

at levels that are high enough to be measured (1 ppq or more). Municipal drinking water does not 

usually contain CDDs because the CDDs do not dissolve in water and primarily stick to particles, 

which are usually filtered out of treated drinking water. This means that using tap water to wash 

clothes or to bathe or shower, or swimming in pools or in uncontaminated lakes, rivers, or at ocean 

beaches will not expose people to significant levels of CDDs. Although CDDs are not usually 

found in filtered, treated drinking water, they have, on occasion, been detected in unfiltered 

groundwater from areas with known CDD contamination. 

Exposure to CDDs can also occur through skin contact with chlorinated pesticides and herbicides, 

contaminated soils, or other materials such as PCP-treated wood and PCB transformer fluids. 

Background levels of CDDs in soil are higher than background levels in both air and water. 

Background levels of CDDs detected in uncontaminated soils in the United States are generally 

very low or not detectable. 2,3,7,8-TCDD is not usually found in rural soil, but is typically found 

in soil in industrialized areas at levels ranging from 0.001 to 0.01 ppb. However, higher levels of 

2,3,7,8-TCDD may be found in areas where CDDs have contaminated the soil. For example, 

contaminated soil at Times Beach, Missouri, had levels of 2,3,7,8-TCDD ranging from 

4.4–317 ppb. 

If CDDs are present at all in outdoor air in rural areas, they are generally present at very low 

levels or at concentrations near the detection limits for testing equipment. In winter, because of 

the burning of wood and other fuels for home heating, CDD levels may be slightly higher than 

during other seasons. In general, the background air levels of CDDs in urban areas are higher than 

in rural areas. Typical levels of CDDs in outdoor air in urban areas and industrial areas averaged 

2.3 picograms per meter cubed (pg/m3). 2,3,7,8-TCDD is not usually found in rural or urban air, 

but it is found in air near urban waste incinerators and high-traffic areas. The air around people 

who are smoking cigarettes may also have CDDs at levels above background levels. Although 

breathing contaminated air is a minor route of exposure for most people, exposure may be greater 

in areas near these CDD sources. 

CDDs have been found in all samples of adipose tissue and blood (serum lipids) from individuals 

with no known previous exposure. This indicates that all people are exposed to small amounts of 

CDDs. Levels of 2,3,7,8-TCDD in serum from the general population typically range from 3 to 
CDDs 7 

1. PUBLIC HEALTH STATEMENT 

7 ppt (on a lipid basis), and rarely exceed 10 ppt. Typically, lower levels of CDDs are found in 

less industrialized countries and in younger people. 

The production, use, and disposal of pesticides and phenoxy herbicides, disposal of production 

waste containing 2,3,7,8-TCDD, industrial accidents involving 2,4,5-trichlorophenol (2,3,5-TCP), 

and the consumption of CDD-contaminated food, have all led to increased potential for excess 

exposure of some groups of people. 2,3,7,8-TCDD has been detected at 91 of the 126 hazardous 

waste sites on the NPL that have been reported to contain CDDs. People living around these sites 

may be exposed to above-background levels of 2,3,7,8-TCDD and other CDDs. Elevated levels of 

CDDs have been reported in fish, shellfish, birds, and mammals collected in areas surrounding 

various chemical production facilities, various hazardous waste sites, and pulp and paper mills 

using the chlorine bleaching process. Sometimes these findings have resulted in closure of these 

areas for the purpose of fishing. People who eat contaminated food from these contaminated areas 

are at risk of increased exposure to CDDs. 

Occupational exposure to CDDs generally occurs through breathing contaminated air, or through 

skin contact with materials containing CDDs. Workers with the potential to be exposed to aboveaverage levels of CDDs include those involved in the production or handling of certain 

chlorinated phenols (such as 2,4,5-TCP, PCP) or chlorinated pesticides or herbicides (such as 

2,4,5-T, 2,4-D, hexachlorophene, Silvex®), and those involved in application of chlorinated 

pesticides containing CDDs as impurities. Workers whose jobs involve pressure treatment of 

wood with PCP and the handling of PCP-treated wood products, chlorination processes at pulp 

and paper mills, or operation of municipal solid waste or hazardous waste incinerators may have 

increased exposure to CDDs. Finally, workers involved in hazardous waste clean-up or clean-up 

of PCB transformer and/or capacitor fires including emergency service personnel like fire fighters 

and police who respond to such fires are also at additional risk of exposure to CDDs. Most of 

these occupational exposures have been significantly reduced in recent years. 

In general, workers involved in the manufacture of 2,4,5-TCP and subsequent products were 

exposed to far greater levels of 2,3,7,8-TCDD than those involved in the handling and application 

of chlorinated pesticides containing CDDs. Current serum lipid levels of 2,3,7,8-TCDD in a small 

number of U.S. Air Force veterans who were directly involved in the aerial spraying of herbicides 

(Agent Orange contaminated with 2,3,7,8-TCDD) in Vietnam as part of Operation Ranch Hand, 
CDDs 8 

1. PUBLIC HEALTH STATEMENT 

are up to 3 times higher than the general population. However, while studies on blood or fatty 

tissue 2,3,7,8-TCDD levels in U.S. Army ground combat Vietnam veterans also found some 

individuals with 2,3,7,8-TCDD levels higher than those of the general population, overall, most 

Vietnam veterans and Vietnamese living in Vietnam studied to dated have blood and fatty tissue 

2,3,7,8-TCDD levels comparable to members of the general U.S. population. 

For more information on exposure to CDDs, see Chapter 5. 

1.4 HOW CAN CDDs ENTER AND LEAVE MY BODY? 

CDDs can enter your body when you breathe contaminated air, eat contaminated food, or have 

skin contact with contaminated soil or other materials. The most common way CDDs can enter 

your body is by eating food contaminated with CDDs. 

If you breathe air that contains CDDs, the CDDs can enter your body through your lungs and pass 

into the blood stream, but we do not know how fast or how much of the CDDs will enter the blood 

stream. If you swallow food or water containing CDDs, most of the CDDs will enter your body 

and pass from the intestines to the blood stream. Smaller amounts of highly chlorinated CDDs 

will enter your body compared to the less chlorinated 2,3,7,8-TCDD. If you swallow soil 

containing CDDs, a small amount of the CDDs will pass through the intestines into the blood 

stream. If soil contaminated with CDDs comes into contact with your skin, some of the CDDs 

will enter the body but we do not know how fast they will enter the blood stream. 

Once in your body, CDDs can be found in most tissues with the highest amounts found in the liver 

and body fat (adipose tissue). Body fat and possibly the liver can store CDDs for many years 

before eliminating them from the body. CDDs with chlorine atoms in the 2, 3, 7, and 8 positions 

and highly chlorinated dioxins, such as OCDD, are generally found in higher concentrations in the 

fat than other CDDs. 

Little is known about CDDs breakdown in the human body. Studies in animals show that some of 

the 2,3,7,8-TCDD from food is slowly broken down. There is evidence from animals suggesting 

that the break-down products are less harmful than the unchanged 2,3,7,8-TCDD. 
CDDs 9 

1. PUBLIC HEALTH STATEMENT 

For people, the average time it takes to remove one-half of the 2,3,7,8-TCDD from the body is
 

highly variable and may take from 7 to 12 years. There is less information on the other CDDs, but
 

what information exists suggests 5 to 15 years. CDDs are eliminated from the body primarily in
 

the stool, and only a small amount leaves the body in the urine. Some CDDs will leave the body
 

in the breast milk of nursing mothers. 

Much less is known about how much other CDD compounds will enter the body, how much will
 

be stored in the body and for how long, and how they are removed from the body.
 

For more information about how CDDs can enter and leave your body, see Chapter 2.
 

1.5 HOW CAN CDDs AFFECT MY HEALTH? 

Many studies have looked at how CDDs can affect human health. Most of these studies examined 

workers exposed during the manufacture of chemicals and pesticides contaminated with 

2,3,7,8-TCDD. Other studies have looked at American Vietnam veterans and Vietnamese 

populations exposed to Agent Orange and populations exposed to 2,3,7,8-TCDD as a result of an 

accident. The workers and Vietnam veterans were most likely exposed to 2,3,7,8-TCDD mainly 

through breathing and skin contact. People who were accidentally exposed to 2,3,7,8-TCDD in 

Seveso, Italy, or Times Beach, Missouri, were probably exposed through eating and drinking 

contaminated food and milk, breathing contaminated particles and dust, through skin contact with 

contaminated soil and through unintentional hand-to mouth activity. Epidemiology is an inexact 

science and many of the human studies have many shortcomings which make it difficult for 

scientists to establish an association between 2,3,7,8-TCDD exposure levels and health effects. A 

common problem with most of the human studies is that the people are exposed to a number of 

chemicals at the same time. In most human health studies, we do not know how much 

2,3,7,8-TCDD people were exposed to or how long the exposure lasted. In other studies, the 

people were examined many years after they were exposed and some of the effects may have not 

have been present at the time of examination or the effects observed may not have been caused by 

2,3,7,8-TCDD. Some of the more recent studies have measured 2,3,7,8-TCDD levels in the blood 

or fat tissue of exposed populations. The levels of 2,3,7,8-TCDD in the blood or fat tissue can be 

used to estimate the extent of past exposures. 
CDDs 10 

1. PUBLIC HEALTH STATEMENT 

A number of effects have been observed in people exposed to 2,3,7,8-TCDD levels which are at 

least 10 times higher than background levels. The most obvious health effect in people exposed to 

relatively large amounts of 2,3,7,8-TCDD is chloracne. Chloracne is a severe skin disease 

characterized by acne-like lesions. Chloracne generally occurs on the face and upper body, but 

may occur elsewhere on the body. Unlike common acne, severe chloracne is harder to cure and 

can be more disfiguring. In milder cases, the lesions heal several months after exposure ends. In 

more severe cases, the lesions may last for many years after exposure. Most of the chloracne 

cases have been attributed to accidental exposure to high doses of 2,3,7,8-TCDD. Other effects to 

the skin, such as erythematous or red skin rashes, discoloration, and excessive body hair, have 

been reported to occur in people following exposure to high concentrations of 2,3,7,8-TCDD. 

Changes in blood and urine that may indicate liver damage have been observed in people. 

Alterations in the ability of the liver to metabolize (or breakdown) hemoglobin, lipids, sugar, and 

protein have been reported in people exposed to relatively high concentrations of 2,3,7,8-TCDD. 

Most of the effects are considered mild and were reversible. However, in some people these 

effects may last for many years. Slight increases in the risk of diabetes and abnormal glucose 

tolerance have been observed in some studies of people exposed to 2,3,7,8-TCDD. We do not 

have enough information to know if exposure to 2,3,7,8-TCDD will result in reproductive or 

developmental effects in people, but animal studies suggest that this is a potential health concern. 

Several studies of workers exposed to high levels (with body burdens more than 50 times higher 

than background body burden levels) of 2,3,7,8-TCDD suggest that exposure to 2,3,7,8-TCDD 

may increase the risk of cancer in people. 

The Department of Health and Human Services (DHHS) has determined that it is reasonable to 

expect that 2,3,7,8-TCDD may cause cancer. The International Agency for Research on Cancer 

(IARC) has determined that 2,3,7,8-TCDD can cause cancer in people, but that it is not possible to 

classify other CDDs as to their carcinogenicity to humans. The EPA has determined that 

2,3,7,8-TCDD is a probable human carcinogen when considered alone and when considered in 

association with phenoxy herbicides and/or chlorophenols. The EPA has determined also that a 

mixture of CDDs with six chlorine atoms (4 of the 6 chlorine atoms at the 2, 3, 7, and 8 positions) 

is a probable human carcinogen. 

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. 
CDDs 11 

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. 

The health effects of some CDDs have been extensively studied in animals. Some CDDs are 

much more toxic than others. 2,3,7,8-TCDD and, to a lesser extent, CDDs with five (penta) or six 

(hexa) chlorine atoms substituted in the 2, 3, 7, and 8 positions, are extremely toxic to animals. 

Other CDDs, which do not have chlorine atoms substituted in the 2, 3, 7, and 8 positions, are 

considered relatively less toxic compared to 2,3,7,8-TCDD. 

2,3,7,8-TCDD has been the most extensively studied CDD and it has been shown to cause a large 

number of adverse health effects in animals. There are always going to be some difficulties in 

using animal data to quantify health risks in people. In general, the doses used in the animal 

studies result in body burdens that are at least 10 times higher than human background body 

burdens, often the animal studies use doses that are over 1,000 times higher than human 

background levels. Some animal species are much more acutely sensitive to 2,3,7,8-TCDD than 

others. For example, it takes several thousand times more 2,3,7,8-TCDD to kill a hamster than a 

guinea pig. The reason for the difference in sensitivity among species is currently being 

investigated. For other effects, such as reproductive toxicity, there is very little difference in 

sensitivity between hamsters and guinea pigs. Another consideration in using animal data to 

predict health effects in people exposed to CDDs in the environment is the design of the animal 

studies. In most of the animal studies, the animals were exposed to only 2,3,7,8-TCDD, the most 

toxic CDD. 2,3,7,8-TCDD is rarely the main CDD found in the environment and people are 

typically exposed to a number of CDDs and compounds with similar toxic actions. Until scientists 

learn more about possible differences between people and animals, levels recommended to be of 

little or no risk to human health are based on the more sensitive species and the assumption that 

effects in animals could occur in people. This approach is further justified on the basis that 

humans are likely to exhibit a wide range of sensitivities to various health effects and the need to 

protect the most susceptible individuals. 
CDDs 12 

1. PUBLIC HEALTH STATEMENT 

In certain animal species, 2,3,7,8-TCDD is especially harmful and can cause death after a single 

exposure to small amounts. Before death, animals may lose as much as 40% or more of their body 

weight following a single dose of 2,3,7,8-TCDD. Exposure to non-lethal levels added in their 

food can cause a variety of adverse effects in animals, such as weight loss, biochemical and 

degenerative changes in the liver. Some animals that were exposed to CDDs in their food had 

effects to the skin such as hair loss, swelling of the face, and moderate to severe chloracne. In 

many species of animals, the immune system appears to be extremely sensitive to 2,3,7,8-TCDD. 

At relatively low levels (approximately 10 times higher than human background body burdens), 

2,3,7,8-TCDD weakens the immune system and causes a decrease in the system's ability to fight 

foreign substances such as bacteria and viruses. 

Exposure to 2,3,7,8-TCDD can cause reproductive damage and birth defects in animals. 

Decreases in fertility, altered levels of sex hormones, reduced production of sperm, and increased 

rates of miscarriages were found in animals exposed to 2,3,7,8-TCDD in food. Rats and mice that 

were exposed to small amounts of 2,3,7,8-TCDD in food for a long time developed cancer of the 

liver and thyroid, and other types of cancer. 

The results of the oral animal studies suggest that the most sensitive effects (effects that will occur 

at the lowest doses) are immune, endocrine, and developmental effects. It is reasonable to assume 

that these will also be the most sensitive effects in humans. 

We know less about the ability of other CDDs to cause adverse health effects. However, it 

appears that all CDDs with chlorine in the 2, 3, 7, and 8 positions have similar effects to 

2,3,7,8-TCDD but the effects occur at higher doses. 

Relatively large amounts of 2,3,7,8-TCDD applied to the skin of some animal species have 

resulted in deaths. Smaller amounts have resulted in weight loss, acne-like sores on the skin, and 

biochemical and degenerative changes in the liver. In addition, mice that had 2,3,7,8-TCDD 

repeatedly applied to their skin developed skin cancer. Although effects in animals following 

exposure through the skin have not been as extensively studied as effects following exposure in 

food, they appear to be quite similar. The ability of other CDDs to cause adverse health effects in 

animals following exposure to the skin has not been well studied. 
CDDs 13 

1. PUBLIC HEALTH STATEMENT 

You can find out more information on the health effects of CDDs in Chapter 2. 

1.6 HOW CAN CDDs AFFECT CHILDREN? 

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

to maturity at 18 years of age in humans. Potential effects on children resulting from exposures of 

the parents are also considered. 

Very few studies have looked at how CDDs can affect children’s health. Chloracne has been 

observed in children exposed to much higher than current background levels of 2,3,7,8-TCDD. 

The children appeared to be more sensitive (effects occurred at a lower body burden) than adults. 

We do not know why children are more sensitive than adults to this effect. It is likely that 

children exposed to higher than background levels will have similar effects as adults. 

We do not know if exposure to CDDs will result in birth defects or other developmental effects in 

people. Birth defects have been observed in animals exposed to higher than background levels of 

2,3,7,8-TCDD. The developing animal is very sensitive to 2,3,7,8-TCDD. In some studies, 

effects were observed at body burdens 10 times higher than human background body burden 

levels. Offspring of animals exposed to 2,3,7,8-TCDD in food during pregnancy often had severe 

birth defects including bleeding, skeletal deformities, kidney defects, weakened immune 

responses, impaired development of the reproductive system, and learning and behavioral 

impairments. Exposure to other CDDs, such as 2,7-DCDD, 1,2,3,7,8-PeCDD, OCDD, and 

HxCDD, can also result in developmental effects in animals. 

We have no information to suggest that there are any differences between children and adults in 

terms of how much CDDs will enter the body, where CDDs can be found in the body, and how 

fast CDDs will leave the body. CDDs from the mother can enter her unborn baby through the 

placenta. It can also be transferred from the mother to infant through breast milk. Because CDDs 

have been measured in human milk, cows milk, and infant formula, nursing infants are also 

exposed to CDDs. In most cases the beneficial aspects (biological and psychological) of breastfeeding outweigh any risks from exposure to CDDs from mother’s milk. 
CDDs 14 

1. PUBLIC HEALTH STATEMENT 

1.7 HOW CAN FAMILIES REDUCE THE RISK OF EXPOSURE TO CDDs? 

If your doctor finds that you have been exposed to significant amounts of CDDs, ask your doctor 

if children may also be exposed. When necessary your doctor may need to ask your state 

Department of Public Health to investigate. 

Structural material used in building homes such as chemically treated lumber for decking and 

plastic PVC pipes used in water pipes and other conduits can release CDDS if they are burned as 

refuse during construction or if there is a structural fire in your home. To avoid exposures from 

some of these sources, construction refuse should not be burned near your home especially when 

children are out playing. 

Children may be exposed to CDDs from ingestion of contaminated soil or by contact of 

contaminated soil with their skin. However, skin contact with contaminated soil will result in 

much less CDDs entering the blood stream than if they ingest contaminated soil. Also, the amount 

of CDDs that will pass to the blood stream after eating contaminated soil will depend on the type 

of soil and on how tight the CDDs are bound to the soil. Children should be restricted from 

playing near any known hazardous waste sites. Some children eat a lot of dirt. Discourage your 

children from eating dirt or from putting their toys or other foreign objects in their mouths that 

may be contaminated with soil. Make sure that your children wash their hands frequently, 

especially before eating. Discourage your children from putting their hands in their mouths or 

other hand-to-mouth activities. 

Older children may be exposed to CDDs if they smoke cigarettes. Younger children and infants 

may be exposed by inhaling the second-hand smoke from their parents or other adult smokers. 

Parents should talk to their children about the dangers of smoking cigarettes. 

You and your children are likely to be exposed to very low amounts of CDDs in the diet 

particularly when you consume meat, milk, other dairy products, and fish. This represents the 

major source of background exposure to CDDs in most people. Children and adults should eat a 

balanced diet preferably containing low to moderate amounts of animal fats including meat and 

dairy products, and fish that contain higher amounts of CDDs and eat larger amounts of fruits, 

vegetables and grains. 
CDDs	 15 

1. PUBLIC HEALTH STATEMENT 

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

certain locations. A number of states have advisories for CDDs in fish and shellfish species; and 

one state has a wildlife advisory in effect for wood ducks. 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 waterbodies have restrictions, and a wildlife advisory will specify which hunting 

areas have restrictions. The advisory will tell you what types and sizes of fish or game are of 

concern. The advisory may completely ban eating fish or game or recommend that you limit the 

number of meals you eat of a certain species. For example, an advisory may tell you to eat a 

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

parts of the fish or game animal and how to prepare or cook the fish or game to decrease your 

exposure to CDDs. Fish and wildlife advisories are often stricter for pregnant women, nursing 

mothers, and young children. To reduce your children’s exposure to CDDs, obey all fish and 

wildlife advisories. Information on Fish and Wildlife Advisories in your state is available from 

your state Public Health Department, or state Natural Resources Department and signs may be 

posted in certain fishing and hunting areas. 

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

EXPOSED TO CDDs? 

Specific tests exist to measure CDD levels in samples of body fat, blood, and breast milk, but 

these tests are not routinely available. All people now have some levels of CDDs in their body fat 

and blood. Levels of 2,3,7,8-TCDD on a lipid basis are generally below 10 pg/g of lipid (ppt) in 

the blood and fatty tissue of the general population of the United States, and usually range from 3 

to 7 ppt. Levels higher than these indicate past exposure to above-normal levels of 2,3,7,8-TCDD. 

Although CDDs stay in the body fat for a long time (see Section 1.4), tests are not used to 

determine when exposure occurred, but can be used to estimate dose of the exposure if the time of 

exposure is known. 

Although exposure to 2,3,7,8-TCDD has been associated with adverse health effects in people, no 

one effect is specifically related to exposure to CDDs. There are laboratory tests which can 

indicate whether you have been exposed to CDDs, but these are costly and take weeks to perform 

and they cannot be used to predict whether you will develop harmful health effects. 
CDDs	 16 

1. PUBLIC HEALTH STATEMENT 

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 

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 CDDs include the following: 

The government has developed regulations and guidelines for 2,3,7,8-TCDD. These are designed 

to protect the public from the potential adverse health effects of the chemical. The Food and Drug 

Administration (FDA) recommends against consuming fish and shellfish with 2,3,7,8-TCDD 

levels greater than 50 ppt. Such levels have resulted in the closing of several commercial fishing 

areas. In addition, EPA has issued guidance to states on how to evaluate health risks to 

recreational and subsistence fishers, and how to issue fish consumption advisories when 

concentrations of CDDs in fish and shellfish pose a risk to these populations. Currently, 66 health 

advisories have been issued by 21 states restricting consumption of fish and wildlife contaminated 

with CDDs. EPA also has recommended limits on how much 2,3,7,8-TCDD can be present in 

drinking water. EPA advises that children should not have more than 1 nanogram 2,3,7,8-TCDD 

per liter of water (ng/L) (ppt) in 1 day, or more than 0.01 ng/L per day for long-term exposure. 
CDDs 17 

1. PUBLIC HEALTH STATEMENT 

For long-term exposure in adults, EPA recommends that there should not be more than 0.04 ng/L 

(ppt) in drinking water. 

Human milk can contain higher levels of CDDs than cow’s milk. Therefore, breast-fed infants can 

be exposed to higher levels of CDDs on a body weight basis than adults. The World Health 

Organization (WHO) has concluded that this risk to infants does not outweigh the positive 

biological and psychological benefits of breast-feeding at general population levels of dioxins. 

However, the specific concentration at which CDD levels in human milk would lead to harmful 

health effects in infants has not yet been determined. 

Regulation of many of the sources of CDDs appears to have been successful in reducing the 

amount of CDDs entering the ecosystem and in decreasing the potential for human exposure. 

EPA and ATSDR listed 2,3,7,8-TCDD as hazardous substance. Many regulations govern its 

destruction and disposal. See Chapter 7 for more information on regulations and guidelines. 

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
 

* Information line and technical assistance 

Phone: 1-800-44701544
 

Fax: (404) 639-6359
 

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. 
CDDs 18 

1. PUBLIC HEALTH STATEMENT 

* To order toxicological profiles, contact: 

National Technical Information Service
 

5285 Port Royal Road
 

Springfield, VA 22161
 

Phone: (800) 553-6847 or (703) 487-4650
 
 

CDDs 19 

2. HEALTH EFFECTS 

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 of the toxicology of chlorinated dibenzo-pdioxins (CDDs). 

CDDs are a class of related chlorinated hydrocarbons that are structurally similar. The basic structure is a 

dibenzo-p-dioxin (DD) molecule comprised of two benzene rings joined via two oxygen bridges at adjacent 

carbons on each of the benzene rings. There are eight homologues of CDDs, monochlorinated through 

octachlorinated. Each homologous class contains one or more isomers or congeners. The family of CDDs 

contains 75 congeners—2 monochlorodibenzo-p-dioxins (MCDD), 10 dichlorodibenzo-p-dioxins (DCDD), 

14 trichlorodibenzo-p-dioxins (TrCDD), 22 tetrachlorodibenzo-p-dioxins (TCDD), 14 pentachlorodibenzop-dioxins (PeCDD), 10 hexachlorodibenzo-p-dioxins (HxCDD), 2 heptachlorodibenzo-p-dioxins 

(HpCDD), and a single octachlorodibenzo-p-dioxin (OCDD). The seven 2,3,7,8-chlorine substituted 

congeners are the most toxic CDD congeners, with 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) 

being one of the most toxic and most extensively studied. This compound is often called "TCDD" or 

merely "dioxin" in the popular literature. Chlorinated dibenzofurans (CDFs) are structurally and 

toxicologically related chemicals as are certain “dioxin-like” PCBs; the reader is encouraged to 

consult the toxicological profile for CDFs (ATSDR 1994) and the toxicological profile for PCBs 

(ATSDR 1996) for information on the health effects associated with exposure to these groups of 

chemicals. 

2.1 HUMAN STUDIES 

This section presents information on human health effects, including those known to be associated and 

those possibly associated with exposure to CDDs (primarily 2,3,7,8-TCDD). Since limited data exist to 

assign a specific route of exposure (inhalation, oral, dermal) to human studies, the information in this 

section is organized by health effects—death, systemic, immunological, neurological, developmental, 

reproductive, 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). 
CDDs 20 

2. HEALTH EFFECTS 

The human studies discussed in this section are of populations known to reside or work in environments 

with above-background levels of CDDs and related compounds. Data on health effects in humans 

following exposure to CDDs have come from studies on accidental, occupational, and residential exposure 

and from studies on the use of 2,3,7,8-TCDD-contaminated pesticides. Because a number of these studies 

examined several end points, brief descriptions of these CDD-exposed populations are included in this 

section. Several factors complicate the interpretation of data regarding health effects in humans following 

exposure to CDDs; these include incomplete exposure data, concomitant exposure to other compounds, and 

a small number of participants, which limits the statistical power of the study to detect adverse health 

effects. Many of the studies on health outcomes following exposure to 2,3,7,8-TCDD and related 

compounds did not monitor exposure levels or internal dose. Surrogates of exposure were often used to 

identify potentially exposed populations and the level of exposure; some of the more commonly used 

surrogates include chloracne (a dermal condition generally indicative of appreciable exposure), potential 

exposure to phenoxy herbicides known to be contaminated with 2,3,7,8-TCDD, living in the vicinity of an 

accidental release of substances containing CDDs and related compounds, or an area with CDDcontaminated soil. Some of the more recent studies have used blood lipid CDD levels as a measure of 

internal dose in order to quantify exposure in individuals. In many of these studies, serum 2,3,7,8-TCDD 

levels were measured a number of years after exposure termination. CDDs are highly persistent lipophilic 

compounds which are resistant to biodegradation and have a great potential to bioaccumulate. Thus, a 

single chemical analysis of blood or adipose tissue represents a measure of past cumulative exposure to 

CDDs. With the assumptions of first-order kinetics for the elimination of 2,3,7,8-TCDD and an elimination 

half-life of 7–12 years, it is possible to extrapolate or adjust the serum or adipose tissue lipid concentration 

of 2,3,7,8-TCDD back to the time of the original excess exposure which may have occurred many years 

earlier, if the time of original exposure is known. Body burden or total dioxin amount can then be 

calculated from the serum 2,3,7,8-TCDD levels using the assumption that the concentration of 

2,3,7,8-TCDD in serum lipid is in equilibrium with total body lipid 2,3,7,8-TCDD concentrations and that 

in an average adult 22% of the body weight is lipid. Body burdens were calculated (see Table 2-1) for the 

human studies reporting serum (or tissue) lipid 2,3,7,8-TCDD concentrations. If only current serum 

2,3,7,8-TCDD levels were reported, then half-life-adjusted serum levels and body burdens were calculated 

using a half-life of 8.5 years (Michalek et al. 1996) and 22% body fat for a 70 kg adult (DeVito et al. 

1995). A number of studies have calculated half-life-adjusted serum 2,3,7,8-TCDD levels; in these cases, 

the reported half-life adjusted serum 2,3,7,8-TCDD levels were used to estimate body burdens. Egeland et 

al. (1994), Jansing and Korff (1994), and Wolfe et al. (1995) calculated half-life adjusted serum 

2,3,7,8-TCDD levels using a half-life of approximately 7 years. The 7.1-year half-life was derived from a 

study of 36 veterans involved in Operation Ranch Hand (Pirkle 


 

CDDs 23 

2. HEALTH EFFECTS 

et al. 1989). In a later study of Ranch Hand personnel, Michalek et al. (1996) calculated a mean serum 

2,3,7,8-TCDD half-life of 8.5 years using blood samples collected in 1982, 1987, and 1992 from more than 

300 veterans. In the studies of populations exposed several decades prior to measuring 2,3,7,8-TCDD 

levels in the blood, a difference of 1.4 years in the half-life can yield a large difference in estimated body 

burdens. For example, in the Manz et al. (1991) study, in which the workers were exposed for 33 years 

prior to 2,3,7,8-TCDD analysis, a body burden of 945 ng/kg at the time of exposure was calculated using a 

half-life of 8.5 years; using the 7.1-year half-life, the body burden would be 1606 ng/kg. See Section 2.3.4 

for a more complete discussion on estimating human body burdens and the overview to Section 2.5 for 

information on background exposure. 

Occupational exposure to CDDs most likely occurs mainly through inhalation of CDD-contaminated 

particles or dust and through dermal contact with solutions containing CDDs. However, data indicate that 

oral exposure to low levels of CDDs from contaminated food (including milk) represents the major route of 

environmental exposure for the general population and for people living in areas with known dioxin 

contamination (Connett and Webster 1987; Schecter et al. 1994a; Travis and Hattemer-Frey 1987). 

Occupational Exposure. Exposures to 2,3,7,8-TCDD, one of the most potent of the CDD congeners, 

have occurred occupationally in workers involved in the manufacture and application of trichlorophenols 

and the chlorophenoxy acid herbicides 2,4-dichlorophenoxyacetic acid (2,4-D) and 2,4,5-trichlorophenoxyacetic acid (2,4,5-T). Holmstedt (1980) has reviewed the history of industrial exposures that have 

occurred between 1949 and 1976, and Kogevinas et al. (1997) summarized recent data on these cohorts. 

The first reported cases of industrial poisoning were in 1949 at a 2,4,5-T producing factory in Nitro, West 

Virginia. 2,3,7,8-TCDD formation resulted from uncontrolled conditions in the reactor producing 

2,4,5-trichlorophenol (2,4,5-TCP) from tetrachlorobenzene in methanol and sodium hydroxide. 

Approximately 228 workers (including production workers, laboratory personnel, and medical 

personnel) were affected. Between 1949 and 1968, 3 other explosive releases were reported: 1 

involved 254 workers at the BASF AG facility in Ludwigshafen, Germany, in 1953 (Goldman 1972; 

Thiess et al. 1982; Zober et al. 1990, 1993); a second similar accident in 1963 involving 106 workers 

at Philips-Duphar facility in Amsterdam, Netherlands was a problem since the seriousness of the 

2,3,7,8-TCDD exposure was not anticipated and cleanup workers were exposed (Holmstedt 1980); 

and the third was an explosion in a 2,4,5-TCP manufacturing facility in Coalite, England, involving 

90 workers (May 1973). Holmstedt (1980) cited papers describing occupational exposure in 24 

additional factories producing TCPs or 2,4,5-T during the same period of time. Exposure data on 

most of these incidents were limited; various numbers of workers were affected, and many of the 
 

CDDs 24 

2. HEALTH EFFECTS 

published reports are anecdotal. Ott et al. (1994) measured serum 2,3,7,8-TCDD levels in 138 of the 254 

exposed workers several decades after the explosion at the BASF facility. More than 35 years after the 

explosion, serum 2,3,7,8-TCDD levels of <1–553 pg/g lipid were found; these correspond to serum levels 

of 3.3–12,000 pg/g lipid (calculated using a 7-year half-life) at the time of the accident. 

Some of the most comprehensive studies on occupational exposure were conducted by the National 

Institute for Occupational Safety and Health (NIOSH). They are cross-sectional studies of workers at U.S. 

chemical facilities involved in the manufacture of 2,3,7,8-TCDD-contaminated products between 1942 and 

1984 (Calvert et al. 1991, 1992; Egeland et al. 1994; Fingerhut et al. 1991; Sweeney et al. 1993). Serum 

2,3,7,8-TCDD levels were measured in the workers at two of the plants. The mean 2,3,7,8-TCDD serum 

lipid level in 281 production workers in the Newark, New Jersey, and Verona, Missouri, plants was 220 ppt 

(range, 2–3,390 ppt) 18–33 years after exposure termination; the referent group of 260 people who had no 

self-reported occupational exposure and were matched by neighborhood, age, race, and sex had a mean 

serum 2,3,7,8-TCDD level of 7 ppt (Calvert et al. 1992; Sweeney et al. 1993). Sweeney et al. (1990) 

estimated current mean lipid-adjusted 2,3,7,8-TCDD levels of 293.4 ppt (range, 2–3,390 ppt) in 103 

production workers at the New Jersey facility and 177.3 ppt (range, 3–1,290 ppt) in 32 workers at the 

Missouri facility; the mean half-life extrapolated levels (using a half-life of 7 years) were 2,664.7 ppt 

(range, 2–30,900 ppt) and 872.3 ppt (range, 3–6,100 ppt) in the two facilities, respectively. It should be 

noted that serum 2,3,7,8-TCDD levels were only measured in workers at these two facilities, and it is not 

known if the levels in these workers are reflective of serum 2,3,7,8-TCDD levels in workers at the other ten 

facilities. 

There are also a number of studies of chlorophenol and phenoxy herbicide applicators. Some of these 

studies used job histories, questionnaires, and interviews to determine which phenoxy herbicides the 

workers had used. Many of the studies did not measure exposure levels or internal doses; rather, 

2,3,7,8-TCDD exposure was assumed if the worker was exposed to a phenoxy herbicide known to be 

contaminated with 2,3,7,8-TCDD, such as 2,4,5-T. However, the level of exposure to these 2,3,7,8-TCDDcontaminated products was generally not determined. 

Residential/Environmental Exposures. Several incidents in which populations were exposed to 

potentially high levels of 2,3,7,8-TCDD include: an industrial accident that occurred during the production 

of 2,4,5-TCP at the ICMESA plant in Seveso, Italy and the spraying of roads and other places with a 

mixture of waste oil, including chemical waste generated during the manufacture of 2,4,5-TCP, in 

Missouri. 
 

CDDs 25 

2. HEALTH EFFECTS 

The most widely studied release of 2,3,7,8-TCDD primarily involving residential exposures occurred in 

Seveso, Italy in 1976. The ICMESA factory produced trichlorophenol by hydrolysis of 1,2,4,5-tetrachlorobenzene with alkali in ethylene glycol. The reactor overheated and the safety valve ruptured releasing a 

cloud containing primarily sodium trichlorophenate but also 2,3,7,8-TCDD. It was estimated that more 

than 1.3 kg of 2,3,7,8-TCDD was released into the atmosphere and that more than 17,000 people in a 

2.8-km2

 area adjacent to the facility were exposed. To investigate this accident, the contaminated area was 

separated into regions A, B, and R based on soil levels of 2,3,7,8-TCDD. The population sizes were 736, 

4,737, and 31,800 in areas A, B, and R, respectively. The respective mean (and maximum) surface soil 

levels of 2,3,7,8-TCDD were 230 (447) μg/m2

, 3 (43.8) μg/m2, and 0.9 (9.7) μg/m2 for areas A, B, and R, 

respectively. Dividing the populations into different zones based on soil levels has been criticized because 

it does not take into consideration actual-exposure levels and differences in within-zone 2,3,7,8-TCDD 

exposure (Mastroiacovo et al. 1988). Blood and tissue samples from exposed individuals have been saved 

and 2,3,7,8-TCDD levels in some of the original samples and in follow-up blood samples have been 

analyzed. Serum 2,3,7,8-TCDD levels ranged from 828 to 56,000 ppt (lipid adjusted) in 19 residents of 

zone A (Mocarelli et al. 1991). 

Various populations in Missouri were exposed to 2,3,7,8-TCDD in 1971 and 1972 as a result of spraying 

approximately 29 kg of 2,3,7,8-TCDD-contaminated waste oil on horse arenas, parking lots, and 

residential roads for dust control (Andrews et al. 1989). The oils originated from an industrial waste 

residue contaminated with 2,3,7,8-TCDD at levels of 305 ppm (Needham et al. 1991). An exposed group 

of 51 adults have been the subject of several studies. Adipose tissue levels, as well as paired human serum 

levels, were measured for 36 of these persons. Sixteen of the individuals were residents of areas where 

roadways had been sprayed and had mean 2,3,7,8-TCDD adipose tissue levels of 21.1 ppt (range, 

1.28–59.1 ppt) in 1985 (Andrews et al. 1989). Eight persons exposed to 2,3,7,8-TCDD at the horse arenas 

had a mean adipose 2,3,7,8-TCDD concentration of 90.8 ppt (5.0–577 ppt). In a comparison population of 

57 people with no known 2,3,7,8-TCDD exposure, 2,3,7,8-TCDD levels in the adipose tissue ranged from 

1.4 to 20.2 ppt, with a mean of 7.4 ppt. Although the population of study was not large, they were 

evaluated in depth for medical effects (Hoffman et al. 1986; Stehr et al. 1986; Webb et al. 1984). 

Exposures in Vietnam. During the Vietnam war, a program of aerial spraying of herbicides, code 

name Ranch Hand, was conducted in 10–20% of the Republic of Vietnam. During the 9 years of the 

program (1962–70), 19 million gallons of herbicides were dispersed. Six herbicides were used, with Agent 

Orange being the primary herbicide used (11 million gallons dispersed) (Wolfe et al. 1985). Agent Orange 
CDDs 26 

2. HEALTH EFFECTS 

was a 1:1 mixture of 2,4-D and 2,4,5-T in diesel oil and contained <1–20 ppm 2,3,7,8-TCDD as a 

contaminant. A number of studies have examined the possible association between Agent Orange exposure 

and adverse health effects in Vietnam veterans and Vietnamese residents living in the area of spraying. The 

results of a study comparing blood 2,3,7,8-TCDD levels in Vietnam veterans and the general U.S. 

population found that on average there was no significant difference between blood 2,3,7,8-TCDD levels 

between Vietnam veterans and comparison populations (CDC 1987). Thus, “service in Vietnam” or selfreported exposure to Agent Orange is not a reliable index of 2,4,5-T or 2,3,7,8-TCDD exposure. Studies of 

Air Force personnel participating in Operation Ranch Hand have found increased serum 2,3,7,8-TCDD 

levels in some of the persons (CDC 1987; USAF 1991). The median level in serum lipids for 888 Ranch 

Hand personnel was 12.4 ppt (range, 0 to 617.7 ppt) in contrast to 4.2 ppt (0-54.8 ppt) in a comparison 

group of 856 matched Air Force personnel (Wolfe et al. 1995). The median and high serum 2,3,7,8-TCDD 

levels would extrapolate to original serum levels of 43 and 3135 ppt, respectively, based on 20 years of 

elapsed time, and a half-life of 8.5 years. Since the tour of duty in Vietnam for the majority of U.S. 

veterans was generally less than 1 year, the military exposure was considered to be of intermediate duration 

if not stated otherwise in the original study. 

No studies were located, however, regarding health effects in humans exposed to CDDs by specific routes of 

exposure (e.g., inhalation, oral, dermal). In this profile, human health effects caused by combined exposure 

through various exposure routes are discussed separately from effects found in animals that were maintained 

under controlled experimental conditions (i.e., route, duration, and levels). 

2.1.1 Death 

None of the studies examining humans acutely exposed to high concentrations of 2,3,7,8-TCDD or other 

CDD congeners (as contrasted with long-term studies) reported acute instances of death. A number of 

epidemiology studies have investigated mortality in populations occupationally or environmentally exposed 

to 2,3,7,8-TCDD or chemicals contaminated with 2,3,7,8-TCDD or other CDD congeners. No significant 

increases in the number of deaths were observed in workers at phenoxy herbicide or chlorophenol 

manufacturing facilities (Cook et al. 1986, 1987b; Fingerhut et al. 1991; Ott et al. 1980, 1987; Zack and 

Suskind 1980) or in workers exposed to 2,3,7,8-TCDD as a result of the accident at the BASF AG facility in 

Germany (Ott and Zober 1996; Thiess et al. 1982; Zober et al. 1990). Additionally no increases in mortality 

were observed in the 10-year period after the Seveso accident (Bertazzi et al. 1989b) or in Vietnam veterans 

involved in Operation Ranch Hand (Wolfe et al. 1985). Although none of these studies found significant 
 

CDDs 27 

2. HEALTH EFFECTS 

increases in the overall mortality rate, several studies found statistically significant increases in causespecific mortality. For example, Flesch-Janys et al. (1995) found a significant risk of cardiovascular disease 

and ischemic heart disease mortality in workers exposed to 2,3,7,8-TCDD and other congeners during the 

BASF AG accident, and Fingerhut et al. (1991) found a significantly increased risk of cancer mortality in 

phenoxy herbicide and chlorophenol production workers. More complete descriptions of significant findings 

in the mortality studies are presented in the appropriate effect portions of Section 2.1. 

2.1.2 Systemic Effects 

The effects of 2,3,7,8-TCDD exposure in humans exposed in occupational or environmental settings have 

been described in several studies. Few studies provided precise exposure levels. However, for some 

cohorts, blood lipid 2,3,7,8-TCDD levels in samples collected shortly after exposure and stored frozen for 

several years have been analyzed. In other studies, the original blood levels of 2,3,7,8-TCDD were 

estimated using 2,3,7,8-TCDD levels measured in recent blood samples, the amount of time between 

exposure and blood sample collection, and a mean serum half-life of 5–12 years. 2,3,7,8-TCDD body 

burdens calculated from available serum lipid 2,3,7,8-TCDD levels are presented in Table 2-1. 

Respiratory Effects. Information regarding respiratory effects of CDDs in humans is limited. Effects of 

acute massive exposure in workers exposed to 2,3,7,8-TCDD in an industrial accident in Germany included 

bronchitis and laryngitis a few days after exposure, and hemorrhagic pleuritis 11 months after exposure 

(Goldman 1973). In an occupationally exposed group, decreased pulmonary function was found in smokers 

10 years after the cessation of manufacture of herbicides contaminated with 2,3,7,8-TCDD as compared with 

nonexposed smokers (Suskind and Hertzberg 1984). In contrast with the results of Suskind and Hertzberg 

(1984), Calvert et al. (1991) found no significant differences in ventilatory function between a group of 281 

workers employed 15 years earlier in the production of NaTCP, 2,4,5-T ester, or hexachlorophene and 260 

referents. At the time of the examination, the lipid-adjusted mean serum 2,3,7,8-TCDD concentration was 

220 ppt in the exposed workers compared to 7 ppt for the referents. In addition, there was no association 

between previous occupational exposure to 2,3,7,8-TCDD contamination and elevation in the incidence of 

chronic bronchitis or in the prevalence of chronic obstructive respiratory disease. Calvert et al. (1991) 

suggested that the disparity between their results and those of Suskind and Hertzberg (1984) may have been 

due to the fact that exposed workers in the Suskind and Hertzberg (1984) study were, on average, 10 years 

older than controls and to the potential exposure to 2,4,5-T acid dust in that study. The 2,4,5-T 
 

CDDs 28 

2. HEALTH EFFECTS 

acid was finished as a liquid as opposed to a powder in the plant studied by Suskind and Hertzberg (1984), 

thus limiting inhalation exposure. 

No respiratory effects were associated with exposure to 2,3,7,8-TCDD-contaminated herbicides in a group of 

Vietnam Air Force veterans involved in Operation Ranch Hand examined more than 10 years after the war 

(Wolfe et al. 1985). In the 1987 follow-up (USAF 1991), no association was found between the initial or 

current serum level of 2,3,7,8-TCDD and incidences of asthma, bronchitis, pleurisy, pneumonia, or 

tuberculosis; abnormal spirometric measurements were often associated with CDD blood levels, but 

according to the authors (USAF 1991), the differences in the mean level between high- and low-exposure 

subjects were not clinically important. The authors suggested that these findings may have been related to 

the association between 2,3,7,8-TCDD and body fat because obesity is known to cause a reduction in vital 

capacity. 

A recent follow-up of the cohort involved in the Seveso accident reported a significant increase in deaths 

(4 deaths) from chronic obstructive pulmonary disease in males from zone A (relative risk [RR]=3.7; 95% 

confidence internal [CI]=1.4–9.9) and in females from zone B (7 deaths; RR=2.4; 95% CI=1.1–5.1) (Pesatori 

et al. 1998). The excess found among zone A males was mainly detected in the first 5 years after the 

accident and mainly affected elderly men. As mentioned below under Cardiovascular Effects among this 

cohort, Pesatori et al. (1998) stated that stress related to the disaster experience could have precipitated early 

deaths among people with pre-existing chronic respiratory disease. The investigators also speculated that 

2,3,7,8-TCDD, through immunotoxic action, may have impaired protection and defense against episodes of 

respiratory infection, which play a major role in the natural history of chronic obstructive respiratory disease. 

The existing information suggests that acute exposure to high levels of CDDs may cause respiratory effects 

mainly as a response to upper respiratory tract irritation, but evidence from the numerous cohorts exposed to 

2,3,7,8-TCDD that have been studied suggests that the respiratory system is not a target for 2,3,7,8-TCDD 

toxicity. 

Cardiovascular Effects. Most earlier data indicated that exposure to CDDs does not induce 

cardiovascular effects (Bond et al. 1983; Moses et al. 1984; Suskind and Hertzberg 1984). In a crosssectional health survey in 1979 of 226 workers who had potential exposure to 2,3,7,8-TCDD from 1948 to 

1969 in 2,4,5-T production, 52% had chloracne (present for a mean of 26 years) which was used as a 

surrogate for heavy exposure. When the chloracne group was compared to workers without chloracne (low 
CDDs 29 

2. HEALTH EFFECTS 

exposure rather than unexposed workers), there was an increased reported incidence of angina and 

myocardial infarction; when these data were age-adjusted, the prevalence was not statistically increased 

(Moses et al. 1984). Examination of West Virginia TCP production workers revealed no increases in the 

prevalence of hypertension or coronary artery disease, abnormal ECG findings, atherosclerotic changes on 

chest X-ray, or blood pressure elevation (Suskind and Hertzberg 1984). 

Cardiovascular examination did not reveal any changes in 17 individuals who were treated for dermal lesions 

following acute exposure to 2,3,7,8-TCDD in the Seveso industrial accident (Reggiani 1980) or in a group of 

Missouri residents living in 2,3,7,8-TCDD-contaminated areas for a chronic period of time (mean 2.8 years 

in one area, 4.9 years in others) (Hoffman et al. 1986). In the 10-year period following the Seveso accident, 

there was a significant increase in the relative risk (RR) of death from chronic ischemic heart disease in men 

(RR=1.56; 95% CI=1.2–2.1), which was predominantly due to the increased risk during the first 5-year 

period (RR=1.76; 95% CI=1.2–2.5) (Bertazzi et al. 1989a). When the residents were divided into 

contamination zones, the relative risks of death from chronic heart disease in zones A, B, and R were 5.16 

(95% CI=1.3–20.9), 1.57 (95% CI=0.6–4.2), and 1.72 (95% CI=1.2–2.5), respectively, for the first 5-year 

period and 3.28 (95% CI=0.8–13.2), 0.96 (95% CI=0.4–2.6), and 1.61 (95% CI=1.2–2.2), respectively, for 

the 10-year period (Bertazzi et al. 1989b). In females, there was an increased risk of death from chronic 

rheumatic heart disease (RR=1.54; 95% CI=0.7–3.2) during the 10-year period (Bertazzi et al. 1989a), which 

was predominately due to the high relative risk in women living in zone A (RR=27.58; 95% CI=8.5–89.9) 

(Bertazzi et al. 1989b). Bertazzi et al. (1989b) noted that increased risk of cardiovascular disease deaths may 

have been due to post-accident stress rather than to 2,3,7,8-TCDD exposure. The results of a 5-year followup were recently published (Pesatori et al. 1998). The recent analysis reported five deaths in males from 

chronic ischemic heart disease (RR=3.0; 95% CI=1.2–7.3); three deaths in females from chronic rheumatic 

heart disease (RR=15.8; 95% CI=4.9–50.4); and three deaths, also in females, from hypertensive vascular 

disease (RR=3.6; 95% CI=1.2–11.4), all from zone A, the most severely affected area. Although these 

observations suggest an association between exposure to 2,3,7,8-TCDD and incidence of cardiovascular 

effects, they do not necessarily show that the effects were caused by 2,3,7,8-TCDD. As previously 

suggested by Bertazzi et al. (1989a), Pesatori et al. (1998) also indicates that the disaster experience with its 

burden of psychosocial stressors may have played a major role in the increased deaths found. 

No cardiovascular effects were observed in a group of Air Force veterans exposed to 2,3,7,8-TCDDcontaminated herbicides during the Vietnam war and examined several years post-exposure (Wolfe et al. 
CDDs 30 

2. HEALTH EFFECTS 

1985). However, a follow-up study of the Ranch Hand cohort reported increased mean diastolic blood 

pressure in those with current serum lipid 2,3,7,8-TCDD levels from 15 to 33.3 ppt, but not in subjects with 

higher 2,3,7,8-TCDD serum levels (USAF 1991). In addition, the proportion of abnormally low peripheral 

pulses in all Ranch Hand veterans, regardless of serum levels, was elevated relative to a comparison group. 

Also, arrhythmias detected on the electrocardiogram were significantly associated with 2,3,7,8-TCDD 

exposure, but there was no consistent dose-response relationship. 

Flesch-Janys et al. (1995) found significant increases in mortality from heart and circulatory diseases in 

workers exposed to 2,3,7,8-TCDD and other CDD congeners during the accident at BASF AG. Relative 

risks for cardiovascular disease and ischemic heart disease mortality were 1.96 (95% CI=1.15–3.34) and 2.48 

(95% CI=1.32-4.66), respectively, for workers with extrapolated serum lipid 2,3,7,8-TCDD levels of 

$348 pg/g (ppt) (current 2,3,7,8-TCDD levels were used to estimate 2,3,7,8-TCDD levels at the end of 

exposure). Additionally, statistically significant dose-response trends for increasing cardiovascular and 

ischemic heart disease deaths were found. The risk for cardiovascular and ischemic heart disease deaths also 

increased as the serum lipid CDD and CDF levels increased. However, the results from the Flesch-Janys et 

al. (1995) study are difficult to interpret since the percentage of chemical workers who died from cardiovascular disease was 38% compared to 49% for a referent group from a gas supply company with no known 

special exposure to CDDs/CDFs. An international study comprising 36 cohorts from 12 countries and a total 

of 21,863 workers exposed to phenoxyacid herbicides and chlorophenols followed from 1939 to 1992 

detected an increased risk for death from cardiovascular disease, especially ischemic heart disease (RR=1.67; 

95% CI=1.23–2.26) among the exposed workers (Vena et al. 1998). Risks did not differ across latency 

categories or by year of first exposure, but increased slightly by duration of exposure except for those with 

20 or more years of exposure. Vena et al. (1998) indicate, however, that the study was hampered by the 

reliance on mortality and the crudeness and inaccuracies of death certificate diagnoses. Furthermore, they 

noted that possible confounding effects from important risk factors for ischemic heart disease such as 

cigarette smoking, high fat diet, blood pressure, obesity, physical inactivity, and serum lipids cannot be ruled 

out. 

In contrast with the positive findings described in the studies summarized above, a recent study of 281 

workers employed 15 years earlier in the manufacturing of 2,4,5-trichlorophenol at two U.S. chemical plants 

and 260 unexposed referents found no significant association between 2,3,7,8-TCDD exposure and adverse 

cardiovascular effects (Calvert et al. 1998). The mean serum 2,3,7,8-TCDD concentrations (on a lipid basis) 

were 220 ppt in the workers and 7 ppt in the referents. Among the workers, the mean 2,3,7,8-TCDD 
 

CDDs 31 

2. HEALTH EFFECTS 

concentration when occupational exposure ceased was estimated to have been 1,900 ppt using a 7-year 

estimated half-life for serum 2,3,7,8-TCDD. Cardiovascular outcomes examined included myocardial 

infarction, angina, cardiac arrhythmias, hypertension, and abnormal, peripheral arterial flow. Calvert et al. 

(1998) indicated that although the study had sufficient statistical power to detect an elevated risk for cardiac 

arrhythmias, hypertension, and abnormal peripheral arterial flow, it had low power (approximately 50%) to 

detect an elevated risk for myocardial infarction and angina and concluded that further examination of the 

association between exposure to 2,3,7,8-TCDD and cardiovascular diseases is necessary. 

In summary, there is suggestive but inconclusive evidence of adverse cardiovascular effects in humans 

exposed to relatively high concentrations of CDDs. Increased deaths from chronic heart disease were 

observed among the Seveso cohort, but psychosocial factors could not be ruled out; no clear dose-response 

relationships were seen among the Ranch Hand cohort; increased deaths from heart and circulatory disease 

were reported among German workers exposed to CDDs; and no evidence of adverse cardiovascular effects 

was detected among U.S. workers exposed to CDDs. 

Gastrointestinal Effects. Earlier studies of individuals with exposure to substances contaminated with 

2,3,7,8-TCDD found significant elevations in self-reported ulcers (Bond et al. 1983; Suskind and Hertzberg 

1984), but a study of Vietnam veterans (USAF 1991) failed to find such effects. A more recent study 

evaluated the gastrointestinal effects of exposure to substances contaminated with 2,3,7,8-TCDD in an 

occupational cohort (Calvert et al. 1992). More than 15 years earlier, the workers were employed in the 

manufacture of trichlorophenol and its derivatives at 2 chemical plants. A total of 281 workers participated 

in the medical study; the control group consisted of 260 unexposed subjects who lived in the same 

communities as the workers. The participants underwent a comprehensive physical examination of the 

abdomen and rectum. The mean serum 2,3,7,8-TCDD level (on a lipid basis) for the workers was 220 ppt 

and was found to be highly correlated with years of exposure to 2,3,7,8-TCDD-contaminated substances; 

controls had a mean serum 2,3,7,8-TCDD concentration of 7 ppt. At the time of examination, the workers 

were not found to be at increased risk for any gastrointestinal diseases. Moreover, neither gastrointestinal 

ulcer disease nor gastritis was found to be significantly associated with any measure of 2,3,7,8-TCDD 

exposure, as determined by logistic regression analysis. The only significant finding from the physical 

examination was a statistically significant association between decreased anal sphincter tone and 

2,3,7,8-TCDD exposure. This, however, was attributed to examiner bias since all those that exhibited 

reduced sphincter tone were examined by the same physician (two physicians conducted the blind 
 

 

CDDs 32 

2. HEALTH EFFECTS 

examinations). The results of these studies suggest that there is no association between occupational 

exposure to 2,3,7,8-TCDD and gastrointestinal disease. 

Hematological Effects. Human studies regarding exposure to 2,3,7,8-TCDD or 2,3,7,8-TCDDcontaminated chemicals did not find any overt hematological effects after intermediate- (Wolfe et al. 1985) 

and chronic-duration exposures (Stehr et al. 1986). 

Contact with 2,3,7,8-TCDD-contaminated soil in Missouri by physical or recreational activities for 6 months 

at 100 ppb or for 2 years at 20–100 ppb resulted in a slight but statistically significant increase in total white 

blood cell (WBC) counts using a prevalence test (5.3% were increased above 10,000 WBC/mm3

 compared to 

0.7% for controls, but the increase was slight) (Hoffman et al. 1986). A follow-up study of the same 

population found no differences in the number of red blood cells, white blood cells, or platelets between 

exposed and nonexposed individuals (Evans et al. 1988). In a similar cohort, Stehr et al. (1986) found no 

consistent differences in hematology parameters in a high-risk group (68 persons) compared to a low-risk 

group (36 persons) except a slightly elevated platelet count. No significant differences in total leukocyte, 

granulocyte, or lymphocyte levels were observed between workers with high serum lipid CDD and CDF 

levels and workers with lower serum CDD and CDF levels (Neubert et al. 1993). 

A health study of Vietnam veterans involved in Operation Ranch Hand indicated an association between 

high initial and current serum 2,3,7,8-TCDD levels and increased erythrocyte sedimentation (Wolfe et al. 

1995), and an earlier study by Wolfe et al. (1985) indicated an increase in mean corpuscular volume; 

however, these changes were minor and were not observed in the 1991 follow-up (USAF 1991). Higher 

serum 2,3,7,8-TCDD levels were also associated with positive dose-response trends for increases in white 

blood cell and platelet levels. 

The existing information suggests that CDDs, at the body burdens seen in the studied populations, do not 

cause adverse hematological effects. 

Musculoskeletal Effects. The only information available comes from two anecdotal reports. In one of 

them, two individuals exposed to 2,3,7,8-TCDD in a horse arena that was sprayed with waste oil for dust 

control complained of painful joints (arthralgia) (Kimbrough et al. 1977). In the second case, a chemist 

exposed to 2,3,7,8-TCDD and 2,3,7,8-tetrabromo-p-dibenzo dioxin (2,3,7,8-TBDD) complained of muscle 
 

CDDs 33 

2. HEALTH EFFECTS 

pain in the lower extremities and back (Schecter and Ryan 1991). The role that 2,3,7,8-TCDD played in 

these cases, if any, is unknown. No further information was located. 

Hepatic Effects. Two of three laboratory workers synthesizing or working with 2,3,7,8-TCDD in the 

laboratory developed chloracne 8 weeks after potential acute exposures (Oliver 1975). Blood cholesterol 

levels (the only biochemical change) were elevated in all three workers and remained elevated for two years. 

Biochemical examinations were conducted on 55 male workers in Prague who were admitted into a hospital 

in 1968 and 1969 suffering from chronic 2,3,7,8-TCDD intoxication from exposure in a plant producing 

2,4,5-T (Pazderova-Vejlupkova et al. 1981). The first symptoms of intoxication included chloracne (present 

in the majority of the workers) and neurological symptoms; levels of exposure were never measured. 

Hypercholesterolemia was seen in 56% of the patients, hyperlipemia in 67%, hyperphospholipidemia in 

42%, diabetes mellitus in 8%, a low glucose tolerance level in 19%, increased α and γ globulins in 42% and 

uroporphyria in 21% (the study did not include a referent group) (Jirasek et al. 1976; Pazderova-Vejlupkova 

et al. 1981). Liver biopsies revealed mild steatosis, periportal fibrosis, and activated Kupffer cells in those 

examined. At 10 years postexposure, most of the biochemical changes were not detected; only cholesterol 

levels remained high (Pazderova-Vejlupkova et al. 1981). Transient alterations of liver function tests were 

reported in workers exposed to 2,3,7,8-TCDD following an industrial accident in Great Britain (May 1973). 

Jennings et al. (1988) found nonsignificant increases in serum cholesterol and triglycerides but no effect on 

gamma-glutamyltransferase (GGT), alanine aminotransferase (ALT), or creatinine in a group of 18 workers 

in England who had been exposed 17 years previously, when comparisons were made with a group of 15 

carefully matched controls. 

Mocarelli et al. (1986) conducted a 6-year study on clinical laboratory parameters of children exposed to 

2,3,7,8-TCDD following the Seveso accident. ALT, aspartate aminotransferase (AST), GGT, alkaline 

phosphatase, cholesterol, and triglycerides in plasma and delta amino levulinic acid in urine were monitored 

yearly in exposed and control groups beginning in June, 1977, approximately 1 year after the incident. The 

children were 6–10 years old at the time of the accident; 69, 528, and 874 resided in the A, B, and R zones, 

respectively. Chloracne was seen in 19, 0.7, and 4.6%, of the children in areas A, B, and R, respectively. 

Blood samples were drawn from 69, 83, and 221 children in areas A, B, and R, respectively. A slight 

increase in GGT and ALT occurred in the highest exposure group (based on zone of residences) compared to 

controls, but the values were not considered abnormal and returned to baseline levels within 3 years of the 

initial exposure. 2,3,7,8-TCDD blood levels have been more recently analyzed in about 30 of the subjects 

(Mocarelli et al. 1991). Similarly altered biochemical values (mainly increased serum transaminases and 
CDDs 34 

2. HEALTH EFFECTS 

GGT) were reported much earlier in individuals residing in an area of Seveso with average soil 

2,3,7,8-TCDD concentrations of 580.4 μg/m2

 (Pocchiari et al. 1979). A study of clinical chemistry 

parameters and urinary porphyrins in Missouri residents living in a 2,3,7,8-TCDD-contaminated area did not 

indicate any definitive changes that were of clinical importance in 154 persons (high-risk, based on soil 

levels of dioxins) exposed for up to 11 years (Webb et al. 1989). Multivariate regression analysis with 

number of years of residence as a surrogate of dose gave a positive trend for GGT and alkaline phosphatase. 

A medical survey of workers employed more than 15 years earlier in the manufacture of sodium trichlorophenol and its derivatives at two chemical plants found no evidence of an elevated risk for clinical hepatic 

disease at the time of examination (Calvert et al. 1992). The cohort consisted of 282 workers and 260 

unexposed matched controls. Exposure was assessed by measuring lipid-adjusted serum 2,3,7,8-TCDD 

levels. The mean serum 2,3,7,8-TCDD level in the workers was 220 ppt, compared with 7 ppt in the control 

group. The results from abdominal and rectal examination were unremarkable. Similarly, the results from 

blood and urine tests measuring liver function showed no statistically significant differences between 

exposed workers and controls, with the exception of a statistically significantly higher mean GGT level in 

workers. Also, workers were found to have a statistically significant elevated risk for an out-of-range GGT 

level compared with referents. However, multivariate analysis with logistic regression showed a statistically 

significant interaction between 2,3,7,8-TCDD exposure and lifetime alcohol consumption, indicating that the 

elevated risk for an out-of-range GGT was confined to those workers with a history of alcohol consumption 

and that the risk among the alcohol-consuming workers for an out-of-range GGT increased with increasing 

2,3,7,8-TCDD level. 

In a follow-up study, Calvert et al. (1996) examined the association between exposure to 2,3,7,8-TCDD and 

serum lipids. In the follow-up the authors chose not to adjust the 2,3,7,8-TCDD serum concentrations for 

total lipids to avoid the problems of interpretation that would arise when adjusting a covariate by the 

dependent variable. Consequently, the results obtained in this study cannot be compared directly with those 

from the Operation Ranch Hand study (see below). The median serum 2,3,7,8-TCDD concentration among 

the workers was 406.6 femtograms/g serum (fg/g) compared with 36.9 fg/g among the referents. The results 

of logistic regression analyses revealed an association between serum 2,3,7,8-TCDD and risk for an 

abnormally decreased HDL cholesterol concentration that approached statistical significance (p=0.09) after 

controlling for body weight index, use of beta-blocker medication, age, diabetes, and employment at the two 

plants. The concentration of 2,3,7,8-TCDD was not associated with having either an abnormal total 

cholesterol concentration or an abnormal total cholesterol/HDL cholesterol ratio. When the workers were 
 

CDDs 35 

2. HEALTH EFFECTS 

stratified into quartiles according to their serum 2,3,7,8-TCDD concentration, and after controlling for 

important confounders, those with the highest 2,3,7,8-TCDD concentrations (1,516–19,717 fg/g) had an 

elevated risk for an abnormal HDL cholesterol concentration (odds ratio [OR]=2.2; 95% CI=1.1–4.7). A 

small but statistically significant association between triglyceride concentration and serum 2,3,7,8-TCDD 

was also found (p=0.05) after controlling for gender, plant location, body weight index, cumulative cigarette 

consumption, use of beta-blocker medication, race, and diabetes. However, in the logistic analyses, 

abnormal triglyceride concentration was not associated with serum 2,3,7,8-TCDD concentration (p=0.21). 

Analysis by quartiles showed that workers with the highest 2,3,7,8-TCDD serum concentration had a 

statistically significant elevation in mean triglyceride concentration compared with the referent group. 

However, no significant trend was observed in the quartile analyses that evaluated risk for an abnormal 

triglyceride level. Calvert et al. (1996) concluded that the associations of serum 2,3,7,8-TCDD 

concentration with triglycerides and HDL (high density lipoprotein) cholesterol were small when compared 

with the influence of many other factors. 

A health study in Vietnam veterans involved in Operation Ranch Hand found no liver diseases linked to 

2,3,7,8-TCDD exposure, but biochemical examinations revealed a pattern suggestive of a subclinical effect 

on lipid metabolism (USAF 1991). Blood triglycerides showed a strong positive association with both the 

initial levels of 2,3,7,8-TCDD and the current serum levels; the authors indicated that this variable is highly 

sensitive to body fat, which was increased in the more highly exposed individuals. Cholesterol, HDL, and 

the cholesterol-HDL ratio also showed significant associations with 2,3,7,8-TCDD. 

In conclusion, hepatotoxic effects, such as elevated GGT levels and small alterations in lipid profile, have 

sometimes been observed in humans following exposure to high 2,3,7,8-TCDD levels. In general, the effects 

are mild and in some cases appear to have been transient. 

Information regarding hepatic effects observed in infants exposed perinatally to CDDs and structurally 

related compounds is presented in Section 2.5 under Developmental Effects. 

Renal Effects. A child who played in a sand box contaminated with waste oils containing 2,3,7,8-TCDD 

developed hemorrhagic cystitis and focal pyelonephritis (Kimbrough et al. 1977). Since chloracne was not 

seen and levels of 2,3,7,8-TCDD in the sand were not provided, the effects cannot be definitely attributed to 

2,3,7,8-TCDD exposure. No renal effects were reported in other individuals exposed at the same location. 

An early study in Missouri residents chronically exposed to a 2,3,7,8-TCDD-contaminated environment 
 

CDDs 36 

2. HEALTH EFFECTS 

found increased incidence of self-reported urinary problems, leukocyturia, and microscopical hematuria 

(Webb et al. 1984). However, the results of urinalysis on this group did not indicate any kidney effects 

(Hoffman et al. 1986; Stehr et al. 1986). No renal effects were found in a group of Vietnam veterans 

exposed to 2,3,7,8-TCDD in Agent Orange based on case histories and evaluation of five laboratory 

variables comparing Ranch Hand veterans and the various comparison groups (USAF 1991; Wolfe et al. 

1985, 1990). Kidney lesions have not been reported in any of the several studies on occupational exposure 

or for exposed cohorts in Seveso, Italy. These studies suggest that the kidney is not a target organ of 

2,3,7,8-TCDD toxicity in humans. 

Endocrine Effects. Jennings et al. (1988) examined thyroid function in a group of 18 workers exposed 

to 2,3,7,8-TCDD as a result of an industrial accident during the manufacture of 2,4,5-T. At the time of the 

study, 17 years after the accident, all the workers appeared healthy. No measure of exposure was provided. 

An unexposed group of 15 subjects served as controls. The end points monitored were serum thyroxine 

(T4), triiodothyronine (T3), and thyroid stimulating hormone (TSH). Without providing further details, the 

authors indicated that none of the subjects studied had biochemical evidence of thyroid dysfunction. A 

35-year follow-up study of workers exposed to 2,3,7,8-TCDD during the BASF accident found a significant 

increase in the incidence of thyroid disease, as compared to an age-matched referent group (Zober et al. 

1994). The workers were divided into two groups based on back-calculated (using a 7-year half-life) serum 

lipid 2,3,7,8-TCDD levels of $1,000 ppt and <1,000 ppt. For both groups of 2,3,7,8-TCDD-exposed 

workers, the incidence of thyroid disease was significantly higher than for the referent group, but did not 

differ between the two groups of workers. 

Endocrine function was assessed in Vietnam veterans involved in Operation Ranch Hand. A strong positive 

association was found between glucose intolerance or increased risk of diabetes and 2,3,7,8-TCDD serum 

levels (USAF 1991). The diabetes finding remained significant even after adjusting for body fat. Furthermore, subclinical effects in thyroid function (significant decrease in mean T3 uptake and increases in mean 

TSH) were reported for Operation Ranch Hand veterans with high current 2,3,7,8-TCDD serum levels of 

$33.3 ppt (USAF 1991). However, the magnitude of the differences was not considered physiologically 

significant. A follow-up study of Operation Ranch Hand veterans provides further information on the 

association between serum 2,3,7,8-TCDD levels and the incidence of diabetes mellitus and glucose and 

insulin levels (Henriksen et al. 1997). The cohort consisted of 989 exposed subjects and 1,276 comparison 

individuals who served in Southeast Asia (54) during the same period but who were not involved with 

spraying herbicides. Initial dioxin levels were computed using a first-order pharmacokinetic model with a 
 

CDDs 37 

2. HEALTH EFFECTS 

constant half-life of 8.7 years. Four exposure categories were defined: 1) comparisons, with current dioxin 

levels of #10 ppt; 2) background Operation Ranch Hand veterans, with current dioxin levels of #10 ppt; 

3) low category, with initial dioxin levels exceeding 10 ppt but #94.2 ppt; and 4) high category, with initial 

dioxin levels >92.4 ppt. Adjustments were made for age, race, and military occupation. The current median 

dioxin levels in the low and high categories were 15.0 and 46.2 ppt, respectively. The results of the analysis 

showed an increase in glucose abnormalities (RR=1.4; 95% CI 1.1, 1.8), diabetes prevalence (RR=1.5; 

CI 1.2, 2.0), and use of oral medications to control diabetes (RR=2.3; CI 1.3, 3.9) and a decrease in the timeto-diabetes onset with dioxin exposure. Serum insulin abnormalities increased in nondiabetics. Henriksen et 

al. (1997) pointed out that although some unknown confounder may not have been adjusted for, the strengths 

of the study included high participation and low attrition rates and accurate serum dioxin measurement and 

that, taken together, the results indicated a possible relation between dioxin exposure and diabetes mellitus, 

glucose metabolism, and insulin production. A follow-up evaluation of a cohort from the Seveso accident 

population found a significant increase in deaths from diabetes among women from zone B (RR=1.9; 95% 

CI=1.1–3.2) (Pesatori et al. 1998). Thirteen deaths were reported and 9 out of the 13 occurred in the second 

decade after the accident (RR=3.1; 95% CI=1.6–6.1). Pesatori et al. (1998) indicated that the fact that only 

women were affected might be explained by the systematically higher 2,3,7,8-TCDD concentrations in 

females than in males. 

In summary, the evidence available from epidemiological studies suggests that exposure to high 

concentrations of CDDs may induce long-term alterations in glucose metabolism and subtle alterations in 

thyroid function. 

Information regarding endocrine effects observed in infants exposed perinatally to CDDs and structurally 

related compounds is presented in Section 2.5 under Developmental Effects. 

Dermal Effects. The most commonly observed effect of 2,3,7,8-TCDD exposure in humans is chloracne 

(Jirasek et al. 1976; Kimbrough et al. 1977; May 1973; Oliver 1975; Reggiani 1980). Chloracne is 

characterized by follicular hyperkeratosis (comedones) occurring with or without cysts and pustules (Crow 

1978). Unlike adolescent acne, chloracne may involve almost every follicle in an involved area and may be 

more disfiguring than adolescent acne (Worobec and DiBeneditto 1984). Chloracne usually occurs on the 

face and neck, but may extend to the upper arms, back, chest, abdomen, outer thighs, and genitalia. In mild 

cases, the lesions may clear several months after exposure ceases, but in severe cases they may still be 

present 30 years after initial onset (Crow 1978; Moses and Prioleau 1985). In some cases lesions may 
CDDs 38 

2. HEALTH EFFECTS 

resolve temporarily and reappear later. Scarring may result from the healing process. Other chlorinated 

organic chemicals can also cause chloracne. 

Acute exposure to 2,3,7,8-TCDD in a chemical laboratory induced the development of chloracne in two of 

three individuals within 8 weeks of the exposure (Oliver 1975). Chloracne occurred in workers 

occupationally exposed to 2,3,7,8-TCDD during the manufacture of herbicides (Bond et al. 1989b; Moses 

and Prioleau 1985; Poland et al. 1971) and after industrial accidents in several locations throughout the 

world (Goldman 1973; May 1973; Moses et al. 1984; Pocchiari et al. 1979; Suskind and Hertzberg 1984). 

Accidental exposure to 2,3,7,8-TCDD in a 1949 explosion in a trichlorophenol plant in Nitro, West Virginia, 

resulted in an outbreak of severe chloracne. Moses et al. (1984) conducted a cross-sectional survey of 

workers in this plant in 1979. In reviewing the impact of the accident, the authors indicated that 117 workers 

had severe chloracne as a result of the explosion; however, 111 additional workers were found to have had 

chloracne prior to the explosion. A cross-sectional study of 226 workers in 1979 indicated that 52% had 

chloracne which persisted for 26 years, and in 29 subjects it was still present after 30 years. Blood levels 

were not measured, but the air dust in the plant was suspected to have contained 2,4,5-T contaminated with 

6 ppm 2,3,7,8-TCDD compared to 0.1 ppm in later years. Similarly, high incidences of chloracne were also 

found in other facilities (Jirasek et al. 1976; May 1973; Poland et al. 1971; Vos et al. 1978). Appearance of 

chloracne after accidental occupational exposure may be immediate or delayed; since workers may not 

always be removed from the work environment, the duration of exposure and total exposure is difficult to 

assess. 

Skin lesions from environmental exposures to 2,3,7,8-TCDD have been most thoroughly studied in the 

population exposed in Seveso, Italy. Reggiani (1980) described dermal lesions for 17 persons (primarily 

children) hospitalized shortly after the accidental release in Seveso. Acute lesions probably due to alkali and 

burns were observed immediately and had a duration of up to 2 months; chloracne in children occurred 

within 2 weeks (earliest occurrence was 3 days) and usually persisted for 8–26 months. Irritative lesions 

(characterized by erythema and edema of exposed areas, vesiculobollus and necrotic lesions, and 

papulonodular lesions) were observed in 447 people in Seveso 20–40 days after the accident, and 34 of these 

individuals later developed chloracne (Caputo et al. 1988). In 1976 and 1978, there were 193 childhood 

cases of chloracne and 17 of the most severe were in zone A where soil levels were the highest. Bisanti et al. 

(1980) reported that in zone A, 46 early cases (3–6 months) and 15 late cases (7–10 months) of chloracne 

were seen, and in zone B, 9 delayed cases were observed. In all zones, 50 early- and 143 late-appearing 
CDDs 39 

2. HEALTH EFFECTS 

cases of chloracne were reported (Caputo et al. 1988). In the 193 people with chloracne, the comedones and 

cysts progressively decreased in the 2 years following the accident (Caputo et al. 1988). In the most severe 

cases, regression of the lesions began at the end of 1978. All affected children were clear of lesions by 1982. 

Histological examination of the lesions from the limbs of severe chloracne patients revealed orthokeratotic 

hyperkeratosis with loss of adhesiveness, particularly near the follicular ostia; dilated follicular ostia filled 

with cornified lamellae; acanthosis; horny metaplasia with possible acrosyringeal cyst formation in the 

dermal and intradermal eccrine duct; and foreign body granulomas around the detached wall of the excretory 

ducts of some eccrine sweat glands (Caputo et al. 1988). Thirty of the 30,000 samples of serum collected 

and frozen in 1976 (10 zone A residents with the most severe cases of chloracne types 3 and 4 [chloracne 

was rated as type 1 for the mildest form to type 4 for the most severe cases], 10 former zone A residents who 

did not develop chloracne, and 10 controls from non-contaminated zones) were analyzed by Mocarelli et al. 

(1991). 2,3,7,8-TCDD blood levels (lipid adjusted) of 12,100–56,000 ppt were observed in 6 children with 

type 4 chloracne and levels of 828, 1,690, 7,420 ppt were found in 3 children with type 3 chloracne. In 

adults, levels of 1,770–10,400 ppt were associated with no chloracne. No chloracne was observed in 

Missouri residents who had adipose 2,3,7,8-TCDD levels of 5.2–59.1 ppt 16 years after exposure (using a 

half-life of 8.5 years, peak tissue levels of 6–204 ppt can be estimated) or in Operation Ranch Hand veterans. 

While there is a higher incidence of this disorder in those with higher serum 2,3,7,8-TCDD levels, 

interindividual variability makes it difficult to specify a dose that will result in chloracne (Needham et al. 

1991). 

The results of a further examination of Operation Ranch Hand veterans was recently published (Burton et al. 

1998). The cohort consisted of 930 exposed subjects and 1,200 comparison individuals who served in SEA 

during the same period but who were not involved with spraying herbicides. The authors examined the 

associations between serum dioxin levels and a) chloracne, b) occurrence of acne relative to the tour of duty 

in SEA, and c) anatomical location of acne after service in SEA. Initial dioxin levels were computed using a 

first-order pharmacokinetic model with a constant half-life of 8.7 years. Four exposure categories were 

defined: 1) comparisons, with current dioxin levels of #10 ppt; 2) background Operation Ranch Hand 

veterans, with current dioxin levels of #10 ppt; 3) low category, with current dioxin levels exceeding 10 ppt 

but #94.2 ppt; and 4) high category, with dioxin levels >92.4 ppt. Adjustments were made for age, race, and 

military occupation. The range of initial dioxin levels in the low and high categories was 27.7–94.1 ppt and 

94.2–3,290 ppt, respectively. Because physicians did not find any cases of chloracne among Operation 

Ranch Hand veterans at any physical examination and no cases were found via medical record review, the 

analysis was restricted to cases of acne. The results showed that among Operation Ranch Hand veterans 
CDDs 40 

2. HEALTH EFFECTS 

who had acne only after their service in SEA, the prevalence of acne at any location was increased in the 

high-exposure category, but the adjusted odds ratio relating acne in the eye-ear-temple location and dioxin 

category was increased for all three Operation Ranch Hand exposure categories. The increase was greatest 

in the background exposure category (OR=1.3; 95% CI=0.8–2.2). According to Burton et al. (1998), the 

results suggest that the Operation Ranch Hand exposure to dioxin, which was much lower than the Seveso 

exposure, was insufficient for the production of chloracne or that the exposure may have caused chloracne 

that resolved and was currently undetectable. 

The incidence of chloracne was examined in a group of 3 men and 4 women who were among 231 workers 

exposed to dioxins at a chemical factory in Ufa, Russia, approximately 25 years prior to blood collection in 

1991 and 1992 (Schecter et al. 1993). Five of the seven (three males and two females) were diagnosed with 

chloracne after working in the manufacture of 2,4,5-T contaminated with 2,3,7,8-TCDD between 1965 and 

1967. Blood analysis showed 2,3,7,8-TCDD levels (on a lipid basis) ranging from 36 to 291 ppt (mean 

185 ppt) in 1991 and 1992 compared with a mean of 4.4 ppt from a sample of 68 subjects from the general 

Russian population. Polychlorinated dibenzofurans and “dioxin-like” polychlorinated biphenyls (PCBs) 

were also detected, but it was estimated that in the workers, 2,3,7,8-TCDD contributed over 60% of the total 

dioxin equivalents (2,3,7,8-TCDD plus “dioxin-like” CDDs and PCBs). One of the workers diagnosed with 

chloracne had the lowest 2,3,7,8-TCDD blood concentration of the group, whereas two workers with higher 

levels did not display chloracne. This suggested that the presence of chloracne indicates exposure to dioxin 

(or similar chlorinated chemical), but its absence does not preclude such exposure, as noted by others 

(Mocarelli et al. 1991). Schecter et al. (1993) estimated that in the workers, the dioxin toxic equivalents 

(TEQ) in 1967 ranged from 226 to 1,707 ppt, assuming a 10-year half-life, and from 1,173 to 9,366 ppt 

assuming a 5-year half-life (see Section 2.5 for a detailed explanation on TEQs and toxicity equivalent 

factors [TEFs]). They also estimated the total 2,3,7,8-TCDD body burden for the workers to have been 

between 22 and 172 μg using a 5-year half-life and 4–30 μg using a 10-year half-life (mean present body 

burden was 3.2 μg versus 0.072 μg for general population). According to Schecter et al. (1993), this is the 

first reported incidence of chloracne in females with elevated dioxin blood levels from occupational 

exposure. 

A group of 8 individuals who had contracted chloracne between 1973 and 1976 while working in the 

manufacture of TCP or in the maintenance of a TCP plant were examined 15 years after the exposure 

(Jansing and Korff 1994). Slight residual chloracne was diagnosed in two subjects, but otherwise the 

workers were healthy. 2,3,7,8-TCDD levels in blood ranged from 163 to 1,935 ppt (lipid basis), and by 

assuming a half-life of 7 years, the authors estimated that the blood concentration during the exposure had 
 

 

CDDs 41 

2. HEALTH EFFECTS 

ranged from 545 to 9,894 ppt. It was found that the concentration of 2,3,7,8-TCDD in blood correlated well 

(r=0.93) with duration of chloracne if 2 subjects with a disposition to hypersensitive skin reactions were not 

included in the analysis. 

Other effects manifested as dermal changes have also been noted to accompany chloracne. In addition to 

chloracne, hyperpigmentation and hirsutism (also known as hypertrichosis or abnormal distribution of hair) 

were also reported in 2,3,7,8-TCDD-exposed workers (Jirasek et al. 1976; Oliver 1975; Poland and Smith 

1971; Suskind and Hetzberg 1984). In the cohort examined by Suskind and Hetzberg (1984), hypertrichosis 

was observed 25 years after exposure, particularly among workers with persistent chloracne upon clinical 

examination. In contrast, Moses et al. (1984) found no evidence of hypertrichosis, even though 31% of the 

exposed workers had evidence of residual chloracne. Webb et al. (1989) observed three cases of 

hypertrichosis, but not hyperpigmentation, among Missouri residents, one with serum levels of <20 pg/g and 

two with levels between 20 and 60 pg/g. However, neither condition was noted on examination among 

residents of the Quail Run Mobile Home Park (Hoffman et al. 1986). Actinic or solar elastosis was also 

observed among a group of workers diagnosed with active chloracne at the time of their examinations in 

1979 (Suskind and Hertzberg 1984). 

In conclusion, dermal effects, particularly chloracne, are the most commonly reported effects of 

2,3,7,8-TCDD exposure in humans because they are easy to identify. Additional information is needed to 

determine the level and frequency of 2,3,7,8-TCDD exposure needed to cause chloracne and whether 

individual susceptibility plays a role in the etiology. Also, chloracne in humans indicates CDD exposure, 

but lack of chloracne does not indicate that exposure has not occurred. Other dermal conditions reported 

include hypertrichosis, hyperpigmentation, and solar elastosis. 

Ocular Effects. Eye irritation, which correlated with severity of chloracne, was reported by Poland et al. 

(1971) among workers employed in a 2,4,5-T factory; however, the role of 2,3,7,8-TCDD, if any, cannot be 

determined. 

Body Weight Effects. Limited information was located regarding body weight effects in humans 

following exposure to CDDs. A transient weight loss was reported in a laboratory worker following an acute 

exposure to 2,3,7,8-TCDD (Oliver 1975). Weight loss associated with severe cases of chloracne was 

mentioned in a study among herbicide-manufacturing workers (Jirasek et al. 1976), but further information 

regarding weight loss was not provided. 
CDDs 42 

2. HEALTH EFFECTS 

2.1.3 Immunological Effects 

A limited number of studies have examined the immunotoxicity of 2,3,7,8-TCDD in humans. Most of these 

studies have found potential alterations in lymphocyte populations (e.g., T cells, B cells), cell surface 

markers (e.g., CD4RO+

, CD8+), or lymphoproliferative responses. The interpretation of these studies is 

limited by the lack of data correlating changes in immune function measurements and changes in host 

resistance to disease challenges (Kerkvliet 1995). Based on medical insurance records, Zober et al. (1994) 

found a significant increase in the incidence of infectious and parasitic disease in 2,3,7,8-TCDD-exposed 

workers in the 35-year period after the BASF accident. When the workers were divided into groups based 

on the severity of chloracne or back-calculated serum lipid 2,3,7,8-TCDD levels (assuming a half-life of 

7 years), the increase in infectious disease was significant only in the group with severe chloracne and the 

group with 2,3,7,8-TCDD levels of $1,000 ppt. Among the workers with severe chloracne, one disease 

subcategory, intestinal infections, accounted for the increased incidence of infectious diseases. A two-fold 

increase in the incidence of upper respiratory tract infections was also observed in the cohort. Dividing the 

workers into various groups did not result in evidence of increased respiratory infections in a particular 

group. Zober et al. (1994) also found a significantly higher incidence of appendicitis in 2,3,7,8-TCDD 

workers; it is not known if this effect was the result of immunotoxicity or a direct effect on the appendix. 

Although the results of this study suggest a relationship between 2,3,7,8-TCDD exposure and an increased 

risk of infection, the authors note that the difference may reflect differences in medical care use between the 

workers and the referent group. Jennings et al. (1988) examined immunological parameters in a group of 18 

workers 17 years after an industrial accident during the manufacture of 2,4,5-T in Coalite, England. At the 

time of the study all members of the cohort were apparently healthy. An unexposed group of 15 subjects 

served as controls. No measure of exposure was provided. The exposed group had a significantly increased 

number of natural killer (NK) cells, and concentration of antinuclear antibodies and immune complexes. 

Total lymphocytes, B cells, T cells, T-helper cells, T-suppressor cells, the lymphoproliferative response to 

phytohemagglutinin, and serum levels of immunoglobulins were similar between exposed and control 

groups. An immunological assessment of 41 subjects exposed to 2,3,7,8-TCDD-contaminated soil in Times 

Beach, Missouri was conducted by Webb et al. (1989). Sixteen participants had 2,3,7,8-TCDD adipose 

tissue levels <20 ppt, 13 had levels between 20 and 60 ppt, and 12 had levels >60 ppt. The highest level was 

750 ppt. Results from multiple regression analysis showed that increased 2,3,7,8-TCDD levels correlated 

with an increased percentage and total number of T lymphocytes. The increase was due to CD8+

 and T11+ T 

cells; CD4+

 T cells were not altered in percentage or number. The lymphoproliferative responses to T cell 

mitogens or tetanus toxoid were not altered, and neither was the cytotoxic T cell response. Serum 
CDDs 43 

2. HEALTH EFFECTS 

immunoglobulin A (IgA) was increased, but IgG was not. The study subjects did not exhibit any clinical 

disease associated with the 2,3,7,8-TCDD levels. In contrast with Webb’s findings, depressed cell-mediated 

immunity was found in residents from the Quail Run Mobile Home Park in Missouri (Hoffman et al. 1986); 

however, repeated examination of the group failed to confirm the observation (Evans et al. 1988). The levels 

of 2,3,7,8-TCDD in adipose tissue from this group were unknown. 

No indications of immune disease were found in a group of 8 subjects who had worked in a TCP 

manufacturing plant 15 years earlier and had elevated blood levels of 2,3,7,8-TCDD (163–1,935 ppt) at the 

time of the examination (Jansing and Korff 1994). The only significant observation was that individuals 

with the greater exposure (judged by 2,3,7,8-TCDD blood levels and duration of chloracne) showed a 

tendency of lower gamma-globulin levels. Neubert et al. (1993) examined surface receptors on lymphocyte 

subpopulations of workers with moderately increased body burden of 2,3,7,8-TCDD and of other CDDs and 

CDFs. The group consisted of 89 volunteers involved in decontamination work at a chemical plant in 

Hamburg, Germany. The volunteers were grouped according to their body burden, as defined by the CDD 

concentration in blood (on a lipid basis). Four groups were formed: a low-, medium-, and higher-level 

reference group, and the exposed group. Their respective median 2,3,7,8-TCDD blood concentrations were 

2, 5, 11, and 41.5 ppt. 2,3,7,8-TCDD was a minor contributor to the total dioxin equivalents. Regression 

analysis of the data showed some slight trends for some of the biomarkers, such as CD45R0+

. Except for 

one, all the trends were increases. The slight increase in the percentage of CD4+

CD45R0+ cells remained 

significant even after accounting for age-related changes. The authors concluded that altogether, the data did 

not provide any evidence for a decrease in cellular components of the human immune system in subjects 

with moderately increased CDD/CDF body burden. They also pointed out that adult humans appear to be 

less susceptible to this action of CDDs than adolescent marmoset monkeys. In a follow-up study, Neubert et 

al. (1995) examined lymphocyte proliferation responses (measured as 3

H-thymidine incorporation) in the 

same volunteers. They found no decrease in the capacity of 3

H-thymidine incorporation with any of the 

proliferation stimulators in the group with the increased 2,3,7,8-TCDD body burden, compared with the 

other groups. A recent study examined the long-term effects of 2,3,7,8-TCDD on immune function in 11 

industrial workers in Germany who had been exposed for several years to high doses of 2,3,7,8-TCDD 

20 years earlier (Tonn et al. 1996). Current 2,3,7,8-TCDD blood concentrations (lipid basis) ranged from 43 

to 874 ppt, compared with about 4 ppt as the average for the German population. Ten matched unexposed 

subjects served as controls. End points monitored included determination of lymphocyte subsets, and 

immunocompetence of T- and B-lymphocytes by mitogen-induced lymphoproliferation assays and by assays 

using sensitive mixed-lymphocyte cultures. At the time of the study, the workers were generally 
CDDs 44 

2. HEALTH EFFECTS 

healthy, although five persons still exhibited chloracne. Analysis of lymphocyte subsets showed no 

differences between the control and the 2,3,7,8-TCDD-exposed group. Moreover, there was no statistically 

significant difference in the response to mitogen stimulation between the two groups, and no correlation was 

found between individual 2,3,7,8-TCDD levels in the blood or the age of the person and the respective 

proliferative capacity of their lymphocytes. However, exposed subjects showed a reduced response to 

human lymphocyte antigen-allogeneic lymphocytes and interleukin-2-boosted proliferation. According to 

Tonn et al. (1996), this suppression is indicative of a reduced T-helper cell response, although the actual 

number of T-helper cells was not altered by 2,3,7,8-TCDD. The authors concluded that 2,3,7,8-TCDD 

immunosuppression is more likely mediated by a reduced functionality of individual cells rather than by a 

reduction in numbers of cells circulating in the blood. Tonn et al. (1996) further noted that the changes in 

immunocompetence observed did not correlate with obvious diseases related to severe immunodeficiency 

such as certain cancers and infections. 

In a study of 192 persons exposed to 2,3,7,8-TCDD (and CDFs) in a pesticide-producing factory in 

Germany, there was also no correlation between the levels of 2,3,7,8-TCDD in blood from exposed workers 

and the frequency of infectious diseases (Jung et al. 1998). The investigators also conducted a number of 

assays such as immunoglobulins, serum electrophoresis, monoclonal bands, surface markers, autoantibodies, 

lymphocyte proliferation, the rise of tetanus antibody concentrations after vaccination, and the in vitro 

resistance of lymphocytes to chromate to evaluate the morphologic and functional state of the immune 

system. A subgroup of 29 most highly exposed workers was compared to a control group of 28 subjects not 

exposed to above background levels of 2,3,7,8-TCDD. The median concentration of 2,3,7,8-TCDD in the 

workers was 217 pg/g blood lipid (range, 33.6–2,252) compared to 3.9 pg/g in the controls (range, 2.9–6.0). 

There was no significant correlation between the current 2,3,7,8-TCDD concentrations and alterations in any 

of the immune parameters among the entire exposed group. In addition, the results of the tetanus vaccination 

and the chromate resistance test were not correlated with exposure to 2,3,7,8-TCDD. The only significant 

finding was that the chromate resistance of lymphocytes stimulated with phytohemagglutinin of highly 

exposed persons was significantly lower than that for the control group. This, according to Jung et al. 

(1998), suggested that the function of lymphocytes can be stressed and possibly impaired by high exposure 

to 2,3,7,8-TCDD. A separate report on the same group of workers found no significant exposure-related 

alterations in the phenotype and function of peripheral blood mononuclear cells (PBMC) as judged by the 

proportions of CD3, CD4, or CD8+

 T-lymphocytes; of CD16+ natural killer cells; and of CD19+ Blymphocytes (Ernst et al. 1998). However, in the 2,3,7,8-TCDD exposed workers, the proportion of CD8+ 

memory T-cells (CD45RO+) was significantly higher, and that of lymphocytes with naive phenotype 
CDDs 45 

2. HEALTH EFFECTS 

(CD45RA+) was significantly lower than in PBMC of the control group. Also, in vitro tests of T-cell 

activation showed a significantly reduced interferon γ release in diluted whole blood cultures but not in 

isolated PBMC cultures of the TCDD-exposed cohort when T-cells were stimulated with tetanus toxoid. 

Based on these results, Ernst et al. (1998) suggested that exposure to high concentrations of 2,3,7,8-TCDD 

can partially impair in the blood milieu those T-cell/monocyte interactions that are essential for antigenspecific T-cell responses, whereas isolated PBMC in the same donor appeared functionally less affected. 

The immune status of children exposed to 2,3,7,8-TCDD in the Seveso incident was also examined 

(Mocarelli et al. 1986). The group consisted of 44 children, 20 of whom had chloracne. The results of the 

testing showed no abnormalities in serum immunoglobulin concentrations, levels of circulating complement, 

or lymphoproliferative responses to T- and B-cell mitogens. However, a different cohort of 2,3,7,8-TCDDexposed children examined 6 years after the explosion showed a significant increase in complement protein 

levels, which correlated with the incidence of chloracne (Tognoni and Bonaccorsi 1982). The children also 

had increased numbers of peripheral blood lymphocytes and increased lymphoproliferative responses. No 

specific health problems were correlated with exposure to 2,3,7,8-TCDD in these children. The findings 

from the Tognoni and Bonaccorsi (1982) study suggest that chloracne is a more sensitive toxicological 

endpoint than immunological effects because alteration in complement levels is a subclinical effect and 

correlated with the incidence of chloracne. 

A health study of Vietnam veterans involved in operation Ranch Hand did not find any correlations between 

clinically significant immunological alterations and serum 2,3,7,8-TCDD levels (USAF 1991). The only 

significant positive association with exposure to 2,3,7,8-TCDD was an increase in serum IgA levels. The 

authors suggested that this alteration was indicative of a subtle inflammatory process, but there was no other 

evidence for an inflammatory response. 

Parameters of immunocompetence were assessed in a group of 23 men with high fish consumption from the 

Baltic Sea (Svensson et al. 1994). Twenty men with almost no fish consumption served as controls. The 

parameters examined included WBC, lymphocyte levels, serum immunoglobulin levels, and lymphocyte 

subsets. The mean dioxin equivalent concentration (TEQ, includes CDDS, CDFs, and dioxin-like PCBs) in 

blood (lipid basis) from fish eaters (n=7) was 64 pg TEQ/g (range, 18–88 pg/g) compared to 21 pg TEQ/g 

(18–33 pg/g) for controls (n=4). CDDs and CDFs were the major contributors to the total dioxin equivalents 

in blood. Of all the parameters examined, only the level of NK cells was reduced in fish eaters, but the 

difference between groups was not statistically significant. No correlation was found between blood levels of 
CDDs 46 

2. HEALTH EFFECTS 

2,3,7,8-TCDD and the reduction in NK levels, but weak correlations existed between the latter and some nonortho-PCB congeners and p,pN-DDT. 

In conclusion, while some studies are suggestive, no consistent exposure-related immunological effects have 

been observed in human populations exposed to levels of CDDs several orders of magnitude higher than 

background exposure. This may in part be due to the lack of functional assays of immune competence in 

humans. 

Information regarding immunological effects observed in infants exposed perinatally to CDDs and 

structurally related compounds is presented in Section 2.5 under Developmental Effects. 

2,3,7,8-TCDD human body burdens calculated from available serum lipid 2,3,7,8-TCDD levels are presented 

in Table 2-1. 

2.1.4 Neurological Effects 

Symptoms of intoxication including lassitude, weakness of the lower limbs, muscular pains, sleepiness or 

sleeplessness, increased perspiration, loss of appetite, headaches, and mental and sexual disorders were 

reported in several of the 117 workers with severe chloracne who had been exposed to 2,3,7,8-TCDD in an 

occupational setting (Moses et al. 1984; Suskind 1985). Neurological symptoms persisted in these individuals 

for up to 10 years based on an increased incidence of sensory findings. Similar symptoms of intoxication 

were observed in a trichlorophenol factory in Czechoslovakia (Jirasek et al. 1976) for which a 10-year followup of 55 of the 80 affected workers was conducted (Pazderova-Vejlupkova et al. 1981). At autopsy, damage 

to peripheral neuron Schwann cells was confirmed in a worker who died (Jirasek et al. 1976). 

Polyneuropathy and encephalopathy were found in 23 and 7% of the surviving workers, respectively (study 

did not include a referent group). Most patients suffered from peripheral neuronal lesions of the lower 

extremities (confirmed by electromyography). Encephalopathy developed in older individuals (aged 50 and 

above) and was accompanied by an organic psychosyndrome due to atherosclerosis of cerebral arteries. The 

most-affected patient developed paroxysms of temporal epilepsies and external and internal hydrocephalus. 

Patients with polyneuropathy did not improve during 10 years postexposure (Pazderova-Vejlupkova et al. 

1981). Similarly, neurological effects that included peripheral neuropathy, sensory impairment, tendency to 

orthostatic collapse, and reading difficulties were reported in workers exposed to 2,3,7,8-TCDD in an 

industrial accident in Germany (Goldman 1973). 
CDDs 47 

2. HEALTH EFFECTS 

Several studies have investigated the neurological effects of 2,3,7,8-TCDD in Seveso residents. Residents in 

zone A and zones B and R participated in a neurological screening test in 1977 (Pocchiari et al. 1979). Of the 

446 residents in zone A, 6.7% and 3.1% had evidence of “idiopathic clinical neurologic damage” or 

“idiopathic subclinical neurologic damage” respectively, as compared to 1.2% and 1.2%, respectively, in the 

255 residents in zones B or R. No definitions of clinical or subclinical damage were provided. The authors 

did note that the most frequently observed effects occurred in the peripheral nervous system (particularly 

reduced nerve conduction velocity, which was considered a subclinical effect). In 1978, 205 residents in zone 

A were re-tested. At this time, 11.7% had clinical damage and 4.9% had subclinical damage. No relationship 

between chloracne and neurological symptoms was found (Pocchiari et al. 1979). In another study of zone A 

residents, 22 cases of peripheral neuropathy were observed in the 470 residents examined in 1977 (prevalence 

rate of 8.9%, 95% confidence interval [CI] of 6.2 to 11.6) (Filippini et al. 1981). Peripheral neuropathy was 

diagnosed based on the occurrence of neurological symptoms (paresthesia, hypesthesia, pain, and 

hyposthenia), clinical signs (superficial and deep sensory impairment, muscular weakness, and tendon hypoor areflexia), and/or electrophysiological alterations. During a re-evaluation in 1978, 26 of the 308 subjects 

had neurological symptoms, clinical signs, and/or altered electrophysiological readings and 16/308 had two or 

more electrophysiological abnormalities (at least one being altered nerve conduction velocity). The 42 

subjects with evidence of peripheral neuropathy were divided into three groups: 1) residents with existing 

predisposing factors (e.g., excess alcohol consumption, diabetes, nutritional diseases), 2) residents with 

potential high exposure to 2,3,7,8-TCDD (assessed via increased serum levels of several hepatic enzymes 

[GGT, AST, ALT] and/or chloracne), and 3) remaining residents. The prevalence rate ratio was significantly 

higher in the first two groups (2.6, 95% CI=1.2–5.6 for predisposing factors group and 2.8, 95% CI=1.2–6.5 

for high 2,3,7,8-TCDD exposure group) as compared to the third group. In a 6-year follow-up of 152 subjects 

with chloracne in Seveso, Barbieri et al. (1988) found no clear-cut peripheral neuropathy but an increase in 

clinical and electrophysiological signs of peripheral nervous system involvement when compared to 123 ageand sex-matched controls. In 1985, 141 of these subjects were re-examined (Assennato et al. 1989). No 

statistically significant alterations in motor nerve conduction velocity of the median or peroneal nerves or 

sensory nerve conduction velocity of the sural nerve, as compared to 167 matched controls, were observed. 

No significant increases in neurological effects (based on self-reported neurological effects and neurological 

examination) were observed in 68 Missouri residents with potential high risk exposure to 2,3,7,8-TCDD as 

compared to 36 residents with low-risk exposure (Stehr et al. 1986). Abnormal neurological symptoms were 

observed in a group of 41 Missouri residents with measured 2,3,7,8-TCDD serum lipid levels (Webb et al. 
CDDs 48 

2. HEALTH EFFECTS 

1989). The symptoms included abnormal pain sensation in lower extremities, abnormal vibratory sensation, 

and abnormal reflexes. However, the distribution of these effects among residents with serum lipid 

2,3,7,8-TCDD levels of <20 ppt, 2–60 ppt, or >60 ppt was not dose-related. 

Nerve conduction velocities were measured in 55 employees of a 2,4,5-T and 2,4-D manufacturing facility in 

Jacksonville, Arkansas (Singer et al. 1982). Statistically significant decreases in median motor nerve and 

sural nerve conduction velocities were observed, as compared to a control group of workers not exposed to 

phenoxy herbicides. No effect on median sensory nerve conduction velocity was found. Sural nerve 

conduction velocity was significantly inversely correlated with duration of employment. 

Psychological effects have been associated with 2,3,7,8-TCDD exposure in some human studies. Personality 

changes were reported following acute exposure (Oliver 1975). Depression (Levy 1988; Wolfe et al. 1985), 

hypochondria, hysteria, and schizophrenia (Wolfe et al. 1985) were found more often in Vietnam veterans 

exposed to 2,3,7,8-TCDD-contaminated herbicides than in the control group of veterans. A battery of tests 

used for the psychological evaluation included Minnesota Multiphasic Personality Inventory, Cornell, 

Wechsler Memory Scale I, Wechsler Adult Intelligence Scale, Wide Range Achievement Test, and HalsteadReitan Neuropsychological Battery. Peper et al. (1993) reported that the results of neuropsychological testing 

of 19 persons living in an area with high concentration of dioxins in soil in Germany were within the range of 

values expected from standardized age samples. However, increased levels of dioxins in blood (but not 

substantially different from a national sample) were associated with a reduction of cognitive performance in 

verbal conceptualization, mnemonic organization of verbal and visual stimuli, psychomotor slowing, and a 

variety of subjective complaints. The concentration of CDDs (including CDFs) in blood (lipid basis) ranged 

from 16.1 to 80.4 ppt (TEQ), with a mean of 31 ppt. The authors recognized, however, that given the small 

number of subjects and the relatively low amount of exposure (monitored by blood levels) the results must be 

interpreted with caution. 

A health study in Vietnam veterans involved in Operation Ranch Hand reported that elevated serum 

2,3,7,8-TCDD levels were not associated with any neurological disease and were not related to verified 

psychological or sleep disorders (USAF 1991). In a more recent study, Sweeney et al. (1993) compared the 

prevalence of chronic peripheral neuropathy in a group of workers employed 15 years earlier in the 

manufacture of sodium trichlorophenol and its derivatives at two chemical plants. The cohort consisted of 

265 exposed workers and an unexposed matched comparison group of 244 subjects. Exposure was assessed 

by measuring lipid-adjusted serum 2,3,7,8-TCDD levels. The neurologic status was evaluated through a 
CDDs 49 

2. HEALTH EFFECTS 

standardized neurological examination and electrophysiologic measurements and quantitative sensory tests of 

thermal and vibratory sensitivity. The results showed that the workers had a significantly higher mean serum 

2,3,7,8-TCDD level (220 ppt) compared to controls (7 ppt). In both worker and referent groups, 32% met the 

case definition for peripheral neuropathy; however, logistic regression analyses revealed that serum 

2,3,7,8-TCDD level was not related to peripheral neuropathy. The results suggested that despite continued 

elevated serum 2,3,7,8-TCDD levels, peripheral neuropathy is not a long-term sequela of exposure to TCDDcontaminated chemicals. Nevertheless, the authors (Sweeney et al. 1993) indicated that the study could not 

preclude the occurrence and subsequent resolution of acute effects caused by high exposure, as observed in 

Seveso (Fillippini et al. 1981; Pocchiari et al. 1979) and possibly in early case reports (Goldman 1973; Jirasek 

et al. 1976; Oliver 1975). 

The overall evidence from case reports and epidemiological studies showed that exposure to CDDs is 

associated with signs and symptoms of both central and peripheral nervous system shortly after exposure. In 

some cases, the effects lasted several years. However, evaluation of individuals 5 to 37 years after the last 

exposure has not revealed any long-lasting abnormalities. 

Information regarding neurological effects observed in infants exposed perinatally to CDDs and structurally 

related compounds is presented in Section 2.5 under Developmental Effects. 

2.1.5 Reproductive Effects 

A number of studies have investigated the possible association between 2,3,7,8-TCDD exposure and 

reproductive toxicity in humans. A common limitation of many of these studies, particularly those conducted 

prior to the development of assays to quantify serum and adipose levels of 2,3,7,8-TCDD, is the lack of 

adequate exposure data (Sweeney 1994). 

The effects of 2,3,7,8-TCDD exposure on gonadal function (production of germ cells and secretion of sex 

hormones) has not been extensively investigated. A health study in Vietnam veterans involved in Operation 

Ranch Hand found a significant association between decreased testicular size and serum 2,3,7,8-TCDD 

levels, but no association was found for low serum testosterone levels (USAF 1991). When the Operation 

Ranch Hand cohort was re-examined in 1992 using ultrasound methodology, no significant alterations in 

testes size were found (Henriksen et al. 1996). No alterations in sperm count or the percentage of abnormal 

sperm were observed in Vietnam veterans involved in Operation Ranch Hand (Wolfe et al. 1985). No 
CDDs 50 

2. HEALTH EFFECTS 

consistent alterations in testosterone levels, follicle-stimulating hormone (FSH) levels, luteinizing hormone 

(LH) levels, testicular abnormalities, sperm abnormalities, and sperm counts were found in a follow-up study 

of the Operation Ranch Hand cohort; reproductive parameters were assessed in 1982, 1987, and 1992 

(Henriksen et al. 1996). In workers at two 2,4,5-trichlorophenol manufacturing facilities, serum 

2,3,7,8-TCDD levels were positively correlated with follicle stimulating hormone and luteinizing hormone 

levels and inversely correlated with total testosterone levels (Egeland et al. 1994). However, the magnitude 

of the change in hormone levels per unit of increase in serum 2,3,7,8-TCDD levels was small. The 

prevalence of high luteinizing hormone levels, high follicle stimulating hormone levels, and low testosterone 

levels was significantly increased in workers with half-life extrapolated serum lipid 2,3,7,8-TCDD levels of 

$140 pg/g, $1,860 pg/g, and $140 pg/g, respectively (based on 2,3,7,8-TCDD levels extrapolated at the time 

occupational exposure ceased and assuming a 7.1-year half-life). The study authors note that both low 

testosterone and high LH levels were not observed in the same individuals. Although the number of workers 

with elevated or depressed hormone levels was significantly higher than in the referent group, the adjusted 

mean hormone levels (adjusted for age, body mass index, alcohol consumption, smoking, and diabetes 

mellitus) were within 20% of referent values. 

A number of studies have examined pregnancy outcomes following paternal exposure or paternal and 

maternal exposure to 2,3,7,8-TCDD. No significant alterations in the incidence of spontaneous abortions 

were found in several studies of Vietnam veterans. In a case-control study conducted by Aschengrau and 

Monson (1989), no association was observed between paternal military service in Vietnam and the risk of 

spontaneous abortion (odds ratio [OR] of 0.88, 95% confidence interval [CI] of 0.42–1.86). A limitation 

of this study is that service in Vietnam is not an adequate exposure surrogate for 2,3,7,8-TCDD exposure; 

CDC (1988) found that 2,3,7,8-TCDD body burdens in Vietnam veterans were not significantly different 

than background levels. In a study of Air Force personnel involved in Operation Ranch Hand, no 

relationship between paternal 2,3,7,8-TCDD exposure (as measured by serum 2,3,7,8-TCDD levels) and 

the occurrence of spontaneous abortions (relative risk [RR] of 1, 95% CI=0.7–1.3 for veterans with halflife adjusted serum lipid 2,3,7,8-TCDD levels of >110 ppt) or stillbirths (RR of 1.8, 95% CI=0.7–4.7 for 

veterans with half-life adjusted serum 2,3,7,8-TCDD levels of <110 ppt, current levels >10 ppt; and RR of 

0.3, 95% CI=0–2.3 for veterans with half-life adjusted serum 2,3,7,8-TCDD levels of >110 ppt) were 

observed (Wolfe et al. 1995). No significant alterations in the relative risk of stillbirths (3 stillbirths 

observed in 2,3,7,8-TCDD-exposed group compared to 0 in control group) or miscarriages (RR of 1.19, 

90% CI=0.58–2.45) were observed in the wives of New Zealand 2,4,5-T applicators (Smith et al. 1982). It 

should be noted that many of the wives were occasionally exposed while helping with spray activities and 
CDDs 51 

2. HEALTH EFFECTS 

while washing contaminated clothing. An increased incidence of spontaneous abortions was reported in 

women living close to a herbicide manufacturing factory in Sweden (Forsberg and Nordstrom 1985). The 

residents were exposed to phenoxy acids, chlorophenols, 2,3,7,8-TCDD, and dibenzofurans which were 

released into the soil and groundwater. The small number of cases in the exposed cohort and concomitant 

exposure to several other chemicals limits the conclusions which can be drawn from this study. 

Several studies have reported alterations in the sex ratio of children of men and women exposed to high 

levels of CDDs. Mocarelli et al. (1996) observed decreases in the sex ratio of children born to parents 

living in area A at the time of the accident in Seveso, Italy. More females than males (48 females versus 

26 males; normal ratio is 100 females to 106 males) were born between April 1977 (9 months after the 

accident) and December 1984. Between 1985 and 1994, there was no significant alteration in sex ratio (64 

females, 60 males). In nine families in which both parents lived in area A at the time of the accident, 

serum lipid 2,3,7,8-TCDD levels (blood samples collected at the time of the accident) ranged from 126 to 

1,650 ppt and 104 to 2,340 ppt for the mothers and fathers, respectively. Basharova (1996) reported an 

alteration in sex ratio (more females than males) in children of workers exposed to 2,3,7,8-TCDDcontaminated 2,4,5-T at a production facility in Ufa, Russia. No additional information on the percentage 

of male and female children or statistical analysis of data was provided. Similarly, more females than 

males (51.4% versus 48.6%; 19,675 births) were born to 9,512 male workers exposed to chlorophenate 

wood preservatives contaminated with CDDs (Dimich-Ward et al. 1996). James (1997) statistically 

analyzed the results of this study and found that the sex ratio was statistically significant, as compared to 

the expected Caucasian live birth sex ratio of 0.514. Stockbauer et al. (1988) did not find an alteration in 

the sex ratio (53.% of children were males versus 51.2% in nonexposed controls) among the children of 

mothers potentially exposed to CDDs in the Times Beach incident. This retrospective cohort study did 

not monitor where the mothers or fathers lived at the time of conception. Although three studies have 

found alterations in the sex ratio among the offspring of exposed individuals, a causal relationship between 

CDD exposure and alterations in the sex ratio cannot be inferred at this time. Further investigation of this 

positive association is clearly warranted. 

In an inadequately reported study, Phuong et al. (1989a) reported a statistically significant increase in the 

incidence of hydatidiform mole in families living in southern Vietnam potentially exposed to 2,3,7,8-TCDD 

contaminated herbicides as compared to a group of Ho Chi Minh City residents presumably never exposed to 

2,3,7,8-TCDD-contaminated herbicides. In contrast, a case-control study by Ha et al. (1996) did not find a 
CDDs 52 

2. HEALTH EFFECTS 

significant association between exposure to 2,3,7,8-TCDD-contaminated herbicides and the occurrence of 

complete hydatidiform mole or choriocarcinoma among women living in southern Vietnam. 

The results of the human reproductive toxicity studies are inconclusive. In studies which measured 

2,3,7,8-TCDD levels (Egeland et al. 1994; Henriksen et al. 1996), mixed results were found for alterations in 

hormone levels. The Egeland et al. (1994) study of the NIOSH cohort found significant alterations in 

testosterone and gonadotropins, but the alterations were small, frequently not found in the same individuals, 

and it is not known if they would adversely affect reproductive performance. In contrast, the Henriksen et al. 

(1996) study of the Operation Ranch Hand cohort did not find any alterations, but the members of this cohort 

were exposed to lower concentrations of 2,3,7,8-TCDD and for shorter durations than the NIOSH cohort. 

Studies which examined pregnancy outcomes following paternal exposure (Aschengrau and Monson 1989; 

Smith et al. 1982) did not find increases in the incidence of spontaneous abortions and/or stillbirths. Forsberg 

and Nordstrom (1985) found an increased incidence of spontaneous abortions in residents likely exposed to 

2,3,7,8-TCDD. However, the Aschengrau and Monson (1989), Smith et al. (1982), and Forsberg and 

Nordstrom (1985) studies did not measure 2,3,7,8-TCDD levels, and it is difficult to determine the level of 

2,3,7,8-TCDD exposure from the data reported. Three studies found altered sex ratios among offspring of 

exposed individuals (Basharova 1996; Dimich-Ward et al. 1996; Mocarelli et al. 1996), but the role of CDDs 

could not be determined with any certainty. Without exposure information, relationships between 

2,3,7,8-TCDD exposure and the risk of adverse pregnancy outcomes cannot be established. 

2,3,7,8-TCDD body burdens calculated from available serum lipid 2,3,7,8-TCDD levels are presented in Table 

2-1. 

2.1.6 Developmental Effects 

The potential for 2,3,7,8-TCDD to induce developmental effects has been examined in several populations: 

residents exposed to 2,3,7,8-TCDD during aerial spraying of 2,4,5-T or from accidental releases of 

2,3,7,8-TCDD or 2,3,7,8-TCDD-contaminated chemicals, workers involved in manufacturing or application of 

phenoxy herbicides and/or chlorophenols, and Vietnam veterans. In most of the human studies, exposure was 

poorly characterized. 

In residents of Seveso, Italy, a significant rise in the incidence of birth defects, as compared to pre-accident 

levels, was observed the year after the accident (Bisanti et al. 1980). A variety of birth defects were 
CDDs 53 

2. HEALTH EFFECTS 

observed, but the incidence for any particular defect was not elevated. The authors suggest that the rise in 

birth defects may not be related to 2,3,7,8-TCDD exposure. Prior to 1976, birth defects in Italy were usually 

under reported; the authors note that the reported incidences of birth defects after the accident (23 per 1,000 

births) were similar to incidences reported in other western countries. Thus, the increased incidence may be 

reflective of the increased reporting rather than an increased number of birth defects. In a study which 

assessed the risk of birth defects for the 6-year period after the Seveso accident, no increases were observed 

for the risk of total defects (RR of 1.2, 90% CI of 0.88–1.64 for zones A and B and RR of 0.97, 90% 

CI=0.83–1.13 for zones A, B, and R), major defects RR of 1.02, 90% CI=0.64–1.61 for zones A and B and 

RR of 0.83, 90% CI=0.67–1.04 for zones A, B and R), and minor defects RR of 1.44 90% CI=0.92–2.24 for 

zones A and B and RR of 1.14, 90% CI=0.92–1.42 for zones A, B and R) (Mastroiacovo et al. 1988). The 

small number of observed birth defects limits the statistical power of this study to detect significant increases 

in a specific defect. 

In a study of residents of Northland, New Zealand exposed to 2,4,5-T during aerial spraying, no significant 

alterations in the total number of birth defects were observed in children born between 1973 and 1976, as 

compared to the incidence in children born between 1959 and 1960 (before the aerial 2,4,5-T spraying began) 

(Hanify et al. 1981). Stockbauer et al. (1988) studied the Missouri cohort and found no statistically 

significant excess risk of birth defects among infants from exposed mothers (n=410) compared to an 

unexposed referent group (n=820). However, a significant increase in the incidence of talipes (incidence ratio 

of 1.66, 90% CI=1.2–2.29) was observed in children born after the spraying program began. The relationship 

between 2,4,5-T usage and the incidence of facial clefts was investigated in residents of Arkansas exposed 

during the spraying of rice acreage (Nelson et al. 1979). The population was divided into areas of high, 

medium, and low potential exposure based on herbicide application rates. Increasing trends over time in 

facial clefts for both the high- and low-exposure groups were observed. The authors attributed this to better 

case-ascertainment rather than 2,4,5-T exposure. 

In the offspring of male workers at a chlorophenol manufacturing facility, no significant increases in the 

incidence of infant deaths, health defects, or congenital malformations were observed (Townsend et al. 1982). 

The adjusted odds ratio (95% CI) for infant deaths, health defects, and congenital malformations were 0.63 

(0.27–1.39), 0.85 (0.6–1.21), and 0.85 (0.53–1.35), respectively, for workers exposed to any dioxin and 0.82 

(0.3–2.09), 0.93 (0.6–1.43), and 1.08 (0.63–7.83) for workers exposed to 2,3,7,8-TCDD only. 
CDDs 54 

2. HEALTH EFFECTS 

Several studies investigated the outcome of pregnancies fathered by Vietnam veterans potentially exposed to 

2,3,7,8-TCDD-contaminated herbicides. Two case-control studies (Aschengrau and Monson 1990; Erickson 

et al. 1984) have examined the risk of Vietnam veterans having a child with birth defects. The Erickson et al. 

(1984) study used a cohort of 7,133 infants with birth defects registered by the Metropolitan Atlanta 

Congenital Defects Program and 4,246 control infants; information on military service and possible exposure 

to Agent Orange was obtained during interviews with the mother and father. The overall risk of having a 

child with birth defects was not significantly increased in the Vietnam veterans (OR of 0.97, 95% 

CI=0.83–1.14). However, Vietnam veterans fathered a higher proportion of the children with some birth 

defects (spina bifida, cleft lips, and congenital tumors including dermoid cysts, teratomas, hepatoblastomas, 

central nervous system tumors, and Wilm's tumors) (Erickson 1984). The case group (857 infants with 

congenital anomalies, 61 stillbirths, and 48 neonatal deaths) and control group (998 infants) for the 

Aschengrau and Monson (1990) study consisted of infants delivered between August 1977 and March 1980. 

No significant increase in the risk of fathering a child with birth defects was observed for the Vietnam 

veterans (OR of 1.3, 95% CI=0.7–2.4). Among the children with birth defects, an increased risk of having 

one or more major systemic malformation (OR of 1.8; 95% CI=1–3.1) was reported in infants fathered by 

Vietnam veterans. The largest increases were reported for malformations of the nervous system, 

cardiovascular system, genital organs, and urinary tract. No pattern of multiple malformations was found; the 

only pattern of multiple malformations observed in more than one infant was ventricular septal defect and 

talipes. The results of these two case-control studies (Aschengrau and Monson 1990; Erickson et al. 1984) 

should be interpreted cautiously because there is no documentation of 2,3,7,8-TCDD exposure. CDC (1988) 

found that in Vietnam veterans self-reporting exposure to Agent Orange, the levels of serum 2,3,7,8-TCDD 

were not significantly different than levels found in a control population. 

In a study of Vietnam veterans participating in Operation Ranch Hand (Wolfe et al. 1995), an increase in 

nervous system defects with increasing paternal serum lipid 2,3,7,8-TCDD levels was observed (statistical 

analysis was not performed due to the small number of defects: 3/981 in comparison group, 0/283 in Ranch 

Hand veterans with current 2,3,7,8-TCDD levels of #10 ppt, 2/241 in veterans with current 2,3,7,8-TCDD 

levels of >10 ppt and initial levels of #110 ppt, and 3/268 in veterans with current 2,3,7,8-TCDD levels of 

>10 ppt and initial levels of >110 ppt). However, the authors caution that this relationship is based on a 

limited amount of data. No relationships between paternal 2,3,7,8-TCDD exposure (based on serum 

2,3,7,8-TCDD levels) and the prevalence of other birth defects were observed. In an earlier study by Wolfe et 

al. (1985) of Air Force personnel involved in Operation Ranch Hand, a significant increase in the number of 

reported neonatal deaths (no additional details provided), as compared to a comparison group of Air Force 
CDDs 55 

2. HEALTH EFFECTS 

military employees not stationed in Vietnam, was observed. The incidence of major defects, prematurity, 

learning disabilities, or infant deaths was not increased in the Ranch Hand personnel. A significant increase 

in the incidence of minor health effects such as birth marks, rashes, and neonatal jaundice was reported by the 

Ranch Hand veterans. It should be noted that the pregnancy outcomes were self-reported, and this finding 

was not corroborated by the follow-up study (Wolfe et al. 1995) which used birth certificates, medical 

records, and death certificates to assess possible relationships between paternal exposure to 2,3,7,8-TCDD 

and developmental effects in offspring. Michalek et al. (1998) examined birth records of children born 

between 1959 and 1992 to Operation Ranch Hand veterans. A slight increase in the incidence of preterm 

births (not statistically significant) was observed in the low (current CDD level of #10 ppt) and high 

(extrapolated initial CDD level of >79 ppt) exposure groups but not in the medium (extrapolated initial CDD 

level of #79 ppt) exposure group. An increase in the relative risk of infant deaths was observed in all three 

groups, as compared to the referent group of veterans in SEA not exposed to Agent Orange); the relative risks 

in the low, medium, and high groups were 3.2 (95% CI=1.0–10.3), 1.5 (95% CI=0.3–7.5), and 4.5 (95% 

CI=1.5–14.0). Short gestation and low birth weight were the most common causes of infant deaths. No 

adverse effect on intrauterine growth was observed. Michalek et al. (1998) concluded that the increased 

infant mortality may not be due to paternal 2,3,7,8-TCDD exposure because the risk was increased in 

Operation Ranch Hand cohort members with essentially background current 2,3,7,8-TCDD levels (low 

exposure group) and in the highest exposure group. In Vietnamese families potentially exposed to 

2,3,7,8-TCDD-contaminated herbicides during the Vietnam War, a statistically significant increase in the 

incidence of unspecified congenital anomalies was observed as compared with a nonexposed population 

(Phuong et al. 1989a). Serum lipid 2,3,7,8-TCDD levels were not measured and the extent of exposure was 

based on subject recall of how many times they were exposed to herbicides during the Vietnam war. 

The results of the available developmental studies in humans were inconclusive. The lack of exposure data, 

small sample sizes, and the lack of reliable data for birth defect rates prior to 2,3,7,8-TCDD exposure limit the 

power of the human studies to determine if an association between 2,3,7,8-TCDD exposure and 

developmental toxicity exists. 

2.1.7 Genotoxic Effects 

Data regarding genotoxic effects in humans exposed to CDDs are inconclusive. A statistically significant 

increase in the incidence of cells with chromosomal aberrations and a greater number of aberrations were 

found in fetal tissues from induced abortions in women possibly exposed to 2,3,7,8-TCDD after the Seveso 
CDDs 56 

2. HEALTH EFFECTS 

accident (Tenchini et al. 1983). The results from cytogenetic analysis of maternal tissues were comparable to 

those of the control group. Furthermore, no increase in the frequency of chromosomal aberrations was found 

in 17 individuals who were treated for chloracne following the Seveso accident (Reggiani 1980). An 

increased incidence of chromosomal aberrations was found in a group of 10 Vietnam veterans (Kaye et al. 

1985); however, in another study, no increases in chromosomal aberrations or sister chromatid exchanges 

were reported in 15 Vietnam veterans (Mulcahy et al. 1980). None of these studies included 2,3,7,8-TCDD 

dosimetry and all were limited by using exposed groups that were relatively small (less than 20 individuals) to 

have the statistical power to reliably assess the cytogenetic damage. A more recent study examined the 

incidence of chromosomal aberrations and of sister chromatid exchanges in human lymphocytes in 27 

workers whose current 2,3,7,8-TCDD concentrations in blood were above 40 ppt, and in 28 age-comparable 

referents (Zober et al. 1993). The results showed no statistically significant differences between the two 

groups in the percentages of gaps, chromatid or chromosome exchanges, chromatid or chromosome 

breaks/fragments/deletions, multiple aberrations, or the overall percentage of aberrations including or 

excluding gaps. In the exposed group there was an increased rate of sister chromatid exchanges per cell and a 

higher percentage of cells with more than 10 sister chromatid exchanges. However, these associations were 

no longer significant when smoking status was included as covariate. Moreover, neither current nor backcalculated 2,3,7,8-TCDD concentration was a significant predictor of these parameters. Zober et al. (1993) 

indicated that some limitations such as the small number of individuals studied, a possible selection effect, 

and the possibility that some effects were transient should be considered in the interpretation of the results. 

The human data on the genotoxicity of 2,3,7,8-TCDD is inconsistent and inconclusive. The lack of exposure 

data, small sample sizes, and the inconsistent results precludes drawing conclusions from these studies. 

2.1.8 Cancer 

The carcinogenicity of 2,3,7,8-TCDD in humans has been assessed in numerous case-control and mortality 

cohort studies of chemical manufacturing and processing workers and phenoxy herbicide and chlorophenols 

applicators, Vietnam veterans exposed to Agent Orange, and residents of Seveso, Italy. A major weakness in 

many of these studies is the lack of adequate exposure data. Exposure levels or 2,3,7,8-TCDD body burdens 

were not measured, rather surrogates of exposure such as exposure to chemicals contaminated with 

2,3,7,8-TCDD or chloracne were used to identify subjects likely exposed to 2,3,7,8-TCDD. Another major 

weakness of most of the human cancer data is concomitant exposure to other compounds. The focus of this 

discussion on the carcinogenic potential of 2,3,7,8-TCDD and other CDDs will be on studies that have 

documented exposure by measuring blood levels or in which exposure can be reasonably presumed. The 
CDDs 57 

2. HEALTH EFFECTS 

section is divided into four parts: 1) the effect of CDD exposure on overall cancer risk in workers involved in 

the manufacture or application of phenoxy herbicides or chlorophenols followed by a discussion of specific 

types of cancer in this group, 2) cancer risks in Vietnam veterans, 3) cancer risks in Seveso residents, and 4) 

conclusive statement. 

Increases in the overall cancer risk were observed in a number of large cohort mortality studies of chemical 

manufacturing workers and phenoxy herbicide applicators (Becher et al. 1996; Fingerhut et al. 1991; 

Hooiveld et al. 1998; Kogevinas et al. 1993, 1997; Manz et al. 1991; Ott and Zober 1996; Zober et al. 1990). 

Most of the subjects in these studies were males working in chlorophenoxy herbicide or trichlorophenol 

manufacturing facilities. In one of the few studies assessing the carcinogenicity of 2,3,7,8-TCDD in women, 

Kogevinas et al. (1993) found a significantly elevated risk for cancer in women probably exposed to 

2,3,7,8-TCDD during the production or application of chlorophenoxy herbicides and/or chlorophenols. The 

Zober et al. (1990), Fingerhut et al. (1991), Manz et al. (1991), Ott and Zober (1996) and Hooiveld et al. 

(1998) studies used current serum 2,3,7,8-TCDD levels in surviving workers to estimate exposure. The 

results of these studies, as well as two large multinational cohort mortality studies (Kogevinas et al. 1997; 

Saracci et al. 1991), are described below. 

Saracci et al. (1991) examined cancer mortality in workers on the International Register of Workers Exposed 

to Phenoxy Herbicides and Their Contaminants. The registry consists of 18,390 workers (16,863 males and 

1,527 females) distributed among 20 cohorts from 10 countries. Based on information obtained from 

questionnaires, factory or spraying records, and job histories, the workers were classified as exposed (13,482 

workers), probably exposed (416), exposure unknown (541), or non-exposed (3,951). The exposed workers 

were workers who sprayed chlorophenoxy herbicides or worked in factories producing chlorophenoxy 

herbicides or chlorinated phenols. The probably exposed workers worked at facilities producing 

pentachlorophenol (145 workers) or 2,4-D, 2,4-(dichlorophenoxy)butanoic acid, (4-chloro-2-methylphenoxy)acetic acid, and (4-chloro-2-methyl)propanoic acid (275 workers). For 4 of the 20 cohorts, a 

minimum employment duration of 1 to 12 months was required for inclusion in the registry, for the 

remaining cohorts, the criterion for inclusion was “ever employed in production or spraying of phenoxy 

herbicides.” The average follow-up period for the entire cohort was 17 years. Cancer mortalities were not 

significantly increased in exposed and probably exposed workers (SMR=101; 95% CI=93–110). Duration of 

exposure or time since first exposure did not appear to influence the number of cancer deaths. The lack of 

both a clear definition of exposure and uniformity of exposure classification between and within cohorts 

makes the results difficult to interpret and lessens the confidence in the results. 
CDDs 58 

2. HEALTH EFFECTS 

After publication of the Saracci et al. (1991) study, the IARC cohort was expanded to include 12 

manufacturing facilities in the United States (also examined by Fingerhut et al. 1991) and 4 plants in 

Germany (also examined by Becher et al. 1996). This expanded cohort, studied by Kogevinas et al. (1997), 

which included almost all workers world-wide who ever produced phenoxy herbicides, comprised 21,863 

workers (20,851 males and 1,012 females) in 12 countries. 2,3,7,8-TCDD levels were measured in 573 

workers at 10 facilities (approximately 50% of these workers were part of the NIOSH cohort examined by 

Fingerhut et al. 1991). The levels were 2-34 pg/g blood lipid (measured in 1990, mean not reported), 98–659 

(1990, mean of 389 pg/g), 1.9–194 (1993, 53 pg/g), 3.0–131 (1988, 53.3 pg/g), 9–37 (1992, 17 pg/g), 

1.3–6.49 (1996, estimated mean of 3.2 pg/g), 3–2,252 (1985–1994, estimated mean of 141 pg/g), 23–1,935 

(1989–1992, estimated mean of 401.7 pg/g), and 2–3,400 (1987–1988, mean of 233 pg/g). Mortality rates for 

the cohort were compared to national mortality rates calculated using data from the WHO mortality data bank. 

There was a significant increase in the SMRs for all cancers for male workers (1,083 deaths; SMR=1.07; 95% 

CI=1.01–1.13) but not among female workers (44 deaths; SMR=0.93; 95% CI=0.68–1.25). When mortality 

rates were calculated for the 13,831 workers (males and females) exposed to phenoxy herbicides 

contaminated with 2,3,7,8-TCDD or higher chlorinated dioxins, the all-cancer SMR increased to 1.12 (710 

deaths; 95% CI=1.04–1.21). The SMR for all cancer deaths was not elevated in workers not exposed to 

2,3,7,8-TCDD or higher chlorinated dioxins. 2,3,7,8-TCDD-exposed workers were divided into groups based 

on years since first exposure, duration of exposure, and year of first exposure. The mortality rate appeared to 

be related to years since exposure and the related variable of year of first exposure. Significantly increased 

SMRs were observed in workers with a $20 year latency period (394 deaths; SMR=1.20; 95% CI=1.09–1.33), 

workers employed before 1955 (335 deaths; SMR=1.12; 95% CI=1.00–1.25), or workers employed between 

1955 and 1964 (242 deaths; SMR=1.17; 95% CI=1.03–1.25), but these differences were overall small. 

The cancer mortality experience of 247 male workers who were exposed to 2,3,7,8-TCDD during an 

accidental uncontrolled decomposition reaction and subsequent clean-up activities in a German 2,4,5-TCP 

production plant (BASF AG facility) in 1953 and followed for 34 years was studied by Zober et al. (1990). 

Three subcohorts were defined based on qualitative exposure information: Subcohorts 1 (n=69) included 

workers known to be exposed to 2,3,7,8-TCDD during the accident, and Subcohort 2 (n=84) and 3 (n=94) 

included workers considered exposed to amounts of 2,3,7,8-TCDD less than in Subcohort 1. Chloracne (114 

cases) or erythema (13 cases) developed in 127 (51%) of the main cohort. Subcohorts 1, 2 and 3 contained 69 

cases (21 severe, 27 extensive, 21 moderate), 17 cases (1 severe, 3 extensive, 13 moderate) and 28 cases (0 

severe, 3 extensive, 25 moderate) of chloracne, respectively, and erythema affected 4 subjects in 
CDDs 59 

2. HEALTH EFFECTS 

Subcohort 2 and 9 subjects in Subcohort 3. Blood 2,3,7,8-TCDD levels were measured in 28 subjects in 

Subcohorts 1 (median 24.5 ppt, n=11), 2 (median 9.5 ppt, n=7), and 3 (median 8.4 ppt, n=10). Subjects with 

chloracne/erythema had higher average serum 2,3,7,8-TCDD levels (15 ppt, n=16) than those without 

(5.8 ppt, n=12). There was no clear increase in all malignant neoplasms or site-specific cancer in the entire 

cohort or either subcohort based on comparison with national mortality rates in the Federal Republic of 

Germany. When the 127 workers with chloracne/erythema were examined separately, the standardized 

mortality ratio (SMR) for all malignant neoplasms was not significantly elevated overall (SMR=139; 95% 

CI=87–211), although the SMR for all malignancies was significantly increased when analysis was restricted 

to workers whose first exposure was $20 years earlier (SMR=201; 95% CI=122–315, p<0.05). 

In a follow-up to the Zober et al. (1990) study, the cohort of workers exposed at the BASF AG facility in 

Germany, was followed through 1992 (Ott and Zober 1996). The workers were divided into 3 subcohorts 

based on half-life extrapolated 2,3,7,8-TCDD body burdens of <0.01 μg/kg body weight, 0.1–0.99 μg/kg, and 

$1 μg/kg. 2,3,7,8-TCDD half-lives were estimated from repeated blood samples from 29 people with initial 

serum lipid 2,3,7,8-TCDD levels of 29–553 pg/g. The mean half-life was 5.8 years, but the half-life increased 

with higher percentages of body fat. Half-life estimates of 5.1 and 8.9 years were used for workers with 20 

and 30% body fat, respectively. There was suggestive evidence that 2,3,7,8-TCDD exposure affected the 

occurrence of deaths from cancer (all sites combined) and deaths from respiratory or digestive cancer. The 

number of cancer deaths increased with increasing body burdens, SMRs of 0.8 (8 deaths; 95% CI=0.4–1.6), 

1.2 (8 deaths; 95% CI=0.5–2.3), and 1.6 (15 deaths; 95% CI=0.9–2.6) in the cohorts with body burdens of 

<0.1, 0.1–0.99, and $1 μg/kg, respectively. No increases in cancer deaths were observed among nonsmokers; 

among current smokers, there were increases in cancer risks for workers with 2,3,7,8-TCDD body burdens of 

1.0–1.99 μg/kg (6 deaths; SMR=3.0; 95% CI=1.1–6.5) and $2.00 μg/kg (6 deaths; SMR=4.0; 95% 

CI=1.5–8.6) but not for workers with lower 2,3,7,8-TCDD levels. Ott and Zober (1996) concluded that past 

body burdens of $1.0 μg/kg 2,3,7,8-TCDD were consistent with a 2,3,7,8-TCDD-induced carcinogenic effect. 

At the same time, they stated that with such a small cohort, the risk estimates are not very stable and could be 

affected by selection and confounding. 

Becher et al. (1996) examined a cohort of 2,479 men employed at 4 German facilities involved in the 

production of phenoxy acid herbicides and chlorophenols. The workers were divided into 4 subcohorts: 

1,144 male workers at facility 1 with a mean duration of employment of 7.7 years, 135 male workers at 

facility 2 with a mean duration of employment of 21.5 years, 520 male workers at facility 3 with a mean 

duration of employment of 9.1 years, and 680 workers at facility 4 with a mean duration of employment of 
CDDs 60 

2. HEALTH EFFECTS 

18.5 years. 2,3,7,8-TCDD blood levels in groups of workers in subcohorts 1 (112 males and 18 females) and 

2 (8 workers, all with a history of chloracne) were 3–2,252 and 163–1,935 pg/g blood lipid, respectively. The 

study authors noted that 2,3,7,8-TCDD exposure was probably lower in subcohorts 3 and 4 because 

2,3,7,8-TCDD-contaminated products only made up a small percentage of the total products manufactured at 

these facilities. Messerer et al. (1998) measured serum CDD and CDF levels in 19 of the current employees 

at facility 4. The mean CDD and CDF levels were slightly higher than background levels, even in the most 

exposed workers (involved in synthesis); the mean 2,3,7,8-TCDD and TEQ levels in the 7 synthesis workers 

were 3.8 and 35.4 ppt, respectively, as compared to 3.2 and 25.0 ppt in controls. The SMR for all cancer 

mortalities was significantly increased in the entire cohort (138 deaths; SMR=119; 95% CI=100–141); 

subcohort 1 was the only subcohort with a significant increase in all cancer mortalities (97 deaths; SMR=134; 

95% CI=109–164). When the entire cohort was divided into groups based on time since first exposure, there 

were no increases in SMRs for all cancer deaths during the three time periods (0 to <10 years, 10 to <20 

years, and $20 years). 

In the Fingerhut et al. (1991) study, cancer mortality was evaluated in 5,172 male workers involved in the 

production of 2,3,7,8-TCDD-contaminated chemicals in 12 U.S. plants, as well as in subcohorts with low 

(<1 year, n=1,516) or high ($1 year, n=1,520) exposures and $20 years latency. Exposure was documented 

by reviewing job descriptions and records of 2,3,7,8-TCDD levels in industrial hygiene samples, and 

measuring lipid-adjusted serum 2,3,7,8-TCDD levels in 253 of the workers from two of the facilities; it was 

assumed that a similar relationship would exist at the other 10 facilities. Duration of exposure was used as a 

surrogate for cumulative 2,3,7,8-TCDD exposure on the basis of a high correlation (r=0.72, p<0.001) between 

log serum 2,3,7,8-TCDD level and log number of years of exposure. Mean lipid-adjusted serum 

2,3,7,8-TCDD levels were 233 ppt (range 2–3,400) for all 253 workers, 418 ppt for 119 workers with $1 year 

exposure and 7 ppt in a comparison group of 79 unexposed persons. When compared to the U.S. population, 

mortality from all cancers was significantly increased (p<0.05) in the overall cohort (SMR=115; 95% 

CI=102–130) and the high exposure subcohort (SMR=146; 95% CI=121–176). Cancer mortality increased 

with increasing latency. The number of deaths were too small to allow meaningful analysis according to 

duration of exposure. 

A cancer mortality follow-up study of 1,583 workers (1,184 men, 399 women) who were employed in a 

German chemical plant that produced trichlorophenol, 2,4,5-T, and other herbicides known to be 

contaminated with 2,3,7,8-TCDD and other polychlorinated dioxins and furans (this appears to be the same 

facility as subcohort 1 in the Becher et al. [1996] study) was reported by Manz et al. (1991). Production of 
CDDs 61 

2. HEALTH EFFECTS 

these chemicals was discontinued during 1954–57 after an outbreak of chloracne. Cohort members worked 

for at least 3 months during 1952–1984 and were followed through 1989. SMRs were calculated using 

national mortality rates for West Germany and deaths in a cohort of male gas supply company workers. 

Exposures of cohort members were classified as high (n=496), medium (n=901), or low (n=186) based on an 

analysis of production processes in the plants where they had worked. Some validation for these categories 

was provided by measurement of adipose 2,3,7,8-TCDD levels in 48 males in 1985 (mean concentrations 

were 296 ng/kg lipid in 37 subjects in the high-exposure group compared with 83 ng/kg for 11 subjects in the 

medium- and low-exposure groups combined). When compared with national rates, mortality from all 

cancers combined was increased in the entire cohort (93 deaths, SMR=1.24; 95% CI=1.00–1.52), especially 

among men in the high exposure subgroup with $20 years employment (8 deaths, SMR=2.54; 95% 

CI=1.00–5.00) or who began employment before 1955 (18 deaths, SMR=2.11; 95% CI=1.25–3.34). The 

greatest increase in risk was found in high exposure men with time of entry before 1955 and $20 years 

employment (8 deaths, SMR=3.5; 95% CI=1.51–6.9). The aforementioned findings were corroborated when 

the gas workers were used as the comparison group. In a follow-up study (Flesch-Janys et al. 1998) in which 

the cohort was followed through 1992, the SMR for all cancers was 1.4 (124 deaths; 95% CI=1.17–1.68). 

When the workers were divided into four exposure groups, a significant increase in SMR (36 deaths; 

SMR=1.73; 95% CI=1.21–2.40) was only found in the highest exposure group (2,3,7,8-TCDD levels of 

$2,503 ng/kg blood lipid × years, a measure of cumulative lifetime exposure); the trend toward increasing 

SMR with increasing dose was also statistically significant. 

The cohort in the Manz et al. (1991) study served as the basis for additional analysis by Flesch-Janys et al. 

(1995, 1998), who investigated the relation between mortality and quantitative measure of PCDD/F exposure. 

The male chemical workers (n=1,177) were followed for an additional 3 years. Blood levels of each PCDD/F 

congener at the end of exposure were estimated for all members of the cohort based on work histories 

(durations of exposure in particular departments) and tissue levels from a subgroup of male workers, and 

assuming one-compartment first-order elimination kinetics. A total TEQ level was also estimated for all 

measured CDDs and CDFs combined as the weighted sum of PCDD/F congeners using toxicity equivalent 

factor (TEF) values (see Section 2.5 for a detailed explanation on TEFs and dioxin equivalents). In the 

Flesch-Janys et al. (1995) study, risk ratios for the cohort were estimated with year-of-birth stratified Cox 

regression using seven exposure levels (the reference cohort, the first four quintiles and the ninth and tenth 

deciles of the estimated 2,3,7,8-TCDD levels and total TEQ); an external cohort of gas supply workers 

(n=2,158) served as an unexposed control group. The estimated mean 2,3,7,8-TCDD level for the entire 

cohort at the end of employment in the plant was 141.4 ng/kg blood fat (median of 38.2 ng/kg). 
CDDs 62 

2. HEALTH EFFECTS 

The estimated mean total dioxin equivalents, calculated as the weighted sum of combined PCDD/F congeners, 

was 296.5 ng/kg (median of 118.3 ng/kg). There was a dose-dependent increase in cancer mortality with 

increasing levels of 2,3,7,8-TCDD (p=0.01 for trend), predominantly due to increased risk ratio (3.30; 95% 

CI=2.05–5.31) in the highest-dose group (344.7–3,890.2 ng/kg). Similar findings were obtained when total 

TEQs were used as the exposure parameter, or when the two lowest-dose groups in the chemical-worker 

cohort were combined and used for reference. In the Flesch-Janys et al. (1998) study, the mean blood 

2,3,7,8-TCDD level in the subgroup (blood or adipose tissue samples collected from 236 males) was 108.3 

ng/kg blood lipid (range, 2.0–2252 ng/kg), the mean TEQ for CDD/F congeners was 247.5 ng/kg 

(11.7–2,985.8), and the mean TEQ without 2,3,7,8-TCDD was 184.0 ng/kg (9.7–1,263.4 ng/kg). When the 

workers were divided into groups based on cumulative TEQ exposure, statistically significant SMRs were 

found in the second (TEQ between 360.9 and 1,614.4 ng/kg × years) and fourth (TEQ greater than 

5,217.7 ng/kg × years) quartiles; SMRs of 1.64 (34 deaths; 95% CI=1.13–2.29) and 1.64 (34 deaths; 95% 

CI=1.13–2.29), respectively. A significant relationship between cumulative TEQ level and SMR for all 

cancer was not found. Potential sources of error and bias in this study include lack of random sampling in the 

subgroup with PCDD/F assays and the assumption of first-order elimination kinetics. Flesch-Janys et al. 

(1998) compared age of employment and years spent in each product department for workers with measured 

2,3,7,8-TCDD blood levels and workers without blood level data and found no major differences between the 

two groups. 

Workers at two Dutch phenoxy herbicide and chlorophenols production facilities comprised the cohort for the 

Bueno de Mesquita et al. (1993) study of cancer mortality. The cohorts consisted of any worker employed 

between 1955 and 1985 (facility A) or between 1965 and 1986 (facility B). An industrial accident at facility 

A resulted in a release of CDDs, including 2,3,7,8-TCDD. Mortality rates in the workers were compared to 

national rates. In workers at facility A, there was a slight increase in deaths from all cancers; however, the 

increase was not statistically significant (26 deaths; SMR=118; 95% CI=77–173); the SMR for all cancers 

was not increased at facility B or in the combined entire cohort. Likewise, dividing the workers at each 

facility into groups based on time since first exposure or duration of exposure did not result in any significant 

increases in SMRs. Hooiveld et al. (1998) followed the workers at facility A for another 6 years. 

Additionally, serum levels of CDDs, CDFs, and PCBs were measured in a sample of 47 surviving workers 

who were employed for at least 1 year and whose date of first employment was prior to 1975; 14 of these 

workers were exposed during the accident, 17 were workers not involved in the accident, and 16 were not 

exposed to phenoxy herbicides or chlorophenols. The average 2,3,7,8-TCDD serum levels were 105.2, 42.9, 

16.6, and 7.6 ppt (lipid adjusted) in the accident-exposed workers with a history of chloracne (12 workers), 
CDDs 63 

2. HEALTH EFFECTS 

accident-exposed workers without chloracne (2 workers), nonaccident-exposed workers, and nonexposed 

workers, respectively; the mean serum 2,3,7,8-TCDD levels at the time of maximum exposure (extrapolated 

levels) were estimated at 2,014.4, 806.6, and 244.1 ppt in the accident-exposed workers with a history of 

chloracne, accident-exposed workers without chloracne, and nonaccident-exposed workers, respectively. 

There was an increase in deaths from all cancers among all exposed workers (51 deaths; SMR=1.5; 95% 

CI=1.1–1.9), as compared to national rates; a slightly higher mortality rate (20 deaths; SMR=1.7; 95% 

CI=1.1–2.7) was observed in the subcohort of accident-exposed workers. When nonexposed workers were 

used as a comparison group, the relative risk of all cancer deaths was 4.1 (51 deaths; 95% CI=1.8–9.0), the 

relative risk was adjusted for age, calendar year at end of follow-up, and time since first exposure/ 

employment. When the workers were divided into three groups based on model-predicted 2,3,7,8-TCDD 

levels, the relative risk of all cancer deaths was elevated in workers with medium or high exposure, as 

compared to workers with low exposure (adjusted RR=4.8; 95% CI=2.0–11.3 for the medium exposure group 

and adjusted RR=4.4; 95% CI=1.9–10.4 for the high exposure group). 

Case-control studies have been designed to determine if 2,3,7,8-TCDD exposure results in increased risks for 

site-specific cancers. Case-control studies have found significant increases in the risk of soft-tissue sarcomas 

in Swedish agricultural, forestry, and horticultural workers (Eriksson et al. 1981, 1990; Hardell and Eriksson 

1988; Hardell and Sandstrom 1979), workers involved in manufacturing and application of phenoxy 

herbicides (Kogevinas et al. 1995), and New Zealand farmers (Smith et al. 1984a). In the Eriksson et al. 

(1990) study, the risk ratio of soft-tissue sarcoma was 1.80 (95% CI=1.02–3.18) in subjects exposed to 

phenoxyacetic acid herbicides and/or chlorophenols. In subjects exposed to phenoxyacetic acid herbicides 

only or chlorophenols only, the risk ratios were 1.34 (95% CI=0.7–2.56) and 5.25 (95% CI=1.69–16.34), 

respectively. When the phenoxyacetic acid herbicide-only subjects were divided into two groups of subjects 

predominantly exposed to 2,4,5-T and those exposed to phenoxyacetic acid herbicides other than 2,4,5-T, risk 

ratios of 1.81 (95% CI=0.85–3.87) and 0.60 (95% CI=0.18–2.06), respectively, were calculated. Hardell et al. 

(1995) conducted a meta-analysis of their four Swedish case-control studies (Eriksson et al. 1981, 1990; 

Hardell and Eriksson 1988; Hardell and Sandstrom 1979). The odds ratios for workers exposed to 

phenoxyacetic acid herbicides or chlorophenols, phenoxy herbicides only, or chlorophenols only were 2.8 (90 

cases; 95% CI=2.1–4.4), 2.7 (59 cases; 95% CI=1.9–4.7), and 3.3 (34 cases; 95% CI=1.8–6.1), respectively. 

The data from this study suggest that the increased possible risks of soft-tissue sarcomas observed in phenoxy 

herbicide and chlorophenols applicators may be due to exposure to the 2,3,7,8-TCDD (or other CDDs) 

contamination of the mixture. However, the results of the Kogevinas et al. (1995) study suggest that the 

possible risk of soft-tissue sarcoma may not be specific to 2,3,7,8-TCDD-contaminated 
CDDs 64 

2. HEALTH EFFECTS 

phenoxy herbicides. An increase in the possible risk of soft-tissue sarcoma (OR of 10.3; 95% CI=1.2–90.6) 

was found in multinational workers involved in the production and spraying of phenoxy herbicides 

(Kogevinas et al. 1995). Exposure to chlorophenols was not associated with an increase in the possible risk of 

soft-tissue sarcomas (OR=1.29; 95% CI=0.24–6.91). Using job histories, the workers were divided into 

groups based on the type of phenoxy herbicide used. Excess possible risks of soft-tissue sarcoma were 

observed in workers predominantly exposed to 2,4-dichlorophenoxyacetic acid (2,4-D, OR=5.7; 95% 

CI=1.1–29), 2,4,5-T (OR=4.3, 90% CI=0.7–26), 4-chloro-2-methylphenoxyacetic acid (MCPA, OR=11.3; 

95% CI=1.3–98), exposure to any CDDs or CDFs (OR=5.6; 95% CI=1.1–28), and exposure to 2,3,7,8-TCDD 

(OR=5.2, CI=0.9–32). Although this study suggests that exposure to 2,3,7,8-TCDD increases the possible 

risk of soft-tissue sarcoma, it also suggests that exposure to phenoxy herbicides on their own may also 

increase the possible risk of soft-tissue sarcoma. 

Cohort mortality studies have also found increases in the incidences of soft-tissue sarcomas. Fingerhut et al. 

(1991) found significant increase in deaths from soft-tissue sarcomas (SMR=922; 95% CI=190–2,695) in the 

high-exposure cohort, although this was only based on 3 deaths. In the Saracci et al. (1991) multinational 

cohort, an increase in deaths from soft-tissue sarcoma was observed in phenoxy herbicide sprayers (3 deaths, 

SMR=297; 95% CI=61–868) and in workers dying 10–19 years after first exposure (4 deaths, SMR=606; 

95% CI=165–1,552). Similarly, the Kogevinas et al. (1997) study of the IARC cohort found 3 cases of softtissue sarcoma in workers exposed for 10–19 years (SMR=6.52; 95% CI=1.35–19.06); shorter exposure 

durations did not result in significantly elevated SMRs. Additionally, when workers were divided into 

latency groups, the SMRs were not elevated. Duration of probable exposure to 2,3,7,8-TCDD did not appear 

to influence soft tissue sarcoma cancer risks. 

Case-control studies and/or cohort mortality studies have also found significant increase in the possible risk of 

malignant lymphoma (RR=4.8; 95% CI=2.9–8.1) in Swedish agricultural, forestry, and horticultural workers 

Hardell et al. 1981) and German phenoxy herbicide and chlorophenols manufacturer workers (SMR=239; 

95% CI=119–427) (Becher et al. 1996); non-Hodgkin’s lymphoma in Wisconsin farmers (OR=1.22; 95% 

CI=0.98–1.51) (Cantor 1982), multinational phenoxy herbicide manufacturers and applicators (OR of 1.25; 

CI=0.54–2.90) (Kogevinas et al. 1995), and German workers at a phenoxy herbicide and chlorophenols 

facility (SMR=375; 95% CI=101–957) (Becher et al. 1996); and stomach cancer in railroad workers (rate 

ratio of 6.1) (Axelson et al. 1980) and workers at a German trichlorophenol facility (observed/expected ratio 

of 3/0.52) (Thiess et al. 1982). 
CDDs 65 

2. HEALTH EFFECTS 

Cohort mortality studies by Fingerhut et al. (1991), Zober et al. (1990), Manz et al. (1991) (including the 

Flesch-Janys et al. [1998] follow-up data), and Kogevinas et al. (1997) found significant increases in risk of 

respiratory tract cancer. In the Manz et al. (1991) study, a slightly increased risk of lung cancer was observed 

in the entire cohort compared to the risk in the West German population (30 deaths, SMR=1.41; 

95% CI=0.95–2.01) or the reference gas workers (26 deaths, SMR=1.67; 95% CI=1.09–2.44); the increase in 

lung cancer risk was statistically significant when the gas workers were used as the referent group (Manz et 

al. 1991). Smoking does not appear to be a major confounder because available data (partial cohort) suggest 

that the percentage of smokers in the study cohort and gas-worker control group were similar. In the FleschJanys et al. (1998) follow-up of this cohort, the SMRs for lung (38 deaths) and respiratory (44 deaths) cancer 

were 1.51 (95% CI=1.07–2.08) and 1.71 (95% CI=1.24–2.29), respectively. When the cohort was divided 

into four cumulative exposure categories, the SMRs were not significantly elevated for any exposure group. 

In this cohort, blood 2,3,7,8-TCDD levels were not correlated with smoking status (Flesch-Janys et al. 1995). 

The study authors additionally noted that when lung cancer deaths were removed from total cancer deaths, 

there was a stronger relationship between 2,3,7,8-TCDD levels and cancer rates, suggesting that smoking was 

not a strong confounder. Death from cancers of the respiratory tract (SMR=142; 95% CI=103–192) were 

significantly increased in the high-exposure subcohort of the Fingerhut et al. (1991) study. The expected 

number of lung cancers was adjusted for smoking using smoking-prevalence data for a small subset of 

workers; the adjusted SMR for the high-exposure subcohort was 137 (95% CI=98–187). The authors 

concluded that the increased lung cancer risk was probably not due to smoking because the incidences of 

smoking-related diseases were not higher than expected in the subcohort and mortality from non-malignant 

respiratory disease was lower than expected. However, when smoking was included as a confounding factor, 

the increased risk for lung cancer was no longer statistically significant. Zober et al. (1990) found a 

borderline increase in mortality from cancer of the trachea bronchus/ lung (SMR=252; 95% CI=99–530) in 

workers with chloracne/erythema and $20 years latency. In the follow-up to the Zober et al. (1990) study, Ott 

and Zober (1996) found an increased number of deaths from respiratory cancer in the subcohort with 

2,3,7,8-TCDD body burdens of $1 μg/kg (SMR=2.4; 95% CI=1.0–5.0). However, in 10 of the 11 cases, the 

worker smoked, which makes it difficult to determine if the cancer deaths were due to 2,3,7,8-TCDD 

exposure. The Kogevinas et al. (1997) examination of the expanded IARC cohort did not find a significant 

elevation in deaths from lung cancer (225 deaths; SMR=1.12; 95% CI=0.98–1.28) among workers exposed to 

phenoxy herbicides contaminated with 2,3,7,8-TCDD or higher chlorinated dioxins, but it did find a rise in 

deaths from other respiratory organ cancers (9 deaths; SMR=3.20; 95% CI=1.46–6.08). 
CDDs 66 

2. HEALTH EFFECTS 

Using regression analysis based on Cox’s proportional hazards model, Ott and Zober (1996) found evidence 

of association between 2,3,7,8-TCDD exposure and digestive cancer (conditional risk ratio of 1.46; 

95% CI=1.13–1.89); the primary tumor sites were the liver, stomach, and pancreas. 

A number of studies have looked at cancer incidences among Vietnam veterans to determine if exposure to 

Agent Orange with its 2,3,7,8-TCDD contamination resulted in a higher cancer risk. Many of these studies 

compared cancer incidences in Vietnam veterans to Vietnam-era veterans stationed outside of Vietnam. A 

limitation of this study design is that not all veterans in Vietnam were exposed to Agent Orange and exposure 

was lower than that of occupational workers. CDC (1988) found that the levels of 2,3,7,8-TCDD in Vietnam 

veterans were usually similar to a comparison group. Thus, studies which examined cancer incidences in 

“Vietnam veterans” may not be adequate to assess the carcinogenicity of 2,3,7,8-TCDD. Wolfe et al. (1985) 

focused on Air Force personnel involved in Operation Ranch Hand. 2,3,7,8-TCDD levels in 888 Ranch Hand 

personnel was 12.4 ppt as compared to 4.2 ppt in a referent group of Air Force personnel (CDC 1987; USAF 

1991). No significant alterations in the incidence of systemic malignancies were observed in the Ranch Hand 

personnel, as compared to a group of veterans flying cargo in Southeast Asia during the Vietnam war. A 

significant increase in non-melanomic skin cancer (predominantly basal cell carcinoma) was found in the 

Ranch Hand personnel; however, cancer incidences were not adjusted for sun exposure. No significant 

alterations in systemic malignancy indices were found. In a similar study of Air Force veterans involved in 

Operation Ranch Hand, a significant increase in benign systemic neoplasms was observed (USAF 1991). No 

alterations in the risk of malignant neoplasm were observed. No increases in the risk of Hodgkin’s disease, 

non-Hodgkin’s lymphoma or soft-tissue sarcoma were observed; however, the statistical power of the study to 

detect significant risk ratios for site-specific cancers was limited by the small number of cancers and the small 

sample size. The incidence of benign neoplasms (primarily lipomas) was highest in veterans with the highest 

blood dioxin levels. The incidence of basal cell skin neoplasms was not positively associated with serum 

dioxin levels except among enlisted flyers with basal cell carcinomas at sites other than the ear, face, head, or 

neck. 

Increases in the risk of several types of cancer have been observed in residents of Seveso, Italy. In the 

residents with the highest exposure (zone A), no increases in the risk ratio of all malignancies were observed 

(Bertazzi et al. 1993). However, the small number of zone A residents (724) limits the statistical power of the 

analysis. Among residents living in zone B (4,824 people), significant increases were observed for the risk of 

hepatobiliary cancer (risk ratio of 3.3; 95% CI=1.3–8.1) and multiple myeloma (risk ratio of 5.3; 95% 

CI=1.2–22.6) in women and lymphoreticulosarcoma in men (RR of 5.7; 95% CI=1.7–19). In zone R 
CDDs 67 

2. HEALTH EFFECTS 

(31,647 residents), the risk ratio of soft-tissue sarcomas in men (2.8; 95% CI=1–7.3) was significantly 

increased. In both zone B and R, the risk ratios of all malignancies was not significantly altered. Estrogendependent cancers (breast cancer and corpus uteri cancer) were consistently decreased in the women living in 

zone A, B, and R. It should also be noted that the latency period of 10 years may be too short to find 

increases in other types of cancer. A study of children (aged 0–19 years at the time of the accident) exposed 

to 2,3,7,8-TCDD during the accident in Seveso found increased risks of Hodgkin’s lymphoma [RR =2; 95% 

CI=0.5–7.6), myeloid leukemia (RR=2.7; 95% CI=0.7–11.4), and thyroid cancer (RR=4.6; 95% CI=0.6–32.7) 

(Pesatori et al. 1993). However, the differences in RRs for these cancer types between the Seveso residents 

and the control population did not reach statistical significance. The small number of detected cancers and 

the relatively short latency period (10 years) limits the interpretation of the results of this study. Similar 

results were found in the 15-year follow-up study (Bertazzi et al. 1997). No significant increases in cancer 

mortality were found in zone A residents (805 residents, 70 deaths). In zone B, death from all cancers was 

not significantly elevated in males (104 deaths; RR=1.1; 95% CI=0.9–1.3) or females (48 deaths; RR=0.9; 

95% CI=0.7–1.2). However, significant increases in site-specific cancers were observed, including rectal 

cancer in males (7 deaths; RR=2.9; 95% CI=1.2–5.9), pleural cancer in males (31 deaths; RR=5.3; 95% 

CI=1.1–5.5), lymphohemopoietic cancer in males (12 deaths; RR=2.4; 95% CI=1.2–4.1), leukemia in males 

(7 deaths; RR=3.1; 95% CI=1.3–6.4), and myeloma in females (4 deaths; RR=6.6; 95% CI=1.8–16.8). The 

cancer with an elevated incidence among zone R residents was bone cancer in females (7 deaths; RR=2.4; 

95% CI=1.0–4.9). 

The available epidemiology data suggest that 2,3,7,8-TCDD may be a human carcinogen. Statistically 

significant increases in risks for all cancers were found in highly exposed workers with longer latency 

periods. Although the estimated SMRs are low, they are consistent across studies with the highest exposures. 

The evidence for site-specific cancers is weaker, with some data suggesting a possible relationship between 

soft-tissue sarcoma, non-Hodgkin’s lymphoma, or respiratory cancer with 

2,3,7,8-TCDD exposure. It should be emphasized that some of the human studies do not provide adequate 

exposure data and were confounded by concomitant exposure to other chemicals. 

2,3,7,8-TCDD body burdens calculated from available serum lipid 2,3,7,8-TCDD levels are presented in 

Table 2-1. 
CDDs 68 

2. HEALTH EFFECTS 

2.2. ANIMAL STUDIES 

This section contains descriptions and evaluations of studies and presents levels of significant exposure for 

CDDs based on toxicological studies. 

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, developmental, reproductive, 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-adverseeffect levels (LOAELs) reflect the actual doses (levels of exposure) used in the studies. LOAELs have been 

classified into "less serious" or "serious" effects. These distinctions are intended to help the users of the 

document identify the levels of exposure at which adverse health effects start to appear. They should also 

help to determine 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 tables and figures may differ depending on the user's 

perspective. For example, physicians concerned with the interpretation of clinical findings in exposed 

persons may be interested in levels of exposure associated with "serious" effects. Public health officials and 

project managers 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 (LOAEL) or exposure levels 

below which no adverse effects (NOAEL) have been observed. Estimates of levels posing minimal risk to 

humans (Minimal Risk Levels, MRLs) may be of interest to health professionals and citizens alike. 

Levels of exposure associated with carcinogenic effects (Cancer Effect Levels, CELs) of CDDs are indicated 

in Tables 2-2, 2-3, and 2-4 and Figures 2-1 and 2-2. 

Estimates of exposure levels posing minimal risk to humans (MRLs) have been made (see Section 2.5), where 

data were believed reliable, for the most sensitive noncancer effect for each exposure duration. MRLs include 

adjustments to reflect human variability and extrapolation of data from laboratory animals to humans. 
CDDs 69 

2. HEALTH EFFECTS 

Although methods have been established to derive these levels (Barnes et al. 1988; EPA 1989c), 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. 

2.2.1 Inhalation Exposure 

No studies were located regarding the following health effects in animals after inhalation exposure to CDDs: 

2.2.1.1 Death 

2.2.1.2 Systemic Effects 

2.2.1.3 Immunological Effects 

2.2.1.4 Neurological Effects 

2.2.1.5 Reproductive Effects 

2.2.1.6 Developmental Effects 

2.2.1.7 Genotoxic Effects 

Genotoxicity studies are discussed in Section 2.5. 

2.2.1.8 Cancer 

No studies were located regarding cancer in animals after inhalation exposure to CDDs. 

2.2.2 Oral Exposure 

Information regarding adverse health effects in animals exposed to CDDs via the oral route was located 

for the following congeners: 2-monochlorodibenzo-p-dioxin (2-MCDD), 2,3-dichlorodibenzo-p-dioxin 

(2,3-DCDD), 2,7-dichlorodibenzo-p-dioxin (2,7-DCDD), 1,2,3-trichlorodibenzo-p-dioxin (1,2,3-TrCDD), 

1,2,3,4-tetrachlorodibenzo-p-dioxin (1,2,3,4-TCDD), 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(2,3,7,8-TCDD), 1,2,3,7,8-pentachlorodibenzo-p-dioxin (1,2,3,7,8-PeCDD), 1,2,3,4,7,8-hexachlorodibenzo-p-dioxin (1,2,3,4,7,8-HxCDD), 1,2,4,7,8-pentachlorodibenzo-p-dioxin (1,2,4,7,8-PeCDD), 
CDDs 70 

2. HEALTH EFFECTS 

1,2,3,6,7,8-hexachlorodibenzo-p-dioxin (1,2,3,6,7,8-HxCDD), 1,2,3,7,8,9-hexachlorodibenzo-p-dioxin 

(1,2,3,7,8,9-HxCDD), 1,2,3,4,6,7,8,-heptachlorodibenzo-p-dioxin (1,2,3,4,6,7,8,-HpCDD), and 

1,2,3,4,6,7,8,9-octachlorodibenzo-p-dioxin (OCDD). Some of the animal studies used a mixture of 

1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD. Of all the CDD congeners, 2,3,7,8-TCDD has been the one 

most extensively studied. 

2.2.2.1 Death 

Numerous studies provided doses associated with death following exposure to CDDs in animals. LD50 (lethal 

dose, kill for 50% of dosed animals during a certain time interval) values for each congener varied not only 

among species, but also among different strains. 

LD50 values following a single oral dose of 2,3,7,8-TCDD were calculated as 22 μg/kg (males) and 45 μg/kg 

(females) in Sherman rats (Schwetz et al. 1973); and 164 μg/kg, 297 μg/kg, 303 μg/kg, and 340 μg/kg in 

Fischer 344 rats from Charles River Breeding Laboratories, Charles River CD, Frederick Cancer Research 

Center, and Harlan Industries, respectively (Walden and Schiller 1985); 165 μg/kg (males) and 125 μg/kg 

(females) in Osborne Mendel rats (NTP 1982b); 43 μg/kg in male Sprague-Dawley rats (Stahl et al. 1992), 

and 60 and 100 μg/kg in female and male Long Evans rats, respectively (Fan and Rozman 1995). A single 

gavage dose of 100 μg/kg caused death in 95% of exposed male Fischer 344 rats (Kelling et al. 1985), and a 

dose of 25 μg/kg led to the death of 25% of exposed male Sprague-Dawley rats (Seefeld et al. 1984a). 

Furthermore, the reported LD50 values were 4.2 μg/kg in minks (Hochstein et al. 1988), 115 μg/kg in New 

Zealand albino rabbits (Schwetz et al. 1973b), 1.75 μg/kg in male Hartley guinea pigs (McConnell et al. 

1984), 0.6 μg/kg (males) and 2.1 μg/kg (females) in Hartley guinea pigs (Schwetz et al. 1973b), and 

1,157 μg/kg (Olson et al. 1980a) or 5,051 μg/kg (Henck et al. 1981) in Syrian hamsters. A 42-day LD50 of 

2.5 μg/kg was calculated for female Hartley guinea pigs when 2,3,7,8-TCDD was administered in corn oil and 

19 μg/kg when administered in methyl cellulose (Silkworth et al. 1982). No effect on survival was observed 

after a single oral dose of 200 μg/kg in B6C3F1 mice (NTP 1982b), but 69% of C57BL/6 mice died following 

exposure to 360 μg/kg (Kelling et al. 1985), and an LD50 was calculated as 146 μg/kg 2,3,7,8-TCDD in male 

C57BL mice (Smith et al. 1981). An acute LD50 in excess of 3,000 μg/kg was reported for male DBA/2J mice 

(Weber et al. 1995). Increased lethality was observed in Hartley guinea pigs exposed to 0.03 μg/kg/day 

2,3,7,8-TCDD in the feed for 11 days (DeCaprio et al. 1986) and in pregnant rabbits following 10 daily 

doses of 1 μg/kg during gestation (Giavini et al. 1982). Beagle dogs survived a single dose of 

300 μg/kg but not 3,000 μg/kg (Schwetz et al. 1973). In addition, 3 of 12 pregnant rhesus 
CDDs 71 

2. HEALTH EFFECTS 

monkeys died following a single dose of 1 μg/kg (McNulty 1984). It is evident from the above results that 

guinea pigs were the most sensitive species, while hamsters were the most resistant (up to 5,000 times greater 

lethal doses). In all studies cited above, the animals died following a latency period of several days (mean 

values varied from 9 to 43). In almost all laboratory animals, a pronounced wasting syndrome appears to be a 

major contributor to lethality. 

In the intermediate-duration experiments, increased lethality was observed in Osborne Mendel rats exposed to 

2,3,7,8-TCDD by gavage in oil vehicle at 0.56 μg/kg/day for up to 13 weeks (NTP 1982b). Mortality of 5% 

(no deaths in controls) was observed in Sprague-Dawley rats administered 2,3,7,8-TCDD by gavage at a rate 

of approximately 0.8 μg 2,3,7,8-TCDD/kg/day for 13 weeks (Viluksela et al. 1994); the first death occurred 

on day 57. Four of 7 male Sprague-Dawley rats dosed by gavage with approximately 1.6 μg 

2,3,7,8-TCDD/kg/day died in a 10-week study (Li and Rozman 1995); the mean time to death was 53.5 days. 

Increased mortality was reported in Hartley guinea pigs exposed daily for up to 60 days to diets that provided 

0.03 μg/kg/day (DeCaprio et al. 1986); 4 of 10 males died by day 42 and 4 of 10 females by day 59. In a 

dietary study, all male Sprague-Dawley rats that received the diet that provided the highest doses 

(3.4 μg/kg/day or more) died within 4 weeks (Van Miller et al. 1977). C57BL/6 mice had decreased survival 

following exposure by gavage to 3 μg/kg/day of 2,3,7,8-TCDD 3 days a week for 25 weeks (Umbreit et al. 

1987). Two monkeys were exposed intermittently by gavage to 0.6 μg/kg/day of 2,3,7,8-TCDD for 3 weeks 

and both died (McNulty 1984); 5 of 8 monkeys died within 2 months following exposure to diets that 

provided 0.02 μg/kg/day (Hong et al. 1989); also, 5 of 8 monkeys died within 9 months of dietary exposure to 

0.011 μg/kg/day (Allen et al. 1977). In all species, severe weight loss and body fat depletion were 

experienced prior to death, but usually no other overt toxic signs were observed. Pancytopenia, a secondary 

effect, was the cause of death in monkeys. 

Decreased survival was reported after chronic exposure to CDDs. Chronic dietary exposure to 2,3,7,8-TCDD 

increased the mortality over controls in Sprague-Dawley rats at 0.1 μg/kg/day (Kociba et al. 1978a). 

Increased mortality also occurred in Swiss mice given 2,3,7,8-TCDD by gavage at 1.0 μg/kg/day (Toth et al. 

1979) and in B6C3F1 mice at 0.36 μg/kg/day (Della Porta et al. 1987). In both studies, the mice were dosed 

once a week for 1 year and followed for the rest of their lives or until 110 weeks of age. No treatment-related 

effects on survival were observed in Osborne-Mendel rats or in B6C3F1 mice administered up to 0.25 μg 

2,3,7,8-TCDD/kg/day, 2 days a week by gavage for 104 weeks(0.071 μg/kg/day for rats and male mice; 

0.3 μg/kg/day for female mice) (NTP 1982b). 
CDDs 72 

2. HEALTH EFFECTS 

Increased mortality occurred after acute exposure to other congeners. After a single oral dose of a mixture of 

1,2,3,7,8,9-HxCDD and 1,2,3,6,7,8-HxCDD, LD50 values were calculated as 1,800 μg/kg and 800 μg/kg in 

male and female Osborne-Mendel rats, respectively, and 750 μg/kg and 500 μg/kg in male and female 

B6C3F1 mice, respectively (NCI/NTP 1980). In addition, LD50 values were calculated for several congeners 

in guinea pigs (29,444 μg/kg for 1,2,3-TrCDD, 1,125 μg/kg for 1,2,4,7,8-PeCDD, 3.1 μg/kg for 

1,2,3,7,8-PCDD, 70–100 μg/kg for 1,2,3,6,7,8-HxCDD, 60–100 μg/kg for 1,2,3,7,8,9-HxCDD, and 

72.5 μg/kg for 1,2,3,4,7,8-HxCDD) and in mice (825 μg/kg for 1,2,3,4,7,8-HxCDD and 337.5 μg/kg for 

1,2,3,7,8-PCDD) following a single oral exposure by gavage in oil vehicle (McConnell et al. 1978b). In male 

Sprague-Dawley rats, the oral LD50 for 1,2,3,7,8-PCDD, 1,2,3,4,7,8-HxCDD, and 1,2,3,4,6,7,8-HpCDD 

administered in corn oil/acetone (95/5) was 206, 887, and 6,325 μg/kg, respectively (Stahl et al. 1992). Other 

CDD congeners have a much lower order of toxicity, as evidenced by data showing no effects on mortality at 

much higher doses than those of 2,3,7,8-TCDD, TrCDD, HxCDD, or PCDD that cause death. No deaths were 

observed after a single oral dose of 1×106

 and 2×106 μg/kg 2,7-DCDD in Sprague-Dawley rats and Swiss 

Webster mice, respectively (Schwetz et al. 1973). In addition, rats and mice survived acute oral doses of 

1×106

 and 4×106 μg/kg OCDD, respectively (Schwetz et al. 1973). The relative species differences in 

sensitivity for 2,3,7,8-TCDD also applied for other congeners. 

Mortality rates of 15 and 50% were reported in groups of male Sprague-Dawley rats administered 73 and 

110 μg 1,2,3,4,6,7,8-HpCDD/kg/day by gavage for 13 weeks, respectively (Viluksela et al. 1994). At the 

highest dose, the first death occurred on day 31; at the 73 μg/kg/day dose, on day 41. Fifteen out of 20 

female Sprague-Dawley rats died during a 13-week treatment period with daily doses of approximately 2.6 μg 

1,2,3,7,8-PeCDD/kg (total dose was 233 μg/kg) (Viluksela et al. 1998a). The first death occurred on day 16. 

The same mortality rate was observed in males treated with approximately 3.8 μg/kg/day (total dose was 

350 μg/kg). In the same study, administration of approximately 10.3 μg 1,2,3,4,7,8-HxCDD resulted in a 

25% death rate (5/20, first death on day 61) in female rats; the same death rate was seen among male rats 

treated with approximately 15.4 μg/kg/day (first death on day 24). The main causes of death were wasting 

syndrome, hemorrhage, and anemia (Viluksela et al. 1998a). No effects on survival were observed following 

chronic dietary exposure of Osborne-Mendel rats and B6C3F1 mice to 5×105

 μg/kg/day of 2,7-DCDD and to 

1.3×106

 μg/kg/day of 2,7-DCDD, respectively (NCI/NTP 1979a), or following chronic gavage dosing with a 

mixture of 1,2,3,7,8,9-HxCDD and 1,2,3,6,7,8-HxCDD at 0.34 μg/kg/day and 0.7 μg/kg/day, respectively 

(NCI/NTP 1980). 
 

CDDs 73 

2. HEALTH EFFECTS 

In conclusion, 2,3,7,8-TCDD was the most toxic of all congeners tested, and doses on the order of 

several μg/kg body weight have led to death in all species tested, except hamsters and dogs, in acute-exposure 

experiments. In contrast, of the congeners tested, 2,7-DCDD and OCDD were the least toxic as tested 

animals survived very high doses (g/kg body weight). The wasting syndrome was the major toxic effect of 

acute- and intermediate-duration exposure to CDDs in most species. It was characterized by body weight 

loss, adipose tissue depletion, and eventual death. In most of the chronic duration studies the cause 

of death was not determined. 

The LD50 values and all reliable representative LOAEL values for death in each species and duration category 

for each congener tested are recorded in Tables 2-2 and 2-3 and plotted in Figures 2-1 and 2-2. 

2.2.2.2 Systemic Effects 

The highest NOAEL values and all reliable representative LOAEL values for each systemic effect in each 

species and duration category for each congener tested are recorded in Tables 2-2 and 2-3 and plotted in 

Figures 2-1 and 2-2. 

Respiratory Effects. Few studies have examined the respiratory system in animals following oral 

exposure to CDDs. However, serious respiratory effects have been observed in monkeys that died from 

2,3,7,8-TCDD exposure. 

Bleeding from the nose was reported in rhesus monkeys exposed via gavage to 0.1 μg/kg/day, 3 days a week 

for 3 weeks (McNulty 1984). Hemorrhage, hyperplasia, and metaplasia of the bronchial epithelium (as well 

as at other organ sites that had mucous-secreting cells) developed in monkeys exposed to diets providing 

0.011 μg/kg/day for 9 months (Allen et al. 1977); 5 of 8 monkeys died with this dose level. Focal alveolar 

hyperplasia and squamous metaplasia and carcinoma were reported in Sprague-Dawley rats chronically 

exposed to 0.1 μg/kg/day 2,3,7,8-TCDD in the feed (Kociba et al. 1978a). Since powdered feed containing 

2,3,7,8-TCDD was given to the rats, there is a distinct possibility that the respiratory effects were attributable 

to inhalation exposure rather than oral systemic absorption. In contrast, no respiratory effects were observed 

in rats or mice chronically exposed by gavage to 2,3,7,8-TCDD at approximately 0.071 μg/kg/day or 

0.3 μg/kg/day, respectively (NTP 1982b). 






































































 

CDDs 144 

2. HEALTH EFFECTS 

Similarly, no respiratory effects were found in rats and mice chronically exposed by diet to 5×105 and 

1.3×106

 μg/kg/day of 2,7-DCDD, respectively (NCI/NTP 1979a). In contrast, rats exposed chronically by 

gavage to a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD at 0.18, 0.34, and 0.7 μg/kg/day had a 

dose-related increased incidence of adenomatous hyperplastic lesions in terminal bronchioles and adjacent 

alveoli of both males and females; no such effects were found in mice exposed chronically to 0.7 μg/kg/day 

of that same mixture (NCI/NTP 1980). The existing information suggests that in animals, the respiratory 

system is not a sensitive target for CDDs toxicity via oral exposure. 

Cardiovascular Effects. Cardiovascular effects have been detected in animals following acute-, 

intermediate-, and chronic-duration oral exposure to 2,3,7,8-TCDD. These included changes in heart weight, 

pathophysiological effects, and degenerative changes. However, exposures at or near a lethal dose were 

required to elicit these effects. 

Decreased absolute heart weight was reported in minks 28 days after a single oral dose of 5 μg/kg, but not at 

2.5 μg/kg (Hochstein et al. 1988). A reduction of absolute heart weight which is attributed to weight loss was 

also found in monkeys at 70 μg/kg (relative heart weight was increased) (McConnell et al. 1978a). 

Histological examinations of the heart were normal in the monkeys. This examination was not performed in 

minks. Doses in both species were near the lethal dose. 

Kelling et al. (1987) assessed the effects of 2,3,7,8-TCDD on cardiac function tests in male Sprague-Dawley 

rats 7 days after single oral doses of 6.25, 25, or 100 μg/kg. At 100 μg/kg (near-lethal dose), an increased 

sensitivity to the inotropic (left atrium) and chronotropic (right atrium) effects of isoproterenol were observed. 

Three daily oral doses of 40 μg/kg caused decreased heart rate, depressed blood pressure, and increased 

myocardial peroxidase activity in rats (Hermansky et al. 1988). All of these effects may have been secondary 

to the modulation of adenylate cyclase activity at β-adrenergenic receptors as a result of hypothyroidism 

(Hermansky et al. 1987). 

In intermediate-duration experiments, monkeys that died after exposure to diets providing 0.011 μg/kg/day of 

2,3,7,8-TCDD (lethal dose) had hemorrhages in the epicardium, myocardium, and endocardium (Allen et al. 

1977). Myocardial degenerative changes and periarteritis were reported in Sprague-Dawley rats chronically 

exposed to a diet providing a lethal dose of 0.1 μg/kg/day of 2,3,7,8-TCDD, but not in those receiving 

0.01 μg/kg/day (Kociba et al. 1978a). In contrast, no histopathological lesions were observed in 
 

CDDs 145 

2. HEALTH EFFECTS 

the hearts of rats and mice chronically exposed by gavage to approximately 0.071 and 0.3 μg/kg/day of 

2,3,7,8-TCDD, respectively (NTP 1982b). 

No histopathological lesions were observed in the hearts of rats and mice chronically exposed in the diet to 

5×105

 and 1.3×106 μg/kg/day of 2,7-DCDD, respectively (NCI/NTP 1979a); or exposed for 104 weeks by 

gavage to approximately 0.34 and 0.7 μg/kg/day of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD, 

respectively (NCI/NTP 1980). 

Gastrointestinal Effects. One of the major 2,3,7,8-TCDD-induced effects in various animal species is 

the wasting syndrome and hypophagia which occur after a single near-lethal dose or after repeated dosing 

(discussed under Body Weight Effects). Studies of effects on the gastrointestinal system have been carried 

out to investigate the mechanism of this starvation-like syndrome. Ulceration of the gastrointestinal tract and 

bloody stools were observed in minks after a single oral exposure to 5 μg 2,3,7,8-TCDD/kg (3 of 4 mink 

died) but not at a dose of 2.5 μg/kg/day. The response of the antral mucosa of the rat stomach to 

2,3,7,8-TCDD has been studied by Theobald et al. (1991). In Sprague-Dawley rats, a single oral dose of 

100 μg 2,3,7,8-TCDD/kg caused a 7–10-fold increase in serum gastrin (secreted by G-cells in the antrum) that 

was not detected until 14 days after dosing, whereas control rats fed a restricted diet had atrophic changes in 

the antral mucosa and no increase in gastrin (Theobald et al. 1991). The number of G-cells in the antral 

mucosa was not affected by treatment with 2,3,7,8-TCDD or paired-feed restriction, indicating that 

hypergastrinemia in treated rats is not due to reduced feed intake or antral G-cell hyperplasia. In 

2,3,7,8-TCDD-treated rats, both gastrin and somatostatin (which inhibits gastrin release) levels in the antral 

mucosa were significantly decreased, and these changes were observed a week earlier than the 

hypergastrinemia. Moreover, the ED50 values (half maximum effect level of 2,3,7,8-TCDD) for the decrease 

in antral mucose content and concentration of gastrin (29 and 22 μg/kg, respectively) and somatostatin (24 

and 19 μg/kg, respectively) was less than that for hypergastrinemia (46 μg/kg). This suggested that 

hypergastrinemia in 2,3,7,8-TCDD-treated rats is not a consequence of reduced antral levels of gastrin or 

somatostatin. Epithelial hyperplasia of the stomach occurred in rhesus monkeys following a single oral dose 

of 70 μg/kg; this response is unique to monkeys and cows and is not seen in rats, mice, or guinea pigs 

(McConnell et al. 1978b). Monkeys undergo a similar wasting syndrome as rodents after a single oral lethal 

dose. Moderate to severe ileitis, characterized by hyperplasia of the mucosal epithelium with hemorrhaging 

and necrosis, and peritonitis were observed in hamsters that died after oral administration of $1000 μg/kg 

2,3,7,8-TCDD (Olson et al. 1980a). 
 

CDDs 146 

2. HEALTH EFFECTS 

Repeated dosing of rats at 3.4 μg/kg/day or higher caused gastrointestinal hemorrhaging in rats that died in a 

chronic oral-dosing study (Van Miller et al. 1977). Metaplasia of the gastric mucosa was found in rhesus 

monkeys exposed to 0.1 μg/kg/day of 2,3,7,8-TCDD for 3 weeks (McNulty 1984), and gastric ulcers 

developed after exposure to 0.011 μg/kg/day for 9 months in the feed (Allen et al. 1977). No gastrointestinal 

effects were observed in rats and mice chronically exposed by gavage to approximately 0.071and 

0.3 μg/kg/day of 2,3,7,8-TCDD, respectively (NTP 1982b) or in rats on diets that provided 0.1 μg/kg/day 

(Kociba et al. 1978a). 

Gastrointestinal lesions were not observed following exposure of rats and mice to 5×105

 and 

1.3×106

 μg/kg/day of 2,7-DCDD, respectively, in the diet (NCI/NTP 1979a) or to 0.34 and 0.7 μg/kg/day of a 

mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD, respectively, by gavage for 104 weeks (NCI/NTP 

1980). 

The above studies demonstrated that monkeys are more sensitive to gastrointestinal effects of 2,3,7,8-TCDD 

than rodents. 

Hematological Effects. Hematological effects were reported in some animals following exposure to 

lethal or near-lethal doses of 2,3,7,8-TCDD. Increases in erythrocyte counts, hemoglobin, and hematocrit 

were observed 10–14 days after CD rats received a single oral dose of 10 μg/kg 2,3,7,8-TCDD. Increases in 

total leukocyte count and neutrophil counts, and a decrease in platelet counts were also observed, but bleeding 

time and megakaryocytes were not altered (Weissberg and Zinkl 1973). Reduction of germinal centers and 

increased hemosiderin deposits were seen histologically in the spleen of Sprague-Dawley rats after a single 

oral dose of 25 μg/kg (Christian et al. 1986a). Mild anemia developed in rhesus monkeys after a single oral 

dose of 70 μg/kg (McConnell et al. 1978a). No effects were found in minks exposed acutely to a lethal dose 

of 2,3,7,8-TCDD (7.5 μg/kg) (Hochstein et al. 1988) or in B6C3F1 mice exposed to 1 μg/kg/day for 14 days 

(Holsapple et al. 1986a). Reversible changes (suppression of progenitor cells, decreases in leukocyte and 

lymphocyte counts) were reported in CD-1 mice at doses between 1 and 10 μg/kg 2,3,7,8-TCDD (Zinkl et al. 

1973). 

Hematological effects have also been reported following intermediate-duration exposures to 2,3,7,8-TCDD. 

Decreased white blood cell counts were reported in guinea pigs exposed by gavage to 0.008 μg/kg/day 

2,3,7,8-TCDD for 8 weeks (Vos et al. 1973), but no hematological changes were observed following dietary 

exposure to 0.005 μg/kg/day 13 weeks (DeCaprio et al. 1986). Exposure to higher doses 
CDDs 147 

2. HEALTH EFFECTS 

(3.4 μg/kg/day or more) caused splenic atrophy in Sprague-Dawley rats that died during the first 4 weeks of 

exposure in a chronic-duration dietary study (Van Miller et al. 1977). In contrast, no hematological changes 

were found in rats exposed to 0.71 μg/kg/day of 2,3,7,8-TCDD for 6 weeks (Vos et al. 1973). One month of 

intermittent exposure to 0.1 μg/kg/day 2,3,7,8-TCDD induced thrombocytopenia in CD rats (Zinkl et al. 

1973); exposure to 1 μg/kg/day caused increased erythrocyte counts and hemoglobin levels. Administration 

of 2,3,7,8-TCDD by gavage for 13 weeks to male Sprague-Dawley rats at doses equivalent to 0.8 μg/kg/day 

(only dose level tested) produced a significant decrease in platelet counts, and in some animals, increased 

prothrombin times (Viluksela et al. 1994). Anemia and bone marrow hypoplasia were observed in rhesus 

monkeys exposed to 0.1 μg/kg/day of 2,3,7,8-TCDD by gavage 3 days a week for 3 weeks (McNulty 1984). 

The changes were more severe with longer exposure; pancytopenia and bone marrow atrophy developed in 

monkeys exposed to 0.011 μg/kg/day (a lethal dose) in the feed for 9 months (Allen et al. 1977). 

In chronic-duration studies, reduced erythrocyte counts were found in Sprague-Dawley rats at dietary doses of 

0.1 μg/kg/day of 2,3,7,8-TCDD but not at 0.01 μg/kg/day (Kociba et al. 1978a). No hematological effects 

were observed in Osborne-Mendel rats or B6C3F1 mice chronically exposed by gavage to approximately 

0.071or 0.3 μg/kg/day of 2,3,7,8-TCDD, respectively (NTP 1982b). Results from a more recent study 

showed that 2,3,7,8-TCDD administered by gavage to female C57BL/6 mice in gavage doses equivalent to 

approximately 0.03 μg/kg/day (0.2 μg/kg once/week) for 14–15 months produced no significant effects on the 

total number of circulating red or white blood cells or in white blood cell differentials (Oughton et al. 1995). 

Hematological effects have been reported in some animals following exposure to other CDDs. No hematological 

effects were observed in rats after 2 weeks of intermittent exposure to 50 μg/kg/day OCDD (Couture et al. 1988), 

but increased neutrophils, decreased mean cell volume, and hemoglobin (Couture et al. 1988), and mild anemia 

was observed at the same exposure level after 13 weeks of intermittent exposure (Birnbaum et al. 1989a). A 

dose-dependent decrease in platelet counts was observed in male Sprague-Dawley rats following administration 

by gavage of doses equivalent to 73 or 110 μg 1,2,3,4,6,7,8-HpCDD/kg/day for 13 weeks (Viluksela et al. 1994); 

no such effect was observed with doses #24 μg/kg/day. Some rats administered the highest dose also showed 

increased prothrombin times. Administration of doses equivalent to 2.6 μg 1,2,3,7,8-PCDD/kg/day or 10.3 μg 

1,2,3,47,8-HxCDD/kg/day for 13 weeks resulted in decreased hematocrit and reduced platelet count in female 

Sprague-Dawley rats (Viluksela et al. 1998a); these doses also caused mortality. 
 

 

CDDs 148 

2. HEALTH EFFECTS 

Splenic hyperplasia was observed in rats exposed by gavage to a mixture of 1,2,3,6,7,8-HxCDD and 

1,2,3,7,8,9-HxCDD at 7.1 μg/kg/day, but not at 1.4 μg/kg/day for 13 weeks (NCI/NTP 1980). No 

hematological effects were observed in Osborne-Mendel rats or B6C3F1 mice chronically exposed to 5×105

 and 

1.3×106

 μg/kg/day of 2,7-DCDD, respectively, in feed (NCI/NTP 1979a) or exposed to 0.34 and 0.7 μg/kg/day 

of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD, respectively, 2 days a week for 104 weeks by 

gavage (NCI/NTP 1980). 

The above results demonstrated hematological effects in animals following CDD exposure; however, the 

observed changes in the red and white blood cell counts were nonspecific and were probably due to the broad 

systemic toxicity of 2,3,7,8-TCDD rather than to a direct effect on the hematological system. 

Musculoskeletal Effects. The musculoskeletal system does not appear to be a major target of toxicity in 

animals exposed to CDDs. Only one study reported hemorrhages in the musculoskeletal system of severely 

debilitated monkeys following dietary exposure to 0.011 μg/kg/day of 2,3,7,8-TCDD for an intermediate 

duration (Allen et al. 1977). 

No musculoskeletal effects were observed in Sprague-Dawley rats exposed to 0.1 μg/kg/day in the diet for 

2 years (Kociba et al. 1978a) or in Osborne-Mendel rats and B6C3F1 mice chronically exposed 2 days a week 

by gavage to 0.071and 0.3 μg/kg/day of 2,3,7,8-TCDD, respectively (NTP 1982b). 

Chronic experiments with other congeners showed no musculoskeletal effects in Osborne-Mendel rats and 

B6C3F1 mice exposed in the diet to 5×105

 and 1.3×106 μg/kg/day of 2,7-DCDD, respectively (NCI/NTP 1979a) 

or by gavage to approximately 0.34 and 0.7 μg/kg/day of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9­

HxCDD, respectively (NCI/NTP 1980). 

Hepatic Effects. Effects on the liver are seen after acute oral exposure or after intermediate and chronic 

exposure to CDDs. Alterations in metabolism, biochemical changes, and increases in liver weights (without 

histologic changes) are sensitive markers of effects, but they are not clearly overt toxic effects; they may 

predict a toxic or histopathologic effect that will occur at higher doses or after longer exposure. Likewise 

induction of mixed-function oxidases (MFO) and cytochrome P-450s are generally considered adaptive 

effects; they may be associated with increased liver weight, but are not necessarily associated with 

histopathologic changes. However, alterations in cytochrome P-450 (e.g., CYP1A1) may lead to altered 

metabolism and/or toxicity of other xenobiotics and endogenous compounds. Increases in liver enzymes 
CDDs 149 

2. HEALTH EFFECTS 

such as AST and ALT in serum are an indication of cell death or necrosis. Effects on the liver that occur at 

near-lethal doses or when animals are debilitated and approaching moribundity are secondary effects and are 

not specific to the action of CDDs on the liver. The liver should not be implicated as a target organ in these 

cases. The types of histological changes caused by CDDs and their severity vary widely between species and 

strains of laboratory animals and the doses administered. 

Histological changes in the liver were observed after acute-, intermediate-, and chronic-duration exposures. 

Enlarged hepatocytes with finely vacuolated cytoplasm in the centrilobular region were observed 20 days 

after administration of a single oral lethal dose of 75 μg/kg 2,3,7,8-TCDD to adult Sprague-Dawley rats 

(Christian et al. 1986a); cellular degeneration in the centrilobular region was observed in Fischer 344 rats 

20 days after a single oral dose of 1000 μg/kg (Kelling et al. 1985). Three daily doses of 40 μg/kg caused 

centriacinar necrosis and enlarged hepatocytes in Sprague-Dawley rats (Hermansky et al. 1988). In all the 

above citations, the changes were secondary to the wasting syndrome and occurred after severe weight loss. 

Focal areas of mild hydropic degeneration of the liver associated with increased liver weight was seen in 

B6C3F1 mice receiving 1 μg/kg/day 2,3,7,8-TCDD for 14 days (non-lethal) (Holsapple et al. 1986a). 

Swelling of hepatocytes, disruption of cell membranes, and dilation of sinusoids in the central vein area of the 

liver accompanied by cellular necrosis were observed after a single lethal dose of 360 μg/kg in C57B46 mice 

(Kelling et al. 1985); central necrosis in the livers of hairless A2G hr/+ mice dosed once at 75 μg/kg was 

reported (Greig 1984; Greig et al. 1987). Guinea pigs only showed minimal focal necrosis at 42 days after a 

single oral dose of 0.1 μg/kg 2,3,7,8-TCDD (Turner and Collins 1983). Hypertrophy, steatosis, cytoplasmic 

degeneration, and hyaline-like cytoplasmic inclusion bodies were observed at non-lethal and lethal doses (0.5 

to 20 μg/kg); no between group qualitative differences in this histological alteration were found (Turner and 

Collins 1983). No degenerative changes were observed by Kelling et al. (1985) prior to death in guinea pigs 

given a single oral lethal dose of 2 μg/kg. In minks, a sensitive species, pale and mottled livers were observed 

at gross necropsy 28 days after a single lethal oral dose of 5 μg/kg (Hochstein et al. 1988). 

Mild-to-moderate hepatic effects were also seen after intermediate-duration exposure to 2,3,7,8-TCDD. 

Unspecified histopathological hepatic lesions were reported following intermittent exposure to 

2,3,7,8-TCDD in female Osborne-Mendel rats at 0.07 μg/kg/day and in female B6C3F1 mice at 

0.7 μg/kg/day for 13 weeks (NTP 1982b). Cytomegaly with focal necrosis was observed in C57BL/6J mice 

exposed to 10 μg/kg/day by gavage 1 day/week for 4 weeks (Thigpen et al. 1975). Increase relative liver 

weight and hepatocellular inclusions were found in guinea pigs exposed to 0.005 μg/kg/day in the feed 
CDDs 150 

2. HEALTH EFFECTS 

for 90 days (DeCaprio et al. 1986); this dose level also significantly reduced serum ALT activity in females 

and increased triglycerides in males. Biliary epithelial hyperplasia has been reported in monkeys following 

exposure to lethal levels (0.011 μg/kg/day) in feed for 9 months (Allen et al. 1977) and after intermittent 

exposure at 0.1 μg/kg/day by oral gavage for 3 weeks (McNulty 1984). 

Liver necrosis occurred in Sprague-Dawley rats that died during the first 4 weeks of dietary exposure to 

2,3,7,8-TCDD at 3.4 μg/kg/day in a chronic-exposure experiment, while no non-cancerous liver effects 

were found in rats chronically exposed to 0.286 μg/kg/day (Van Miller et al. 1977). Toxic hepatitis 

characterized by lipidosis and hydropic degeneration of hepatocytes with proliferation of bile ductules and 

by mild fibrosis was observed in 14/50 male and 32/49 female Osborne-Mendel rats following exposure by 

gavage to approximately 0.071 μg/kg/day of 2,3,7,8-TCDD administered for 104 weeks and in 44/50 male 

B6C3F1 mice receiving 0.071μg/kg/day or 34/47 female mice receiving 0.3 μg/kg/day (NTP 1982b). In 

addition, cytoplasmic vacuolation, hyperplasia, hepatocellular degeneration, and liver necrosis occurred in 

Sprague-Dawley rats chronically exposed to diets providing doses of 0.001 (females) and 0.01 (both 

sexes) μg/kg/day 2,3,7,8-TCDD, respectively (Kociba et al. 1978a). 

Biochemical changes indicating liver effects following acute oral exposure to 2,3,7,8-TCDD included 

hypoglycemia and increased serum triglycerides and cholesterol 10 days after a single sublethal oral dose of 

45 μg/kg in Fischer 344 rats (Walden and Schiller 1985); earlier, Albro (1978) found increased triglycerides 

and decreased sterol esters after a non-lethal dose of 2,3,7,8-TCDD and increased cholesterol and free fatty 

acids after a lethal dose. Reduced retinol storage in the liver was found in Sprague-Dawley rats exposed to 

a single dose of 1 μg/kg 2,3,7,8-TCDD (Thunberg 1984). Reduction of hepatic retinol by 2,3,7,8-TCDD 

was greater (87%) in younger rats with lower initial weights (Thunberg et al. 1984) than in more mature rats 

(60%) (Thunberg et al. 1979, 1980). Significant and maximum induction of hepatic ethoxyresorufin-Odeethylase (EROD, marker for CYP1A1 activity) activity and dose-related decrease in liver 

phosphoenolpyruvate carboxykinase (PEPCK, a key enzyme of gluconeogenesis) was observed in female 

Long Evans rats 4 days after a single gavage dose of 5.3–60 μg 2,3,7,8-TCDD/kg (Fan and Rozman 1995). 

Hepatic activity of tryptophan 2,3-dioxygenase (TdO, a key enzyme of tryptophan metabolism) was 

elevated by 2,3,7,8-TCDD-treatment (significantly at 5.3, 12, and 18 μg/kg but not 60 μg/kg), whereas 

serum tryptophan levels were not altered. EROD activity had diminished considerably 90 days after 

dosing, although it was still 10 times the control values, and PEPCK and TdO activities had returned to 

control values. The authors concluded that gluconeogenesis is inhibited by 2,3,7,8-TCDD in Long Evans 

rats by reducing liver PEPCK activity (Fan and Rozman 1995). 
CDDs 151 

2. HEALTH EFFECTS 

Hepatic porphyria has been found after acute oral dosage of mice with 2,3,7,8-TCDD. A single oral dose of 

150 μg/kg or 4 weekly doses of 25 μg/kg caused an accumulation of porphyrins and induction of delta 

aminolevulinic acid synthetase in B6C3F1 mice (Goldstein et al. 1973). These were lethal doses and caused 

severe histologic liver damage. Single oral doses as high as 30 μg/kg did not produce porphyria acutely or 

within 16 weeks (Goldstein 1982). Elevation of serum levels of ALT and sorbitol dehydrogenase, which 

are indicative of subclinical toxic effects on the liver, have been found in mice after a single oral non-lethal 

dose (Greig 1984; Rosenthal et al. 1989; Smith et al. 1981). Smith et al. (1981) compared the sensitivity of 

C57BL/10 mice and DBA/2 mice, and found that the DBA/2 mouse was 20 times less sensitive than the 

C57 strain to 2,3,7,8-TCDD-induced porphyria. Recent results from Weber et al. (1995) suggested that 

acute toxicity of 2,3,7,8-TCDD occurs between 37.5 and 235 μg/kg in male C57BL/6J mice and between 

375 and 3,295 μg/kg in male DBA/2J mice, as judged by decreases in liver PEPCK and glucose6-phosphatase (G-6-Pase, also a key enzyme of gluconeogenesis) activities, reduction in blood glucose, and 

changes in relative liver weight 8 days after a single gavage dose of 2,3,7,8-TCDD. The ED50 for induction 

of hepatic EROD activity in male C57BL/6J mice was estimated at 1.1 μg/kg compared with 16 μg/kg in 

DBA/2J mice. Also there was no evidence of a reduction of liver TdO activity or of elevation of serum 

tryptophan levels over the dose range tested (0.03–235 μg/kg in C57BL/6J mice and 1–3,295 μg/kg in 

DBA/2J mice). Dose-dependent induction of EROD has also been observed in the liver from female 

B6C3F1 mice after single (Diliberto et al. 1995) and repeated (DeVito et al. 1994) oral 2,3,7,8-TCDD doses. 

In the repeated-dosing study (DeVito et al. 1994), both EROD and acetanilide-4-hydroxylase (marker for 

CYP1A2) activities were induced with doses as low as 1.5 ng 2,3,7,8-TCDD/kg/day. Pegram et al. (1995) 

found no differences in the dose-response curves for hepatic EROD induction between young and old male 

C57BL/6N mice 8 days after a single dose of 0.015–15 μg 2,3,7,8-TCDD/kg. However, induction of 

acetanilide-4-hydroxylase was significantly greater in old than in young mice. Also a trend of greater 

relative liver weight with increasing dioxin dose was observed in young mice, whereas liver weight was not 

altered in old mice (Pegram et al. 1995). 

Increased liver weights were reported in pregnant mice that received 25 μg/kg/day (Courtney 1976), 

3 μg/kg/day (Smith et al. 1976), or 0.5 μg/kg/day (Silkworth et al. 1989b) for 10 days during gestation, and 

in monkeys receiving a single oral dose of 70 μg/kg (McConnell et al. 1978a). The ED50 values for liver 

enlargement following a single oral dose of 2,3,7,8-TCDD were calculated as 100 μg/kg 2,3,7,8-TCDD in 

Sprague-Dawley rats, 1,000 μg/kg in C57BL/6 mice, and 14 μg/kg in Syrian hamsters (Hanberg et al. 

1989). These ED50 values approach and exceed the LD50 values for rats and mice (see Section 2.2.2.1). 
CDDs 152 

2. HEALTH EFFECTS 

In a 1-week dietary study in female Sprague-Dawley rats, a dose of 1 μg 2,3,7,8-TCDD/kg/day induced a 

significant increase in absolute liver weight and a lower dose of 0.32 μg/kg/day significantly increased 

relative liver weight (Van Birgelen et al. 1995). Other hepatic effects observed in this study included 

significantly dose-related increased hepatic microsomal activities of EROD and acetanilide-4-hydroxylase, 

beginning at the lowest dose tested (0.014 μg/kg/day), and dose-related decrease in hepatic retinol at 

$0.014 μg/kg/day. In agreement with the results of Van Birgelen et al. (1995), Viluksela et al. (1994) also 

reported increases in absolute and relative liver weights in male Sprague-Dawley rats administered 

2,3,7,8-TCDD by gavage for 13 weeks at doses that supplied approximately 0.8 μg 2,3,7,8-TCDD/kg/day 

(the only dose level tested). An increase in liver EROD activity and decrease in liver PEPCK activity were 

also reported in the study. Liver TdO activity and total serum tryptophan were not significantly altered in 

surviving rats, but TdO activity was significantly decreased in moribund animals with signs of wasting 

syndrome, whereas serum tryptophan levels were doubled in these animals. Li and Rozman (1995) 

examined the reversibility of 2,3,7,8-TCDD-induced changes in some liver enzymes in male SpragueDawley rats treated by gavage with doses equivalent to 0.003–1.6 μg/kg/day for 10 weeks and allowed to 

recover for an additional 6-week period. As reported by others, there was a dose-dependent decrease in 

TdO activity with a concurrent increase in serum tryptophan levels (both significant at the highest-dose 

level) and a decrease in PEPCK activity (significant at $1 μg/kg/day). These dose responses were very 

similar to the dose response for body weight reduction. EROD was induced even at the lowest dose and 

maximum induction was attained at $35 μg/kg/day. After the 6-week recovery period, PEPCK and TdO 

activities, as well as serum tryptophan levels, returned to near-control levels; however, EROD still remained 

induced. The authors (Li and Rozman 1995) indicated that the results supported the hypothesis that 

subchronic toxicity of 2,3,7,8-TCDD is similar to its acute toxicity when the dose is corrected for 

pharmacokinetics. In other words, toxicity is determined by the body burden represented by the cumulative 

dose minus the portion of the dose already eliminated. 

Hepatic effects have also been reported in animals following exposure to other CDDs. Pale, friable livers 

were observed in 3/20 Sprague-Dawley rat dams exposed to 100 μg/kg/day of mixed HxCDD during 

gestation (incidence in control group was not reported), but not in those exposed to 10 μg/kg/day (Schwetz 

et al. 1973). No effects on the liver were observed at 10 μg/kg/day of 2,7-DCDD in B6C3F1 mice 

(Holsapple et al. 1986b), at 20 μg/kg/day of OCDD in pregnant CD-1 mice (Courtney 1976), or at 

1,000 μg/kg/day of 1,2,3,4-TCDD in pregnant CD-1 mice (Courtney 1976). 
 

CDDs 153 

2. HEALTH EFFECTS 

Data for hepatic effects after intermediate-duration exposure to the other congeners were available for 

1,2,3,7,8-PCDD, 1,2,3,4,7,8-HxCDD, a mixture of 1,2,3,7,8,9-HxCDDs and 1,2,3,6,7,8-HxCDD, 

1,2,3,4,6,7,8-HpCDD, and OCDD. Mild hepatotoxicity (not otherwise specified) was recorded in rats 

exposed to 0.71 μg/kg/day of a mixture of 1,2,3,7,8,9-HxCDD and 1,2,3,6,7,8-HxCDD, and in mice 

exposed to the same mixture by gavage for 13 weeks to 1.4 μg/kg/day (NCI/NTP 1980). No effects were 

seen at 0.36 μg/kg/day and at 0.71 μg/kg/day in rats and mice, respectively. Absolute liver weight was 

significantly increased in male Sprague-Dawley rats administered $24 μg 1,2,3,4,6,7,8-HpCDD/kg/day by 

gavage for 13 weeks (Viluksela et al. 1994). Relative liver weight was increased at $4 μg/kg/day. Liver 

activity of PEPCK was significantly decreased with the 2 highest-dose levels tested, 73 and 110 μg/kg/day, 

whereas hepatic EROD activity was dose-dependently induced over the dose range tested, 0.3 to 

110 μg/kg/day. Liver TdO activity and serum total tryptophan were not significantly altered at 

#24 μg/kg/day; however, TdO was decreased and serum tryptophan was increased in rats that died at the 

two highest dose levels. Similar results were reported in rats treated with 1,2,3,7,8-PCDD 

(2–4 μg/kg/day) or 1,2,3,4,6,7-HxCDD (10–15 μg/kg/day) (Viluksela et al. 1998b). Cytoplasmic 

vacuolization of hepatocytes (Couture et al. 1988) and liver hypertrophy with induced hepatic enzymes 

(Birnbaum et al. 1989a) were reported in rats gavaged with 50 μg/kg/day OCDD 5 days a week for up to 

13 weeks. 

Toxic hepatitis was reported in Osborne-Mendel rats and in B6C3F1 mice following gavage exposure to 

0.18 μg/kg/day and to 0.34 μg/kg/day of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD, 

respectively, for 104 weeks (NCI/NTP 1980). The corresponding NOAEL values are recorded in 

Table 2-3. Furthermore, fatty changes in the liver were found in rats chronically exposed to 2,7-DCDD at a 

dose of 2.5×105

 μg/kg/day in the feed (NCI/NTP 1979a). In contrast, no liver effects were observed in mice 

following chronic exposure to 1.3×106

 μg/kg/day of 2,7-DCDD in the feed (NCI/NTP 1979a). 

In conclusion, the above studies demonstrated that the liver is a primary target of CDD toxicity. 

2,3,7,8-TCDD was the most toxic congener, but other congeners were also capable of inducing hepatic 

effects. The induced effects were dose-related and species- and strain-related. It also appeared that for 

some hepatic end points, and after repeated dosing, toxicity is determined by the body burden represented 

by the cumulative dose minus the portion of the dose eliminated. 

Renal Effects. Mild-to-moderate renal effects have been reported in some mature animals exposed to 

lethal or near-lethal levels of 2,3,7,8-TCDD. Acute exposure to 2,3,7,8-TCDD caused dilation of 
 

CDDs 154 

2. HEALTH EFFECTS 

convoluted tubules and Bowman's spaces at 25 μg/kg in Sprague-Dawley rats (Christian et al. 1986) and 

epithelial hyperplasia in the renal pelvis at 70 μg/kg in rhesus monkeys (McConnell et al. 1978a). Similar 

findings were reported in monkeys exposed to 0.011 μg/kg/day of 2,3,7,8-TCDD for 9 months (Allen et al. 

1977). Kidney weights were not affected in male C57BL/6J and DBA/2J mice treated with a single 

0.03–235 μg/kg or 1–3,295 μg/kg 2,3,7,8-TCDD dose, respectively (Weber et al. 1995). Rats administered 

doses of $0.047 μg/kg/day for 13 weeks exhibited an increase in relative kidney weight; a dose of 

0.026 μg/kg/day was without effect (Van Birgelen et al. 1995). Chronic exposure of B6C3F1 mice by 

gavage to approximately 0.071 μg/kg/day of 2,3,7,8-TCDD induced renal inflammatory changes; no effects 

were found at 0.0071 μg/kg/day (NTP 1982b). In contrast, no renal effects were found in Osborne-Mendel 

rats exposed to 0.071 μg/kg/day of 2,3,7,8-TCDD for 104 weeks (NTP 1982b) or in Sprague-Dawley rats 

exposed to 0.1 μg/kg/day of 2,3,7,8-TCDD in the feed for 2 years (Kociba et al. 1978a). 

Studies with other congeners reported no renal effects following chronic exposure to 0.34 μg/kg/day and 

0.7 μg/kg/day of a mixture of 1,2,3,7,8,9-HxCDD and 1,2,3,6,7,8-HxCDD by gavage in rats and mice, 

respectively (NCI/NTP 1980) or 5×105

 and 1.3×106 μg/kg/day of 2,7-DCDD in the feed in rats and mice, 

respectively (NCI/NTP 1979a). 

The above data suggest that the observed renal effects in mature animals may be secondary to the general 

response to 2,3,7,8-TCDD toxicity with the exception of the epithelial hyperplasia reported in monkeys. 

However, developmental studies clearly show that the ureteral epithelium is altered by in utero exposure to 

CDDs as manifested by hyperplasia of ureteral lining epithelial cells leading to hydronephrosis (see 

Section 2.2.2.6). 

Endocrine Effects. Blood corticosterone levels were decreased to 29 and 26% of control values in 

male Sprague-Dawley rats at 14 and 21 days after a single oral dose of 25 μg/kg 2,3,7,8-TCDD, 

respectively (Balk and Piper 1984). Since 11-β-hydroxyprogesterone levels were elevated, the authors 

suggested that 2,3,7,8-TCDD produced a block at the 21-hydroxylase step in the synthesis of 

corticosterone. This was directly demonstrated in a follow-up study in which the authors observed a 35% 

decrease in 21-hydroxylase activity 7 days after a single oral dose of 50 μg/kg 2,3,7,8-TCDD (Mebus and 

Piper 1986). Corticosterone serum levels from samples taken late in the light phase decreased up to 40% in 

male Sprague-Dawley rats administered a single 50 μg 2,3,7,8-TCDD/kg dose (DiBartolomeis et al. 1987). 

The effect was attributed to 2,3,7,8-TCDD-induced inhibition of cholesterol side-chain cleavage. In 

samples taken early in the light cycle, corticosterone levels increased 4-fold relative to controls; however, 
CDDs 155 

2. HEALTH EFFECTS 

this increase was shown to result from nutritional deprivation rather than from a direct effect of 

2,3,7,8-TCDD. The possibility that altered levels of corticosterone result from a 2,3,7,8-TCDD-induced 

effect on adrenocorticotropin (ACTH) was examined by Bestervelt et al. (1993). ACTH serum levels were 

significantly increased in male Sprague-Dawley rats over a 14-day period following administration of a 

single dose of 50 μg 2,3,7,8-TCDD/kg; maximum increases were observed on days 3 and 14. Plasma 

corticosterone levels were significantly increased on days 1 and 5, but were reduced below control levels 

on days 10 and 14. Treatment with 2,3,7,8-TCDD did not affect the activity of the rate-limiting enzyme for 

adrenal steroidogenesis, mitochondrial cytochrome P-450 cholesterol side chain cleavage. Basal 

corticosterone concentration in adrenal glands from 2,3,7,8-TCDD-treated rats was significantly lower than 

in controls on days 5, 7, and 14 after dosing; however, secretion of corticosterone induced by stimulation 

with exogenous ACTH was not altered by treatment with 2,3,7,8-TCDD. Based on these results, the 

authors concluded that 2,3,7,8-TCDD may interfere with secretion or synthesis of appropriate, bioactive 

ACTH from the anterior pituitary gland, which could compromise adrenal steroidogenesis. 

The effects of 2,3,7,8-TCDD on thyroid function has been extensively studied. For example, a single 

gavage dose of 25 μg 2,3,7,8-TCDD/kg significantly decreased serum levels of T4 and increased serum 

levels of triiodothyronine (T3) in male hooded rats 9 days after dosing (Bastomsky 1977). The decrease in 

T4 appeared to be the result of an increased biliary excretion of T4-glucuronide, and this was attributed to 

induction of UDP-glucuronyltransferase (UDPGT) by 2,3,7,8-TCDD. UDPGT catalyzes glucuronidation of 

T4 and clearance. The increase in T3 was consistent with increased thyroid secretion from thyrotropin 

(TSH) stimulation. Administration of a single dose of 6.25–100 μg 2,3,7,8-TCDD/kg by gavage to adult 

male Sprague-Dawley rats produced a significant dose-related decrease in serum T4 levels (50% of control 

with the lowest dose) 7 days after dosing (Potter et al. 1986). Serum levels of T3 were elevated in a doserelated manner, whereas levels of TSH achieved a maximum increase with the lowest dose. Potter et al. 

(1986) also observed a small 2,3,7,8-TCDD-related increase in thyroid weight, but no consistent pattern of 

histological alterations. Hermansky et al. (1988) reported a 65% decrease in serum T4 levels in female 

Sprague-Dawley rats 6 days after administration of 3 doses of 40 μg 2,3,7,8-TCDD/kg; however, in this 

study the authors observed a 9% increase in serum T3. A dose-related decrease in serum total T4 was 

observed in female Long Evans rats 4 days after a single dose of 5.3–60 μg 2,3,7,8-TCDD/kg; statistical 

significance was achieved with a 12 μg/kg dose (Fan and Rozman 1995). Total serum T3 was not 

significantly altered. However, 90 days after single doses of 27–60 μg 2,3,7,8-TCDD/kg total serum T4 

and T3 were elevated, which led the authors to suggest that 2,3,7,8-TCDD triggers adaptive responses 

which persist after most of the chemical has cleared the organism (Fan and Rozman 1995). In male mink, 
CDDs 156 

2. HEALTH EFFECTS 

relative thyroid gland weight was increased over a 4-week period after administration of a single dose of 

7.5 μg 2,3,7,8-TCDD/kg, but a dose of 5 μg/kg was without effect (Hochstein et al. 1988). Hochstein et al. 

(1988) also reported a significant dose-related increase in relative adrenal gland weight over a 2.5–7.5 μg 

2,3,7,8-TCDD/kg dose range. However, when the weights were expressed as a percentage of brain weight, 

only the increase in the adrenal gland at 7.5 μg/kg was significant. The concentrations of plasma cortisol 

and free and bound T3 and T4 were slightly reduced as a result of 2,3,7,8-TCDD treatment (2.5 μg/kg, 

only dose level tested), but the differences relative to controls were not significant. 

Acute effects of 2,3,7,8-TCDD on thyroid function have been also reported in mice. In contrast with 

observations in rats, in which 2,3,7,8-TCDD appears to have independent effects on T4 and T3 levels, 

serum T4 and T3 levels were decreased in a dose-dependent fashion in male C57BL/6J mice 8 days after a 

single gavage dose of 0.03–235 μg 2,3,7,8-TCDD/kg (Weber et al. 1995). A similar effect was observed in 

male DBA/2J mice treated with a single dose of 1–3,295 μg/kg. In C57BL/6J mice, maximum depression 

of thyroid hormones (35% of controls) was achieved with a dose of 133 μg/kg. In male DBA/2J mice, 

maximum reductions in T3 and T4 levels (40 and 20% of controls, respectively) were attained with the 

highest dose level (Weber et al. 1995). It should be noted that the Weber et al. (1995) study did not include 

statistical analysis of the results. 

A significant decrease in serum total T4 was observed in male Sprague-Dawley rats administered 

2,3,7,8-TCDD by gavage for 13 weeks at doses equivalent to 0.8 μg/kg/day, the only dose level tested 

(Viluksela et al. 1994). Serum total T3 was not significantly altered and neither was the relative or 

absolute weight of the pituitary. Similar results on thyroid function were reported in female SpragueDawley rats administered 2,3,7,8-TCDD by gavage for 30 weeks at doses equivalent to 

0.0001–0.125 μg/kg/day (Sewall et al. 1995). The dose-related decrease in serum T4 was statistically 

significant beginning at the 0.035 μg/kg/day dose level. Serum levels of T3 were not significantly altered 

by treatment. Sewall et al. (1995) also reported that serum levels of TSH were increased about 3-fold in 

the highest-dose group. Treatment with 2,3,7,8-TCDD also induced UDP-glucuronosyltransferase-1. 

Administration of 2,3,7,8-TCDD in the diet also affected thyroid function as demonstrated by Van 

Birgelen et al. (1995) who found a significant dose-related decrease in plasma total T4 in female SpragueDawley rats at dietary doses of $0.047 μg/kg/day for 13 weeks; plasma total T3 was not altered with 

2,3,7,8-TCDD doses of up to 1 μg/kg/day. Li and Rozman (1995) examined the reversibility of the 

2,3,7,8-TCDD-induced decrease in serum T4 in male Sprague-Dawley rats. The rats were gavaged once a 

week for 10 weeks with doses equivalent to approximately 0.003–1.6 μg 2,3,7,8-TCDD/kg and this was 
 

CDDs 157 

2. HEALTH EFFECTS 

followed by a 6-week recovery period. Serum T4 levels were significantly depressed with 2,3,7,8-TCDD 

doses of $0.03 μg/kg in a dose-dependent fashion and remained low during the recovery period. Based 

on these results, the authors suggested that the ED50 for this dose-response is close to a total cumulative 

dose of 1 μg/kg. 

No significant non-neoplastic lesions were observed in the thyroid, parathyroid, adrenal, and pituitary 

gland from male and female Sprague-Dawley rats maintained for 2 years on a diet that supplied 0.001, 

0.01, or 0.1 μg 2,3,7,8-TCDD/kg/day (Kociba et al. 1978a). Similar results were obtained in male and 

female Osborne-Mendel rats and in male B6C3F1 mice administered up to approximately 0.071 μg 

2,3,7,8-TCDD/kg/day by gavage for 104 weeks, and in female B6C3F1 mice given up to 0.3 μg 

2,3,7,8-TCDD/kg/day (NTP 1982b). 

Information regarding other CDD congeners is limited. Administration of 1,2,3,4,6,7,8-HpCDD by 

gavage for 13 weeks to male Sprague-Dawley rats in doses equivalent to 24–110 μg/kg/day produced a 

dose-related decrease in serum total T4 (Viluksela et al. 1994). Doses of #4 μg/kg/day were without 

significant effect. Serum levels of total T3 were not significantly affected by treatment. A more recent 

study reported a 69% decrease in serum T4 levels in male Sprague-Dawley rats administered doses 

equivalent to 3.8 μg 1,2,3,7,8-PeCDD/kg/day or 15.4 μg 1,2,3,4,7,8-HxCDD/kg/day for 13 weeks 

(Viluksela et al. 1998b). After an additional 13-week CDD-free period, T4 levels returned to near control 

levels. In females administered doses of 2.6 μg/kg/day of the penta-CDD or 10.3 μg/kg/day of the hexaCDD, T4 serum levels were 40% below control levels at the end of the dosing period and 62% below 

controls at the end of the additional 13-week period. Serum T3 levels were not significantly affected by 

treatment with either congener (Viluksela et al. 1998b). 

In summary, CDDs were shown to alter endocrine parameters mostly in rodent studies. One of the better 

characterized effects was a decrease in serum T4, caused apparently by CDD-induced T4 metabolism and 

excretion. Alterations in T3 levels were less consistent. Results from additional studies suggested that 

2,3,7,8-TCDD may interfere with secretion and synthesis of ACTH in the pituitary. 

Dermal Effects. A number of changes in the skin have been observed in rodents and monkeys. In 

monkeys, skin lesions seen after a single oral dose or repeated dosing resemble the chloracne observed in 

humans. Distinctive changes in rhesus monkeys included swelling and inflamed eyelids, nail loss, and 

facial hair loss with acneform lesions following acute exposure to a single dose of 70 μg/kg (McConnell et 
 

 

CDDs 158 

2. HEALTH EFFECTS 

al. 1978a). Monkeys had hair loss due to squamous metaplasia and keratinization of the sebaceous glands 

and hair follicles, and periorbital edema following intermediate-duration exposure to 0.011 μg/kg/day of 

2,3,7,8-TCDD in the diet or exposure to 0.1 μg/day, 3 days a week for 3 weeks, but not in those exposed to 

0.02 μg/kg/day (Allen et al. 1977; McNulty 1984). Rough hair coats were described in Syrian hamsters 

exposed to a single dose of 1,000 μg/kg 2,3,7,8-TCDD, but not in those exposed to 600 μg/kg (Henck et 

al. 1981). Skin thickening was observed in A2G-hr/+ mice exposed to 75 μg/kg 2,3,7,8-TCDD (Greig 

1984). Chronic exposure by gavage to 2,3,7,8-TCDD induced dermatitis in B6C3F1 mice at 

0.36 μg/kg/day (Della Porta et al. 1987) and amyloidosis in Swiss mice at 0.001 μg/kg/day (Toth et al. 

1979). In the B6C3F1 mice, dermatitis regressed after discontinuation of treatment (Della Porta et al. 

1987). In contrast, no dermal effects were observed in Osborne-Mendel rats and in B6C3F1 mice 

following chronic exposure to 0.71 μg/kg/day and 0.3 μg/kg/day of 2,3,7,8-TCDD, respectively, by 

gavage for 104 weeks (NTP 1982b). 

No dermal effects were found in Osborne-Mendel rats and B6C3F1 mice gavaged with approximately 

0.34 μg/kg/day and 0.7 μg/kg/day of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD, 

respectively, for 104 weeks (NCI/NTP 1980). However, male and female Sprague-Dawley rats treated 

with doses equivalent to 2.6–3.8 μg 1,2,3,7,8-PeCDD/kg/day or 10.3–15.4 μg 1,2,3,4,7,8-HxCDD/kg/day 

for 13 weeks exhibited occasional hair loss and sores in the ears, nose, neck, tail, and feet (Viluksela et al. 

1998a). No effects were observed following chronic exposure of Osborne-Mendel rats and B6C3F1 mice 

to 5×105

 μg/kg/day and 1.3×106 μg/kg/day of 2,7-DCDD, respectively, in the feed (NCI/NTP 1979a). 

Ocular Effects. No ocular effects were observed in Osborne-Mendel rats and in B6C3F1 mice 

following chronic exposure to 0.071 μg/kg/day and 0.3 μg/kg/day of 2,3,7,8-TCDD, respectively, by 

gavage for 104 weeks (NTP 1982b). Also no ocular effects were found in Osborne-Mendel rats and 

B6C3F1 mice gavaged with approximately 0.34 μg/kg/day and 0.7 μg/kg/day of a mixture of 1,2,3,6,7,8­

HxCDD and 1,2,3,7,8,9-HxCDD, respectively, for 104 weeks (NCI/NTP 1980). Similarly, no effects were 

observed following chronic exposure of Osborne-Mendel rats and B6C3F1 mice to 5×105

 μg/kg/day and 

1.3×106

 μg/kg/day of 2,7-DCDD, respectively, in the feed (NCI/NTP 1979a). 

Body Weight Effects. A characteristic effect of exposure to 2,3,7,8-TCDD in animals is the wasting 

syndrome. This is observed following exposure in all duration categories. 
CDDs 159 

2. HEALTH EFFECTS 

Weight loss or decreased weight gain were recorded in Sprague-Dawley rats following a single dose of 

6.25 μg/kg (Moore et al. 1985), 10.6 μg/kg (Roth et al. 1988), 15 μg/kg (Seefeld and Peterson 1984), and 

25 μg/kg 2,3,7,8-TCDD (Christian et al. 1986a), and in Fischer 344 rats following a single oral dose of 

100 μg/kg (Kelling et al. 1985). Furthermore, about 40% weight loss was recorded in the range of LD50 

values (164–340 μg/kg) for Fischer 344 rats from different breeding stations (Walden and Schiller 1985). 

None of the studies provided a NOAEL value. Acute exposure (10–14 days) to lower doses of 

2,3,7,8-TCDD caused reduced weight gain in rats at 0.5 μg/kg/day, but not at 0.125 μg/kg/day (Giavini et 

al. 1983; Sparschu et al. 1971b). 

Decreased body weight gain was observed in guinea pigs after a single dose of 6 μg/kg 2,3,7,8-TCDD in 

oil vehicle but not after 12 μg/kg in soil (Umbreit et al. 1985). A decreased weight gain was recorded in 

pregnant rabbits exposed during gestation to 0.25 μg/kg/day (Giavini et al. 1982) and in hamsters exposed 

to 1,000 μg/kg (Henck et al. 1981). ED50 values (doses causing a 50% decrease in a measurable parameter 

relative to the control value) for reduced body weight gain were calculated for 2,3,7,8-TCDD as 1.8 μg/kg 

for Hartley guinea pigs, 89 μg/kg for Sprague-Dawley rats, 890 μg/kg for C57BL/6 mice, and 1,000 μg/kg 

for Syrian hamsters (Hanberg et al. 1989). Single doses of $75 μg 2,3,7,8-TCDD/kg produced a slight 

reduction in body weight in male C57BL/6J mice 8 days after dosing (Weber et al. 1995); feed intake was 

not affected during this period. In the same study, it was found that body weights of male DBA/2J mice 

dosed with 1–3,295 μg 2,3,7,8-TCDD/kg were significantly reduced at $1,500 μg 2,3,7,8-TCDD/kg. It 

should be noted, however, that in mice, decreases in body weight resulting from 2,3,7,8-TCDD exposure 

do not become evident until 5–7 days after dosing (Shen et al. 1991), and that in C57BL/6J mice reduction 

of feed intake is insignificant during the first week after dosing (Kelling et al. 1985). Weight loss (28%) 

was also found in monkeys after a single dose of 70 μg/kg 2,3,7,8-TCDD (McConnell et al. 1978a). 

Decreases in body weight gain or body weight loss have been consistently reported in animals following 

intermediate-duration exposures to 2,3,7,8-TCDD. A decreased weight gain was observed in OsborneMendel rats exposed intermittently for 13 weeks by gavage to 0.07 μg/kg/day (NTP 1982b) or in SpragueDawley rats treated with $0.2 μg/kg/day for 10–13 weeks (Li and Rozman 1995; Viluksela et al. 1994), in 

guinea pigs exposed to 0.005 μg/kg/day in the feed (DeCaprio et al. 1986), and in mice intermittently 

exposed to 20 μg/kg/day by gavage (Thigpen et al. 1975). A weight loss was recorded in rhesus monkeys 

after 0.1 μg/kg/day of 2,3,7,8-TCDD for 3 weeks of intermittent exposure (McNulty 1984). However, 

weight loss occurred with longer exposure of 9 months at 0.011 μg/kg/day (Allen et al. 1977). A recent 

study reported a 10% decrease in body weight gain in female Sprague-Dawley rats fed a diet that supplied 
CDDs 160 

2. HEALTH EFFECTS 

a daily dose of 0.047 μg 2,3,7,8-TCDD/kg for 13 weeks (Van Birgelen et al. 1995). At the highest 

exposure level (1 μg/kg/day) terminal body weights were reduced to 72% of controls; this group consumed 

32% less food than controls. 

In chronic-duration experiments with 2,3,7,8-TCDD, decreased body weight gain was reported in SpragueDawley rats exposed to 0.1 μg/kg/day (Kociba et al. 1978a) and 0.286 μg/kg/day (Van Miller et al. 1977) 

in the feed; Osborne Mendel rats exposed to approximately 0.0014 μg/kg/day by gavage for 104 weeks 

(NTP 1982b); and in B6C3F1 mice exposed to 0.36 μg/kg/day by gavage for 52 weeks (Della Porta et al. 

1987), but not in C57BL/6 mice gavaged once per week for 14–15 months with 0.03 μg 2,3,7,8-TCDD 

(Oughton et al. 1995). 

Experiments with other congeners showed milder effects. Acute exposure during gestation caused a 

decreased maternal weight gain in Sprague-Dawley rats exposed to 10 μg/kg/day mixed HxCDD (Schwetz 

et al. 1973). Decreased weight gains were observed in rats and mice gavaged for intermediate-duration 

with 0.71 μg/kg/day and 0.18 μg/kg/day of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD, 

respectively (NCI/NTP 1980). Chronic-duration exposure induced decreased weight gain in Osborne 

Mendel rats exposed to 0.18 μg/kg/day of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD by 

gavage for 104 weeks (NCI/NTP 1980). In contrast, no effects on body weight were observed in mice 

exposed to 0.7 μg/kg/day of a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD for 104 weeks 

(NCI/NTP 1980). Male Sprague-Dawley rats administered 1,2,3,4,6,7,8-HpCDD by gavage at dose levels 

equivalent to 73 and 110 μg/kg/day for 13 weeks exhibited a 5.1% and 19.3% reduction in body weight 

gain, respectively, at the end of the study period (Viluksela et al. 1994). No significant effect was 

observed with doses #24 μg/kg/day. Relative to controls, the body weight of male Sprague-Dawley rats 

administered doses equivalent to 3.8 μg 1,2,3,7,8-PeCDD/kg/day or 15.4 μg 1,2,3,4,7,8-HxCDD/kg/day 

for 13 weeks was reduced by 27% and 24%, respectively, at the end of the dosing period (Viluksela et al. 

1998a). In females, doses equivalent to 2.6 μg 1,2,3,7,8-PeCDD/kg/day for 13 weeks resulted in an 18% 

reduction in body weight relative to controls, whereas doses of approximately 10.3 μg 1,2,3,4,7,8­

HxCDD/kg/day for 13 weeks were without significant effect (Viluksela et al. 1998a). 

No effect on the body weight of CD-1 mice was observed after 14 daily doses of OCDD at 1 μg/kg/day or 

2,7-DCDD at 1,000 μg/kg/day (Courtney 1976). Chronic-duration exposure induced decreased weight 

gain in Osborne Mendel rats and in B6C3F1 mice exposed to 2.5×105

 μg/kg/day and 6.5×105 μg/kg/day of 

2,7-DCDD, respectively, in the feed (NCI/NTP 1979a). 
CDDs 161 

2. HEALTH EFFECTS 

As summarized above, body weight effects were consistently observed in all species exposed to CDDs. 

Effects occurred after intermittent exposure by gavage and after exposure in a diet. In acute- and 

intermediate-duration exposure experiments, the wasting syndrome seemed to be the primary cause of 

death. 

2.2.2.3 Immunological Effects 

An effect of sublethal exposures (acute, intermediate-term, or chronic) to 2,3,7,8-TCDD common to all 

species studied is thymic atrophy. Depletion of lymphocytes results in suppression of T-cell immunity. 

The T-cell responses studied have included delayed hypersensitivity responses, rejection of skin allografts, 

and in vitro mutagen responses of lymphoid cells. T-cell immunotoxicity is probably the most sensitive 

end point. Effects on T-cells can occur at levels of exposure three orders of magnitude lower than the 

effects on thymus cellularity. B-lymphocytes are also affected by 2,3,7,8-TCDD, but higher exposure 

levels are necessary for suppression of humoral immunity. CDDs suppress resistance to different 

infectious agents by various mechanisms (see Section 2.4 for more detailed information). 

Acute ED50 values for thymic atrophy following a single dose of 2,3,7,8-TCDD were calculated as 

26 μg/kg in Sprague-Dawley rats, 0.8 μg/kg in Hartley guinea pigs, 280 μg/kg in C57BL/6 mice, and 

48 μg/kg in Syrian hamsters (Hanberg et al. 1989). A significant dose-related reduction in absolute 

thymus weight was reported in young male Wistar rats administered single doses of $1 μg/kg 

2,3,7,8-TCDD; this effect was paralleled by a significant decrease in thymic cellularity (De Heer et al. 

1994b). Thymic atrophy was shown to be initiated in the thymus cortex on day 4 after a single dose of 

25 μg/kg 2,3,7,8-TCDD (De Heer et al. 1994a). The initial lymphodepletion in the cortex was followed by 

a secondary depletion of medullary thymocytes on day 6, and on day 10, a preferential depletion of cortical 

thymocytes was no longer observed. Decreased thymus weight was reported in pregnant C57BL/6J mice 

exposed to 0.5 μg/kg/day 2,3,7,8-TCDD for 10 days (Silkworth et al. 1989b). Offspring of C57BL/6J 

mice similarly exposed to 1.5 μg/kg/day had severe thymic atrophy, cellular depletion and altered 

thymocyte antigen expression, and immune function (Holladay et al. 1991). In contrast, similar changes 

were observed in DBA/2J mice only after exposure to higher doses of 8 μg/kg/day. Furthermore, thymic 

atrophy was observed in rhesus monkeys after a single dose of 70 μg/kg (McConnell et al. 1978a) and in 

guinea pigs after a dose of 6 μg/kg (Umbreit et al. 1985). 
CDDs 162 

2. HEALTH EFFECTS 

Treatment of rats with daily doses of 0.72 μg 2,3,7,8-TCDD/kg/day by gavage for 14 days did not alter 

spontaneous NK-cell activity in the lung, but significantly suppressed influenza virus-augmented NK 

activity (Yang et al. 1994). A significantly higher virus titer was observed on days 2, 3, and 4 in whole 

lung homogenate from rats treated with a single dose of 10 μg/kg (Yang et al. 1994). Decreased resistance 

to infection, as evidenced by increased mortality, was observed in B6C3F1 mice infected with 

Streptococcus pneumoniae and administered 1 μg/kg/day 2,3,7,8-TCDD for 14 days (White et al. 1986), 

and in B6C3F1 mice infected with influenza A virus and administered a single gavage dose of 0.01, 0.05, 

or 0.1 μg/kg 2,3,7,8-TCDD (Burleson et al. 1996). The Burleson et al. (1996) study identified a NOAEL 

of 0.005 μg/kg for this effect. Acute exposure to 2,3,7,8-TCDD reduced polymorphonuclear activity in 

B6C3F1 mice at 5 μg/kg (no effect was seen in DBA/2N mice) (Ackermann et al. 1989). Suppressed 

antibody response to sheep erythrocytes (SRBC) was reported in B6C3F1 mice that were given a single 

gavage dose of 1 μg/kg; no such effect was found after a single dose of 0.5 μg/kg (Holsapple et al. 1986a). 

However, suppression of the antibody response occurred after 14 daily doses of 0.1 μg/kg/day. In rats, a 

single dose of 20 μg 2,3,7,8-TCDD/kg administered 5 days before immunization significantly enhanced 

the primary antibody response to SRBC as judged by a significant increase in serum IgG levels 7 days 

after immunization (Fan et al. 1996). However, serum IgM levels were not significantly affected by doses 

of 2,3,7,8-TCDD of up to 40 μg/kg. Fan et al. (1996) also observed that cell-mediated immunity, tested 

with a delayed-type hypersensitivity (DTH) assay, exhibited a U-shaped response to treatment with 

2,3,7,8-TCDD, as doses of 1–20 μg/kg increased the DTH response, whereas doses of 30–90 μg/kg 

decreased it, even below control levels. 

Suppressed total serum complement activity was observed in female B6C3F1 mice exposed to a single 

gavage dose of 14 μg/kg or 14 daily doses of 0.01 μg/kg/day (White et al. 1986). Serum levels of 

complement component C3 were also suppressed at doses of $0.5 μg/kg 2,3,7,8-TCDD (White et al. 1986). 

Subsequent studies by the same group showed that the 2,3,7,8-TCDD-induced reduction in serum C3 is not 

the result of a decrease in C3 production by hepatocytes but, at least in part, may be due to increased 

catabolism (Lin and White 1993). Single gavage doses of $2.5 μg 2,3,7,8-TCDD/kg suppressed cytotoxic 

T-lymphocyte (CTL) activity in mice challenged with a tumor allograft by a mechanism that did not involve 

elevation in plasma glucocorticoid levels (De Krey and Kerkvliet 1995). This was directly correlated with 

reduced numbers of splenic CTL effector cells (Kerkvliet et al. 1996). In these same animals, a suppression 

of the alloantibody response was correlated with a decreased expansion of the B-cell splenocyte population. 

This dose of 2,3,7,8-TCDD also initially induced interferon-γ, interleukin-2, and tumor necrosis factor 

production, but the normal increase of these in response to the tumor allograft was 
CDDs 163 

2. HEALTH EFFECTS 

not observed. Based on these and additional studies, the authors concluded that these effects are due to 

TCDD initially interfering with the activation of CD4+

 T cells and possibly T helper-B cell interactions. A 

recent study from the same group of investigators presented evidence that immune 2,3,7,8-TCDD-induced 

suppression in C57BL/6 mice is not caused by direct alterations in the production of immunomodulatory 

metabolites of arachidonic acid (Lawrence and Kerkvliet 1997). The above results indicate that 

immunological effects occur after moderate-to-low single doses or after repeated low doses that accumulate 

in the body, suggesting that the total dose of 2,3,7,8-TCDD is important. As shown in Figure 2-1, 

immunotoxicity was a very sensitive end point; the lowest LOAEL for immune effects is 0.01 μg/kg/day 

(Burleson et al. 1996; White et al. 1986). In the Burleson et al. (1996) study, decreased resistance to 

infection was observed in mice receiving a single gavage dose of 0.01 μg/kg, and no effects were observed 

at 0.005 μg/kg. Reduced serum complement levels were observed in mice exposed to 0.01 μg/kg/day for 

14 days (White et al. 1986); no NOAEL was identified in this study. The NOAEL of 0.005 μg/kg/day 

identified in the Burleson et al. (1996) study was used to derive an acute oral MRL for 2,3,7,8-TCDD of 

2×10-4 μg/kg/day as described in the footnote to Table 2-2, Section 2-5, and in Appendix A. 

Several immunological effects were observed following intermediate-duration exposure to 2,3,7,8-TCDD. 

Decreased thymus weight after 2,3,7,8-TCDD exposure was observed in rats dosed by gavage with 

0.71 μg/kg/day for 6 weeks (Vos et al. 1973), in the F3 generation of rats receiving 0.01 μg/kg/day 

(Murray et al. 1979), and in guinea pigs receiving 0.005 μg/kg/day or 0.03 μg/kg/day (thymic atrophy) in 

the feed for 90 days (DeCaprio et al. 1986). A significant reduction in absolute and relative thymus weight 

was observed in male Sprague-Dawley rats administered 2,3,7,8-TCDD by gavage at doses equivalent to 

0.8 μg/kg/day (only dose level tested) for 13 weeks (Viluksela et al. 1994). Spleen weight was not 

significantly altered. Similar results were reported in female Sprague-Dawley rats fed for 13 weeks a diet 

that supplied doses of $0.014 μg 2,3,7,8-TCDD/kg/day (Van Birgelen et al. 1995). Relative spleen weight 

was increased at $0.047 μg 2,3,7,8-TCDD/kg/day. Decreased cell-mediated immunity was found in mice 

and guinea pigs exposed by gavage to 0.71 μg/kg/day for 4 weeks and 0.03 μg/kg/day for 8 weeks, 

respectively (Vos et al. 1973). Guinea pigs seem to be especially sensitive to 2,3,7,8-TCDD toxicity; an 

intermediate-duration exposure to 0.001 μg/kg/day reduced the lymphocyte counts, and exposure to 

0.03 μg/kg/day caused decreased humoral immunity and thymic atrophy (Vos et al. 1973). A recent study 

examined the effect of low-level dietary exposure to 2,3,7,8-TCDD to young adult male Leeds strain rats 

(Badesha et al. 1995). A 30-day exposure to approximately 0.1 μg/kg/day (or a total dose of 

approximately 3 μg/kg) resulted in an exposure duration-dependent reduction of in vitro lipopolysaccharide-induced production of interleukin-1 in cultures of their splenic macrophages. A 180-day 
CDDs 164 

2. HEALTH EFFECTS 

exposure to approximately 0.017 μg/kg/day suppressed the production of interleukin-2 by either concanavalin A or phorbol ester/calcium ionophore stimulation, and reduced the lectin-induced proliferation of 

splenic T cells. The authors concluded that exposure to a low dietary dose of 2,3,7,8-TCDD suppresses 

the functions of several T-cell subsets. The highest NOAEL value for immunological effects (decreased 

thymus weight) was 0.0007 μg/kg/day 2,3,7,8-TCDD given to the most sensitive species, guinea pigs, in 

the diet (DeCaprio et al. 1986). The NOAEL value of 0.0007 μg/kg/day was used to derive an 

intermediate-duration oral MRL for 2,3,7,8-TCDD of 2×10-5 μg/kg/day as described in the footnote to 

Table 2-2, Section 2.5, and in Appendix A. 

Increased mortality that was indicative of altered immunity was also observed in C57BL/6Jfh mice 

challenged with Salmonella bern following exposure to $1 μg/kg/day of 2,3,7,8-TCDD by gavage once a 

week for 4 weeks (Thigpen et al. 1975); no significant effects were observed at 0.5 μg/kg/day. In the same 

study, using the same experimental design, doses of up to 20 μg/kg/day of 2,3,7,8-TCDD had no 

significant effect on mortality in mice infected with Herpesvirus suis (Thigpen et al. 1975). Exposure to 

0.5 μg/kg/day 2,3,7,8-TCDD once a week for 5–8 weeks caused suppression of humoral activity in 

C57BL/6 mice (Vecchi et al. 1983a). In addition, lymph node atrophy was reported in monkeys exposed 

to a lethal dose of 0.011 μg/kg/day in the feed for 9 months (Allen et al. 1971). 

Administration of 2,3,7,8-TCDD at approximately 0.071 μg/kg/day to Osborne-Mendel rats or at about 

0.3 μg/kg/day to B6C3F1 mice by gavage for 104 weeks produced no histological alterations in the spleen 

or thymus (NTP 1982b). Chronic exposure to 2,3,7,8-TCDD in food induced thymic atrophy in SpragueDawley rats at 0.1 μg/kg/day in a 2-year study (Kociba et al. 1978a) with the highest NOAEL of 

0.01 μg/kg/day. Furthermore, rhesus monkeys exposed chronically to 0.002 μg/kg/day 2,3,7,8-TCDD in 

the feed exhibited degeneration of the bone marrow and lymphoid tissues (Hong et al. 1989). A recent 

study examined the effect of long-term exposure to 2,3,7,8-TCDD on various immune cell phenotypes of 

female C57 BL/6 mice (Oughton et al. 1995). The mice were administered 0.2 μg 2,3,7,8-TCDD/kg once 

per week for 14–15 months; this resulted in a cumulative dose of 12–13 μg/kg (approximately 

0.03 μg/kg/day) and a concentration of 2,3,7,8-TCDD in adipose tissue of 1.27 ng/g abdominal fat. There 

were no significant 2,3,7,8-TCDD-related effects on thymus and spleen weight or in the cellularity of these 

tissues. Exposure to 2,3,7,8-TCDD induced subtle changes in thymic phenotypes which, according to the 

authors, were of questionable biological relevance given the age-related decrease in thymic cellularity 

observed. 2,3,7,8-TCDD did not alter the frequencies of the major leukocyte subpopulations, but 

significantly altered functionally discrete subpopulations within the T-cell compartment. The most notable 
CDDs 165 

2. HEALTH EFFECTS 

change was a decrease in the frequency of memory T helper cells, with a concomitant increase in the 

proportion of naive T helper cells. Oughton et al. (1995) also presented preliminary data suggesting that 

phenotypic changes in spleen cells correlated with similar changes in blood cells. 

Other CDD congeners also appear to affect the immune system. Significant dose-related decreases in 

absolute and relative thymus weight were observed in male Sprague-Dawley rats administered doses 

equivalent to 4–110 μg/kg/day 1,2,3,4,6,7,8-HpCDD for 13 weeks by gavage (Viluksela et al. 1994). A 

dose level of 0.3 μg/kg/day was without significant effect. Treatment with 1,2,3,4,6,7,8-HpCDD had no 

significant effect on spleen weight. Suppressed antibody response was reported in B6C3F1 mice after 

2 weeks of exposure to 0.1 μg/kg/day of 2,7-DCDD, but not after exposure to 10 μg/kg/day of OCDD 

(Holsapple et al. 1986b). Depressed antibody response was found in C57BL/6 mice exposed to a single 

dose of 33 μg/kg/day 1,2,3,4,6,7,8-HpCDD (Kerkvliet and Brauner 1987). Suppressed serum complement 

activity was found in B6C3F1 mice following 2 weeks of exposure to 1 μg/kg/day 1,2,3,6,7,8-HxCDD 

(White et al. 1986). Splenic hyperplasia was observed in Osborne-Mendel rats after exposure to a mixture 

of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD at 7.1 μg/kg/day, 1 day/week for 13 weeks (NCI/NTP 

1980). 

In conclusion, the immunological system was a sensitive target of CDD toxicity under experimental 

conditions in animals. Effects on all types of mediated immunity were seen at doses of 2,3,7,8-TCDD as 

low as 0.01 μg/kg. Doses of 2,3,7,8-TCDD that were well below the lethal dose affect humoral immunity. 

Thymic atrophy occurs as single or multiple doses approach those that may increase lethality. Neonates 

and young animals are much more sensitive than adults to most of the immunological responses. 

The highest NOAEL values and all reliable LOAEL values for immunological effects in each species and 

duration category for each congener are recorded in Table 2-2 and 2-3 and plotted in Figures 2-1 and 2-2. 

2.2.2.4 Neurological Effects 

Limited information was obtained regarding neurological effects in animals. Decreased motor activity was 

observed in Sprague-Dawley rats after a single dose of 5 μg/kg of 2,3,7,8-TCDD that was not associated 

with mortality (Seefeld et al. 1984a) and after 14 daily doses of 2 μg/kg/day to pregnant females that were 

sacrificed on day 21 of gestation for developmental effects evaluation (Giavini et al. 1983). The NOAEL 

value was 0.01 μg/kg/day. Administration of 2,3,7,8-TCDD by gavage to male and female Osborne­
CDDs 166 

2. HEALTH EFFECTS 

Mendel rats and male B6C3F1 mice at doses of up to 0.071 μg/kg/day for 104 weeks did not result in 

significant histological alterations in the brain, spinal cord, or sciatic nerve (NTP 1982b). The same was 

found for female B6C3F1 dosed with up to 0.3 μg/kg/day for the same time period (NTP 1982b). 

Although motor effects have been described in rats dosed with 2,3,7,8-TCDD, in most studies, the 

neurological system was not specifically examined; therefore the issue of whether CDDs have a direct 

effect on the nervous system of animals has not been conclusively resolved. 

The highest NOAEL values and all reliable LOAEL values for neurological effects in each species and 

duration category are recorded in Tables 2-2 and 2-3 and plotted in Figures 2-1 and 2-2. 

2.2.2.5 Reproductive Effects 

A number of reproductive effects have been observed in animals orally exposed to 2,3,7,8-TCDD, 

including reduced fertility, pre- and post-implantation losses, decreases in gonad weights, decreased 

androgen levels, and altered estrus cycle and ovulation. Increased pre- and postimplantation losses were 

observed in CRCD rats exposed to 0.5 μg/kg/day for 2 weeks before mating (Giavini et al. 1983). 

Increased resorptions were found in Sprague-Dawley rats exposed to 0.125 μg/kg/day (Sparschu et al. 

1971a), in CF-1 mice exposed to 1.0 μg/kg/day (Smith et al. 1976), and in NMRI mice exposed to 

9 μg/kg/day (Neubert and Dillmann 1972) on gestation days (Gd) 6–15. In rabbits, increased 

postimplantation losses were recorded in a group exposed to 0.25 μg/kg/day, but not in those exposed to 

0.1 μg/kg/day on Gd 6–15 (Giavini et al. 1982). Furthermore, increased abortions (10 of 12) were 

observed in monkeys after a single gavage dose of 1 μg/kg (McNulty 1984). 

Reproductive toxicity has also been observed in non-pregnant female rats. Significant decreases in ovarian 

weight, ovulation rate, and the number of ova released have been observed in female Sprague-Dawley rats 

receiving a single gavage dose of $10 μg/kg (Li et al. 1995a, 1995b). Effects on hormone levels were also 

observed in the rats. Within 24 hours after dosing, significant increases in LH and follicle stimulating 

hormone levels were observed; prolactin levels were not altered (Li et al. 1995a). Following the 

administration of 17β-estradiol, LH and follicle stimulating hormone levels dropped below control levels. 

2,3,7,8-TCDD is also a reproductive toxicant in males. Decreased seminal vesicle weight was reported in 

Sprague-Dawley rats after a single dose of 4.5 μg/kg, with decreased androgen levels detected only on day 
CDDs 167 

2. HEALTH EFFECTS 

6 postexposure (Moore et al. 1985). Inflammation of the epididymis with sperm granuloma formation was 

reported in Wistar rats exposed to 4 μg/kg/day for 7 days, and decreases in the weight of male reproductive 

organs together with reduced levels of serum testosterone and dihydrotestosterone (compared with the 

pair-fed controls) were seen in Sprague-Dawley rats after a single gavage dose of 12.5 μg/kg 

2,3,7,8-TCDD (Khera and Ruddick 1973). An ED50 for altered regulation of LH levels was calculated as 

10 μg/kg 2,3,7,8-TCDD in Sprague-Dawley male rats (Bookstaff et al. 1990a). No dominant lethality was 

reported when male Wistar rats were given 12 μg/kg/day 2,3,7,8-TCDD for 7 days before mating (Khera 

and Ruddick 1973). 

In intermediate-duration studies with 2,3,7,8-TCDD, increased mortality was found in the offspring of 

Swiss Webster mice that were kept on a diet providing 0.35 μg/kg/day 2,3,7,8-TCDD for 4 weeks before 

mating, during gestation, and for 3 weeks of lactation (Thomas and Hinsdill 1979). Blocked estrous cycle 

was observed in female C57BL/6 mice exposed by gavage to 3 μg/kg/day, 3 days a week for 25 weeks 

(Umbreit et al. 1987), but no reproductive effects were seen in male mice exposed 1 day/week for 

30 weeks to the same dose (Umbreit et al. 1988). However, reduced spermatogenesis was found in 

Sprague-Dawley rats exposed for 4 weeks to 3.4 μg/kg/day in the feed, but not in those similarly exposed 

to 0.286 μg/kg/day (Van Miller et al. 1977). Exposure of female rhesus monkeys to 0.1 μg/kg/day 3 days 

a week by gavage for 3 weeks caused abortions in 3 of the 4 monkeys; 1 of the 4 monkeys administered 

0.02 μg/kg/day aborted (McNulty 1984). In a 3-generation study with 2,3,7,8-TCDD, significantly 

reduced fertility was observed among F1- and F2-generation rats exposed before mating to 0.01 μg/kg/day 

in the feed for 90 days, but not in those exposed to 0.001 μg/kg/day (Murray et al. 1979). 

In chronic-duration studies, increased abortions and reduced reproduction rates were reported in monkeys 

exposed to 0.00064 μg/kg/day of 2,3,7,8-TCDD in the feed (Bowman et al. 1989b; Hong et al. 1989; 

Schantz et al. 1992). No reproductive effects were found at 0.00012 μg/kg/day. No changes were 

observed in the reproductive organs of Sprague-Dawley rats chronically exposed to 0.1 μg/kg/day in the 

feed (Kociba et al. 1978a), or Osborne-Mendel rats and B6C3F1 mice exposed by gavage to approximately 

0.0.71μg/kg/day and 0.3 μg/kg/day of 2,3,7,8-TCDD, respectively (NTP 1982b). 

Rier et al. (1993) found a dose-related increase in the incidence and severity of endometriosis in monkeys 

chronically exposed to 0.00012 or 0.00064 μg/kg/day of 2,3,7,8-TCDD in the diet. Surgical-induced 

endometriosis was enhanced by 2,3,7,8-TCDD exposure in rats and mice. In a surgically induced 

endometriosis model, significant increases in the diameter of the endometriotic site and an acceleration of 
CDDs 168 

2. HEALTH EFFECTS 

growth were observed in rats (Cummings et al. 1996) and mice (Cummings et al. 1996; Johnson et al. 

1997), respectively. In this model, the animals received a gavage dose of 2,3,7,8-TCDD every 3 weeks 

(first dose was administered 3 weeks prior to surgical induction of endometriosis) for a total of five doses. 

Mice appear to be more sensitive than rats in terms of the magnitude of the effect on endometrial site 

diameter and adverse effect levels (endometriosis promotion was observed at 1, 3, and 10 μg/kg in mice 

[Cummings et al. 1996; Johnson et al. 1997] and at 10 μg/kg [Cummings et al. 1996]; no effects were 

observed in rats at 3 μg/kg). In contrast to these results, Foster et al. (1997) found that 2,3,7,8-TCDD 

exposure suppressed endometrial growth in mice. In their model, the mice were not pre-exposed to 

2,3,7,8-TCDD prior to the induction of endometriosis. Foster et al. (1997) notes that pre-exposure to 

2,3,7,8-TCDD results in endometriosis development due to immune suppression rather than an estrogenresponsive disease. 

Acute-duration studies examining reproductive effects have been conducted with other congeners. 

Increased resorptions were found in Sprague-Dawley rats exposed to 10 μg/kg/day mixed HxCDD during 

gestation but not in those exposed to 1 μg/kg/day (Schwetz et al. 1973). No reproductive effects were 

found in rats exposed to 1×105

 μg/kg/day 2,7-DCDD or 5×105 μg/kg/day OCDD during gestation. 

Similarly, no reproductive effects were found in rats exposed to 800 μg/kg/day 1,2,3,4-TCDD or 

2,000 μg/kg/day 2-MCDD, 2,3-DCDD, or 2,7-DCDD on Gd 6–15 (Khera and Ruddick 1973). 

The above data demonstrate that exposure to CDDs caused reproductive effects in animals. 2,3,7,8-TCDD 

was the most potent congener. The effects included increased pre- and postimplantation losses in females, 

morphological and functional changes in male and female reproductive organs, and hormonal imbalance in 

both sexes. 

The highest NOAEL values and all reliable representative LOAEL values for reproductive effects in each 

species and duration category for each congener are recorded in Tables 2-2 and 2-3 and plotted in 

Figures 2-1 and 2-2. 

2.2.2.6 Developmental Effects 

A number of developmental effects have been observed in animals acutely exposed to 2,3,7,8-TCDD by 

the oral route. The types of effects observed in the offspring of animals exposed to 2,3,7,8-TCDD include 

structural malformations-cleft palate and kidney anomalies, functional alterations-damage to the immune 
CDDs 169 

2. HEALTH EFFECTS 

system and impaired development of the reproductive system, decreased growth, and fetal/newborn 

mortality. 

Cleft palate and other skeletal anomalies have been observed in the offspring of CRCD rats exposed to 

2 μg/kg/day for 2 weeks prior to conception (Giavini et al. 1983), in Long Evans rats exposed to 5 μg/kg 

on Gd 8 (Huuskonen et al. 1994), in rabbits exposed to 0.1 μg/kg/day during Gd 6–15 (Giavini et al. 

1982), in C57BL/6N mice exposed to 12 μg/kg once on Gd 10 (Weber et al. 1985), in C57BL/6 mice 

exposed to 6 μg/kg once on Gd 10 or 12 (Abbott and Birnbaum 1989a), in C57BL/6J mice exposed once 

to 15 μg/kg on Gd 9 (Dasenbrock et al. 1992), in CD-1 mice exposed to 25 μg/kg/day during Gd 7–16 

(Courtney 1976), in CF-1 mice exposed to 1 μg/kg/day on Gd 6–15 (Smith et al. 1976), in DBA2J mice 

exposed to 150 μg/kg on Gd 9 (Dasenbrock et al. 1992), and in NMRI mice exposed to 3 μg/kg/day during 

Gd 6–15 (Neubert and Dillmann 1972). The incidence of cleft palate was not significantly altered in 

Han/Wistar rats exposed to 1 or 10 μg/kg on Gd 8 (Huuskonen et al. 1994), in Long Evans rats exposed to 

1 μg/kg on Gd 8 (Huuskonen et al. 1994), or in C57BL/6N mice exposed to 3 μg/kg/day on Gd 10–13 

(Abbott et al. 1992). 

Kidney anomalies, mainly hydronephrosis, were found in the offspring of CRCD rats exposed to 

2 μg/kg/day for 2 weeks prior to conception (Giavini et al. 1983), in Han/Wistar rats exposed to 10 μg/kg 

on Gd 8 (Huuskonen et al. 1994), in C57BL/6N mice exposed to 12 μg/kg on Gd 10 (Abbott et al. 1987a, 

1987b), in C57BL/6N mice exposed to 1 μg/kg on Gd 10 (Moore et al. 1973), in CD-1 mice exposed 

during Gd 7–16 to 25 μg/kg/day (Courtney 1976), in C57BL/6J and DBA/2J mice exposed to 

0.5 μg/kg/day 2,3,7,8-TCDD on Gd 6–15 (Silkworth et al. 1989b), in C57BL/6N mice exposed postnatally 

through contaminated mothers' milk (Couture-Haws et al. 1991b), and in Golden Syrian hamsters exposed 

to 1.5 μg/g on Gd 7 or 9 (Olson and McGarrigle 1992). An increase in the severity of nephrosis was 

observed in 4–5-month-old Syrian hamsters receiving a single dose of 2 μg/kg in utero on Gd 11(Gray et 

al. 1995). No significant increases in the incidence of hydronephrosis or dilatation of renal pelvis were 

observed in Long Evans rats exposed to 1 or 5 μg/kg on Gd 8 (Huuskonen et al. 1994) or Han/Wistar rats 

exposed to 1 μg/kg on Gd 8 (Huuskonen et al. 1994). 

The immune system effects include thymic atrophy and immunosuppression. Thymic atrophy was found 

in pups of Sprague-Dawley rats exposed to a single dose of 10 μg/kg 2,3,7,8-TCDD on lactation day 1 

(Håkansson et al. 1987) and in Long Evans and Han/Wistar rats exposed to 5 or 10 μg/kg, respectively, on 

Gd 8 (Huuskonen et al. 1994). At lower doses, the thymic atrophy may be transitory; thymic atrophy was 
CDDs 170 

2. HEALTH EFFECTS 

observed on Gd 19 in the offspring of F344 rats exposed to 3.0 μg/kg on Gd 14 but not on Gd 22 (Gehrs et 

al. 1997a). Similarly, transient thymus atrophy was observed in offspring of BALB/cGa mice exposed to 

10 μg/kg on Gd 14 (Fine et al. 1989). A dose-related decrease in relative thymus weights was seen in 

offspring of rats dosed at levels of 0.005–0.35 μg/kg 2,3,7,8-TCDD on Gd 16 (Madsen and Larsen 1989). 

Severe thymic atrophy and cellular depletion occurred in offspring of C57BL/6N mice exposed to 1.5 or 

3 μg/kg/day on Gd 6–14 (Blaylock et al. 1992; Holladay et al. 1991). Thymus size was not affected in the 

offspring of Long Evans or Han/Wistar rats exposed to 1 μg/kg on Gd 8 (Huuskonen et al. 1994). 

Reversible suppression of cell-mediated immunity was reported in pups of Fischer 344 rats exposed to 

2,3,7,8-TCDD through the dosing of dams on lactation day 0, 7, and 14 with 5 μg/kg/day (Faith and 

Moore 1977). Increased neutrophils were found in pups of B6C3F1 mice exposed to 1 μg/kg/day 

2,3,7,8-TCDD on Gd 14 and lactation days 1, 7, and 14 (Luster et al. 1980). Furthermore, increased 

lymphocytes and decreased erythrocytes and hematocrit were recorded in groups exposed to 5 μg/kg/day. 

Alterations in thymocyte phenotypes have also been observed following in utero and/or lactational 

exposure. A decrease in the percentage of CD3-

/CD4-CD8-, CD3+/CD4-CD8-, and CD3+/CD4+CD8+ 

thymocytes and an increase in CD3+/CD4-CD8+ thymocytes were observed in the offspring of F344 rats 

exposed to 1.0 or 3.0 μg/kg on Gd 14 (Gehrs et al. 1997a). A decrease in CD4-

/CD8- thymocytes was 

observed following in utero, lactation only, or in utero and lactational exposure to 1.0 μg/kg (administered 

on Gd 14) (Gehrs et al. 1997b). In utero and lactational exposure also resulted in an increase in the 

percentage of CD4-

/CD8+ lymphocytes; this was not observed in the in utero only or lactation only groups. 

Gehrs et al. (1997b) also found a suppression of the delayed hypersensitivity response to BSA in 5-monthold male offsprings receiving in utero and lactational exposure. 

A number of studies have found impaired development of the reproductive system in male and female 

animals exposed to 2,3,7,8-TCDD during gestation. 2,3,7,8-TCDD affects androgen levels, secondary sex 

organs, spermatogenesis, fertility, and sexual behaviors. Effects have been observed in male and female 

offspring, although most of the studies have focused on males. Malformations of external genitalia 

(clefting, hypospadias, and vaginal thread), delayed vaginal opening (only significant in rats exposed on 

Gd 15; Gray and Ostby 1995), decreased number of ovarian follicles (only tested in Gd 15-exposed rats), 

and decreased fertility have been observed in female offspring of Holtzman and Long Evans rats exposed 

to a single dose of 1 μg/kg on Gd 8 or 15 (Gray and Ostby 1995; Flaws et al. 1997; Heimler et al. 1998). 

Gd 8 exposure also resulted in accelerated onset of constant estrus, shortened reproductive lifespan, and 

increased incidences of cystic hyperplasia of the endometrium. Malformations of the external genitalia 

were also observed in the offspring of Long Evans rats exposed to 0.20 or 0.80 μg/kg on Gd 15 but not at 
CDDs 171 

2. HEALTH EFFECTS 

0.05 μg/kg (Gray et al. 1997a). The fertility rate was not adversely affected in the offspring, but there was 

an increase in time to pregnancy in the 0.80 μg/kg group. In a cross-fostering experiment (Gray et al. 

1997a), similar morphological reproductive alterations were observed following in utero exposure to 

1.0 μg/kg (Gd 15) but not after lactation-only exposure. Chaffin et al. (1996) found significant decreases 

in serum estrogen levels in the female offspring of Holtzman rats receiving a single dose of 1 μg/kg on Gd 

15. This study also found an increase in estrogen receptor mRNA in the hypothalamus, uterus, and ovary 

and a decrease in the pituitary; an increase and a decrease in estrogen receptor binding DNA were found in 

the uterus and hypothalamus, respectively. 2,3,7,8-TCDD exposure did not alter gonadotropin secretion; 

no alterations in serum FSH, LH, or androstenedione levels were observed on postnatal day 21 (Chaffin et 

al. 1997). 

In male Holtzman rats exposed to 1 μg/kg on Gd 15, significant decreases in plasma testosterone were 

observed on Gd 18–21 (Mably et al. 1992a). Additionally, the normal surge in testosterone levels that 

occurs 2 hours after birth was delayed until 4 hours after birth and the amplitude of the surge was lower in 

2,3,7,8-TCDD-exposed male offspring. Bjerke and Peterson (1994) observed significant decreases in 

plasma testosterone levels in Holtzman rats at age 63 days following exposure to 1 μg/kg on Gd 15. But 

studies by Mably et al. (1992a) and Gray et al. (1995) did not find significant alterations in plasma 

testosterone levels in pre- and post-pubescent rats; although Mably et al. (1992a) did report a tendency 

toward dose-related decreases in plasma testosterone and 5α-dihydrotestosterone levels in Holtzman rats at 

age 32, 49, 63, and 120 days following perinatal exposure to 0.064–1 μg/kg on Gd 15. Roman et al. 

(1995) found similar decreases at age 32 and 49 days, but at day 63 the testosterone levels were similar to 

controls. Most studies found large inter-animal variations in plasma testosterone levels; this coupled with 

small sample sizes may have been a contributing factor in the conflicting results that were found. 

Luteinizing hormone levels (primary hormone stimulating testosterone production) were decreased in the 

32-day-old male offspring of dams receiving a dose of 1 μg/kg on Gd 15; no significant alterations in LH 

levels were observed in 49-, 63-, or 120-day-old rats (Mably et al. 1992a). Additional effects on 

androgenic status (androgen concentrations and androgen-dependent structures and functions) have been 

observed. Significant decreases in anogenital distance (corrected for differences in body size by dividing 

by crown-rump length) were observed in male Holtzman rats exposed to $0.16 μg/kg on Gd 15 (Mably et 

al. 1992a); however, Bjerke and Peterson (1994), Bjerke et al. (1994a), and Gray et al. (1995) did not find 

significant alterations in anogenital distance (also corrected for body size) in Holtzman and Long Evans 

rats exposed to 0.7 or 1 μg/kg on Gd 15. More consistent results were found for other measures of 

androgenic status. Significant alterations included a delay in testis descent (Bjerke et al. 1994a), delay in 
CDDs 172 

2. HEALTH EFFECTS 

preputial separation (external indicator of delayed puberty) (Bjerke et al. 1994a; Gray et al. 1995, 1997b), 

decreased ventral prostate weight (Bjerke and Peterson 1994; Mably et al. 1992a), and decreased seminal 

vesicle weight (Bjerke and Peterson 1994; Bjerke et al. 1994a; Mably et al. 1992a). The results of the 

Mably et al. (1992a) study suggest that the decrease in ventral prostate weight may be the most sensitive 

indicator of 2,3,7,8-TCDD-induced toxicity on androgen status. Decreases were observed in the rats 

exposed to $0.064 μg/kg on Gd 15; the other effects were observed in rats exposed to $0.16 μg/kg. 

Decreases in absolute testis weight and cauda epididymis weight were observed in juvenile (Mably et al. 

1992c), pubertal (Mably et al. 1992c), postpubertal (Bjerke and Peterson 1994; Mably et al. 1992c), and 

sexually mature (Gray et al. 1995; Mably et al. 1992c) rats prenatally exposed to 2,3,7,8-TCDD on Gd 15. 

The lowest LOAEL for these effects was 0.064 μg/day (decreased testes weight), identified in the Mably et 

al. (1992c) study. Recent studies from the same group of investigators have focused on evaluating the 

potential role of the Ah receptor (see section 2.4.2 for a detailed discussion on the Ah receptor-mediated 

mechanism of action of CDDs) on the developmental alterations of the male reproductive tract (Roman 

and Peterson 1998; Roman et al. 1998a, 1998b). These studies are summarized in Section 2.5. 

Significant decreases in daily sperm production (Bjerke and Peterson 1994; Mably et al. 1992c; Sommer et 

al. 1996), the amount of mature sperm stored in the cauda epididymis (Bjerke and Peterson 1994; Gray et 

al. 1995; Mably et al. 1992c), and the amount of sperm ejaculated (Gray et al. 1995, 1997b) were observed 

in gestationally exposed rats. These adverse effects on spermatogenesis occurred at doses of $0.05 μg/kg 

(Gray et al. 1997b). Sommer et al. (1996) suggest that observed decreases in cauda epididymal sperm 

number is likely due to a decrease in daily sperm production and an increase in sperm phagocytosis in the 

excurrent duct system. Sommer et al. (1996) and Wilker et al. (1996) did not find alterations in sperm 

epididymal transit time. Significant decreases in follicle stimulating hormone levels (necessary for the 

initiation of spermatogenesis) have been observed in 32-day-old male rats receiving a perinatal dose of 

0.064, 0.40, or 1 μg/kg on Gd 15, but not in 49-, 63-, or 120-day-old rats (Mably et al. 1992c). In 70- and 

120-day-old males exposed to #1 μg/kg and mated with unexposed females, no significant alterations in 

reproductive outcomes (fertility index, gestational index, survival index) were observed (Mably et al. 

1992c); however, a non-significant decrease in fertility index was observed in the 0.4 and 1 μg/kg males. 

Gray et al. (1995) found a significantly decreased number of implants when males exposed to 1 μg/kg on 

Gd 15 were mated with unexposed females. Altered fertility was not observed in the male offspring of rats 

exposed to 2,3,7,8-TCDD on Gd 8 (Gray et al. 1995). Several studies have found a demasculinization of 

sexual behavior (prolonged intromission latency, increased number of intromissions prior to ejaculation) 

(Bjerke et al. 1994b; Mably et al. 1992b) and partial feminization of sexual behavior (increased intensity of 
CDDs 173 

2. HEALTH EFFECTS 

lordosis and lordosis quotient) in male offspring of rats dosed with 2,3,7,8-TCDD on Gd 15 (Bjerke and 

Peterson 1994; Bjerke et al. 1994b; Mably et al. 1992b). Mably et al. (1992b) identified 0.16 μg/kg on 

Gd 15 as the lowest LOAEL for demasculinized and feminized behaviors with a NOAEL of 0.064 μg/kg. 

A demasculinization and feminization of male rats has also been observed in 2,3,7,8-TCDD-exposed males. 

In perinatally exposed castrated rats, no significant alteration in plasma LH levels were observed following 

injection of estradiol benzoate. However, the response to progesterone administration in the rats receiving 

0.40 or 1 μg/kg on Gd 15 was similar to that seen in unexposed ovariectomized females, and the plasma LH 

levels were significantly higher than in control males (Mably et al. 1992b). Impaired development of the 

reproductive system has also been observed in male Syrian hamsters exposed to 2 μg/kg on Gd 11 (Gray et 

al. 1995). Decreased epididymal sperm reserves, decreased testis and cauda epididymides weight, and 

delayed puberty were observed. 

Alterations in mammary gland differentiation (less differentiation) were observed in 50-day-old offspring 

from rats treated by gavage with 1 μg 2,3,7,8-TCDD/kg on gestation day 15 (Brown et al. 1998). 

Specifically, treatment with 2,3,7,8-TCDD resulted in significantly more terminal end buds and fewer 

lobules II. However, no such effect was seen in 21-day-old rats. Prenatal treatment with 2,3,7,8-TCDD 

also resulted in an increased number of mammary adenocarcinomas in the offspring in response to 

dimethylbenz[a]anthracene (DMBA) relative to rats treated with DMBA alone. The authors speculated that 

the decreased differentiation may have rendered the gland more susceptible to mammary cancer. 

Intestinal hemorrhage (Khera and Ruddick 1973; Sparschu et al. 1971a), subcutaneous edema, and 

hemorrhages in brain (Khera and Ruddick 1973) were observed in the offspring of Wistar rats treated with 

0.125 or 0.25 μg/kg/day 2,3,7,8-TCDD during Gd 6–15, and gastrointestinal hemorrhage was observed in 

Han/Wistar rats exposed to 10 μg/kg on Gd 8 or 12 (Huuskonen et al. 1994). Decreases in mammary gland 

size due to inhibition of cell proliferation and gland development were observed in female Sprague-Dawley 

rats dosed with 2.5 μg/kg/day at ages 25, 27, 29, and 31 days (Brown and Lamartiniere 1995). Exposure to 

0.10 μg/kg/day 2,3,7,8-TCDD during gestational days 10–16 resulted in significant decreases in T4 levels 

in female Sprague-Dawley rat pups. Thyroxine levels were not significantly altered in male rats or in 

females exposed to 0.025 μg/kg/day, and no alterations were observed in triiodothyronine or TSH values in 

males and females exposed to either dose (Seo et al. 1995). A decrease in core body temperature was 

observed in the offspring of Long Evans rats exposed to 1 μg/kg on Gd 15; no effect on metabolic rate or 

evaporative heat loss was observed (Gordon et al. 1995). 
CDDs 174 

2. HEALTH EFFECTS 

Decreases in fetal and newborn body weight were observed in Holtzman rats exposed to 0.7 or 1 μg/kg on 

Gd 15 (Bjerke and Peterson 1994; Bjerke et al. 1994a). No body weight effects were observed in 

C57BL/6N mouse fetuses exposed to maternal doses of 3 μg/kg on Gd 10–13 (Abbott et al. 1992). Crownrump length was also decreased in Holtzman rats exposed to 1 μg/kg on Gd 15 (Bjerke and Peterson 1994). 

Several studies have reported increased mortality in the offspring of rats and monkeys exposed to 

2,3,7,8-TCDD during gestation. Fetal/newborn deaths have occurred at doses which were either non-toxic 

or minimally toxic to the mothers. Increased newborn mortality was observed in Holtzman rat pups 

exposed to 0.7 or 1 μg/kg on Gd 15 (Bjerke and Peterson 1994; Bjerke et al. 1994a); and decreased 

numbers of live fetuses, caused by increased resorption and fetal deaths, were observed in monkeys after a 

single exposure to 1 μg/kg on Gd 25, 30, 35, or 40 (McNulty 1984) and in Long Evans or Han/Wistar rats 

exposed to 5 or 10 μg/kg, respectively, on Gd 8 (Huuskonen et al. 1994). After exposure of mouse dams to 

12.5 μg/kg/day 2,3,7,8-TCDD on Gd 14–17, 75% lethality was observed in the pups (Nau et al. 1986). 

In intermediate-duration exposure experiments, decreased neonatal survival was found in the F1 and F2 

generations of Sprague-Dawley rats exposed via the feed to 0.01 μg/kg/day, but not to 0.001 μg/kg/day, of 

2,3,7,8-TCDD in a 3-generation study (Murray et al. 1979). Thymic atrophy was found in offspring of 

Swiss Webster mice that were kept on a diet providing 0.35 μg/kg/day 2,3,7,8-TCDD for 4 weeks before 

mating, during gestation, and for 3 weeks of lactation (Thomas and Hinsdill 1979). No developmental 

effects were found in the offspring of C57BL/6 male mice treated with 3 μg/kg/day of 2,3,7,8-TCDD by 

gavage (in oil or soil vehicle) for 30 weeks (Umbreit et al. 1988). No fetal abnormalities were found in the 

3 fetuses of rhesus monkeys administered 0.02 μg/kg/day 2,3,7,8-TCDD 3 days a week for 3 weeks 

(McNulty 1984). At the higher dosages (0.1 and 0.6 μg/kg/day) only 1 fetus (from the 0.1 μg/kg/day 

group) was not aborted. 

Developmental effects of 2,3,7,8-TCDD after chronic exposure were studied in rhesus monkeys. Decreased 

offspring survival was found when mothers were exposed continuously during pregnancy to 

6.4×10-4 μg/kg/day in the feed (Bowman et al. 1989b). In addition, alterations in peer-group behavior 

(Bowman et al. 1989b; Schantz et al. 1992) and cognitive deficits were observed in the offspring of rhesus 

monkeys exposed to 1.2×10-4 μg/kg/day in the diet for 7 months prior to mating and during mating and 

lactation (16 months total duration). Significant alterations were observed in play behavior, displacement, 

and self directed behavior. Exposed monkeys tended to initiate more rough-tumble play bouts and retreated 
CDDs 175 

2. HEALTH EFFECTS 

less from play bouts than controls, were less often displaced from preferred positions in the playroom than 

the controls, and engaged in more self-directed behavior than controls. Cognitive function was altered as 

evidenced by impaired-reversal-learning performance in the absence of impaired delayed-spatial-alterations 

performance (Bowman et al. 1989b; Schantz and Bowman 1989); no NOAEL was identified for these 

effects. Schantz et al. (1986) also found increased and prolonged maternal care of these infants. The 

LOAEL of 1.2×10-4 μg/kg/day identified for neurobehavioral effects identified in the Schantz et al. (1992) 

study was used to derive a chronic oral MRL of 1×10-6 μg/kg/day, as described in the footnote to Table 2-2, 

Section 2.5, and in Appendix A. 

Other CDD congeners have also been found to induce developmental toxicity. Rat pups exposed in utero to 

2,000 μg/kg/day 2,7-DCDD had edematous separation of the cardiac myofibrils (Khera and Ruddick 1973). 

Schwetz et al. (1973) found no developmental effects in fetuses of rats exposed to 100,000 μg/kg/day 

2,7-DCDD during gestation, but histological examinations of soft tissues were not performed. Decreased 

thymic weight was found in the offspring of rats exposed once on Gd 16 to 0.125 μg/kg 1,2,3,7,8-PCDD 

(Madsen and Larsen 1989). Subcutaneous edema was found in the offspring of Sprague-Dawley rats 

exposed to 1 μg/kg/day of mixed HxCDD during Gd 6–15 (Schwetz et al. 1973). Furthermore, decreased 

fetal body weight, reduced crown-rump length, delayed ossification, and dilated renal pelvis were observed 

at 10 μg/kg/day, and an increased incidence of cleft palate was found at 100 μg/kg/day. The NOAEL for 

the mixture of HxCDD isomers was 0.1 μg/kg/day. Subcutaneous edema was also reported in fetuses of 

rats exposed to 5×105

 μg/kg/day of OCDD during Gd 6–15; no effects were found in the 1×105 μg/kg/day 

OCDD-exposure group (Schwetz et al. 1973) or in mice exposed to 20 μg/kg/day of OCDD during Gd 7–16 

(Courtney 1976). In contrast to most experiments with 2,3,7,8-TCDD, the 1,2,3,4-TCDD isomer did not 

induce developmental effects in the offspring of Wistar rats treated on Gd 6–15 with 800 μg/kg/day (Khera 

and Ruddick 1973) or CD-1 mice exposed to 1,000 μg/kg/day during gestation (Courtney 1976). No 

developmental effects were seen in the offspring of Wistar rats exposed to 2,000 μg/kg/day 2,3-DCDD or 

2-MCDD on Gd 6–15 (Khera and Ruddick 1973). 

In conclusion, studies in rodents and monkeys demonstrated that oral exposure to CDDs induced 

developmental effects with congeners primarily having chlorine atoms at the 2, 3, 7, and 8 positions. 

Following CDD exposure of dams, the primary effects observed in the offspring included cleft palates, 

hydronephrosis, impaired development of the reproductive system, immunotoxicity, and death. Effects 

were observed in offspring from dams exposed before mating and/or during gestation; transfer of CDDs via 

maternal milk also resulted in adverse developmental effects. Mice appear to be particularly sensitive to 
CDDs 176 

2. HEALTH EFFECTS 

the induction of cleft palate; this alteration also occurred in rats, but at dose levels that were maternally 

toxic. Alterations to the immune system from offspring (mostly rats and mice) included thymic atrophy, 

alterations in cell-mediated immunity, and changes in lymphocyte surface cell markers. The development 

of the reproductive systems of male and female rats was also affected by parental exposure to CDDs. 

Chronic exposure of monkeys starting before mating and continuing throughout gestation and lactation 

resulted in neurobehavioral alterations in the infants; this effect was used to derive a chronic oral MRL. 

The highest NOAEL values and all reliable representative LOAEL values for developmental effects in each 

species and duration category for each congener are recorded in Tables 2-2 and 2-3 and plotted in 

Figures 2-1 and 2-2. 

2.2.2.7 Genotoxic Effects 

Mostly negative results were obtained in animal studies following oral exposure to 2,3,7,8-TCDD. 

Cytogenetic analysis of the bone marrow did not reveal any increase in chromosomal aberrations in 

CD-COBS rats exposed to 1 μg/kg 2,3,7,8-TCDD by gavage once a week for 45 weeks (Loprieno et al. 

1982), but an increased incidence was reported in Osborne-Mendel rats exposed to 4 μg/kg twice a week for 

13 weeks (Green et al. 1977). However, chromosomal aberrations were not increased in peripheral 

lymphocytes of monkeys exposed to 0.001 μg/kg/day in the feed for 4 years (Lim et al. 1987). 

Furthermore, 7 daily doses of 12 μg/kg did not induce dominant lethality in Wistar rats (Khera and Ruddick 

1973). In addition, an intermediate-duration exposure to 1 μg/kg/week of 2,3,7,8-TCDD or 1,2,3,7,8­

PCDD for up to 6 months did not enhance the formation of deoxyribonucleic acid (DNA) adducts in rat 

hepatocytes (Randerath et al. 1989). In conclusion, CDDs were not genotoxic in most animal studies. 

Other genotoxicity studies are discussed in Section 2.5. 

2.2.2.8 Cancer 

The carcinogenicity of CDDs has been demonstrated in several experiments in animals. Chronic exposure 

of male Osborne-Mendel rats to approximately 0.0071 μg 2,3,7,8-TCDD/kg/day by gavage significantly 

increased the incidence of thyroid follicular cell adenoma; in females, doses of approximately 

0.071 μg/kg/day increased the incidence of neoplastic nodules in the liver and of hepatocellular carcinoma 

(NTP 1982b). Exposure to 0.00014 μg/kg/day 2,3,7,8-TCDD in the feed resulted in an increase in the 

number of tumor bearing male Sprague Dawley rats (5/6 versus 0/4 in control) (Van Miller et al. 1977). 
CDDs 177 

2. HEALTH EFFECTS 

The types of neoplasms found were ear duct carcinoma, lymphocytic leukemia, kidney adenocarcinoma, 

peritoneal malignant histiocytoma, skin angiosarcoma, and Leydig cell adenoma. Each tumor-bearing rat 

had different tumor types, and one rat had 2 types of tumors. The high mortality in the control group, 

inadequately reported results, and small group sizes (10/group) limit the interpretation of these results. 

Exposure to 0.1 μg/kg/day in the feed induced hepatocellular carcinoma, squamous cell carcinoma of lungs, 

and hard palate and tongue in Sprague-Dawley rats (Kociba et al. 1978a). A significant increase in 

hepatocellular hyperplastic nodules was observed in the female rats exposed to 0.01 or 0.1 μg/kg/day. 

Females were more affected by 2,3,7,8-TCDD exposure than males. 2,3,7,8-TCDD was also carcinogenic 

in mice exposed chronically by gavage. Hepatomas and hepatocellular carcinomas were induced in Swiss 

mice exposed to 0.1 μg/kg/day for 1 year (Toth et al. 1979). Increased incidence of hepatocellular 

carcinomas was observed in male B6C3F1 mice administered 2,3,7,8-TCDD at approximately 

0.071 μg/kg/day by gavage for 104 weeks (NTP 1982b); females exhibited an increase in thyroid follicular 

cell adenomas and in histiocytic lymphoma at a dose of approximately 0.3 μg/kg/day (NTP 1982b). 

Hepatocellular carcinomas (males and females) and adenomas (females) were found in B6C3F1 mice 

exposed to 0.36 μg/kg/day given by gavage for 1 year (Della Porta et al. 1987). 

Experiments with other congeners showed that chronic exposure to a mixture of 1,2,3,6,7,8-HxCDD and 

1,2,3,7,8,9-HxCDD by gavage induced hepatocellular carcinoma, adenoma, and neoplastic nodules at 

approximately 0.34 μg/kg/day in female Osborne-Mendel rats and at 0.71 μg/kg/week in B6C3F1 male mice 

(NCI/NTP 1980). Therefore, HxCDD is approximately 1/20 as potent a liver carcinogen as 2,3,7,8-TCDD. 

Furthermore, chronic exposure to 6.5×105

 μg/kg/day of 2,7-DCDD in the feed caused leukemias, 

lymphomas, hemangiosarcomas, hemangiomas, and dose-related increased incidences of hepatocellular 

adenomas and carcinomas in male B6C3F1 mice (NCI/NTP 1979a). In contrast, no cancer effects were 

observed following chronic exposure of Osborne-Mendel rats to 5×105

 μg/kg/day of 2,7-DCDD (NCI/NTP 

1979a) in the feed. 

In conclusion, the tested congeners (2,3,7,8-TCDD, mixed HxCDDs) were carcinogenic in rodents. 

2,7-DCDD was carcinogenic in mice, but not in rats that received a lower dose. 

The cancer effect levels (CELs) for each species and duration category for each congener are recorded in 

Tables 2-2 and 2-3 and plotted in Figures 2-1 and 2-2. 
 

 

CDDs 178 

2. HEALTH EFFECTS 

2.2.3 Dermal Exposure 

2.2.3.1 Death 

Information regarding mortality following dermal exposure to CDDs in animals is limited. For 

2,3,7,8-TCDD, a dermal LD50 value in rabbits was calculated as 275 μg/kg (Schwetz et al. 1973). Deaths 

occurred within 12–22 days, but the cause of death was not specifically indicated. Decreased survival was 

observed in Swiss Webster mice exposed 3 days a week to 2,3,7,8-TCDD at 0.05 μg for 13 weeks and 

0.001 μg for chronic duration (NTP 1982a). In the subchronic study (NTP 1982a), male mice exhibited a 

higher mortality rate than females; lethal doses in males caused marked effects in lymphoid and hematopoietic tissues as well as on the liver and lung. The cause of death in the chronic study (NTP 1982a) was 

not specified. No increase in lethality was reported in HRS/J hairless mice dermally exposed to 0.0025 μg, 

2 days a week, for 20 weeks (Hebert et al. 1990). 

The LD50 value for rabbits and LOAEL values in mice for increased mortality for the intermediate- and 

chronic-duration categories are recorded in Tables 2-4 and 2-5. 

2.2.3.2 Systemic Effects 

The highest NOAEL values and all reliable LOAEL values for each systemic effect in each species and 

duration category for each congener are recorded in Table 2-4 and 2-5. 

Respiratory Effects. Bronchiolar adenomatoid changes were found in Swiss Webster mice exposed 

3 days a week to 0.05 μg 2,3,7,8-TCDD for 13 weeks, but no respiratory effects were observed in mice 

exposed 3 days per week to 0.01 μg for a chronic exposure period (NTP 1982a). 

Cardiovascular Effects. Information regarding cardiovascular effects in animals after dermal 

exposure to CDDs is limited. Chronic dermal exposure of Swiss Webster mice to 2,3,7,8-TCDD at 

0.005 μg, 3 days per week, did not induce any cardiovascular changes observable under histopathological 

examination (NTP 1982a). 






 

 

 

 

 

 

CDDs 185 

2. HEALTH EFFECTS 

Gastrointestinal Effects. Information regarding gastrointestinal effects in animals after dermal 

exposure to CDDs is limited. No histopathological changes were observed in the gastrointestinal tract of 

Swiss Webster mice chronically exposed to 0.005 μg 2,3,7,8-TCDD 3 days per week (NTP 1982a). 

Hematological Effects. Hematological examination of Swiss Webster mice chronically exposed to 

0.005 μg 2,3,7,8-TCDD 3 days per week did not reveal any differences between exposed and control 

groups (NTP 1982a). 

Musculoskeletal Effects. No studies were located regarding musculoskeletal effects in animals after 

dermal exposure to CDDs. 

Hepatic Effects. Hepatic effects have been observed in animals after dermal exposure to 

2,3,7,8-TCDD. Necrosis, peripheral fibrosis, and bile duct proliferation were observed in rabbits acutely 

exposed to 2,3,7,8-TCDD on the ear surface (Kimbrough et al. 1977). Increased liver/body weight ratio 

and hepatocellular hypertrophy were seen in HRS/J hairless mice topically exposed to 0.0025 μg 

2,3,7,8-TCDD twice a week for 20 weeks (Hebert et al. 1990). Fatty degeneration and hepatocellular 

necrosis were observed in the livers of Swiss Webster mice exposed to 0.005 μg 2,3,7,8-TCDD 3 days a 

week for 13 weeks (NTP 1982a). No hepatic effects were found in mice chronically exposed to 

0.005 μg/day 2,3,7,8-TCDD (NTP 1982a). The data indicated that 2,3,7,8-TCDD induced hepatotoxic 

effects similar to those observed after oral exposure. 

Renal Effects. Information regarding renal effects in animals after dermal exposure to 2,3,7,8-TCDD is 

limited. No histopathological changes were found in Swiss Webster mice exposed to 0.005 μg 

2,3,7,8-TCDD 3 days per week for 99–104 weeks (NTP 1982a). 

Dermal Effects. Dermal effects of several CDD congeners have been studied in animals. Acute dermal 

exposure to 0.01 μg (newborn) and 0.1 μg 2,3,7,8-TCDD (adult) per animal caused hyperkeratosis and 

epidermal hyperplasia in hairless HRS/J mice (Puhvel and Sakamoto 1988). An involution of sebaceous 

glands was found in both (haired and hairless) strains. Similar results were found following intermediateduration exposure (Puhvel et al. 1982). Furthermore, acne-like lesions in the ears were found in CD-1 mice 

following exposure to 0.1 μg 2,3,7,8-TCDD applied on the pre-shaved back 2 days a week for 30 weeks 

(Berry et al. 1978, 1979). In contrast, no dermal effects were observed in Swiss Webster mice exposed to 

0.005 μg 2,3,7,8-TCDD/application, 3 days a week for up to 104 weeks (NTP 1982a). 
 

 

CDDs 186 

2. HEALTH EFFECTS 

Ocular Effects. A single application of 2,000 μg 2,7-DCDD, 2,3,7,8-TCDD, mixed HxCDD, or OCDD 

into the conjunctival sac of rabbits caused transient pain and conjunctival inflammation (Schwetz et al. 

1973). Delayed conjunctival chemosis was observed with 2,3,7,8-TCDD. None of the CDDs caused 

corneal injury or iritis. 

Body Weight Effects. In animal studies, decreased body weight was observed in HRS/J and Skjh:HR1 mice following intermediate-duration dermal exposure to 0.1 μg 2,3,7,8-TCDD (Puhvel et al. 1982) and 

in Swiss Webster mice following chronic exposure to 0.005 μg 2,3,7,8-TCDD 3 days per week (NTP 

1982a). 

2.2.2.3 Immunological Effects 

The only information regarding immunological effects in animals after dermal exposure to CDDs was 

obtained from an intermediate-duration study in HRS/J mice (Hebert et al. 1990). Mice dermally exposed 

to 0.01 μg 2,3,7,8-TCDD 2 days per week for 20 weeks had decreased thymus/body weight ratio. No 

effects were observed at 0.005 μg. 

The NOAEL and LOAEL values for immunological effects in mice after intermediate-duration exposure are 

recorded in Table 2-4. 

2.2.3.4 Neurological Effects 

No studies were located regarding neurological effects in animals after dermal exposure to CDDs. 

2.2.3.5 Reproductive Effects 

Data regarding reproductive effects following dermal exposure in animals are scarce. No treatment-related 

changes were observed in the reproductive system of Swiss Webster mice after chronic exposure to 

0.005 μg 2,3,7,8-TCDD per application (NTP 1982a). 

2.2.3.6 Developmental Effects 

No studies were located regarding developmental effects in animals after dermal exposure to CDDs. 
CDDs 187 

2. HEALTH EFFECTS 

2.2.3.7 Genotoxic Effects 

No studies were located regarding genotoxic effects in animals after dermal exposure to CDDs. 

Genotoxicity studies are discussed in Section 2.5. 

2.2.3.8 Cancer 

Acute- and intermediate-duration studies in animals investigated the interactions of 2,3,7,8-TCDD with 

known carcinogens. A single dermal pretreatment of CD-1 Charles River mice with 0.01 μg 2,3,7,8-TCDD 

inhibited the development of skin papillomas otherwise initiated by 1,3-dimethylbenz(o)anthracene 

(DMBA) (Berry et al. 1979). In intermediate-duration experiments, 2,3,7,8-TCDD did not promote skin 

tumors initiated by DMBA (Berry et al. 1978, 1979). In contrast, the promoting ability of 2,3,7,8-TCDD at 

0.0025 μg/day (and higher), 2 days a week, for 20 weeks, was reported in HRS/J hairless mice following 

the initiation with N-methyl-N-nitro-N-nitrosoguanidine in intermediate-duration experiments (Hebert et al. 

1990; Poland et al. 1982). The effect was not observed in mice heterozygous for the hairless trait (Poland et 

al. 1982). In a chronic study, significantly increased incidence of fibrosarcoma of the integumentary system 

was found in Swiss Webster female mice following dermal exposure to 2,3,7,8-TCDD at 0.005 μg, 3 days a 

week for 2 years (NTP 1982a). The cancer effect level (CEL) from this study is shown in Table 2-4. 

2.3 TOXICOKINETICS 

Data regarding toxicokinetics of CDDs in humans are limited to information derived from exposures that 

occurred after industrial accidents, exposures of Vietnam veterans, and ingestion of 2,3,7,8-TCDD by a 

volunteer. Humans can absorb CDDs by the inhalation, oral, and dermal routes of exposure. CDDs, when 

administered orally, are well absorbed by experimental animals, but they are absorbed less efficiently when 

administered by the dermal route. Limited data in rats showed that transpulmonary absorption of 

2,3,7,8-TCDD may be at least as efficient as oral absorption. In a human volunteer, >86% of the 

administered single oral dose appeared to have been absorbed. In general, absorption is vehicle-dependent 

and congener-specific. Passage across the intestinal wall is predominantly limited by molecular size and 

solubility. These parameters are most significant for hepta- and octachlorinated congeners, which exhibit 

decreased absorption in mammals. The predominant CDD carriers in human plasma are serum lipids and 
CDDs 188 

2. HEALTH EFFECTS 

lipoproteins, but chlorine substitution plays a role in the distribution in these fractions. For most 

mammalian species, the liver and adipose tissue are the major storage sites of CDDs; in some species, skin 

and adrenals also can act as primary deposition sites. 2,3,7,8-Substituted CDDs are the predominant 

congeners retained in tissues and body fluids. Tissue deposition is congener-specific and depends on the 

dose, the route of administration, and age. CDDs are very slowly metabolized by the microsomal 

monooxygenase system to polar metabolites that can undergo conjugation with glucuronic acid and 

glutathione. The major routes of excretion of CDDs are the bile and the feces; smaller amounts are excreted 

via the urine. In mammalian species, lactation is an effective way of eliminating CDDs from the liver and 

other extrahepatic tissues. Physiologically based pharmacokinetic (PBPK) models have been developed to 

describe disposition of 2,3,7,8-TCDD in humans and animals. Some of these models included parameters 

to describe complex interactions of 2,3,7,8-TCDD with cellular proteins that lead to specific biological 

responses. 

2.3.1 Absorption 

2.3.1.1 Inhalation Exposure 

No quantitative data were located regarding absorption of CDDs in humans following inhalation exposure. 

However, based on data from studies with structurally related chemicals it is reasonable to assume that 

CDDs are absorbed by this route. Furthermore, data on levels of CDDs in blood from populations with 

above-background exposures (occupational, accidental) also suggest that transpulmonary absorption occurs 

in humans; see Section 2.1 for more information. 

Systemic effects (hepatic aryl hydrocarbon hydroxylase [AHH] and cytochrome P-450 induction, hepatic 

histological alterations) were observed in rats following a single intratracheal instillation of 2,3,7,8-TCDD 

in a corn oil vehicle or as a laboratory-prepared contaminant of gallium oxide particles (Nessel et al. 1990). 

In a subsequent study, the same group of investigators (Nessel et al. 1992) using a similar protocol found 

that the relative pulmonary bioavailability of 2,3,7,8-TCDD on respirable soil particles was 100% as 

compared to the gallium oxide vehicle. One and 7 days post-treatment, 13.9 and 11.9% of the administered 

dose were detected in the liver, respectively, and this was similar to the percentage found after instillation of 

contaminated gallium oxide particles. Twenty-eight days after treatment, 5.2% of the administered dose 

was detected in the liver from soil-treated rats and 2.9% in liver from gallium oxide-treated rats suggesting 

that redistribution and retention of 2,3,7,8-TCDD differed in the two treatment groups. Recently, Diliberto 
CDDs 189 

2. HEALTH EFFECTS 

et al. (1996) reported that 3 days after intratracheal application of a single dose of 0.32 μg 2,3,7,8-TCDD/kg 

to male Fischer 344 rats, 95% of the applied dose was absorbed, suggesting that inhalation can be an 

effective route of exposure. The extent of inhalation absorption was higher than when the same dose was 

administered orally (88%) or dermally (40%). The available data suggest that inhaled CDDs will be 

absorbed. However, the degree of absorption and the rate will depend on the media on which the CDDs are 

adsorbed and the degree of chlorination. 

2.3.1.2 Oral Exposure 

The absorption of 2,3,7,8-TCDD was estimated to be >87% in a human volunteer following ingestion of a 

single radioactively labeled dose of 0.00114 μg 2,3,7,8-TCDD/kg in corn oil (Poiger and Schlatter 1986). 

Data regarding absorption of CDDs from breast milk in nursing infants are provided in Section 2.3.4.4. 

Gastrointestinal absorption of radiolabeled 2,3,7,8-TCDD has been investigated in rodents. About 73.5% of 

the total dose of 2,3,7,8-TCDD (administered by gavage in corn oil vehicle) was absorbed in Syrian 

hamsters, the species most resistant to acute 2,3,7,8-TCDD toxicity (Olson et al. 1980b). In SpragueDawley rats given a single gavage dose of 50 μg/kg 2,3,7,8-TCDD in corn oil, at least 70% was absorbed 

(Piper et al. 1973). Rose et al. (1976) found a mean of 84% of a single oral gavage dose of 1 μg/kg 

absorbed within a day in a similar study and a steady-state body burden was achieved after dosing with 

0.01, 0.1, or 1 μg/kg in corn oil, 5 days a week for 7 weeks. When 14C-2,3,7,8-TCDD was fed to SpragueDawley rats at 0.35 μg/kg/day or 1 μg/kg/day in the diet for 42 days, about 60% of the consumed dose was 

absorbed (Fries and Marrow 1975). Intestinal absorption of 2,3,7,8-TCDD did not vary with age of 

Fischer 344 rats (13 weeks, 13 or 26 months) when in vivo absorption was studied with an in situ intestinal 

perfusion technique (Hebert and Birnbaum 1987). When ICR/Ha Swiss mice were given a single dose of 

radioactively labeled 2,3,7,8-TCDD, 67–76% of the administered dose was excreted in feces and 1–2% in 

urine within the first 24 hours (Koshakji et al. 1984). The authors (Koshakji et al. 1984) concluded that 

most of the dose was not absorbed. The more highly chlorinated CDD congeners are absorbed from the 

gastrointestinal tract to a lesser extent than 2,3,7,8-TCDD. 

Gastrointestinal absorption of 2,3,7,8-TCDD may differ depending on the vehicle used. When hepatic 

concentrations were used as a measure of absorbed dose, the levels observed in rats 24 hours after 

2,3,7,8-TCDD administration in 50% ethanol were higher than in an aqueous suspension of soil (Poiger and 

Schlatter 1980). Use of activated carbon as a vehicle almost completely eliminated 2,3,7,8-TCDD 
CDDs 190 

2. HEALTH EFFECTS 

absorption. It was further demonstrated that the absorption of 2,3,7,8-TCDD from the gastrointestinal tract 

of rats was .50% less from contaminated soil than from corn oil (Lucier et al. 1986), which is supported by 

the finding that 2,3,7,8-TCDD-contaminated soil was less toxic to guinea pigs than an equivalent amount of 

2,3,7,8-TCDD in oil (Umbreit et al. 1985). Gastrointestinal absorption of OCDD was <10% of the 

administered dose in Sprague-Dawley and Fischer 344 rats following single or repeated (3-week) exposures 

by gavage in oil vehicle (Birnbaum and Couture 1988; Norback et al. 1975). Low doses (50 μg/kg) in a 

o-dichlorobenzene:corn oil (1:1) vehicle were found to give the best oral bioavailability for this extremely 

insoluble compound (Birnbaum and Couture 1988). The bioavailability of CDDs (2,3,7,8-TCDD, 

1,2,3,7,8-PCDD, 1,2,3,6,7,8-HxCDD, and 1,2,3,7,8,9-HxCDD) to rats was lower on fly ash (0.4% for 

2,3,7,8-TCDD) as compared to extracts of the same fly ash administered in an oily vehicle (45% for 

2,3,7,8-TCDD) (Van den Berg et al. 1983, 1987c). The differences in hepatic levels between fly ash- and 

extract-treated rats were greater for the more highly chlorinated congeners. 

2.3.1.3 Dermal Exposure 

No quantitative data were located regarding absorption of CDDs in humans following dermal exposure. 

However, based on data from studies with structurally related chemicals it is reasonable to assume that 

CDDs are absorbed by this route. Furthermore, data on levels of CDDs in blood from populations with 

above-background exposures (i.e., occupational, accidental) also suggest that dermal absorption occurs in 

humans. Due to the relatively low vapor pressure and high lipid solubility, dermal uptake of 2,3,7,8-TCDD 

in the workplace may be a significant source of occupational exposure (Kerger et al. 1995). 

Kerger et al. (1995) examined the potential contribution of dermal exposure to 2,3,7,8-TCDD for three 

different occupational exposure scenarios: 1) trichlorophenoxy herbicide manufacturing worker (20-year 

exposure), 2) contract maintenance mechanic exposed by repairing a trichlorophenol reactor after an 

explosion accident (6-week exposure), and 3) trichlorophenoxy applicator handling only diluted 

trichlorophenoxy herbicides (seasonal exposure for 20 years). In their evaluation, the authors used a 

conceptual model of workplace exposure, dermal bioavailability/uptake calculations, and simple 

pharmacokinetic modeling techniques (details of the model were not provided). The contribution of 

background uptake of 2,3,7,8-TCDD from dietary sources in the United States was accounted for in the 

estimates of steady-state adipose concentrations. The results of the modeling showed that considerable 

occupational uptake can occur following both long-term continuous exposure and short-term high exposure. 

In the former case, occupational uptake can be distinguished from background exposures when 
 

CDDs 191 

2. HEALTH EFFECTS 

body burden is measured within a 10-year period following cessation of exposure. In contrast, seasonal 

exposure to dilute 2,3,7,8,-TCDD residues may result in little or no change in 2,3,7,8-TCDD body burden. 

The in vitro penetration of 3

H-labeled 2,3,7,8-TCDD into human cadaver skin was studied at concentrations 

of 6.5 and 65 ng 2,3,7,8-TCDD/cm2

 of skin (Weber et al. 1991). Two vehicles were used: acetone to 

simulate exposure to 2,3,7,8-TCDD as a dry material, and mineral oil to simulate exposure in an oily 

medium. The experiments were conducted in intact skin and in skin with stripped stratum corneum and 

penetration was monitored for 30, 100, 300, and 1,000 minutes. The results showed that acetone as a 

vehicle allowed 2,3,7,8-TCDD to penetrate deeply into the loose surface of the lamellae of the stratum 

corneum, but there was little further penetration. On the other hand, mineral oil appeared to compete with 

lipophilic constituents of the stratum corneum for 2,3,7,8-TCDD, thus slowing its penetration even more. 

Removal of the stratum corneum increased the amount of 2,3,7,8-TCDD absorbed into layers of the skin. 

Rates of absorption were calculated in two ways: a worst case scenario where 2,3,7,8-TCDD absorbed into 

any layer of skin including the stratum corneum was used for analysis; and a physiological approach where 

only the amount of 2,3,7,8-TCDD which had penetrated beyond the epidermis into the region of dermal 

vascularization was considered absorbed. In the former case, the stratum corneum appeared to mediate 

dermal absorption of 2,3,7,8-TCDD since the rates decreased when stripped skin was exposed to 

2,3,7,8-TCDD. With the physiological approach, the rate of absorption was a function of the amount 

applied suggesting that the rate of absorption per unit time was a first-order function. The amount of 

2,3,7,8-TCDD that penetrated the skin also correlated with exposure duration. The rate of 2,3,7,8-TCDD 

penetration with acetone as vehicle ranged from 100 to 800 pg 2,3,7,8-TCDD per hour-cm2

 (worst-case 

scenario), or 6–170 pg per hour-cm2

 with the physiological approach. The corresponding values with 

mineral oil as a vehicle were 20–220 pg and 1.4–18 pg per hour-cm2

, respectively. 

Data regarding dermal absorption of CDDs in animals are limited. When 200 pmol 2,3,7,8-TCDD was 

applied to the skin of Fischer 344 rats, absorption followed first-order kinetics with an absorption rate 

constant of 0.005 hour-1 (Banks and Birnbaum 1991). Within 120 hours postexposure, about 0.026 μg 

2,3,7,8-TCDD was absorbed (less than 50% of the applied dose); at each interval of measurement, about 

70% of detected radioactivity on the skin could be removed by swabbing with acetone. About 15% of the 

dose was detected in the liver of rats 24 hours after dermal exposure to 26 ng of 2,3,7,8-TCDD in 50% 

methanol (Poiger and Schlatter 1980). It was estimated that the amount absorbed from the dermal 

exposure represents .40% of the amount absorbed from an equivalent oral dose. Absorption of 

2,3,7,8-TCDD was significantly reduced by application in Vaseline or polyethylene glycol and practically 
CDDs 192 

2. HEALTH EFFECTS 

eliminated in soil or activated carbon. Dermal absorption of radioactively labeled 2,3,7,8-TCDD in soil 

vehicle was reported to be only 1% of the administered dose during a 24-hour contact in rats (Shu et al. 

1988). The dermal absorption of 2,3,7,8-TCDD after 4 hours of contact was about 60% of that after 

24-hour contact. The uptake was not influenced by the 2,3,7,8-TCDD concentration in soil, nor were 

there any differences between normal and hairless rats. 

Dermal absorption in rats was found to be age-related. Banks et al. (1990) found that in Fischer 344 rats, 

the percutaneous absorption was decreased in middle-aged (36-week-old) and senescent (120-week-old) 

rats compared to that in young adults (10-week-old) 72 hours after application of a dose of 40 nmol 

(approximately 12.9 μg) of 3

H-labeled 2,3,7,8-TCDD. The authors suggested a decrease in blood flow 

through the skin between 3 and 4 months of age as a possible explanation for their findings. In a 

subsequent and similar study, the same group of investigators examined the dermal absorption of 

2,3,7,8-TCDD in 3-, 5-, 8-, 10-, and 36-week-old Fischer 344 rats 72 hours after application of 200 pmol 

2,3,7,8-TCDD in acetone (Banks et al. 1993). Dermal absorption was greatest in 3-week-old rats 

(approximately 64% of the applied dose), decreased to about 40% of the applied dose in 5-, 8-, and 

10-week-old rats and to about 22% in 36-week-old rats. In each age group, 70–80% of the radioactivity 

remaining at the application site 72 hours after dosing could be removed with acetone swabs. 

2.3.2 Distribution 

As discussed in Section 2.1, occupational or environmental human exposure to CDDs is not readily 

classifiable as to route of exposure. Human data regarding distribution obtained at autopsy indicated that 

accumulation in the liver following low levels of exposure is based in part on lipid solubility (Leung et al. 

1990a). However, this may not be the case with higher exposure levels that cause hepatic enzyme 

induction (see Section 2.4.1). When human hepatic and adipose tissues were examined for the presence 

of 2,3,7,8-TCDD, the concentration detected in the liver was about 1/10 of that in the adipose tissue on a 

whole-tissue-weight basis. However, on the basis of the total tissue lipid, the concentration in adipose 

tissue lipid was one-half that in the liver lipid (Thoma et al. 1990). It was further demonstrated that over 

a wide range of concentrations, the serum 2,3,7,8-TCDD levels highly correlated with adipose tissue 

2,3,7,8-TCDD levels when both are expressed on a lipid weight basis (Patterson et al. 1988). Adipose 

tissue serves as a storage depot for 2,3,7,8-TCDD in the body, and detectable levels (up to 20.2 ppt) were 

found in the general population with no known risk of high exposure to CDDs (Andrews et al. 1989). An 

average concentration of 2,3,7,8-TCDD in serum lipid of 5.38 ppt has been estimated for the U.S. 
CDDs 193 

2. HEALTH EFFECTS 

population (Orban et al. 1994). The distribution of highly chlorinated CDDs among tissue lipid fractions 

is not equal. For example, the distribution of OCDD is 12:1 (Thoma et al. 1990) between liver and 

adipose tissue lipid factions and 2:1 between serum and adipose tissue lipid fractions (Schecter et al. 

1990). 

Increased adipose tissue levels of CDDs were reported in populations with known high residential or 

occupational exposure (Beck et al. 1989c; Fingerhut et al. 1989; Patterson et al. 1989b; Schecter et al. 

1994). For example, high levels of 2,3,7,8-TCDD were found in fat (42–750 ppt) and serum lipid 

(61–1,090 ppt) of Missouri chemical workers (Patterson et al. 1989b). Measurable CDDs and CDFs 

levels were reported in the liver tissue of human stillborn neonates suggesting that the transplacental 

intrauterine transfer of these persistent chemicals resulted from environmentally exposed mothers 

(Schecter et al. 1990). In addition, CDDs are distributed to human milk (i.e., Fürst et al. 1994; Schecter et 

al. 1987a, 1987b, 1989e) and numerous studies have published concentrations of various congeners in 

human milk samples (see section 5.5.1). Levels of CDDs in human milk have been found to be 

significantly and positively associated with of proximity of residence to waste sites and to dietary fat 

intake per week (Schaud et al. 1995). 

2.3.2.1 Inhalation Exposure 

The tissue distribution of 2,3,7,8-TCDD-derived radioactivity was recently examined in male Fischer 344 

rats 3 days after intratracheal application of a single dose of 0.32 μg 2,3,7,8-TCDD/kg (Diliberto et al. 

1996). The liver and adipose tissue were the major tissue depots for 2,3,7,8-TCDD-derived radioactivity 

with 33 and 15% of the applied dose distributing to these respective tissues. The skin (ear) and muscle 

followed with 4.3 and 1.3%, respectively. All other tissues had less than 0.5% of the administered dose. 

The 2/1 liver/adipose ratio was in contrast to the approximately 1/1 ratio found after gavage 

administration of the same dose. 

2.3.2.2 Oral Exposure 

Following an ingested dose of 3

H-2,3,7,8-TCDD of 0.00114 μg/kg by a volunteer, the concentration of 

2,3,7,8-TCDD in the adipose tissue were 3.09 and 2.86 ppt at 13 and 69 days following exposure, 

respectively (Poiger and Schlatter 1986). The authors estimated that about 90% of the body burden was 

distributed to the fatty tissue. Increased radioactivity was detected in the blood only during the first 2 days 
CDDs 194 

2. HEALTH EFFECTS 

postexposure; no radioactivity was detected in serum lipid after 5 days, but was in the feces for several 

months. 

Studies in animals have shown that 2,3,7,8-TCDD distributes preferentially to the liver and adipose tissue. 

In Sprague-Dawley rats, the highest levels of radioactivity (expressed as percentage of dose per gram of 

tissue) were located in the liver (3.18, 4.49, and 1.33% at days 3, 7, and 21 post-exposure, respectively) and 

adipose tissues (2.6, 3.22, and 0.43% at days 3, 7, and 21, respectively) following a single oral dose of 

labeled 2,3,7,8-TCDD at 50 μg/kg (Piper et al. 1973). Much smaller amounts were found in muscles, 

testes, lungs, stomach, and other organs. In male Fischer 344 rats administered a single gavage dose of 

0.32 μg 2,3,7,8-TCDD/kg, 24.4 and 26.2% of the administered dose was found in the liver and adipose 

tissue, respectively, 3 days after dosing (Diliberto et al. 1996); skin and muscle had 7.3 and 1.8%, 

respectively. 2,3,7,8-TCDD accumulated mainly in the liver and adipose tissue, with smaller amounts in 

the brain of pregnant Wistar rats after 10 daily doses of 2 μg/kg (Khera and Ruddick 1973). Similarly, the 

highest levels of radioactivity were found in the liver, adipose tissue, and the adrenals of Golden Syrian 

hamsters after a single gavage dose of 650 μg/kg labeled 2,3,7,8-TCDD (Olson et al. 1980b). In addition, 

about 36% of the total radioactivity administered remained in the adipose tissue by day 45 postexposure in 

Hartley guinea pigs; only about 7% (each) was found in the liver, muscles, and carcass (Olson 1986). 

Essentially all of the administered dose was unchanged 2,3,7,8-TCDD. When pregnant NMRI mice were 

exposed to a single oral, intraperitoneal, or subcutaneous dose of 2,3,7,8-TCDD, hepatic levels were about 

the same, indicating that there is no major first pass effect after oral 2,3,7,8-TCDD exposure (Nau and Bass 

1981). Liver, then adipose tissue and skin, were the major depots of OCDD in Fischer 344 rats treated with 

single oral doses of this congener (Birnbaum and Couture 1988). 

The dose- and time-dependent tissue distribution of 2,3,7,8-TCDD in mice has been recently examined 

(Diliberto et al. 1995). Female B6C3F1 mice were administered a single dose of 0.1, 1, or 10 μg 

[

3

H]-2,3,7,8-TCDD/kg by gavage in corn oil and the distribution of radioactivity was followed in 18 tissues 

for up to 35 days after dosing. The results showed dose-dependent distribution of 2,3,7,8-TCDD-derived 

radioactivity in all tissues. The highest concentrations of radioactivity were found in liver and adipose 

tissues, and both tissues accounted for 50% of the body burden. Relatively high concentrations of 

2,3,7,8-TCDD-derived radioactivity were also found in skin, adrenal glands, thyroid, pancreas, olfactory 

epithelium, spleen, mesenteric lymph nodes, thymus, lung, and bone marrow. The liver concentration of 

radioactivity increased disproportionally with increasing doses, whereas relative concentration and 

percentage dose/total tissue in extrahepatic tissues decreased with increasing dose and over time. 
CDDs 195 

2. HEALTH EFFECTS 

Liver/adipose tissue concentration ratios were shown to be dose- and time-dependent. At the low-, midand high-dose, the ratios ranged from 0.6 to 0.2, 2.3–0.5, and 3.1–1.4 over time, respectively. This 

variation over time was thought to have been due to redistribution of 2,3,7,8-TCDD between the two 

storage sites and/or hepatic metabolism and subsequent excretion. 

The effect of age of the animal on 2,3,7,8-TCDD tissue distribution has also been examined. Pegram et al. 

(1995) administered a single dose of 0.015, 0.5 or 15 μg [3

H]2,3,7,8-TCDD/kg to 10-week- and 28-monthold male C57BL/6N mice and monitored 2,3,7,8-TCDD-derived radioactivity in blood, liver, skin, kidney, 

and muscle 7 days after dosing. The results showed that in young mice given the low- and high-dose, the 

concentration of 2,3,7,8-TCDD in blood relative to all other tissues was significantly greater than in older 

mice. Also, in older mice, the concentration of 2,3,7,8-TCDD in skin and the percentage of the dose in the 

skin were greater than in the young mice. The same trend was observed in kidney and muscle. The 

concentration of 2,3,7,8-TCDD in liver, as well as the percentage of the dose in the liver, were greater in 

young than old animals at both the mid- and high-doses. In both young and old mice the ratios of liver to 

adipose tissue increased with increasing doses. According to the authors, the higher hepatic concentration 

of 2,3,7,8-TCDD in young mice could be due to the old mice having a larger fat compartment, such that the 

hepatic 2,3,7,8-TCDD sequestering action of CYP1A2 or other inducible binding factors may have been 

less effective in the more obese older mice. In addition, decreased perfusion in the liver and adipose 

compartments in the old mice may have limited the effectiveness of hepatic 2,3,7,8-TCDD accumulation. 

The greater accumulation of 2,3,7,8-TCDD in the skin, muscle, and kidney from old mice were attributed to 

altered perfusion and possibly greater lipid infiltration in these tissues. 

The subcellular distribution of 2,3,7,8-TCDD-derived radioactivity in the liver, lungs, and kidneys from 

female Sprague-Dawley rats and B6C3F1 mice was studied by Santostefano et al. (1996). In the liver of 

rats given a single oral dose of 0.1, 1, or 10 μg [3

H]-2,3,7,8-TCDD/kg, radioactivity accumulated equally in 

the supernatant (S9, cytosol, and microsomes) and pellet (P9, nucleus, lysosomes, and mitochondria) 

fractions; within the S9 fraction, accumulation was predominantly in the microsomal fraction. In contrast, 

in kidneys and lungs radioactivity accumulated preferentially in P9, but radioactivity detected in S9 was 

mostly in the cytosolic fraction. The pattern of distribution of radioactivity in liver and lungs from mice 

was similar to that found in rats, but in mice kidneys, 2,3,7,8-TCDD detected in S9 was equally distributed 

between the microsomal and cytosolic fractions. Accumulation of 2,3,7,8-TCDD in the various fractions in 

this single-dose study was not dose-dependent. The investigators also conducted a 17-week oral dosing 

study in B6C3F1 mice given 1.5 or 150 ng/kg that showed that increasing the dose resulted in equal 
CDDs 196 

2. HEALTH EFFECTS 

accumulation between liver S9 and P9 fractions, whereas the kidney P9 had the most radioactivity 

regardless of the dose. In addition, liver S9 accumulated 2,3,7,8-TCDD in the microsomal fraction, 

whereas kidney S9 did it predominantly in the cytosol. These results are consistent with the hypothesis 

that hepatic microsomal sequestration of 2,3,7,8-TCDD is mediated by cytochrome P-4501A2 (CYP1A2), a 

dioxin-inducible protein. This hypothesis was subsequently confirmed by experiments in transgenic mice 

lacking expression of CYP1A2 (CYP1A2-/-) (Diliberto et al. 1997). These mice, as judged by 

2,3,7,8-TCDD liver/fat concentration ratios, failed to sequester 2,3,7,8-TCDD in the liver after 

administration of a single dose of 2,3,7,8-TCDD. 

Intermediate-duration exposure to 2,3,7,8-TCDD in the feed has been shown to produce higher liver 

accumulation in male (85%) than in female rats (70%) (Fries and Marrow 1975). The percentage retained 

was related to intake, and at steady state, the total amount retained was about 10.5 times the average daily 

intake.

 Intermediate-duration studies have also been conducted with radioactively labeled OCDD. OCDD had 

similar patterns of distribution and similar half-lives as 2,3,7,8-TCDD in Sprague-Dawley (Norback et al. 

1975) and Fischer 344 rats (Birnbaum and Couture 1988; Birnbaum et al. 1989a). Most of the absorbed 

amount (50–97%) was found in the liver and was associated with the microsomal fractions. Skin- and 

adipose-tissue levels were much lower. Radioactivity was also detected in the kidneys, heart, testes, 

skeletal muscle, and serum. 

2.3.2.3 Dermal Exposure 

Male Fischer 344 rats absorbed 40% of a single dermal dose of 0.32 μg of radioactive 2,3,7,8-TCDD/kg 

over a period of 120 hours after dosing (Banks and Birnbaum 1991). The major depots for 2,3,7,8-TCDDderived radioactivity were the liver and adipose tissue. Seventy-two hours after dosing, the liver and 

adipose tissue retained approximately 21 and 8% of the administered dose, respectively. Distribution to the 

liver increased significantly between 4 and 8 hours and between 12 and 72 hours after dosing. Distribution 

in fat increased significantly between 12 and 120 hours after dosing. Skin and muscle accumulated 

considerably less 2,3,7,8-TCDD-derived radioactivity than liver and fat. Within 120 hours of dosing, less 

than 4% of the administered dose was found in either of these tissues. When 2,3,7,8-TCDD was dermally 

applied to HRS/J hairless mice for an intermediate duration, about 5–6% of the total administered dose 

(0.0025–0.01 μg/kg, 2 days a week, for 20 weeks) was detected in the liver (Hebert et al. 1990). 
CDDs 197 

2. HEALTH EFFECTS 

2.3.3 Metabolism 

No data were located regarding metabolic pathways of CDDs in humans. However, there is some evidence 

that 2,3,7,8-TCDD is partially excreted in the feces in the form of metabolites (Wendling et al. 1990). 

Studies in animals indicate that 2,3,7,8-TCDD is metabolized slowly in mammals (Koshakji et al. 1984). 

Metabolic transformation by phase I metabolizing enzymes includes oxidation and reductive dechlorination, 

as well as oxygen bridge cleavage. This is followed by conjugation reactions catalyzed by phase II type 

enzymes, which facilitate excretion by adding more polar groups to the molecule. For example, a study in 

guinea pigs showed that only 28% of the radioactivity in the tissues 45 days following exposure to 

3

H-2,3,7,8-TCDD was in the form of metabolites (Olson 1986). Results from high performance liquid 

chromatography (HPLC) suggested the presence of at least five 3

H-labeled metabolites of 2,3,7,8-TCDD, 

but their structure was not established. The results indicated that in the guinea pig, the metabolites of 

2,3,7,8-TCDD may not leave the body rapidly. In rats and hamsters, metabolism appears to be required for 

urinary and biliary excretion (Olson et al. 1980a). Metabolites of 2,3,7,8-TCDD are not generally detected 

in tissues, suggesting that for most species, 2,3,7,8-TCDD is readily eliminated following metabolism. An 

in vitro study with isolated rat hepatocytes identified 1–hydroxy-2,3,7,8-TCDD and 8-hydroxy2,3,7-TrCDD as metabolites (Sawahata et al. 1982). 2-Hydroxy-1,3,7,8-TCDD was found to be the major 

metabolite of 2,3,7,8-TCDD in dogs but not in rats (Poiger et al. 1982). The metabolites from dogs 

administered to rats were eliminated as conjugates in the bile (Weber et al. 1982). Self induction of 

2,3,7,8-TCDD metabolism was reported in both species (Poiger and Schlatter 1985; Weber et al. 1982). A 

single 10 μg/kg dose of unlabeled 2,3,7,8-TCDD 9 days prior to administration of 3

H-2,3,7,8-TCDD 

resulted in a doubling of the amount of radioactivity eliminated in the bile of dogs. When the 

2,3,7,8-TCDD metabolites, 2-hydroxy-2,3,7-TrCDD and 2-hydroxy-1,3,7,8-TCDD, were synthesized and 

injected intraperitoneal into Wistar rats, no toxic effects were observed (Mason and Safe 1986). This 

supports the observation that the extract from the bile of 2,3,7,8-TCDD-treated dogs is about 100 times less 

toxic to rats and guinea pigs than pure 2,3,7,8-TCDD (Poiger et al. 1982). The lack of toxicity of the 

2,3,7,8-TCDD metabolites suggests that autoinduction of its own metabolism in animals is a detoxification 

mechanism. 

Data regarding other 2,3,7,8-substituted CDDs are limited. Wacker et al. (1986) found at least three 

phenolic radiolabeled metabolites of 14C-1,2,3,7,8-PeCDD in rat bile after treatment with glucuronidase and 

methylation, indicating the probability of formation of hydroxymetabolites. Results from studies in 
CDDs 198 

2. HEALTH EFFECTS 

rats revealed no metabolites of OCDD, as expected from the fully chlorinated molecule (Birnbaum and 

Couture 1988; Tulp and Hutzinger 1978). 

CDDs induce both phase I and phase II drug-metabolizing enzymes including AHH, EROD, ACOH, 

glucuronosyl transferase, glutathione S-transferase, and DT-diaphorase (Van den Berg et al. 1994). These 

enzymes are responsible for the metabolism of a variety of exogenous and endogenous substances. 

Pretreatment of C57BL/6J mice with 2,3,7,8-TCDD increased hepatic accumulation of a subsequent 

radiolabeled dose (total liver burden increased about 50%), whereas distribution to the kidney, fat, heart, 

lung, and gastrointestinal tract were reciprocally decreased (Curtis et al. 1990). The data indicated that an 

inducible, saturable system is involved in 2,3,7,8-TCDD toxicokinetics. The pretreatment, however, did not 

alter the hepatic metabolism of 2,3,7,8-TCDD in exposed mice. Similarly, the rate of metabolism of 

2,3,7,8-TCDD in hepatocytes from 2,3,7,8-TCDD-pretreated (induced) guinea pigs and mice was 

unchanged from that in untreated animals (Olson and Wroblewski 1985; Shen et al. 1989; Wroblewski and 

Olson 1985). In contrast, the rate of metabolism in hepatocytes from 2,3,7,8-TCDD-pretreated rats was 

3.2-fold greater than the rate in hepatocytes from control rats and about 9 times greater than in hepatocytes 

from 2,3,7,8-TCDD-pretreated guinea pigs. The difference between the 2,3,7,8-TCDD ability to induce its 

own rate of metabolism in rats and guinea pigs could be a factor in the difference between the susceptibility 

to 2,3,7,8-TCDD-induced toxicity in these two species, because the parent compound rather than 

metabolites is the toxic agent (Poland and Glover 1979). A generalized scheme of metabolic pathways for 

CDDs based on information from in vivo mammalian studies was proposed by Van den Berg et al. (1994) 

and is presented in Figure 2-3. 

2.3.4 Elimination and Excretion 

A median half-life of 7.1 years was estimated for 2,3,7,8-TCDD in a group of 36 Vietnam veterans (CDC 

1987; Pirkle et al. 1989). The calculation was based on the decrease of 2,3,7,8-TCDD serum levels that 

CDDs 200 

2. HEALTH EFFECTS 

were measured in these individuals in 1982 and again in 1987. The individual half-life values varied from 

2.9 to 26.9 years. In an expanded half-life study of 343 Vietnam veterans participating in Operation Ranch 

Hand, which included the subjects of the Pirkle et al. (1989) study, a half-life estimate of 8.7 years (95% CI 

of 8.0–9.5 years) was calculated (Michalek et al. 1996). The half-life estimate was calculated using 

2,3,7,8-TCDD levels in blood samples collected in 1982, 1987, and 1992. An earlier study of these subjects 

(Wolfe et al. 1994), which used data from two blood collection periods (1982 and 1987) estimated a halflife of 11.3 years (95% CI of 10–14.1 years). This half-life of 11.3 years was considered too high because it 

was based on restricted analysis of veterans with 2,3,7,8-TCDD levels above 10 ppt. By conditioning the 

data to lie above a line with slope equal to the negative of the decay rate, the analysis yielded a revised halflife of 8.7 years. The Michalek et al. (1996) half-life estimate of 8.7 years supersedes other estimates for 

this group of veterans because it includes an additional measurement of serum lipid 2,3,7,8-TCDD levels 

and controls for potential biases. 

Several other studies have calculated 2,3,7,8-TCDD half-lives. A mean half-life of 5.8 years was estimated 

from repeated samples from 29 BASF AG facility workers whose initial 2,3,7,8-TCDD serum lipid 

concentrations ranged from 29 to 553 ppt (Ott and Zober 1996). In a study of 48 German workers at a 

pesticide facility who were exposed to a mixture of CDDs/CDFs, a median half-life of 7.2 years was 

estimated for 2,3,7,8-TCDD (Flesch-Janys et al. 1996). Needham et al. (1994) estimated a half-life of 

8.2 years in 27 Seveso residents with initial serum 2,3,7,8-TCDD levels of 130 to 3,830 ppt. Using data 

from a human subject ingesting a single dose of 1.14 ng/kg 2,3,7,8-TCDD, Poiger and Schlatter (1986) 

calculated a half-life of 2,120 days (5.8 years). Geyer et al. (1986a) noted that they calculated a half-life of 

3.5-6.9 years, but did not describe the basis of this estimation. Overall, there is good agreement between 

the 2,3,7,8-TCDD half-lives estimated in 4 different populations (Vietnam veterans, BASF AG cohort, 

German pesticide workers, and Seveso residents); the half-lives ranged from 5.8 to 8.7 years (Flesch-Janys 

et al. 1996; Michalek et al. 1996; Needham et al. 1994; Ott and Zober 1996). Several studies have found 

correlations between percentage of body fat and 2,3,7,8-TCDD elimination half-times (Flesch-Janys et al. 

1996; Michalek et al. 1996; Ott and Zober 1996; Wolfe et al. 1994). Ott and Zober (1996) estimated halflives of 5.1 and 8.9 years in subjects with 20 and 30% body fat, respectively. 

There are limited data available on the elimination of other CDD congeners in humans. In the Flesch-Janys 

et al. (1996) study of 48 workers at a German pesticide facility, elimination half times were estimated for 

several CDD congeners. The estimated half-lives were 15.7 years for 1,2,3,7,8-PeCDD, 8.4 years for 

1,2,3,4,7,8-HxCDD, 13.1 years for 1,2,3,6,7,8-HxCDD, 4.9 years for 1,2,3,7,8,9-HxCDD, 3.7 years for 
CDDs 201 

2. HEALTH EFFECTS 

1,2,3,4,6,7,8-HpCDD, and 6.7 years for OCDD. In a study of six German workers with high CDD/CDF 

body burdens, elimination half-lives corrected for alterations in body weight ranged from 3.5 years for 

1,2,3,4,6,7,8-HpCDF to 7.9 years for 2,3,7,8-TCDD and 15 years for 1,2,3,4,7,8-HxCDD (Rohde et al. 

1997). In the same study, half-lives for elimination due only to fecal excretion ranged from 10 years for 

OCDD to 22 years for 2,3,7,8-TCDD and 27 years for 1,2,3,7,8-PeCDD. The half-lives for 2,3,4,7,8­

PeCDF in humans exposed to contaminated rice oil in the Yusho incident range from 2 to 30 years, and 

were inversely dependent on adipose tissue concentrations above approximately 10 ng/kg body weight (i.e., 

the higher the body burden, the faster the elimination) (Ryan et al. 1993a). 

Elimination of CDDs through lactation is discussed in Section 2.3.4.4. 

2.3.4.1 Inhalation Exposure 

In male Fischer 344 rats administered a single intratracheal dose of 0.32 μg labeled 2,3,7,8-TCDD/kg, feces 

was the major route of excretion over a 3-day period after dosing (Diliberto et al. 1996). The cumulative 

excretion of 26.3% of the administered dose was observed over 3 days following exposure. Approximately 

4% of the dose was excreted in the feces on day 3. The cumulative urinary excretion was only 1.3% of the 

administered dose. 

2.3.4.2 Oral Exposure 

The half-life for elimination of a single oral dose of 0.00114 μg/kg 3

H 2,3,7,8-TCDD in a human volunteer 

was calculated as 5.8 years (Poiger and Schlatter 1986). The excretion in feces was high during the first 

few days (up to day 6) probably because of elimination of unabsorbed material. During these first few 

days, about 12% of the administered dose was excreted. However, during days 7–125 only about 3.5% of 

the administered dose was eliminated. Urinary levels of radioactivity did not exceed the background levels. 

Studies in animals indicated that elimination of 2,3,7,8-TCDD is a relatively slow process. However, the 

results showed a great variability among species. The half-life for 2,3,7,8-TCDD elimination was 

14.95 days in Syrian hamsters (Olson et al. 1980b), 12 and 14 days in male and female Sprague-Dawley 

rats, respectively (Fries and Marrow 1975), 17 days in male Sprague-Dawley rats in another study (Piper et 

al. 1973), and 94 days in guinea pigs, the most sensitive species to the acute toxicity of 2,3,7,8-TCDD 

(Olson 1986). In contrast, the elimination half-life was 391 days in monkeys chronically exposed to low 
CDDs 202 

2. HEALTH EFFECTS 

doses of 2,3,7,8-TCDD in the feed (Bowman et al. 1989b). A similar half-life of about 1 year was observed 

in monkeys after a single-dose exposure (McNulty et al. 1982). In addition, studies of 2,3,7,8-TCDD halflife in highly exposed rats (Abraham et al. 1988), rhesus monkeys (McNulty et al. 1982) and marmoset 

monkeys showed that rates of excretion decreased with dose. 

The clearance of radioactivity after oral exposure to labeled 2,3,7,8-TCDD followed first-order kinetics in 

most studies. Fecal elimination was the major route, though excretion in the urine, expired air, and milk 

was also reported. 

When Sprague-Dawley rats were given radioactively labeled 2,3,7,8-TCDD, a total of 53% of the 

administered radioactivity was excreted by feces in the first 21 days (Piper et al. 1973). Elimination of 

2,3,7,8-TCDD-derived radioactivity in urine and expired air was 13 and 3% of the administered dose, 

respectively. Thirty-two percent of a single gavage dose of 0.32 μg of 2,3,7,8-TCDD/kg was eliminated in 

the feces of male Fischer 344 rats over a 3-day period after dosing (Diliberto et al. 1996). Only 1.4% of the 

administered dose was excreted in the urine over the same period. About 20–30% of the total oral 

2,3,7,8-TCDD dose was eliminated in the bile of cholecystectomized and cannulated dogs (Poiger et al. 

1982). In addition, excretion of unchanged 2,3,7,8-TCDD in milk was demonstrated in NMRI mice (Nau et 

al. 1986) and in monkeys (Bowman et al. 1989b), after oral exposure. 

Of the other congeners, several have been studied. An elimination half-life of 29.5 days was estimated for 

1,2,3,7,8-PCDD in Sprague-Dawley rats following a single oral exposure (Wacker et al. 1986). OCDD was 

more persistent in Fischer 344 rats with an estimated elimination half-life of 3–5 months following 10 daily 

oral doses (Birnbaum and Couture 1988). These congeners were excreted primarily in the feces following 

biliary elimination as metabolites (1,2,3,7,8-PCDD, at least three phenolic metabolites) or parent 

compound. A 13-week dosing study in which Sprague-Dawley rats were administered various mixtures of 

CDDs estimated liver half-lives of 14.5, 29.3, 45.6, and 100 days for 2,3,7,8-TCDD, 1,2,3,7,8-PCDD, 

1,2,3,4,7,8-HxCDD, and 1,2,3,4,6,7,8-HpCDD, respectively (Viluksela et al. 1998a). 

2.3.4.3 Dermal Exposure 

Within 120 hours after dermal administration of 0.32 μg/kg 2,3,7,8-TCDD to the clipped back skin of male 

Fischer 344 rats, 4% of the administered dose was excreted in the feces and <1% was excreted in the urine 

(Banks and Birnbaum 1991). The rate of 2,3,7,8-TCDD elimination significantly increased over time. 
CDDs 203 

2. HEALTH EFFECTS 

2.3.4.4 Transfer of CDDs Through the Placenta and Breast Milk 

CDDs are lipophilic compounds which can concentrate in maternal milk. Therefore, lactation provides an 

efficient mechanism for decreasing the body burden of these compounds (Schecter and Gasiewicz 1987a). 

CDD levels in breast milk samples from 193 German women ranged from 2.5 to 47 ng TEQ/kg milk fat 

(mean of 13 ng/kg) (Fürst et al. 1989b). More than 50% of the total CDDs detected in samples was 

represented by OCDD, which was detected at a mean concentration of 195 ng/kg milk fat (range, 

13–664 ng/kg). The amounts of other congeners in human milk decreased with decreasing chlorination; the 

mean concentration of 2,3,7,8-TCDD in milk fat was 2.9 ng/kg (range, <1–7.9 ng/kg). A more recent 

analysis of 526 individual milk samples from the German general population revealed a mean 

2,3,7,8-TCDD concentration of 3.2 ng/kg milk fat (Fürst et al. 1994). The analysis also showed the 

presence of only 2,3,7,8-chlorine-substituted CDD congeners. OCDD was the most concentrated congener 

with a mean level of 208 ng/kg milk fat. In general, the levels in milk decreased with decreasing degree of 

chlorination from octa- to tetra-CDD. Schecter and coworkers have published information on levels of 

dioxins in human breast milk from various countries (Schecter et al. 1989d, 1989e) (see also Section 2.6.1). 

In general, milk samples from industrial countries had higher CDD levels than those from less developed 

countries. Representative mean levels of CDDs in samples of human breast milk from various countries are 

presented in Table 2-6. 

Fürst et al. (1989b) also found that the levels of CDDs found in the milk of mothers breast-feeding their 

second child were about 20–30% lower than in those breast-feeding their first child. It was further noted 

that the highest excretion of CDDs was during the first few weeks after delivery. The sharpest decline was 

observed with OCDD; its excretion was reduced by half between the 1st and 5th week of lactation. In 

contrast, there was no significant decline in total HxCDDs in milk during the first year of lactation. The 

concentration of 1,2,3,4,6,7,8-HpCDD in milk fat showed a steady decline over the 1-year period, but its 

levels stayed relatively high. 2,3,7,8-TCDD represented the smallest portion of the total CDDs, and its 

levels in milk continuously declined over the year of lactation. Levels of 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 

1,2,3,4,7,8- and 1,2,3,6,7,8-HxCDD, 1,2,3,4,6,7,8-HpCDD, and OCDD were measured in a mother of twins 

prior to nursing and after 2 years of nursing (Schecter et al. 1996a). There was a 49.5% decrease in the total 

amount of CDDs in the lipid fraction of the breast milk. 2,3,7,8-TCDD had the largest percent decline in 

CDD levels, a decrease of 83.9%. A 52.4% decrease in maternal serum lipid levels of total CDD was also 

observed; the largest percent decline was an 86.8% decline in 1,2,3,7,8-PeCDD levels. 

CDDs 205 

2. HEALTH EFFECTS 

Several studies have shown that CDDs in breast milk are readily absorbed by nursing infants. In a 

19-week-old nursing infant, absorption was estimated as the difference between ingestion and the amount of 

CDDs found in the feces over a period of 12 days (McLachlan 1993). The mother was 32 years old and 

nursing for the first time. Several CDD congeners were determined in the milk: 2,3,7,8-TCDD, 1,2,3,7,8­

PeCDD, three hexacloro-substituted congeners, 1,2,3,4,6,7,8-HpCDD, and OCDD. The percentage of dose 

absorbed ranged from 90 to 95% except for the hepta-substituted congeners and OCDD which exhibited 

absorption rates of 61 and 23%, respectively. The percentage of the dose absorbed increased slightly if 

corrections were made for background levels in the diapers. Similar results were reported by Pluim et al. 

(1993b) who measured the amount of CDDs consumed via breast milk and excreted in the feces in 3 infants 

at the ages of 4, 8, and 12 weeks. Because of the high content of CDDs of the diapers relative to the feces, 

the percentage of dose absorbed was not determined. However, the results showed that, with the exception 

of OCDD, the bioavailability from breast milk was greater than 95%. At 4 weeks of age, the average 

cumulative intake of CDDs from breast milk was 132.1 pg TEQ per kg body weight. Of these, 37.4 

corresponded to 2,3,7,8-TCDD, 46.2 to 1,2,3,7,8-PCDD, and 24.4 to 1,2,3,6,7,8-HxCDD. With the 

inclusion of CDFs, the total TEQ at 4 weeks was approximately 257 pg/kg body weight. Exposure to 

CDDs and CDFs from lactation decreased at 8 and 12 weeks mainly due to a decrease in their concentration 

in whole breast milk which resulted from a reduced fat content of the milk (the depletion of body burden of 

the mother while nursing may have also contributed). Abraham et al. (1994, 1996) and Dahl et al. (1995) 

also reported almost complete absorption of lower chlorinated CDDs and CDFs in breast-fed infants during 

the first year of life. It was also noticed that intake of CDDs and CDFs was up to 50 times higher in breastfed infants compared with a formula-fed infant (Abraham et al. 1996). The latter study further showed that 

despite much lower intake of CDDs and CDFs after weaning, the concentration of these compounds in stool 

fat did not decrease substantially, suggesting that concentration in fecal fat more or less reflect that in body 

fat. Also, at 11 months of age, TEQ concentrations in blood from formula-fed infants were less than 25% 

of maternal values and about 10 times lower than in infants breast-fed for 6-7 months (Abraham et al. 

1996). 

Schecter et al. (1996b) recently presented data on the levels of CDDs and CDFs in human fetuses 

(8–14 weeks gestational age with placenta removed) and in placentas from women from the general 

population who had normal deliveries. On a lipid basis, the total TEQs (CDDs plus CDFs) in a pool of 14 

placentas was 10.1 ng/kg; half this amount (5.3 ng/kg) was measured in a pool of 10 fetuses. In an 

analysis of 43 samples of human milk, Schecter et al. (1996b) found that the total concentration of CDDs 

and CDFs was 16.7 ng/kg (expressed as TEQ). The authors also calculated that the TEQ body burden for 
CDDs 206 

2. HEALTH EFFECTS 

the pooled fetal tissue was 0.034 ng/kg body weight; for pooled placentas, they calculated a total TEQ of 

0.086 ng/kg wet weight. These results suggest that the transfer of CDDs to the fetus may be somewhat 

limited. 

The influence of maternal transfer (placental and via breast milk) of CDDs/CDFs on the body burden of 

newborns and infants was further investigated by Kreuzer et al. (1997). These investigators also 

developed a pharmacokinetic model for 2,3,7,8-TCDD that allowed them to simulate body and tissue 

burden for the entire human lifetime as a function of 2,3,7,8-TCDD uptake from contaminated nutrition. 

On a lipid basis, the concentrations of 2,3,7,8-TCDD in adipose tissue and liver of breast-fed infants who 

died of sudden infant death syndrome were 0.4–4 ppt and 0.5–4 ppt, respectively. The corresponding 

values in nonbreast-fed infants were 0.2–0.8 ppt and 0.3–0.7 ppt. Similar values were detected in adipose 

tissue and livers of three stillborns, confirming the placental transfer of these chemicals to the fetus. The 

model developed by Kreuzer et al. (1997) reflected sex- and age-dependent changes in body weight, 

volumes of liver, adipose and muscle tissue, food consumption, and excretion of feces and was used to 

predict the half-life of elimination of 2,3,7,8-TCDD and its concentrations in adipose tissue, blood, liver, 

and feces at different ages. Also, the influence of breast-feeding on the 2,3,7,8-TCDD burden of the 

mother, her milk, and her child was simulated. The authors used their own data, as well as those from 

others, to validate the model. For nonbreast-fed infants, the model predicted a decrease in the 

concentration of 2,3,7,8,-TCDD in lipids during the first year, and this was supported by the empirical 

data. For infants exclusively breast-fed, the model predicted an increase in 2,3,7,8-TCDD burden 

followed by a decrease after weaning, and this was also confirmed by the measured data. Model 

validation of 2,3,7,8-TCDD concentrations in liver for the 20 infants investigated and in adipose tissue, 

blood, and feces for data in infants published by others showed good agreement between the simulated 

and experimental values. Since one of the model’s assumption was that the concentration of 

2,3,7,8-TCDD in fecal lipids reflected the concentration in lipids of the organism, the good correlation 

between predicted and empirical data validated the assumption. Under the assumption that the 

2,3,7,8-TCDD concentration in lipids of breast milk equals the concentration in the maternal organism, 

the model predicted a value of 2.23 ng 2,3,7,8-TCDD/kg lipids for the beginning of the nursing period. 

The model further predicted that the concentration of 2,3,7,8-TCDD in milk decreases with duration of 

breast-feeding, such that after 6 months of daily nursing the concentration in milk and maternal body 

lipids is approximately 70% of the value at the time of delivery. These predictions were in good 

agreement with published values. Lastly, the investigators modeled the concentration of 2,3,7,8-TCDD in 

lipids or blood of a male subject for a time span of 60 years and compared it with literature values for 

German subjects. One of two curves constructed was computed assuming breast-feeding for the first 
CDDs 207 

2. HEALTH EFFECTS 

6 months of life followed by formula up to 1 year and the other considering feeding only formula for the 

same period of time. In both cases further nutrition was simulated to consist of the common diet. The 

predicted curves differed considerably during the first years of life. For the nonbreast-fed case, 

2,3,7,8-TCDD concentrations decreased during the first year and subsequently increased, reaching a 

maximum at 16 years. For the breast-fed case, the simulation yielded a rapid rise of 2,3,7,8-TCDD in 

lipids followed by a 3-year decrease after weaning and merging at about 7 years with the concentrations 

of nonbreast-fed individuals. Subsequently, 2,3,7,8-TCDD concentrations leveled at between 2 and 3 ng 

2,3,7,8-TCDD/kg body lipids until the end of life. The latter value was in agreement with average 

background levels for the German population. The half-life of nonmetabolic elimination (unchanged 

2,3,7,8-TCDD) was calculated to be 0.42 years in newborns and 9.5 years in 40-year-old adults. The 

half-life of the fraction metabolized by the liver ranged from 1.5 years for newborns to approximately 

10 years for a 40-year-old individual. The three times greater elimination half-life for the metabolized 

fraction relative to the nonmetabolized fraction in infants suggests that metabolic elimination does not 

play a major role in the elimination of 2,3,7,8-TCDD in infants. A key finding from the Kreuzer et al. 

(1997) study is the model prediction that the increased 2,3,7,8-TCDD burden observed as a result of 

breast-feeding does not lead to a raised lifetime value. 

In rodents, placental transfer of CDDs to the fetus is relatively limited, but transfer during sensitive 

periods of organogenesis is biologically important as evidenced by effects on fetuses or offspring 

exposed in utero. Excretion into milk represents a major pathway for maternal elimination of CDDs 

and, therefore, for exposure to offspring. In C57BL/6N mice administered a single oral dose of 30 μg 

14C-2,3,7,8-TCDD/kg on Gd 11 the levels of 2,3,7,8-TCDD-derived radioactivity in the embryos on 

Gd 12, 13, or 14 were below 0.5% of the total 2,3,7,8-TCDD dose (Weber and Birnbaum 1985). In 

the dams, the highest concentration of radioactivity was in the liver (50–67% of total dose), whereas 

embryos had a relatively higher concentration of radioactivity in the heads than in the rest of the 

body. Approximately 0.03% of the administered dose was delivered to each embryo. In a different 

study in NMRI mice, pregnant females were administered a single dose of 25 μg 14C-2,3,7,8-TCDD 

(oral, intraperitoneal, or subcutaneous) on Gd 16 and the distribution of radioactivity was examined in 

the pups on postnatal days 7–36 (Nau et al. 1986). At all times, the highest concentration of 

radioactivity in the pups (per gram of tissue) was found in the liver; extrahepatic tissues such as 

intestines and skin had a concentration of radioactivity that was approximately one order of 

magnitude lower than the liver. During the first postnatal week, 2,3,7,8-TCDD concentrations 

increased considerably in the pups. It was also found that during the first two weeks the pups 

received doses of 2,3,7,8-TCDD through milk which were, on a body weight basis, similar to those 
CDDs 208 

2. HEALTH EFFECTS 

which had been administered to their mothers prior to birth. In pups raised by untreated foster 

mothers, 2,3,7,8-TCDD tissue concentrations decreased rapidly due to organ growth with 

concomitant dilution of 2,3,7,8-TCDD. Abbott et al. (1996) examined the distribution of 

2,3,7,8-TCDD in embryonic tissues of mice at times earlier than previous studies. Pregnant mice 

were treated with 2,3,7,8-TCDD on Gd 12 and embryonic tissues were examined at various times 

from 0.5 to 24 hours after dosing. The rate of accumulation of 2,3,7,8-TCDD reached a maximum in 

placental tissue in about 3 hours and, following a slight decline, remained relatively constant between 

8 and 24 hours. After 24 hours, 0.27% of the maternal dose was detected in the placenta. In 

embryonic liver, 2,3,7,8-TCDD peaked approximately 8 hours after dosing and decreased thereafter, 

as opposed to maternal liver, where it remained constant after achieving an apparent maximum. The 

relative decrease in the rate of concentration in the embryonic liver was attributed to a rapid growth of 

the tissue during that time period. Distribution of 2,3,7,8-TCDD to embryonic palates followed a 

pattern similar to that in embryonic liver. Twenty-four hours after dosing, the secondary palates had 

0.0045% of the administered maternal dose. 

Van den Berg et al. (1987b) examined the transfer of CDDs and CDFs through the placenta and via the 

milk in Wistar rats. Prenatal exposure of the fetus was studied by administering a diet containing a fly 

ash extract from a municipal incinerator to rats from day 8 until 17 of pregnancy, after which time the rats 

were sacrificed. Postnatal transfer was assessed in rats fed the same diet during the first 10 days after 

delivery while nursing their offspring. Of the 49 tetra- to octa-CDDs, only 7 CDD congeners were 

detected and all had a 2,3,7,8-chlorine-substitution pattern. In the fetus, 2,3,7,8-TCDD had the highest 

retention (0.13% of total dose, 0.0092% of the dose/g). Retention decreased with the number of chlorine 

atoms; HpCDDs and OCDD were not detected. In the liver of offspring, 2,3,7,8-TCDD, 1,2,3,7,8-PCDD, 

and the three 2,3,7,8-substituted HxCDDs had the highest retention (5.3–8.1% of total dose, 0.74–1.13% 

of dose/g). The 2,3,7,8-penta- and hexa-substituted congeners had the highest retention in the livers of 

pregnant and lactating rats (53.9–80.2% of total dose, 2.9–5.2% of dose/g). No significant differences 

were found in liver retention of tetra- to octa-chlorinated congeners between pregnant and lactating rats, 

but lactating females stored less CDDs in their adipose tissue. Similar results were reported by Li et al. 

(1995c) in Sprague-Dawley rats. These authors administered a single intravenous dose of 5.6 μg 14C2,3,7,8-TCDD/kg to pregnant rats on Gd 18. Sacrifices were conducted on Gd 19 and 20, and 

postnatal days 1 and 5. Groups of neonates were also cross-fostered between treated and nontreated dams 

to differentially assess transfer of 2,3,7,8-TCDD through the placenta and through nursing. Only about 

0.01% of the dose administered to the dams was found in whole livers of fetuses one and two days after 

dosing (0.04 and 0.07% of dose/g fetal liver), indicating limited placental transfer. In contrast, the 
CDDs 209 

2. HEALTH EFFECTS 

concentration of 2,3,7,8-TCDD in the liver of neonates after 1 day of lactation was 0.65% of the 

administered dose/g liver, and this increased to 2.88% after 4 days of nursing. Four days after nursing, 

the liver concentration of 2,3,7,8-TCDD in neonates from dams dosed 1 day after parturition was 4.1% of 

the administered dose/g of liver, and this was higher than in the dam’s liver (3.32%). As in earlier 

studies, the results from the cross-fostering experiments confirmed that nursing is a major pathway for 

transfer of 2,3,7,8-TCDD to the offspring. 

The transfer of CDDs and CDFs via placenta and through milk was also investigated in a marmoset 

monkey administered a defined mixture of CDDs and CDFs subcutaneously 11 weeks prior to delivery 

(Hagenmaier et al. 1990). Concentrations of CDDs and CDFs were measured in a newborn 1 day after 

birth and in an infant of the same litter after a period of 33 days of lactation. The highest deposition in 

newborn liver was observed for 2,3,7,8-TCDD and 1,2,3,7,8-PCDD (54 and 51 pg/g wet weight, 

respectively) and corresponded to about 0.15% of the administered dose/g tissue. The concentration of all 

other congeners was <10% of the corresponding concentrations in adults. In contrast to liver, the 

concentrations of 2,3,7,8-substituted CDDs in newborn adipose tissue were at least one third the levels in 

adults, and for OCDD, the concentration in adipose tissue was three times higher that in adult adipose 

tissue. Transfer of CDDs through milk was considerable, though selective. The concentration of 

2,3,7,8-TCDD and 1,2,3,7,8-PCDD in the infant’s liver was 395 and 611 pg/g wet tissue, respectively; the 

corresponding concentrations in the mother’s liver were 107 and 326 pg/g. However, the concentration of 

OCDD in infant’s liver was less than 10% that of the mother’s liver. Bowman et al. (1989b) examined 

the transfer of 2,3,7,8-TCDD from mother to offspring in rhesus monkeys. Female monkeys had been 

exposed to 2,3,7,8-TCDD for about 4 years to a diet (5 or 25 ppt) that provided an estimated 

0.0001–0.0006 μg 2,3,7,8-TCDD/kg/day before breeding. Breeding started 10 months after exposure 

ceased. At weaning (4 months), the offspring had a concentration of 2,3,7,8-TCDD in mesenteric fat 4.3 

times higher than in subcutaneous fat from their respective mothers. Bowman et al. (1989b) estimated 

that the mothers excreted between 17 and 44% of their 2,3,7,8-TCDD burden by lactation. Based on 

measurements of 2,3,7,8-TCDD in fat at 4, 12, and 24 months of age, it was found that in the young 

monkeys the decline in 2,3,7,8-TCDD in fat followed first-order, single-compartment kinetics with a halflife of approximately 181 days (Bowman et al. 1990). For the purpose of comparison, the mean half-life 

in 7 adult female rhesus monkeys was 391 days with standard error of 88 days (Bowman et al. 1989b). 

In summary, CDDs can be transferred to the fetus across the placenta and, although the amounts may be 

relatively small, the transfer may have great biological significance if it occurs during critical periods of 
CDDs 210 

2. HEALTH EFFECTS 

organogenesis. Due to their lipophilicity, CDDs can concentrate in human breast milk and can be 

transferred to infants through nursing. In general, the amount of individual congeners in breast milk 

decreases as chlorination decreases. Excretion via milk is highest during the first weeks after delivery. 

Also, the concentration of CDDs in milk is higher in mothers breast-feeding their first child than in those 

breast-feeding their second child. CDDs transferred to infants through nursing are readily absorbed by 

the infants. A pharmacokinetic model predicted that the increased body burden in infants that results 

from breast-feeding does not translate into raised lifetime body burden. Studies in animals have also 

shown transfer of CDDs across the placenta and via mother’s milk. 

2.3.5 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 

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 substance­
CDDs 211 

2. HEALTH EFFECTS 

specific 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 sites) 

based on the results of studies where doses were higher or were administered in different species. 

Figure 2-4 shows a conceptualized representation of a PBPK model. 

If PBPK models for CDDs exist, the overall results and individual models are discussed in this section in 

terms of their use in risk assessment, tissue dosimetry, and dose, route, and species extrapolations. 

PBPK models for 2,3,7,8-TCDD are discussed below. The pharmacokinetic behavior of 2,3,7,8-TCDD, 

especially distribution, has been shown to be dose-dependent and involves protein binding and enzyme 

induction in hepatic tissue. Thus, terms describing these interactions have been included in the animal 

models described below. Furthermore, since induction of these dioxin-binding proteins is a process 

mediated by the interaction of a dioxin-receptor (the Ah receptor) complex with specific binding sites on 

DNA additional terms were included in the models. For a detailed explanation regarding the Ah receptor 

and its involvement in the mechanism of action of 2,3,7,8-TCDD and structurally related halogenated 

aromatic hydrocarbons, see Section 2.4.2. 

CDDs 213 

2. HEALTH EFFECTS 

2.3.5.1 Summary of PBPK Models. 

The elimination of 2,3,7,8-TCDD from humans was evaluated using a PBPK model developed by Kissel 

and Robarge (1988). The steady-state adipose tissue concentration predicted by the model was similar to 

the lipid-based blood levels reported in the general population with no known special exposure to 

2,3,7,8-TCDD. The model was also used to predict elimination of 2,3,7,8-TCDD from Ranch Hand 

Vietnam veterans. The predicted half-lives were similar to an experimental value based on analysis of 

2,3,7,8-TCDD in blood of veterans with adipose burdens >10 ppt. The apparent half-lives increased as the 

adipose tissue concentration approached the steady-state level associated with background exposure. The 

model also predicted reasonably well the elimination of 2,3,7,8-TCDD from a volunteer who ingested a 

single 2,3,7,8-TCDD dose. 

Leung et al. (1988) developed a five compartment PBPK model to describe the time course of 

2,3,7,8-TCDD distribution in tissues of both the Ah-responsive C57BL/6J and Ah-less responsive DBA/2J 

mice (C57BL/6J mice respond to 2,3,7,8-TCDD with an increase in AHH activity, at a dose less than 

required to elicit this response in DBA/2J mice). The model also included binding in blood and two hepatic 

sites, one in the cytosol and the other in microsomes. It was found that the greater accumulation of 

2,3,7,8-TCDD in the liver of C57BL/6J mice, relative to DBA/2J mice, was not attributed to the greater fat 

content in the DBA/2J mice, but to the more avid microsomal binding (CYP1A2) in the liver of the 

C57BL/6J mice. In the concentration range covered in the model simulations, the cytosolic receptor (Ah 

receptor) did not seem to play a major role in determining the overall tissue distribution pattern. 

The same group of investigators (Leung et al. 1990b) developed a PBPK model to describe the tissue 

disposition of 2,3,7,8-TCDD in Sprague-Dawley rats. The description included the same compartments 

used in modeling the behavior of 2,3,7,8-TCDD in mice. The ratio of liver to fat concentration of 

2,3,7,8-TCDD was found to be primarily determined by the dissociation constant of the microsomal 

binding protein (CYP1A2) and the basal and induced concentration of this protein in the liver. In general, 

there was agreement between the simulated data and experimental data from a single-dose study and a 7­

and 13-week repeated-dosing study. However, the model underpredicted the concentration of 

2,3,7,8-TCDD in the fat at low dose and overestimated the concentration at high dose for a 2-year feeding 

study. Induction of microsomal binding protein was necessary to account for the differences in disposition 

at low and high daily doses. Further refinements of Leung et al. (1988, 1990b) models were conducted by 
CDDs 214 

2. HEALTH EFFECTS 

Wang et al. (1997) and Santostefano et al. (1998) and included analyses of distribution and responses to 

2,3,7,8-TCDD exposure at early times and in multiple tissues. 

A receptor-mediated PBPK model was developed by Andersen et al. (1993). The model included 

interactions of the Ah-TCDD complex with DNA and was used to examine the tissue disposition of 

2,3,7,8-TCDD and the induction of the dioxin-binding protein (presumably CYP1A2) and CYP1A1. It 

was found that tumor promotion correlated more closely with predicted induction of CYP1A1 than with 

induction of the hepatic binding proteins (CYP1A2, AhR). More recently, Andersen et al. (1997a, 1997b) 

developed a multicompartment geometric model of the liver that provided a better prediction of both total 

and regional induction of CYP450 proteins within the liver than conventional one-compartment models. 

A mechanistic model (known as the NIEHS model) was constructed to describe 2,3,7,8-TCDD-mediated 

alterations in hepatic proteins in the rat (Kohn et al. 1993). This model included the tissue distribution of 

2,3,7,8-TCDD and its effects on concentrations of CYP1A2 and CYP1A1 and the effects of 2,3,7,8-TCDD 

on the Ah, estrogen, and epidermal growth factor (EGF) receptors over a wide range of 2,3,7,8-TCDD 

doses. The model predictions were compared to experimental data from 2,3,7,8-TCDD promotion studies. 

The biochemical response curves for the proteins examined were hyperbolic, indicating a proportional 

relationship between target-tissue dose and protein concentration at low 2,3,7,8-TCDD doses. Also, the 

model successfully reproduced the observed tissue distribution of 2,3,7,8-TCDD, the concentration of 

CYP1A2 and CYP1A1, and the effects of 2,3,7,8-TCDD on the Ah, estrogen, and EGF receptors over a 

wide dose range. 

Carrier et al. (1995a) developed a model that describes the distribution kinetics of 2,3,7,8-TCDD and 

related chemicals (with chlorine substitutions in positions 2,3,7, and 8) in various mammalian species, 

including humans. Their model takes into account cellular diffusion, binding of the chemicals with the Ah 

receptor and with proteins, and enzyme induction in the liver. The model was used to describe the 

distribution of CDDs between liver and adipose tissue as a function of overall body concentration. Model 

simulations showed that the fractions of the body burden contained in the liver and adipose tissue vary 

nonlinearly as a function of the overall body concentration; this was in agreement with published data in 

rodents, monkeys and humans. The authors further modeled the disposition kinetics of CDDs in liver, 

adipose tissue, and whole body as a function of time (Carrier et al. 1995b). The results showed that the 

rate of change in CDD tissue concentrations varies as a function of total body burden such that whole body 
CDDs 215 

2. HEALTH EFFECTS 

elimination rate decreases as body burden decreases, suggesting nonlinear disposition kinetics. This was 

also in agreement with published data on absorption and elimination kinetics of CDDs in rats and humans. 

2.3.5.2 Comparison of PBPK Models for 2,3,7,8-TCDD. 

Several models that describe the disposition of 2,3,7,8-TCDD in animals and one in humans were 

identified from the literature. In the Leung et al. (1988, 1990b) models in mice and rats, tissue 

2,3,7,8-TCDD concentration ratios, particularly liver concentrations, were related both to intrinsic 

partitioning and to the presence of specific cytosolic and microsomal 2,3,7,8-TCDD-binding proteins. 

This is in contrast to other PBPK models developed for similar, persistent lipophilic chemicals, which 

distributed the chemicals between organs based on experimentally observed concentrations ratios under 

various dosing conditions. This could explain why two organs having the same partition coefficients 

contain very different 2,3,7,8-TCDD concentrations under a particular experimental condition. Carrier et 

al. (1995a, 1995b) developed a similar model, which also included other 2,3,7,8-substituted dioxins and 

furans and simulated experimental data on rodents, monkeys, and humans. The Andersen et al. (1993) 

model extended the Leung et al. (1988, 1990b) models by including induction of binding proteins/enzymes 

and of 2,3,7,8-TCDD metabolism in response to ternary interactions of 2,3,7,8-TCDD, the Ah receptor, 

and DNA binding sites and correlated various tissue dose measures with the promotional efficacy of 

2,3,7,8-TCDD. The five-compartment liver model described by Andersen et al. (1997a, 1997b) provided a 

better description of mRNA production and regional localization of induced proteins, consistent with 

immunohistochemical information, than conventional one-compartment models. The model constructed 

by Kohn et al. (1993) suggested possible biochemical mechanisms which could explain a complex 

response to exposure to 2,3,7,8-TCDD such as cell proliferation in female rats. This model not only 

included enzyme induction and the Ah receptor, but also the estrogen and EGF receptors, all of which 

seem to be involved in a complex sequence of events that lead to cell proliferation as a result of 

2,3,7,8-TCDD exposure. In contrast with the models developed for animals and described above, the 

Kissel and Robarge (1988) fugacity-based model for humans was used to predict vehicle-dependent uptake 

and elimination of 2,3,7,8-TCDD without including any Ah receptor-related terms. (Fugacity is defined as 

the “escaping tendency” of a substance in a phase.) Pharmacokinetic parameters used in the various 

models are listed in Table 2-7. 

 

 

 

CDDs 217 

2. HEALTH EFFECTS 

2.3.5.3 TCDD Models 

The Kissel and Robarge Model 

Description of the Model. The elimination of 2,3,7,8-TCDD from humans was described with a 

fugacity-based model using physiologically based parameters (Kissel and Robarge 1988). In this model, 

transport of 2,3,7,8-TCDD was assumed to be perfusion-limited (flow-limited) and 2,3,7,8-TCDD was 

assumed to be uniformly distributed within each tissue group or fluid phase, and tissue levels were 

considered to be in equilibrium with exiting fluids (blood, urine, bile). Because 2,3,7,8-TCDD appears to 

be poorly metabolized in humans, the model did not include terms for metabolites. Transport between gut 

lumen and gut tissue was described as a diffusive process. Included in the differential equations used to 

solve the system were data for several diets. Body compartment sizes and densities used in the simulations 

of background exposure and of elimination from individuals with body burdens similar to those of Ranch 

Hand veterans were based on reference-man data. Tissue perfusion rates and partition coefficients were 

obtained from the literature. The diet used in all simulations was adapted from the literature and also 

included a typical intake of added fats and oils. The fugacity capacity of the various diet components, 

gastric secretions, and fecal materials were either calculated or obtained from the literature. The model 

was used to predict tissue levels resulting from background exposures, elimination of 2,3,7,8-TCDD from 

Ranch Hand veterans, and elimination of 2,3,7,8-TCDD from a human volunteer. 

Validation and Discussion. The steady-state adipose tissue concentrations predicted by the model, 

assuming no metabolism and a daily background exposure of 50 pg/day in North America, was 7.7 ppt. 

This value was similar to the lipid-based blood tissue levels reported in the general population with no 

known unusual exposure. The body burden projected for an intake of 100 pg/day fell outside the typical 

range associated with background sources. In simulating the elimination of 2,3,7,8-TCDD from Ranch 

Hand veterans the model assumed a background exposure of 50 pg/day and no metabolism. Under these 

conditions, apparent half-lives of 4.4, 5.2, 5.9, 7.2, 9.1, and 20 years were estimated for individuals with 

2,3,7,8-TCDD adipose tissue concentrations of 100, 50, 30, 20, 15, and 10 ppt, respectively. This was in 

good agreement with a half-life of 7.1 years determined by analysis of blood lipids of veterans with adipose 

burdens >10 ppt (Pirkle et al. 1989). The results showed that the apparent half-lives increased greatly as 

tissue concentrations approached the steady-state level associated with background exposure. The model 

also approximated the uptake efficiency and elimination of 2,3,7,8-TCDD from a volunteer as reported by 

Poiger and Schlatter (1986). The fact that the predicted uptake efficiency was similar to that found 
 

 

 

CDDs 218 

2. HEALTH EFFECTS 

experimentally indicated that the estimated gut-lumen/gut-tissue mass transfer coefficient used was in the 

appropriate range. The reported half-life was 5.8 years and the model estimated a value of 6.7 years. 

Overall, the result suggested that a fugacity-based model can provide a viable method for describing overall 

elimination of 2,3,7,8-TCDD from humans, but it does not provide much insight regarding why elimination 

occurs in a particular manner. 

The Leung et al. Model in Mice 

Description of the Model. The model described by Leung et al. (1988) in mice provides quantitative 

descriptions of the time-course of elimination and levels of 2,3,7,8-TCDD in various organs of C57BL/6J 

mice and DBA/2J mice, a less responsive strain with higher body-fat content. The model contains five 

compartments: blood, liver, fat, richly perfused tissues, and slowly perfused tissues. To account for the 

2,3,7,8-TCDD binding to receptor in the liver, the model contained two hepatic binding sites, one 

corresponding to the high affinity/low capacity cytosolic Ah receptor and the other to the inducible, low 

affinity/high capacity microsomal protein (CYP1A2). To simulate the intraperitoneal dose route used by 

Gasiewicz et al. (1983a), 2,3,7,8-TCDD was assumed to be absorbed into the liver compartment by a firstorder uptake process. Bioavailability was assumed to be 100%. Partition coefficients, physiological 

parameters, and biochemical constants were obtained or calculated from the literature for each mouse strain. 

The kidney was assumed to be representative of the richly perfused tissue, whereas the slowly perfused 

tissue consisted mainly of muscle and skin. The binding capacity of the Ah-less responsive DBA/2J mice 

was set to equal that of the Ah-responsive mice even though the binding affinity is extremely low. Blood 

binding was described as a linear process with an effective equilibrium between bound and free 

2,3,7,8-TCDD given by a constant. In blood, only one form of 2,3,7,8-TCDD is exchangeable in the 

tissues, which gives rise to kinetic behavior observed for diffusion-limited uptake into tissues. 

Validation and Discussion. The simulation of the time-course of 2,3,7,8-TCDD concentration in the 

liver and fat of C57BL/6J mice after a single 10 μg/kg intraperitoneal injection generated by the model 

was in good agreement with the empirical data of Gasiewicz et al. (1983a). In trying to simulate the 

3-times-higher liver/fat ratio of 2,3,7,8-TCDD in the C57BL/6L mice than in the DBA/2J mice, Leung et 

al. (1988) varied the fat content parameter in the C57BL/6J mice from 3 to 12% of body weight. The 

rationale was that the difference in hepatic concentration may have been due to greater capacity of the 

DBA/2J mouse to sequester the highly lipophilic 2,3,7,8-TCDD in adipose tissue. However, the results 

showed that 2,3,7,8-TCDD concentration in the liver was relatively insensitive to body fat content, 
 

 

CDDs 219 

2. HEALTH EFFECTS 

indicating that this was not an important factor influencing the disposition of 2,3,7,8-TCDD in the liver 

between the two strains of mice. The authors also found that the distribution of 2,3,7,8-TCDD was 

strongly influenced by the binding characteristics of the microsomal binding protein, especially the 

binding constant. The model gave good simulations of 2,3,7,8-TCDD excretion in both strains of mice. 

The simulation of the time-course of 2,3,7,8-TCDD concentration in the liver and fat of DBA/2J mice after 

a single 10 μg/kg intraperitoneal injection was not as good as that for the C57BL/6J mouse if the input was 

set to be consistent with the uptake and elimination. As with the C57BL/6J mouse, disposition of 

2,3,7,8-TCDD in the liver of DBA/2J mice was greatly influenced by the microsomal protein binding 

constant and rather insensitive to changes in body fat content. The best fit of the empirical data was 

obtained with a binding constant of 75 nM (20 nM for the C57BL/6J mice), indicating that the 2,3,7,8­

binding affinity to the hepatic microsomal protein in the DBA/2J mice was at least 3.5 times lower than 

that of the C57BL/6J mice. 

The Leung et al. Model in Rats 

Description of the model. This model in the Sprague-Dawley rat (Leung et al. 1990b) is an 

extension of the mouse model previously described and contains the same five compartments and two 

types of binding proteins: one corresponding to the high-affinity, low-capacity cytosolic 2,3,7,8-TCDD 

(Ah) receptor, and the other to the inducible, lower-affinity, high-capacity microsomal protein (CYP1A2). 

In the rat model, both types of binding proteins are defined with their own binding capacities and 

dissociation constants. The model was used to analyze experimental data for the single-dose studies of 

McConnell et al. (1984) and Rose et al. (1976), the 7-week Rose et al. (1976) study, the 13-week multipledose study of Kociba et al. (1978b), and the 2-year feeding study of Kociba et al. (1978a). In simulating 

the single-dose gavage study, 2,3,7,8-TCDD was assumed to be absorbed from the gastrointestinal tract by 

a first-order uptake process with a rate constant of 0.2/hour. In simulating the multidosing studies, 

bioavailability was assumed to be 100%. Physiological parameters, partition coefficients, and biochemical 

constants were calculated or obtained from the literature. Since there was no literature value for the 

binding capacity of the microsomal 2,3,7,8-TCDD-binding site in the rat, the value used was approximated 

by assuming it to be 10 times that of the mouse. The total microsomal binding capacity was apportioned 

between a basal level and an induced level. Also, AHH activity was taken to be the sum of a basal and 

induced level. A first-order metabolic rate constant for 2,3,7,8-TCDD metabolism in the liver was 

adjusted to provide a biological half-life of about 25–30 days. 
 

CDDs 220 

2. HEALTH EFFECTS 

Validation and Discussion. When the simulation of the McConnell et al. (1984) data for AHH 

induction included a term for induction of microsomal binding protein there was good agreement between 

the simulation and the empirical data. This had not been the case in an initial fitting which included a 

constant concentration of microsomal binding protein. Rose et al. (1976) examined the accumulation of 

2,3,7,8-TCDD in adipose and liver tissues in rats administered 0.01, 0.1 and 1 μg 2,3,7,8-TCDD/kg/day 

5 days a week for 7 weeks; sampling was done at weeks 1, 3, and 7. Model predictions of 2,3,7,8-TCDD 

concentrations were in good agreement with the experimental data except for concentration in fat at the 

0.01 μg/kg/day dose level, in which case the model overpredicted the tissue concentration. Model 

formulations that had constant microsomal binding capacity overpredicted liver 2,3,7,8-TCDD 

concentrations at the lower-dose rates. Also, model formulations that contained final amounts of 

microsomal binding protein (CYP1A2) very different (much higher or lower) from the basal 200 

nmol/liver could not simulate 2,3,7,8-TCDD concentration in liver at the highest-dose rate. Similar to the 

findings in mice, the liver/fat concentration ratio in rats was extremely sensitive to the dissociation 

constant of the microsomal binding protein. The model simulated well the data from the 7- and 13-week 

studies (Rose et al. 1976; Kociba 1978b), but not as well for data from the 2-year feeding study (Kociba et 

al. 1978a). There was underprediction of 2,3,7,8-TCDD concentration in fat and liver at the low dose 

(0.001 μg/kg/day) and overprediction of the liver concentration at the high-dose level (0.1 μg/kg/day). 

However, the ratios of the concentrations were consistent with those observed experimentally (1/1 at low 

doses, much higher in liver at high doses). According to Leung et al. (1990b), the underprediction at low 

dose may reflect the fact that the low-dose fat concentration in the 2-year study was close to the limit of 

detection and thus, subject to more error. At the high dose, physiological parameters such as tissue 

volume, metabolic constants, and amounts of binding proteins may have been altered by weight loss and 

changes in body composition, known effects of chronic exposure to 2,3,7,8-TCDD. Leung et al. (1990b) 

indicated that the overprediction at high dose could have been due to a loss of microsomal 2,3,7,8-TCDDbinding sites in the chronically exposed rats. The affinity of 2,3,7,8-TCDD for the microsomal binding 

protein appeared to be greater in the Sprague-Dawley rats than in C57BL/6J mice, which could account for 

the higher liver/fat concentration ratio in rats than in mice, assuming that the partitioning between tissues 

is approximately the same in the two species. 

Wang et al. (1997) extended the work of Leung et al. (1988, 1990b) and Andersen et al. (1993) and 

developed an improved model to describe the disposition of 2,3,7,8-TCDD in multiple tissues from female 

Sprague-Dawley rats. The model of Wang et al. (1997) improved previous modeling attempts in some 

specific areas such as 1) providing information on distribution of 2,3,7,8-TCDD at early time points 
 

 

CDDs 221 

2. HEALTH EFFECTS 

(important for determining unique parameters related to mass transfer such as permeability), 2) better 

handling of mass balance when considering 2,3,7,8-TCDD binding to plasma proteins, and 3) improved 

estimation of physical and biochemical parameters. The Wang et al. (1997) model accurately described 

the time course distribution of 2,3,7,8-TCDD following a single oral dose, as well as the concentration of 

2,3,7,8-TCDD in eight target tissues on day 3 after six different doses. The model described by Wang et 

al. (1997) was recently coupled to a biologically-based pharmacodynamic (BBPD) model to quantitatively 

describe the relationship between disposition and response in multiple tissues (Santostefano et al. 1998). 

This later model incorporated both pharmacokinetic and pharmacodynamic events to account for the 

ability to 2,3,7,8-TCDD to induce CYP1A1 and the fact that CYP1A2 is responsible for maintaining high 

concentrations of 2,3,7,8-TCDD in the liver. The results showed that the BBPD model accurately 

described the time course of CYP1A1 protein expression and EROD activity in the liver, skin, and 

kidneys. It also confirmed that EROD activity can be an appropriate marker for CYP1A1 protein 

expression, and the shape of the induction curves supported the hypothesis that similar time-dependent 

mechanism of 2,3,7,8-TCDD-induced CYP1A1 protein expression and associated EROD activity occurs in 

multiple tissues. This, in turn, suggested that parameter estimation in the study accurately described the 

Ah receptor-mediated mechanism on protein expression and enzymatic activities in multiple tissues. 

The Andersen et al. Model 

Description of the Model. This model (Andersen et al. 1993) is an extension of the earlier PBPK 

models developed by Leung et al. (1988, 1990b) for 2,3,7,8-TCDD. Like the earlier models, this model 

consists of five compartments. Each of the four tissue compartments has a specified blood flow, tissue 

compartment volume, and a tissue blood volume. Movement of chemical from blood to tissue was modeled 

to be proportional to the product of a permeation coefficient times surface area for the tissue. When this 

product is lower that the specified blood flow for the tissue, tissue uptake is diffusion-limited. Because of 

the diffusion-limited tissue compartments, the model did not require blood binding to match the time-course 

of tissue uptake. It was assumed that in the liver both the Ah receptor and the inducible binding protein act 

to sequester 2,3,7,8-TCDD through a capacity-limited binding process, and the binding protein was 

assumed to be CYP1A2. Binding interactions with CYP1A2 and CYP1A1 were described by reversible 

equilibrium relationships, which is valid as long as the rate constants for association/dissociation are large. 

It was also assumed that the DNA sites to which the Ah-2,3,7,8-TCDD complex binds are present at much 

lower concentrations than the Ah-ligand complex. For both CYP1A1 and CYP1A2 induction, it was 

assumed that the Ah-ligand complex formation was equivalent, but that the Hill term, n, (a measure of 
 

CDDs 222 

2. HEALTH EFFECTS 

interaction for multiple Ah-ligand complex binding sites) and the Hill binding constant were different for 

the two responses. The model also allowed for autoinduction of metabolism following 2,3,7,8-TCDD 

treatment. Data from Abraham et al. (1988) and Krowke et al. (1989) were analyzed. The former study 

provided dose-response characterization of concentrations of 2,3,7,8-TCDD in liver and of liver CYP1A1 

activity and time-course characterization of 2,3,7,8-TCDD concentration in tissues and enzyme activities in 

female Wistar rats. Krowke et al. (1989) examined liver and fat concentrations in male Wistar rats dosed 

weekly for up to 6 months. In addition, Andersen et al. (1993) examined the potential correlation between 

several measures of dose estimated by the model and the promotional efficacy and carcinogenicity of 

2,3,7,8-TCDD in Sprague-Dawley rats. Cancer data from Kociba et al. (1978a) and Pitot et al. (1980) were 

analyzed. 

Validation and Discussion. Abraham et al. (1988) found that the disposition of 2,3,7,8-TCDD in 

liver and fat from rats administered a single subcutaneous dose (0.001–10 μg/kg) of the chemical was 

highly dose-dependent. The disproportionately higher concentration in the liver at higher doses appeared 

to be due to induction of a dioxin-binding protein, presumably CYP1A2. The model developed by 

Andersen et al. (1993) successfully simulated the experimental data. The affinity of the binding protein 

was estimated to be 6.5 nmol, while a value of 1 for n suggested little interaction among 2,3,7,8-TCDDresponsive DNA-binding sites involved in the expression of CYP1A2. For describing induction of 

CYP1A1, an n of 2.3 was required, which suggested possible interactions among DNA-binding sites for 

the Ah-ligand complex with this gene. The simulation of the time-course of elimination from liver and of 

induction of CYP1A1 was in good agreement with the empirical data, but required the inclusion of timedependent growth parameters over the 100 days of the experiment. The model also successfully simulated 

the data from the repeated-dosing study by Krowke et al. (1989) after small adjustments were made to fat 

and slowly perfused tissue parameters. The measures of dose that were used for comparison with the 

promotional and carcinogenic properties of 2,3,7,8-TCDD were integrated total liver concentration during 

the treatment period, or integrated free liver 2,3,7,8-TCDD concentration. Also, measures of tissue dose 

related to enhanced expression of CYP1A1 and hepatic binding proteins were calculated and examined for 

correlation with promotional activity. Results of the analysis revealed that under the exposure conditions, 

the tumor promotional response of 2,3,7,8-TCDD in the rat liver most closely correlated with integrated 

expression of the CYP1A1 gene. However, Andersen et al. (1993) indicate that since there is no 

expectation of causality between tumor responses and induction of CYP1A1 (or CYP1A2), the correlation 

should be regarded cautiously. Consistent with the findings of Leung et al. (1988, 1990b), the results from 
 

 

CDDs 223 

2. HEALTH EFFECTS 

the Andersen et al. (1993) study showed that over a certain dose range (e.g., at doses several fold above 

background), protein (CYP1A2) induction greatly alters 2,3,7,8-TCDD disposition. 

Recently, Andersen and co-workers developed a model of hepatic enzyme zonation that was combined 

with the PBPK model of protein induction (Andersen et al. 1993) to create a multicompartmental 

representation of the liver architecture that can be used to predict the degree of induction in both the whole 

liver and in specific regions (Andersen et al. 1997a, 1997b). A geometric representation was used to 

divide functional units (based on enzyme distribution) within the liver into five zones. The primary 

objective was to compare model predictions for regional induction with regional protein induction as 

visualized by immuno-histochemistry. The data set modeled included analysis of tissue distribution of 

2,3,7,8-TCDD in the first days or weeks after a single dose, time course studies for about 100 days after a 

single dose, and initiation-promotion studies in rats dosed for up to 6 months. The results showed that the 

five-compartment model was more successful than conventional homogeneous one-compartment liver 

models not only in simulating low-dose behavior for mRNA in whole liver but also in representing 

immunohistochemical observations. Five or more compartments were required to give a sharp boundary 

between induced and noninduced regions of the liver. When the five-compartment liver model was used 

to account for CYP1A1 and CYP1A2 induction and regional distribution of induced enzymes, the lowdose behavior appeared to be nonlinear and was better described, with a large n value (Hill coefficient) and 

a range of affinities in the liver covering about 81-fold differences between centrilobular and periportal 

regions. 

The Kohn et al. (NIEHS) Model 

Description of the Model. Kohn et al. (1993) constructed a mathematical model (the NIEHS model) 

to describe 2,3,7,8-TCDD tissue distribution and 2,3,7,8-TCDD-mediated alterations in hepatic proteins in 

the rat. The model assumed that 2,3,7,8-TCDD mediates increases in liver concentration of transforming 

growth factor-α (TGF-α) by a mechanism which requires the Ah receptor. TGF-α subsequently binds to 

the EGF receptor, a process which is known to cause internalization of the receptor in hepatocytes. This is 

thought to be an early event in the generation of a mitogenic signal. The model included equations for the 

Ah receptor-dependent induction of CYP1A1 and CYP1A2 activity and of the Ah receptor itself. Because 

it was also assumed that estrogen action is required for 2,3,7,8-TCDD-mediated induction of TGF-α, 

production of the estrogen receptor, CYP1A2-catalyzed formation of catechol estrogens, and deactivation 

of estrogens by glucuronidation were included in the model. The model predictions were compared to the 

two-stage initiation-promotion data of Tritscher et al. (1992) and Sewall et al. (1993). Gavage doses 
 

CDDs 224 

2. HEALTH EFFECTS 

equivalent to 3.5–125 ng 2,3,7,8-TCDD/kg/day for 30 weeks were used in these studies. Data from 

Abraham et al. (1988) were also analyzed. Model parameters were obtained from the literature or 

calculated from experimental data and adjusted to make the model reproduce the observations of Tritscher 

et al. (1992) and Sewall et al. (1993). 

Validation and Discussion. The model prediction for the percentage of absorption (>90%) from 

ingestion of 2,3,7,8-TCDD was in good agreement with experimental data of Rose et al. (1976). The model 

also predicted that 92.2% of the metabolite appears in the feces and 7.8% in the urine at a dose of 

125 ng/kg/day. The dose of 2,3,7,8-TCDD did not have a significant effect on these predictions. From the 

fit to the data of Abraham et al. (1988), the model predicted an initial and overall half-time clearance from 

liver of 11.8 and 13.5 days, respectively, which is very close to the experimentally obtained 11.5 and 

13.6 days. Similar good agreement was obtained for half-time elimination from fat (22.3 days versus 

24.5 days). The model predicted a linear relationship between administered dose and the concentration of 

2,3,7,8-TCDD in the liver at doses between 3.5 and 125 ng/kg/day, which was in good agreement with the 

data of Tritscher et al. (1992). The relationship between 2,3,7,8-TCDD dose and induction of both 

CYP1A1 and CY1A2 was best fit by an hyperbolic curve suggesting lack of cooperative interactions among 

binding sites. The hyperbolic curve was consistent with the experimental data for induction of these 

proteins from Tritscher et al. (1992). The model also predicted that the fractional occupancy of the Ah 

receptor by 2,3,7,8-TCDD rises from 13.4% at a dose of 3.5 ng/kg/day to 69.3% at 125 ng/kg/day. The 

model prediction of the degree of internalization of the EGF receptor as a function of the concentration of 

TGF-α was also hyperbolic in shape and successfully reproduced the experimental data of Sewall et al. 

(1993). Kohn et al. (1993) indicate that as this response may be involved in the mechanism of 

tumorigenesis in 2,3,7,8-TCDD-treated rats, it would be expected that it would correlate with tumor 

incidence better than does tissue dose. If so, extrapolation of effects at high dose to low doses may 

underestimate low-dose effects. However, extrapolation from low dose to extremely low dose would still 

be valid. The model predicted that 10 days after administration of a single dose of 1 μg 2,3,7,8-TCDD/kg 

there should be a greater decrease in plasma membrane EGF receptor in female rat liver than in male rat 

liver, which is consistent with the observed lower sensitivity of the male. Consistent with the experimental 

data, the model reproduced the decrease in hepatic estrogen receptor (ER) level resulting from exposure to 

2,3,7,8-TCDD, and the relationship between concentration of 2,3,7,8-TCDD and amount of receptor was 

also hyperbolic. Overall, the model’s success in reproducing the observed responses to 2,3,7,8-TCDD for 

the various proteins included in the model supports the proposed mechanism that internalization of the EGF 
 

 

 

CDDs 225 

2. HEALTH EFFECTS 

receptor in response to induction of TGF-α may be the origin of the mitogenic signal important for 

carcinogenesis. 

The Carrier Model 

Description of the Model. The first part of this model provides a quantitative description of the 

distribution of 2,3,7,8-substituted CDDs (and CDD-like compounds) between liver and adipose tissues as a 

function of overall body concentration at any given time (Carrier et al. 1995a). In a second step, differential 

equations were used to describe the disposition of CDDs in liver, adipose tissues, and whole body as a 

function of time (Carrier et al. 1995b). The first step of the model was based on several hypotheses: 1) 

changes in overall CDD concentration are slow relative to intertissue diffusion exchanges, protein 

induction, and binding of CDDs in the liver; 2) CDDs are mainly in adipose tissue and in the liver, but 

exchanges between these two sites are mediated via the blood; 3) the liver synthesizes proteins that bind 

free CDDs according to standard mass action association-dissociation mechanisms; 4) synthesis of binding 

proteins in the liver is linked to binding of free CDDs to the Ah receptor; 5) CDDs in fat deposits within the 

liver contribute to the overall liver burden and is taken into account; and 6) small amounts of CDDs are 

contained in organs other than the liver and adipose tissues and this fraction is assumed to be constant. In 

the second step, CDDs were assumed to be eliminated mainly by hepatic clearance; elimination by lactation 

or transplacental distribution was not considered. Model simulations of various experimental data sets, as 

specified below, were conducted. When not readily available, anatomical and physicochemical parameters 

were obtained from laboratory or clinical data. 

Validation and Discussion. The model successfully simulated data from Abraham et al. (1988), who 

provided dose-response characterization of concentrations of 2,3,7,8-TCDD in the liver of rats after a single 

dose of the compound. Analysis of the data showed that the higher the body burden, the higher the 

proportion of the burden contained in the liver. However, the model predicted that a plateau is reached 

when body burden is >1 mg 2,3,7,8-TCDD/kg body weight. The model predictions were also in good 

agreement with experimental data from Van den Berg et al. (1986a), who administered a single dose of a 

mixture of CDDs and CDFs to rats and hamsters and with data in monkeys administered a single oral dose 

of 2,3,7,8-TCDD (McNulty et al. 1982). Results from simulations conducted on data from chronic studies 

in rats (Kociba et al. 1978a; Rose et al. 1976) and on human data from Yusho patients also showed that 

increasing the body burden results in an increase in the fraction of the body burden present in the liver and 

in an increase in the liver/adipose concentration ratio. These changes in fractional distributions were 
CDDs 226 

2. HEALTH EFFECTS 

attributed to the affinity of specific liver proteins for binding of free hepatic CDDs and the saturable 

capacity of the binding proteins at high concentration of free CDDs. Model simulations of elimination data 

in rats after single (Abraham et al. 1988) or repeated doses (Kociba et al. 1978a; Rose et al. 1976) of 

2,3,7,8-TCDD, as well as data from a Yu-Cheng patient agreed well with the empirical data and showed 

that disposition kinetics of 2,3,7,8-substituted CDDs are nonlinear (i.e., as body burden decreases with time, 

liver and adipose tissue half-lives increase). According to the model, an additional factor that can influence 

the disposition kinetics of 2,3,7,8-CDDs is a rapid change in body weight and/or adipose tissue mass. A 

rapid loss of adipose tissue whether by dieting or in patients experiencing anorexia, would result in a higher 

concentration of the chemical in the remaining adipose tissue, particularly if the loss of tissue is much faster 

than whole body elimination via the liver. 

2.3.5.4 Risk Assessment. 

In early efforts to model the disposition of persistent halogenated aromatic hydrocarbons, disposition was 

described by simple partitioning between the blood and the various tissues with first-order metabolism in 

the liver. In those studies, the role that extensive tissue binding to particular cellular proteins might play in 

determining the overall disposition of the chemical was not accounted for. In contrast, the descriptions of 

Leung et al. (1988, 1990b) and Carrier et al. (1995a, 1995b) attempted to provide a biochemical basis for 

the observed tissue distribution. The use of this type of model may help explain interspecies differences in 

2,3,7,8-TCDD sensitivity and carcinogenicity. The rodent PBPK model for 2,3,7,8-TCDD revealed very 

consistent behavior between species, and some of the predictions of high dose-low dose behavior were 

verified. 

One advantage of a description that explicitly includes protein binding is the ultimate ability to develop 

pharmacodynamic models for 2,3,7,8-TCDD (and related chemicals) toxicity based on Ah receptor 

occupancy or Ah-TCDD complex concentration in vivo and to realistically couple it with the biologically 

based cancer models (or with models for other 2,3,7,8-TCDD responses). This was attempted by Andersen 

et al. (1993) and Kohn et al. (1993), who included estimates of binding constants between the Ah receptor 

and 2,3,7,8-TCDD and between the Ah receptor-dioxin complex and sites on DNA. Santostefano et al. 

(1998) extended previous modeling attempts by determining parameter values based on time course of 

CYP1A1 and CYP1A2 responses in multiple tissues using a simultaneous PBPK and BBPD models. 

However, as noted by Andersen et al. (1993), in order to develop a complete biologically motivated riskassessment model, these dosimetry models need to be combined with quantitative descriptions of cell and 
CDDs 227 

2. HEALTH EFFECTS 

tissue responses. Kohn et al. (1993) used the NIEHS model to successfully predict tissue concentrations of 

2,3,7,8-TCDD and of various induced proteins involved in the carcinogenic response to 2,3,7,8-TCDD and 

suggested that such a model might permit extrapolation of responses beyond the range obtained from 

experimental data and lead to scientifically sound approaches for estimating risks of adverse health effects 

of exposure to 2,3,7,8-TCDD. The importance of the results of Kohn et al. (1993) can be illustrated by the 

finding that the dose-response curves for various proteins were hyperbolic rather than sigmoid. 

Sigmoidicity in the response requires a higher concentration to produce a given response at low dose than 

does a hyperbolic response having the same concentration for half-maximal effect. This implies that the 

response is approximately linear at very low doses. 

2.4 MECHANISMS OF ACTION 

2.4.1 Pharmacokinetic Mechanisms 

The mechanism of absorption of CDDs by the inhalation and dermal routes of exposure is not known. 

Transfer of CDDs from the aqueous environment of the intestine across cell membranes is predominantly 

limited by molecular size and lipid solubility. The overall evidence indicates that 2,3,7,8-substituted tetraand pentachlorinated congeners are well absorbed. In contrast, OCDD was poorly absorbed from the 

gastrointestinal tract of rats (Birnbaum and Couture 1988), but absorbed more on chronic exposure 

(Birnbaum et al. 1989a). Absorption is also vehicle-dependent (Poiger and Schlatter 1980). Highly 

chlorinated congeners, although absorbed in small amounts, can accumulate in the liver. Results from 

studies in thoracic duct-cannulated rats showed that 2,3,7,8-TCDD was transported primarily via the 

lymphatic route and was predominantly associated with chylomicrons (Lakshmanan et al. 1986). Several 

studies have examined the distribution of CDDs between blood and adipose tissue. Patterson et al. (1989d) 

showed that on a lipid basis the serum/adipose ratio for 2,3,7,8-TCDD in humans was approximately 1:1, 

and this correlation held over a concentration range of almost three orders of magnitude. They also found 

that in blood <10% of 2,3,7,8-TCDD was associated with red blood cells, which according to Patterson et 

al. (1989d), suggested that most of 2,3,7,8-TCDD in blood was bound to serum lipids and lipoproteins. 

However, the distribution between plasma lipid and adipose tissue increased with chlorine substitution, 

which indicated that higher chlorinated congeners have a higher binding affinity for plasma proteins 

(Patterson et al. 1989d; Schecter et al. 1990). Experiments of in vivo binding of CDD congeners to various 

serum fractions revealed that as chlorine content increased, binding to lipoproteins gradually decreased, 

75% of 2,3,7,8-TCDD was found bound to lipoprotein compared to 45% for OCDD (Patterson et al. 
CDDs 228 

2. HEALTH EFFECTS 

1989b). However, binding to other proteins increased with chlorine content (20% for 2,3,7,8-TCDD versus 

50% for OCDD). Also, fewer CDDs (<10%) were bound to the chylomicrons in serum. In studies in vitro 

with human whole blood, 80% of the applied amount of 2,3,7,8-TCDD was associated with lipoproteins, 

15% with proteins, and 5% with cellular components (Henderson et al. 1988). Also, there is some evidence 

that 2,3,7,8-TCDD and related stereoisomers may be associated with plasma prealbumin (McKinney et al. 

1985a; Pedersen et al. 1986). Within the lipoprotein fraction and per mole of lipoprotein, 2,3,7,8-TCDD 

has highest affinity for very low density lipoprotein (VLDL), followed by LDL and HDL (Marinovich et al. 

1983). A study using cultured human fibroblasts presented some evidence that specific binding to LDL and 

the LDL receptor pathway may explain in part the rapid early uptake of 2,3,7,8-TCDD with LDL entry 

(Weisiger et al. 1981). 

2,3,7,8-substituted CDDs are the predominant congeners retained in tissue and body fluids from 

rodents and monkeys (Abraham et al. 1989b; Van den Berg et al. 1983), although minor retention of 

non-2,3,7,8-substituted congeners has been reported in the rat (Abraham et al. 1989b). In general, the 

tissue distribution of CDDs is congener-specific and depends on the dose and route of administration 

(see Van den Berg et al. 1994 for review). In rats, for any particular organ or tissue, distribution 

within 24 hours of dosing depends on blood perfusion rate and relative tissue size, such that relatively 

high initial CDD concentrations are found in the adrenal glands and skeletal muscle (Pohjanvirta et al. 

1990). Shortly thereafter, the liver and adipose tissue become the major storage sites (Allen et al. 

1975; Lakshmanan et al. 1986; Rose et al. 1976). Data from studies in humans, marmoset monkeys 

and rats suggest that the distribution ratio between liver and adipose tissue increases with increasing 

degree of chlorination (Abraham et al. 1989c; Neubert et al. 1990a; Thoma et al. 1990), but also 

depends on the dose, metabolic rate, route of administration, and the time of observation after dosing. 

In non-human primates and in humans, the liver appears to be a less significant storage site than in 

rodents (Van Miller et al. 1976). In mice, the Ah receptor does not appear to play a significant role in 

2,3,7,8-TCDD body distribution for adipose tissue, skin, kidney, and total-body concentration 

(Birnbaum 1986). However, it plays some role in liver retention (Birnbaum 1986; Gasiewicz et al. 

1983a) and this was found to be related to inducibility of cytochrome P-450 (Leung et al. 1988), in 

particular CYP1A2. Distribution of 2,3,7,8-TCDD in mice has been shown to be age-dependent 

(Pegram et al. 1995). The greater fat content of some tissues in old mice enhances partitioning of 

2,3,7,8-TCDD into the tissues, while decreased perfusion prolongs clearance (Pegram et al. 1995). 

Some acute- and chronic-duration studies in rats have demonstrated a disproportionate dose-dependent 

distribution of 2,3,7,8-TCDD in liver and adipose tissue (Abraham et al. 1988; Kociba et al. 1978b). 

The greater the dose, the greater the liver/adipose tissue ratio. A disproportionate 
CDDs 229 

2. HEALTH EFFECTS 

dose-dependent distribution has also been demonstrated in mice (Diliberto et al. 1995). 

2,3,7,8-TCDD-derived radioactivity in the liver was found associated preferentially with the 

microsomal fraction (Allen et al. 1975). Information summarized by Van den Berg et al. (1994) 

suggests that the disproportionately greater hepatic concentration of 2,3,7,8-TCDD after exposure to 

higher doses may be explained in part by the induction of a hepatic-binding species, CYP1A2. This 

distribution parameter is also explained by the Carrier et al. (1995a,b) model in humans. 

There is some experimental data to suggest that the fetal/neonatal period may be more sensitive to the 

toxicity of CDDs than the adult animal. Several studies have shown that limited placental transfer of 

CDDs takes place in rodents (Li et al. 1995c; Nau et al. 1986; Van den Berg et al. 1987b; Weber and 

Birnbaum 1985) and in humans (Schecter et al. 1996a). However, little is known about the 

mechanisms responsible for the transfer of CDDs across the placenta, the dependence of these 

mechanisms on the gestational period, and the distribution of these compounds in fetal tissue. 

However, CDDs and related chemicals are able to concentrate in breast milk, and limited human 

(Abraham et al. 1994; McLachlan 1993; Pluim et al. 1993b) and animal (Nau et al. 1986) data have 

indicated considerable absorption of these compounds by the nursing infant. Thus, while the in utero 

exposure of fetal tissues to CDDs may represent only a small percentage of the maternal body burden 

of CDDs, the breast-fed infants will receive a higher daily dose per body weight than adults. Further 

information regarding placental transfer and elimination of CDDs through breast milk is presented in 

Section 2.3.4.4. 

As mentioned in Section 2.3.3, metabolic transformation of CDDs in vivo includes oxidation and 

reductive dechlorination as well as glutathione conjugation. Studies in two rat strains which differ 

greatly in sensitivity to 2,3,7,8-TCDD did not provide evidence for a role of toxicokinetics and 

metabolism in the difference in sensitivity (Pohjanvirta et al. 1990). Also, studies in various mice 

strains showed no significant Ah receptor-related differences in metabolic pathways (Gasiewicz et al. 

1983a). While in vitro studies have shown similarities between most species regarding metabolite 

profile, the rate of 2,3,7,8-TCDD metabolism and the number of metabolites were reduced in 

hepatocytes in suspension culture from guinea pigs, a highly sensitive species (Wroblewski and Olson 

1985). The overall evidence suggests that 2,3,7,8-TCDD can induce its own metabolism in vivo, but 

only at doses that could cause overt signs of toxicity (Van den Berg et al. 1994). It is important to 

consider the possibility of autoinduction at high doses because data obtained with exposure levels 

associated with a significant induction of CYP1A1 and CYP1A2 may not necessarily reflect 

toxicokinetic behavior at low-exposure levels. 
CDDs 230 

2. HEALTH EFFECTS 

Elimination of 2,3,7,8-substituted CDDs occurs mainly via the bile and the feces as polar metabolites; 

much smaller amounts are excreted via the urine. Moreover, in almost all mammalian species studied, 

the 2,3,7,8-TCDD-derived radioactivity in tissues is associated with the parent compound, suggesting 

that the hydroxylated and/or conjugated metabolites are rapidly eliminated from the body. Studies in 

mice showed that a strain of mice having a low affinity Ah receptor eliminate 2,3,7,8-TCDD at a much 

slower rate than mice with an Ah receptor of high affinity for the ligand (Gasiewicz et al. 1983a). 

These strain differences in body distribution and elimination could be explained not only by the 

differences in adipose tissue content, but also by the presence of a hepatic microsomal binding protein 

(Leung et al. 1988). Further studies in congenic mice suggested that the distribution and excretion of 

2,3,7,8-TCDD is controlled primarily by the total genetic background and not by the allele present at 

the Ah-locus (Birnbaum 1986). Guinea pigs eliminate 2,3,7,8-TCDD considerably slower than other 

rodents (Olson 1986). This may reflect the relatively limited ability of the guinea pig to metabolize 

2,3,7,8-TCDD and may contribute to the greater persistence and greater acute toxicity of 

2,3,7,8-TCDD in this species. Results from a repeated-dosing study in rats showed that the rateconstant defining the approach to steady-state concentrations was independent of the dose over the 

range tested (Rose et al. 1976). This was consistent with evidence suggesting that autoinduction of 

2,3,7,8-TCDD metabolism does not occur following exposure to sublethal doses. Autoinduction of 

metabolism could explain cases of dose-related excretion in which longer half-lives for elimination are 

seen at lower-exposure levels which are not associated with enzyme induction. 

2.4.2 Mechanisms of Toxicity 

The mechanism(s) of toxicity for CDDs is not completely understood but has been extensively studied, 

particularly for 2,3,7,8-TCDD, and numerous reviews are available on this subject (Birnbaum 1994a; 

Goldstein and Safe 1989; Kerkvliet 1995; Landers and Bunce 1991; Okey et al. 1994; Poland and 

Knutson 1982; Safe 1986, 1990; Silbergeld and Gasiewicz 1989; Whitlock 1987, 1993). Many CDDs, 

CDFs, coplanar PCBs, and other structurally related halogenated aromatic hydrocarbons are believed 

to share a common mechanism of action intimately related to similarities in their structural 

configuration. Most of what is known regarding the mechanism of action of these compounds is based 

on research in three main areas: structure-activity relationships for receptor binding and induction of a 

variety of biochemical and toxicological responses; genetic studies using inbred mouse strains; and 

studies at the molecular level which have elucidated key events in the actions of the receptor. Most of 

the studies providing this information used parenteral routes of exposure and/or in vitro tests systems. 

It is beyond the scope of this profile to discuss these studies in detail. 
CDDs 231 

2. HEALTH EFFECTS 

The extraordinary potency of 2,3,7,8-TCDD in evoking a dose-related induction of cytochrome P-450­

associated AHH activity, the stereospecificity among related halogenated aromatic compounds to 

evoke this response, and the tissue specificity of enzyme induction, led Poland and Glover (1973b) to 

postulate the existence of an induction receptor. This receptor, the Ah receptor (Ah for aromatic 

hydrocarbon), was later identified in the cytosol of mouse liver cells (Poland et al. 1976) and, 

subsequently, in hepatic and extrahepatic tissues of a variety of laboratory animals, mammalian cell 

cultures, human organs and cell cultures, and also nonmammalian species (Okey et al. 1994). Results 

from structure-binding relationships for a series of CDD congeners using mouse hepatic cytosol 

showed that not all the congeners had the same affinity for the Ah receptor; affinity was found to be 

determined by the chlorine-substitution pattern (Mason et al. 1986; Poland et al. 1976, 1979). The 

most active compound was 2,3,7,8-TCDD, which is substituted in all four lateral positions. Addition 

of one, two, or four nonlateral chlorine substituents, or removal of lateral chlorine substituents, 

resulted in congeners with lower binding affinities. The stereospecific nature of the binding suggested 

the existence of a cytosolic receptor as a mediator in responses caused by 2,3,7,8-TCDD and related 

compounds. 

2,3,7,8-TCDD and structurally related halogenated aromatic compounds induce a variety of 

microsomal enzymes primarily in the liver. The most widely studied of these responses are induction 

of hepatic AHH and EROD (markers of CYP1A1 activity) in mammalian cell cultures and in 

laboratory rodents (Goldstein and Safe 1989; Poland and Glover 1973a; Safe 1986, 1990). Several 

studies have examined the in vitro and in vivo structure-activity relationships for CDDs as inducers of 

hepatic and extrahepatic CYP1A1 activity (Bradlaw and Casterline 1979; Harris et al. 1990; Mason et 

al. 1986; Poland and Glover 1973a; Poland and Knutson 1982; Poland et al. 1979). The most active 

CDDs were substituted in their 2,3,7, and 8 positions, and the structure-activity relationships for 

induction and receptor binding assay were comparable. The molecular dimensions of the binding site 

was initially estimated to fit ligands that were approximately 3×10D (Poland and Knutson 1982), 

which would accommodate molecules such as 2,3,7,8-TCDD; however, approximate dimensions of 

12x14x5D would be required to accommodate other chemicals (e.g., 3-MC or β-naphthoflavone), 

known to bind (Landers and Bunce 1991; Rannug et al. 1991; Waller and McKinney 1995). Although 

results from these experiments provided further evidence for a receptor-mediated mechanism of action, 

there was not strictly a linear correlation between Ah receptor binding and enzyme induction. Mason 

et al. (1986) suggested that a number of factors, including differential solubilities of the CDDs in the 

assay buffer system at higher concentrations, may contribute to the nonlinearity. They also suggested 

that structure-dependent receptor protein-ligand interactions which occur after the initial binding event 

may have played a role in the nonlinearity of the data sets. 
CDDs 232 

2. HEALTH EFFECTS 

Furthermore, differential rates of metabolism and elimination of particular CDDs also likely account 

for the comparative differences between studies in vivo and in vitro (Birnbaum 1985). 

Numerous studies have examined the structure-toxicity relationships for CDDs. For example, 

examination of lethality data in guinea pigs revealed that the fully lateral-substituted tetra- to 

hexachloro-substituted isomers were the most toxic congeners, and the structure-activity relationships 

were comparable to those observed for their AHH-induction and receptor-binding activities (Eadon et 

al. 1986). Similar results have been reported for responses such as body weight loss and thymic 

atrophy (Mason et al. 1986; Safe 1987). Furthermore, there was an excellent correlation between the 

in vitro AHH induction potencies and the in vivo responses. Additional end points for which structuretoxicity relationships correlate well with structure-induction potencies and/or Ah receptor-binding 

affinities are epidermal responses such as the keratinization of the mouse teratoma XB cells and the 

production of skin lesions in genetically inbred haired and hairless mice (Knutson and Poland 1980, 

1982), suppression of the splenic antibody response to SRBC (Kerkvliet et al. 1985), antiestrogenicity 

(Gierthy et al. 1987; Krishnan and Safe 1993), and teratogenicity (Weber et al. 1985). Taken together, 

these results, and others, strongly supported the role of the Ah receptor in mediating the toxicity of 

2,3,7,8-TCDD and related halogenated aromatic hydrocarbons. 

As previously mentioned, 2,3,7,8-TCDD and structurally related compounds induce a wide range of 

biological responses, including alterations in metabolic pathways, body weight loss, thymic atrophy, 

impaired immune responses, hepatotoxicity, chloracne and related skin lesions, developmental and 

reproductive effects, and neoplasia. The expression of these responses is thought to be initiated by 

the binding of individual congeners (or ligands) with the Ah receptor. However, responsiveness of 

certain mouse strains to aromatic hydrocarbons is inherited in a simple autosomal-dominant mode 

and both enzyme induction and the toxic responses to 2,3,7,8-TCDD appear to segregate with the Ah 

locus (Poland and Glover 1980). For example, certain mouse strains, typified by C57XBL/6J, have 

an Ah receptor protein with a relatively high binding affinity for inducers of AHH such as 

3-methylcholanthrene, β-naphthoflavone, 2,3,7,8-TCDD, and other isostereomers of 2,3,7,8-TCDD, 

and are sensitive to the toxic effects of these chemicals. In contrast, other mouse strains, such as 

DBA/2J, have an Ah receptor protein that has a lower ligand affinity (Okey et al. 1989), and are 

much less sensitive to the toxic effects of these compounds. The use of these mouse strains and 

strains differing only the Ah locus (congenic) has suggested that many of the responses elicited by 

these chemicals (e.g., enzyme induction, thymic involution, cleft palate formation, hepatic porphyria, 

and immunotoxicity) segregate with this Ah locus (Birnbaum et al. 1990; Kerkvliet et al. 1990b; Lin 

et al. 1991a, 1991b; Poland and Knutson, 1982; Swanson and 
CDDs 233 

2. HEALTH EFFECTS 

Bradfield 1993). More recent investigations using an Ah receptor-deficient mouse 

(Fernandez-Salguero et al. 1996) also support the role of this protein in mediating the toxicity of 

2,3,7,8-TCDD and related halogenated aromatic hydrocarbons. 

These genetic data strongly support the role of the Ah receptor in mediating the toxicity of 

2,3,7,8-TCDD and related halogenated aromatic hydrocarbons. However, it has become clear that a 

comparison of the properties of the Ah receptor across species and tissues indicates that it is difficult 

to account for the species-specific sensitivity and diversity of the biological effects of 2,3,7,8-TCDD 

by characteristics of the receptor alone. There are several different forms of this protein in mice 

encoded by several different alleles of the same locus (Poland and Glover 1990; Poland et al. 1994). 

By analogy with the existence of multiple receptor forms in mice, it is reasonable to anticipate that 

the human population will also have different receptor forms. The extent to which these forms in 

mice and humans affect the types of responses elicited and the sensitivity to TCDD is not known. 

As indicated above, the Ah receptor has been identified in several human tissues and cell lines (Cook 

and Greenlee 1989; Harper et al. 1991; Harris et al. 1989a; Lorenzen and Okey 1991; Roberts et al. 

1990). Although the general properties and function of the human Ah receptor (Harper et al. 1991) 

appear to be very similar to the rodent and other species (Denison et al. 1986a; Gasiewicz and Rucci 

1984), some differences exist. For example, the molecular mass from a variety of human cell lines 

or tissues ranges from 106 to 110 kDa (Harper et al. 1991; Poland and Glover 1987; Wang et al. 

1991), compared to approximately 95 kDa for C57XBL/6J mice or from Hepa-1 cells (Landers et al. 

1989; Poland and Glover 1987, 1990; Prokipcak and Okey 1990), and 124 kDa from the hamster 

(Poland and Glover 1987). The same parameter for the nonresponsive DBA/2J mouse is 

approximately 104 kDa (Poland and Glover 1990). There is no known correspondence between 

molecular mass of the protein and its affinity for any ligand and/or ability to mediate a biological or 

toxicological response. Apparent affinity constants (measured under in vitro conditions) for 

2,3,7,8-TCDD-human Ah receptor binding from various cell lines range from 3 to 15 nM compared 

with about 1 nM in cytosol from C57XBL/6J mice, 16 nM for the DBA/2J mouse, 0.1 nM for the 

guinea pig and 0.3 nM for the hamster (Cook and Greenlee 1989; Gasiewicz and Rucci 1984; Harper 

et al. 1991; Okey et al. 1989). 

While the use of structure-activity relationships and mouse genetics are consistent with the notion that the 

binding of 2,3,7,8-TCDD and structurally-related chemicals to the Ah receptor is the initial event that leads 

to the induced synthesis of certain enzymes, it has only been through the work at the cellular and molecular 

biological levels that this has been substantiated. Furthermore, these investigations indicate that complex 

series of events regulate the activity of the receptor and it is likely that the differential regulation of these 
CDDs 234 

2. HEALTH EFFECTS 

may, at least in part, be responsible for the tissue- and species-specific nature of the response observed in 

mammals following the exposure to 2,3,7,8-TCDD and related compounds. Immunohistochemical studies 

have shown that in intact mouse hepatoma cells, the unliganded receptor resides in the cytoplasm, and that 

exposure to 2,3,7,8-TCDD leads to an accumulation of the receptor within the nucleus (Pollenz et al. 1994). 

However, the precise location of the unoccupied (i.e., without 2,3,7,8-TCDD bound) receptor in intact cells 

is still unresolved. The unoccupied AhR exists as a heteromeric complex with 2 molecules of another 

protein called 90 kDa heat-shock protein (hsp90) and another 43-kDa protein (Chen and Perdew 1994). 

Hsp90 appears to be necessary for maintaining the proper folding of the Ah receptor so it can bind ligand 

and limit the presence of another receptor form that is able to bind to DNA (Pongratz et al. 1992). The 

exact role of the 43-kDa protein is not yet known. 

Binding of the Ah receptor by 2,3,7,8-TCDD initiates a series of as yet undefined events resulting in the 

dissociation of hsp90 and nuclear localization (Henry and Gasiewicz 1993; Pollenz et al. 1994; Pongratz et 

al. 1992). Results from experiments in genetically variant cells that respond poorly to 2,3,7,8-TCDD 

revealed a defect in 2,3,7,8-TCDD binding that results in an altered receptor. Other variants exhibited 

normal binding, but the liganded receptors do not bind DNA and do not accumulate in the nucleus 

(Hankinson 1979; Miller and Whitlock 1981). The finding that these variants have a defect in a protein, 

termed Arnt (Ah receptor nuclear transport protein) (Reyes et al. 1992), suggested that 2,3,7,8-TCDD 

responsiveness requires both a ligand-binding protein (the Ah receptor) and a second protein which 

mediates the binding of the liganded receptor to DNA (Whitlock 1993). Furthermore, the ligand-bound Ah 

receptor does not itself bind DNA (Gasiewicz et al. 1991). The Arnt protein does not bind 2,3,7,8-TCDD, 

nor does it bind to DNA in the absence of the liganded Ah receptor protein (Whitelaw et al. 1993). The role 

of the Arnt protein as a translocator of the receptor from cytoplasm to the nucleus has been questioned; 

instead, it has been shown that Arnt interacts with the liganded Ah receptor to form a heterodimeric 

DNA-binding protein complex that can bind DNA and activate gene transcription (Whitlock 1993). Other 

investigations have shown that phosphorylation/dephosphorylation of the Ah receptor and the Arnt protein 

may influence both heterodimerization and the binding of this complex to DNA (Berghard et al. 1993; 

Mahon and Gasiewicz 1995; Okino et al. 1992; Pongratz et al. 1991). 

Most of the information regarding the sequence of events that follow 2,3,7,8-TCDD binding to the Ah 

receptor is based on analyses of induction of AHH activity, which results from enhanced transcription of the 

corresponding cytochrome P-450 1A1 (CYP1A1) gene. Stimulation of transcription occurs within minutes 

and does not require ongoing protein synthesis (Israel et al. 1985). These findings led to the 
CDDs 235 

2. HEALTH EFFECTS 

discovery of a dioxin-responsive regulatory DNA domain which has the properties of a transcriptional 

enhancer (Fisher et al. 1990; Jones et al. 1986; Neuhold et al. 1986). These specific DNA elements have 

been termed dioxin-responsive elements (DREs) and require both receptor protein and Arnt protein for 

enhancer function. DREs function in a chromosomal location distinct from that of the CYP1A1 gene 

(Fisher et al. 1989). In addition to the enhancer, the DNA upstream of the CYP1A1 gene has a second 

control element (a transcriptional promoter), which ensures that transcription is initiated at the correct site. 

Neither enhancer nor promoter functions in the absence of the other (Jones and Whitlock 1990). The fact 

that enhancer function requires both the receptor and Arnt protein, and that the liganded heteromeric form 

of the receptor shows increased affinity for the specific DNA sequence within the enhancer region 

suggested that the activation of the CYP1A1 gene involves the binding of the receptor heteromer to the 

DRE. This has been shown for the purified Ah receptor-Arnt protein complex (Henry et al. 1994). 

Analysis of the interaction of the Ah receptor with specific DNA domains indicates that the heteromer binds 

in a 1:1 ratio to the DRE (Denison et al. 1989). There is, however, no strict relationship between the 

affinity of the receptor heteromer for the DRE and the extent of enhancer activation (Neuhold et al. 1989; 

Shen and Whitlock 1992), which suggests that additional events, including DNA bending (Elferink and 

Whitlock 1990), must take place to activate transcription. 

The use of many in vitro techniques for these studies has required removing the DNA regulatory elements 

from the chromosome environment, and this may produce misleading results. This led researchers to 

examine the protein-DNA interactions at the dioxin-responsive enhancer in intact cells. Results from these 

studies suggested that the inactive enhancer is relatively inaccessible to DNA-binding proteins in vivo and 

that exposure to 2,3,7,8-TCDD leads to a rapid binding of six receptor heteromers and other proteins to the 

enhancer upstream of the CYP1A1 gene (Wu and Whitlock 1993). It has also been shown that the CYP1A1 

promoter, like the enhancer, is inaccessible in uninduced cells, and that exposure to 2,3,7,8-TCDD increases 

its accessibility to constitutively expressed proteins (Durrin and Whitlock 1989; Wu and Whitlock 1992). 

The 2,3,7,8-TCDD-induced change is not dependent on protein synthesis and is receptor- and 

Arnt-dependent. It has been suggested that the inaccessibility of the enhancer/promoter region in 

uninduced cells is due to its organization into nucleosomes (Ko et al. 1996; Morgan and Whitlock 1992). 

The mechanism by which the binding of liganded receptor heteromers to the enhancer alters chromatin 

structure leading to activation of transcription is unknown. Whitlock (1993) suggested that the DRE-bound 

receptor complex affects histones, thereby weakening the histone-DNA interactions and destabilizing 

nucleosomal structures. They also proposed that the receptor-enhancer interaction may alter the DNA 

structure of the enhancer/promoter region stabilizing it in a non-nucleosomal configuration. 
CDDs 236 

2. HEALTH EFFECTS 

The information resulting from the cloning and sequencing of the Ah receptor and Arnt has also expanded 

our knowledge of the molecular mechanisms whereby these proteins influence transcriptional activity 

(Hankinson 1995; Whitlock 1993). Both proteins are members of a class of transcription factors containing 

a basic helix-loop-helix (bHLH) structural motif as well as a PAS (Per-Arnt-Sim) domain. Both of these 

regions are involved in dimerization. The bHLH motif is also required for DNA sequence recognition, 

while the PAS domain contains the ligand-binding site (in the Ah receptor) and interacts with hsp90 

(Coumaileau et al. 1995; Dolwick et al. 1993; Fukunaga et al. 1995; Whitelaw et al. 1993). Both the Ah 

receptor and Arnt have C-terminal regions that function in transcriptional activation, although their relative 

contributions may depend on the gene involved (Ko et al. 1996). Other members of the HLH-PAS family 

include hypoxia-inducible factor 1 alpha (HIF-a) and Drosophila protein Sim (Huang et al. 1993; Wang et 

al. 1995). All of the bHLH proteins identified to date are involved in transcriptional regulation, and have a 

variety of roles in tissue growth and differentiation processes. It is not yet clear whether the 

ligand-activated Ah receptor modulates gene expression only through its interaction with Arnt; it may have 

other dimerization partners. It is known that multiple heterodimerizations occur among several 

transcription factors, and that this multiplicity provides for the recognition of other DNA sequences and 

diversity of regulation of responsive genes. Arnt has been shown to dimerize with HIF-1a and Sim, and 

there appear to be several different isoforms of Arnt, two of which have been shown to interact with the Ah 

receptor (Henry et al. 1994; Hirose et al. 1996; Ireland et al. 1995; Swanson et al. 1995; Wang and Semanza 

1995). However, as of yet, Arnt is the only protein that has been demonstrated to be a functional partner to 

the Ah receptor. Furthermore, the Ah receptor and Arnt appear to be co-expressed in a variety of tissues 

that have been examined (Abbott and Probst 1995; Abbot et al. 1995; Carver et al. 1994), suggesting 

co-dependence. 

As indicated above, much of our understanding of the interaction of 2,3,7,8-TCDD with the Ah receptor 

and how it modulates gene expression has come mainly from the analysis of the regulation of the CYP1A1 

gene. However, other studies have observed the presence of functional DREs in the genes that encode for 

CYP1A2 (Quattrochi et al. 1994), glutathione S-transferase Ya (Paulson et al. 1990), 

aldehyde-3-dehydrogenase (Takimoto et al. 1994), and NAD(P)H:quinone oxidoreductase (Favreau et al. 

1991). In addition, an imperfect DRE is present in the regulatory region of the cathepsin D gene. In this 

case, the Ah receptor-Arnt complex may act as a repressor to prevent the binding of other transcription 

factors to nearby enhancer sequences (Safe 1995). 
CDDs 237 

2. HEALTH EFFECTS 

As discussed below, 2,3,7,8-TCDD-elicited activation of the Ah receptor has been shown to alter the 

transcription of a number of genes. However, only a few of these, as indicated above, are as yet known to 

contain functional DREs (Lai et al. 1996). While in some cases the regulatory regions of all of the genes 

known to be altered by 2,3,7,8-TCDD have not been thoroughly examined; in other cases, the regulatory 

regions are known not to contain the conserved consensus sequence for the DRE. It is possible that other 

dimerization partners exist and that different DNA sequences might be recognized. It is also possible that 

the 2,3,7,8-elicited modulation of many of these genes and processes may be secondary subsequent to the 

induction/repression of a DRE-containing gene. However, there is some evidence to suggest that other Ah 

receptor-dependent pathways may exist for the alteration of gene expression that may not be dependent 

upon the interaction of the Ah receptor with nuclear elements. It has been suggested that the interaction of 

2,3,7,8-TCDD with the Ah receptor may initiate a phosphorylation/dephosphorylation cascade that may 

subsequently activate other transcription factors (Matsumura 1994). Enan and Matsumura (1995) reported 

an increase in protein kinase activity within 1–10 minutes following the addition of 2,3,7,8-TCDD to 

nuclear-free preparations of guinea pig adipose tissue. These results are consistent with previous 

investigations showing increased tyrosine kinase activity within minutes of 2,3,7,8-TCDD exposure 

(Bombick et al. 1988; Clark et al. 1991a; DeVito et al. 1994). Hsp90 has been found associated with a 

protein of 50 kDa (Chen and Perdew 1994; Whitelaw et al. 1993), and both have been shown to regulate the 

activity of pp60v-src, a tyrosine kinase (Brugge et al. 1983; Mimnaugh et al. 1995). c-Src has recently been 

reported to be a component of the unoccupied Ah receptor complex (Enan and Matsumura 1996). Thus, 

2,3,7,8-TCDD may modulate signal transduction processes and gene expression by at least two pathways: 

through the direct interaction of the Ah receptor and its heterodimer partners with gene regulatory elements, 

and from the initiation of a phosphorylation/dephosphorylation cascade and the subsequent modulated 

activity of other nuclear transcription factors. It has yet to be determined which pathways may be more 

important in acute versus chronic responses to these compounds and/or during particular developmental 

periods. Nevertheless, together these data indicate that well regulated and conserved pathways exist for the 

transduction of cellular signals through the binding of 2,3,7,8-TCDD-like chemicals to the Ah receptor. 

Since the modulation of these pathways results in toxicity in response to 2,3,7,8-TCDD and related 

compounds, it is presumed that these chemicals cause these responses by either interfering with the normal 

function of some unknown endogenous ligand, and/or stimulating the signal transduction process at an 

inappropriate time and/or for an inappropriately long period of time. 

As indicated in the preceding sections, cell/tissue death and necrosis are not prominent features of effects 

resulting from 2,3,7,8-TCDD exposure in vivo or in vitro. Hyperplasia, hypoplasia, metaplasia, and 
CDDs 238 

2. HEALTH EFFECTS 

dysplasia are the most common histopathological changes observed in animals (McConnell and Moore 

1979). Likewise, under conditions in vitro, 2,3,7,8-TCDD-like compounds are potent in altering cellular 

differentiation and growth patterns for a number of different cell types including embryonic palatal 

epithelial cells (Abbott et al. 1989), keratinocytes (Gaido and Maness 1994), osteoblasts (Gierthy et al. 

1994) and preimplantation embryos (Blankenship et al. 1993). 

Despite the numerous tissue- and species-specific responses that have been observed and the elegant work 

on the molecular mechanisms mediating some of these, there exists a considerable gap between knowledge 

of these changes and the degree to which they are related to the biological and toxicological end points 

elicited by 2,3,7,8-TCDD and related compounds. These chemicals have been shown to alter the 

transcription and/or translation of a number of genes, including several oncogenes and those encoding 

growth factors, receptors, hormones, and drug-metabolizing enzymes (Birnbaum 1994a, 1994b). More 

recent investigations have noted effects on certain enzymes and proteins (e.g., kinases) involved in various 

signal transduction processes as well as cell cycle control (Birnbaum 1994a, 1994b; Weib et al. 1996). The 

elicited induction of certain drug metabolizing enzymes such as CYP1A1, CYP1A2, and CYP1B1 are some 

of the most sensitive responses observed in a variety of different animal species, including humans. 

Significantly increased levels of CYP1A1 mRNA have been observed as dosages as low as 0.1 ng/kg body 

weight (Kohn et al. 1993). However, the precise 2,3,7,8-TCDD-induced biochemical alterations that are 

causally responsible for the abnormal growth processes observed are not known. This is due predominantly 

to our incomplete understanding of the complex and coordinate molecular, biochemical, and cellular 

interactions that regulate tissue processes during development and under normal homeostatic conditions. 

Nevertheless, there is some evidence that many of these biochemical alterations may be relevant to altered 

growth responses observed. For example, changes in the EGF receptor have been seen in tissues from 

2,3,7,8-TCDD-exposed animals and humans (Abbott and Birnbaum 1990a; Sewall et al. 1993; Sunahara et 

al. 1987). EGF and its receptor possess diverse functions relevant to cell transformation and tumorigenesis, 

and changes in these functions may be related to a number of dioxin-induced responses including neoplastic 

lesions, chloracne, and a variety of developmental effects. Likewise, the known ability of 2,3,7,8-TCDD 

directly or indirectly to alter the levels and/or activity of other growth factors and hormones, such as 

estrogen and thyroid hormone and their respective receptors, as well as enzymes involved in the control of 

cell cycle, may affect growth patterns in cells/tissues leading to adverse consequences. Thus, both the 

biochemical and biological data are consistent with the notion that 2,3,7,8-TCDD and related compounds 

are growth regulators. 
CDDs 239 

2. HEALTH EFFECTS 

2,3,7,8-TCDD and structurally related compounds elicit a wide range of adverse effects. Of the many 

adverse responses observed both in humans and experimental animals after exposure to 2,3,7,8-TCDD, the 

ones that appear at the lowest dose (more sensitive) are perhaps developmental/reproductive effects, 

alterations in the immune response, and neoplasia. An overview of the mechanism(s) involved in these 

effects is presented below. Detailed mechanistic explanations are beyond the scope of this profile. Some of 

the information has been extracted from recent reviews on these subjects (Kerkvliet 1995; Lucier et al. 

1993a; Peterson et al. 1993). 

Some common developmental effects attributed to 2,3,7,8-TCDD exposure in most laboratory mammals are 

thymic hypoplasia, subcutaneous edema, decreased prenatal growth, and prenatal mortality (Couture et al. 

1990). In addition, there are other species-specific effects, such as cleft palate in mice. Any of these effects 

may result from actions on the mother, embryo/fetus, placenta, or any combination of these sites (Peterson 

et al. 1993). In general, developmental effects can be induced at exposure levels that are not maternally 

toxic; however, prenatal mortality appears to be associated with maternal toxicity. Structure-activity results 

for 2,3,7,8-TCDD and related halogenated hydrocarbons for overt fetotoxicity are consistent with an Ah 

receptor-mediated mechanism. Hydronephrosis is the most sensitive developmental response induced by 

2,3,7,8-TCDD in mice and it can be observed at maternal doses that do not cause cleft palate or overt 

maternal toxicity (Abbott and Birnbaum 1989a; Abbott et al. 1987a, 1987b; Couture-Haws et al. 1991b; 

Neubert and Dillman 1972; Weber et al. 1985). Hydronephrosis in vivo is induced by a direct hyperplastic 

action of 2,3,7,8-TCDD on the uretic epithelium. This results in occlusion of the ureter and subsequent 

accumulation of urine in the kidney (Abbott et al. 1987a). As for cleft palate formation, 2,3,7,8-TCDD and 

related compounds seem to allow the palatal shelves to grow and make contact, but prevent the subsequent 

epithelial-to-mesenchyme transformation (Peterson et al. 1993; Pratt et al. 1984). Susceptibility to both 

hydronephrosis and cleft palate formation segregate with the Ah locus, and structure-activity relationships 

for dioxin-like compounds are consistent with those for Ah receptor binding (Safe 1990; Weber et al. 1985). 

Further details on the mechanism of 2,3,7,8-TCDD-induced hydronephrosis and cleft palate formation and 

the involvement of various growth factors in these responses can be found in Section 2.5. 

Another sensitive system for 2,3,7,8-TCDD toxicity is the male reproductive system, and many of the 

effects observed were originally thought to be related to the ability of 2,3,7,8-TCDD to decrease plasma 

androgen concentrations (Mably et al. 1992a, 1992b, 1992c; Moore et al. 1985). The fact that 

2,3,7,8-TCDD is transferred from mother to fetus and to neonates during lactation has a great impact on 
CDDs 240 

2. HEALTH EFFECTS 

the male reproductive system during early development. Testosterone and its metabolite dihydrotestosterone (DHT) are essential prenatally and/or early postnatally for imprinting and development of 

accessory sex organs and for initiation of spermatogenesis. Mably et al. (1992b) suggested that the 

demasculinization and feminization of sexual behavior and feminization by LH secretion is due to the fact 

that perinatal exposure to 2,3,7,8-TCDD impairs sexual differentiation of the central nervous system, which 

is dependent on the presence of androgens during early development. However, results from more recent 

studies suggest that the 2,3,7,8-TCDD-induced effects on the male reproductive system may be related to 

alterations in other systems such as brain amine content or in the expression of growth factors and receptors 

involved in urogenital cell system differentiation and proliferation (Bjerke et al. 1994a; Gray et al. 1995). 

Results from recent studies have also delineated the role of the Ah receptor in the development of 2,3,7,8­

TCDD-induced alterations in the male reproductive tract (Roman and Peterson 1998; Roman et al. 

1998a,b). A more detailed discussion of the mechanisms involved in these responses is presented in 

Section 2.5. 

2,3,7,8-TCDD has been shown to block some estrogenic effects both in vivo and in vitro, and the relative 

potencies of 2,3,7,8-TCDD and related congeners are consistent with their relative binding affinities with 

the Ah receptor (Safe et al. 1991). Estrogens are necessary for normal uterine development and for 

maintenance of the adult uterus. The mechanism of these antiestrogenic effects seems to be related to a 

decrease in gonadal tissue responsiveness to estrogen (DeVito et al. 1992) rather than to increased 

metabolism of estrogen. Studies in cultured MCF-7 cells (estrogen-responsive cells derived from a human 

breast adenocarcinoma) revealed that the antiestrogenic activity of 2,3,7,8-TCDD could result from the 

increased metabolism of estrogens due to Ah receptor-mediated enzyme induction and/or a decreased 

number of estrogen receptors in the nucleus (Gierthy et al. 1987; Harris et al. 1989a, 1990; Safe et al. 1991; 

Zacharewski et al. 1991, 1992). More recent data indicates that in some cases 2,3,7,8-TCDD may block the 

effects of estrogen through the ability of the 2,3,7,8-TCDD-bound Ah receptor-Arnt complex to interfere 

with the estrogen receptors binding to enhancer elements within the regulatory regions of 

estrogen-responsive genes (Krishnan et al. 1995; Safe 1995). Thus, the mechanism by which 

2,3,7,8-TCDD and related compounds may block certain effects of estrogen may be varied depending on 

the particular gene, response, and tissue. Under some conditions, 2,3,7,8-TCDD may also cause 

estrogen-like responses. For example, the treatment of mice with an appropriate dosage of 2,3,7,8-TCDD 

or estrogen results in thymic involution and modulation of particular bone marrow stem cell markers 

(Silverstone et al. 1994). However, the mechanism by which these compounds act are clearly different 

since potent antiestrogens block the effects of estrogen treatment without affecting 2,3,7,8-TCDD-elicited 
CDDs 241 

2. HEALTH EFFECTS 

responses (Frazier et al. 1994). Similarly, the effects of 2,3,7,8-TCDD on the development of external 

genitalia in rats are similar to the effects observed in animals exposed to potent estrogen-like chemicals 

(Gray and Ostby 1995). 

Extensive evidence suggests that the immune system is a sensitive target for toxicity of 2,3,7,8-TCDD and 

structurally related halogenated aromatic hydrocarbons (Kerkvliet 1995). Exposure to 2,3,7,8-TCDD can 

increase susceptibility to bacterial (Thigpen et al. 1975; Thomas and Hinsdill 1979; White et al. 1986), viral 

(Clark et al. 1983; House et al. 1990), parasitic (Tucker et al. 1986), and neoplastic disease (Luster et al. 

1980). However, the specific immunological functions affected by 2,3,7,8-TCDD in most of the hostresistance models have not been fully defined. Thymic involution is characteristic of exposure to 

2,3,7,8-TCDD and structurally related chemicals in all species examined. There is experimental evidence 

showing that immune suppression in rodents occurs at lower doses of 2,3,7,8-TCDD when the animals are 

exposed perinatally as compared with rodents exposed as adults, and that the prenatal effects are selective 

for T-cell-mediated immunity (Clark et al. 1983; Faith and Moore 1977; Vos and Moore 1974). The 

mechanism for 2,3,7,8-TCDD-induced thymic atrophy is not completely understood. There is evidence in 

rats suggesting that the 2,3,7,8-TCDD-induced effect is not mediated by an effect on the pituitary or adrenal 

glands, or from decreased production of thymic hormones (Van Logten et al. 1980; Vos et al. 1978). There 

appear to be multiple mechanisms involving alterations in thymocyte differentiation (Blaylock et al. 1992; 

Cook et al. 1987a; Denker et al. 1985; Greenlee et al. 1985), thymocyte proliferation (Lundberg et al. 

1990), and migration of lymphocyte stem cells (Fine et al. 1990). 

A commonly used assay for immunotoxicity is the suppression of the antibody response to SRBC. The 

magnitude of the anti-SRBC response depends on the interactions of antigen-presenting cells (i.e., 

macrophages), regulatory T-lymphocytes (i.e., helper and suppressor T cells), and B-lymphocytes (i.e., 

antibody-producing cells). Results from experiments in vivo suggested that the target for 2,3,7,8-TCDD in 

the antibody response to either SRBC or tumor cells is the T-cell and/or macrophage components rather 

than the B-cell (Kerkvliet and Brauner 1987; Kerkvliet et al. 1996). Although the effects of 2,3,7,8-TCDD 

on B-cell function in vivo have not been examined, in vitro studies suggest that 2,3,7,8-TCDD inhibits the 

terminal differentiation of B cells via alteration of an early activation event, perhaps increased protein 

phosphorylation and tyrosine kinase activity (Clark et al. 1991a; Kramer et al. 1987; Luster et al. 1988; 

Morris et al. 1991). Macrophage functions, examined ex vivo, generally have been found to be resistant to 

suppression by 2,3,7,8-TCDD (Mantovani et al. 1980; Vos et al. 1978). There is also evidence suggesting 
CDDs 242 

2. HEALTH EFFECTS 

that inflammatory cells may be activated by 2,3,7,8-TCDD via enhanced production of inflammatory 

mediators such as interleukin 1 and tumor necrosis factor (Clark et al. 1991b; Taylor et al. 1992). 

Extensive research has been conducted on the role of the Ah locus in immunotoxicity of 2,3,7,8-TCDD and 

related compounds, and overall, the data linking 2,3,7,8-TCDD-induced immunotoxicity to the Ah receptor 

are convincing. For example, Vecchi et al. (1983a) reported that the antibody response to SRBC was 

greatly suppressed by 2,3,7,8-TCDD in C57BL/6J mice, but not as much in DBA/2J mice. Results from 

structure-activity studies for the SRBC response with CDDs that contaminate technical grade pentachlorophenol supported an Ah receptor-mediated effect (Kerkvliet et al. 1985). The Ah receptor was also found to 

be involved in the suppression of the antibody response to lipopolysaccharide (Kerkvliet et al. 1990a); and 

the cytotoxic T-lymphocyte response, and suppression of the latter by dioxin-like PCBs, correlated with 

relative-binding affinities for the Ah receptor (Kerkvliet et al. 1990b). An additional response found to 

segregate with Ah-responsiveness was the cytotoxic response to activated neutrophils (Ackermann et al. 

1989). It is important to mention that results from some studies suggest that suppression of the in vitro 

antibody response may not be Ah receptor-mediated. For example, Holsapple et al. (1986a) reported that 

the magnitude of the response was comparable using cells from responsive mice relative to nonresponsive 

mice. Also, 2,7-dichlorodibenzo-p-dioxin, a congener with little affinity for the Ah receptor, was 

equipotent with 2,3,7,8-TCDD in suppressing the in vitro response. A similar conclusion was reached by 

Davis and Safe (1991), who found that a series of halogenated aromatic hydrocarbons, which had a 

>14,900-fold difference in in vivo immunotoxic potency, were equipotent in vitro in suppressing the antiSRBC response using cells from either responsive or nonresponsive mice. Although these results suggest a 

possible role of non-Ah receptor mechanisms, the studies fail to rule out a role of the Ah receptor. The 

variable effects of 2,3,7,8-TCDD in vitro may have been due to factors such as media components or 

procedures used to prepare cell suspensions. Kerkvliet (1994) suggested that "the difficulty in 

demonstrating consistent, direct effects of 2,3,7,8-TCDD in vitro on lymphocytes, the dependence of those 

effects on serum components, and the requirements for high concentrations of 2,3,7,8-TCDD are all 

consistent with an indirect mechanism of 2,3,7,8-TCDD on the immune system." One potentially important 

indirect mechanism operates through effects on the endocrine system. Glucocorticoids, sex steroids, T4, 

growth hormone, and prolactin have been shown to regulate immune responses, and 2,3,7,8-TCDD has 

been shown to alter the activity of all of them (see also sections on endocrine and reproductive effects). 

There is sufficient evidence that 2,3,7,8-TCDD is carcinogenic in animals, and the overall epidemiological 

database suggests that the incidence of certain types of cancer may be increased in humans by exposure to 
CDDs 243 

2. HEALTH EFFECTS 

2,3,7,8-TCDD (Hardell et al. 1994; Lucier et al. 1993a). The mechanism of 2,3,7,8-TCDD carcinogenicity 

has not been fully elucidated, but there is considerable evidence indicating that it does not involve direct 

damage to DNA through formation of DNA adducts. The criteria for designating 2,3,7,8-TCDD as a 

nongenotoxic carcinogen are based on the following: studies using extraordinarily sensitive analytical 

methods have been unable to detect DNA adducts in rodent tissue after exposure to 2,3,7,8-TCDD 

(Randerath et al. 1988; Turteltaub et al. 1990), numerous studies have demonstrated that 2,3,7,8-TCDD is 

not mutagenic in the Salmonella/Ames test with or without an activating system (Giri 1986; Wassom et al. 

1978), and 2,3,7,8-TCDD is a potent tumor promoter and a weak initiator or noninitiator in the two-stage 

models for liver (Flodstrom and Ahlborg 1989; Lucier et al. 1991; Pitot et al. 1980) and skin (Poland et al. 

1982). Instead, it has been proposed that 2,3,7,8-TCDD might alter the capacity of both exogenous and 

endogenous substances to damage the DNA by inducing CYP1A1- and CYP1A2-dependent drugmetabolizing enzymes. In some cases, enzyme induction will lead to increased formation of DNAdamaging metabolites, as appears to be the case in the two-stage model in rat liver and mouse skin 

(Flodstrom and Ahlborg 1992; Hebert et al. 1990; Poland and Knutson 1982; Poland et al. 1982). A recent 

study suggested that the induction of CYP1A1 may also lead to an increase in oxygen radicals and 

consequent oxidative DNA damage that could lead to mutation and cancer (Park et al. 1996). However, in 

many cases in which induction leads to increased rate of detoxification, the opposite will occur, as 

demonstrated by Cohen et al. (1979) for benzo[a]pyrene. The protection afforded by preinduction of 

CYP1A1 by 2,3,7,8-TCDD appears to be Ah receptor-mediated since it does not occur in mice deficient 

with low-affinity Ah receptor (Kouri et al. 1978). It should be noted that results from structure-activity 

studies for 2,3,7,8-TCDD and related compounds strongly suggest that the hepatocarcinogenic actions of 

2,3,7,8-TCDD are Ah receptor-dependent (Flodstrom and Ahlborg 1992; Hebert et al. 1990; Poland et al. 

1982; Poland and Knutson 1982). The role of CYP1A2 induction is less clear than for CYP1A1. Some 

have suggested that the liver carcinogenicity of 2,3,7,8-TCDD in intact female rats, but not male rats or 

ovariectomized female rats, could be explained in part by the formation of toxic catechol estrogens from 

17β-estradiol, a reaction catalyzed by CYP1A2 (Lucier et al. 1993a). This is also consistent with the 

finding that CYP1A2 is induced in liver but not in extrahepatic organs. 

The role of the EGF receptor in 2,3,7,8-TCDD-induced carcinogenicity has also been examined. EGF is a 

mitogen that stimulates the generation of mitotic signals in both normal and neoplastic cells, and its receptor 

and ligands have a variety of functions involved in cell transformation and tumorigenesis. It has been 

shown that 2,3,7,8-TCDD decreases the binding capacity of the plasma membrane EGF receptor for its 

ligand without changing the affinity constant (Abbott and Birnbaum 1990a; Hudson et al. 1985; Lin et 
CDDs 244 

2. HEALTH EFFECTS 

al. 1991a; Madhukar et al. 1984). The mechanism involved is not completely understood, but it appears 

that 2,3,7,8-TCDD does not decrease EGF receptor mRNA (Lin et al. 1991a). The effects of 2,3,7,8-TCDD 

on the EGF receptor have been shown to require the Ah receptor (Lin et al. 1991a). The EGF receptor-like 

response produced by 2,3,7,8-TCDD is consistent with the idea that 2,3,7,8-TCDD increases the generation 

of cellular mitotic signals which may, in part, be responsible for the tumor-promoting actions of 

2,3,7,8-TCDD. 

The possible role of UDP-glucuronyltransferases (UDPGT) on the carcinogenicity of 2,3,7,8-TCDD has 

also been studied. UDPGTs are thought to be a deactivation pathway for many environmental chemicals 

and endogenous substances such as steroid hormones by increasing their water solubility, thereby 

facilitating excretion by a conjugation reaction. 2,3,7,8-TCDD induces synthesis of at least one UDPGT 

isozyme (Lucier et al. 1986) by a Ah receptor-mediated mechanism (Bock 1991). For example, the 

oncogenic effect of prolonged stimulation of the thyroid by TSH has been attributed to decreased levels of 

T4 due to UDPGT induction. Decreased T4 levels induce the pituitary gland to respond by secreting 

increased amounts of TSH. This is consistent with results from rodent studies in which 2,3,7,8-TCDD and 

other inducers of UDPGT decreased T4 levels in blood, which is associated with increased TSH levels 

(Henry and Gasiewicz 1987). Lucier et al. (1986) showed that in rats, the shape of the dose-response curve 

for induction of UDPGT by 2,3,7,8-TCDD is similar to that of CYP1A1 induction. Kohn et al. (1996) 

constructed a physiologically based model to investigate the hypothesis that induction of UDPGT by 

2,3,7,8-TCDD may ultimately lead to a tumorigenic response in the thyroid of rats. The model included 

compartments for the thyroid and thyroxine-sensitive tissues, secretion and tissue uptake of thyroid 

hormones, binding of T3 and T4 to proteins in blood and tissues, iodination of iodothyronines, and 

glucuronidation of T4 by hepatic UDPGT. The model accurately predicted the effects of 2,3,7,8-TCDD on 

blood thyroid hormone concentrations, hepatic UDPGT activity, and the consequent increase of serum 

TSH. This was consistent with the observation that induction of UDPGT results in increased 

glucuronidation and biliary excretion of T4. The results of Kohn et al. (1996) provided further support to 

the hypothesis that induction of UDPGT is an early event in the generation of thyroid tumors by 

2,3,7,8-TCDD in the rat. 

There is evidence that some carcinogenic responses to 2,3,7,8-TCDD are related to effects of 

2,3,7,8-TCDD on the estrogen receptor (ER) and on estrogen metabolism. The responses appear to be 

tissue-specific as illustrated by the fact that in rats 2,3,7,8-TCDD increases liver tumor incidence, but 

decreases tumor incidence in mammary glands, the uterus, and pituitary gland (Kociba et al. 1978a). In 
CDDs 245 

2. HEALTH EFFECTS 

rats, a single dose of 2,3,7,8-TCDD decreases the binding capacity of the hepatic ER for estrogens 

(Romkes and Safe 1988; Zacharewski et al. 1991, 1992). This response seems to be Ah receptormediated since 2,3,7,8-TCDD was much more effective in decreasing hepatic ER binding in C57BL/6J 

mice (responsive strain) than in congenic mice with low-affinity Ah receptor (Lin et al. 1991b). 

2,3,7,8-TCDD also decreased rat hepatic ER in a 30-week-duration study (Clark et al. 1991b). The 

single ED50 for decreasing hepatic ER binding is similar to that for CYP1A1 induction, loss of plasma 

membrane EGF receptor, and induction of UDPGT (Lucier et al. 1993a). The relationship, however, 

between changes in concentration and cell proliferation have yet to be fully evaluated. In reproductive 

tract tissues, 2,3,7,8-TCDD decreases tumor incidences by a mechanism possibly involving increased 

estrogen metabolism as a consequence of UDPGT induction. Increased estrogen degradation was also 

observed in the MCF-7 breast cancer cell line after addition of 2,3,7,8-TCDD (Gierthy et al. 1988). 

As indicated above, a substantial body of evidence is consistent with the premise that the Ah receptor 

mediates the biological effects of 2,3,7,8-TCDD. Furthermore, this evidence indicates that a response to 

this chemical requires the formation of a ligand-receptor complex. 2,3,7,8-TCDD-receptor binding 

appears to obey the law of mass action and, therefore, depends on the concentrations of both ligand and 

receptor in the target cell, and the binding affinity of the ligand for the receptor. In principle, and 

according to the law of mass action, some active 2,3,7,8-TCDD-Ah receptor complexes may form even 

at very low levels of exposure. In reality, however, it is likely that at some finite concentration of 

2,3,7,8-TCDD, the formation of 2,3,7,8-TCDD-receptor complexes will be insufficient to elicit 

detectable or biologically relevant effects due to dependence on other factors (e.g., Arnt binding and 

DRE binding) and events (e.g., mRNA transcription and protein synthesis) necessary for the cascade of a 

signal transduction process to occur. Recent studies have indicated no evidence of a threshold for some 

relatively simple biochemical responses to 2,3,7,8-TCDD such as CYP1A1 induction (Kohn et al. 1993). 

However, this cannot yet be interpreted as an absence of a threshold since it is possible that either 

insufficiently low concentrations of 2,3,7,8-TCDD were used, or the background level of 2,3,7,8-TCDD 

equivalents (including any putative endogenous ligand) is already above the threshold level (even in 

experimental animals). Events leading to a toxic response that are subsequent to 2,3,7,8-TCDD-receptor 

binding and the induction of a particular biochemical event such as CYP1A1 activity, may or may not 

exhibit a linear response to 2,3,7,8-TCDD since these events are likely additionally dependent on 

multiple and complex biochemical, cellular, and tissue changes that may or may not be dependent on 

saturable processes. Further information will be required to determine if other responses to 

2,3,7,8-TCDD, both biochemical and biological, do or do not demonstrate threshold behavior. 
CDDs 246 

2. HEALTH EFFECTS 

It is generally accepted that the toxicity of CDDs, including 2,3,7,8-TCDD, is due mainly to the parent 

compound. Hydroxylated metabolites lack the activity of the parent compound, suggesting that 

metabolism is a detoxification process necessary for the biliary and urinary excretion of these 

compounds. For instance, dog metabolites of 2,3,7,8-TCDD administered to guinea pigs had at least 100 

times less acute toxic potency than the parent compound (Weber et al. 1982). In another study, two 

hydroxylated metabolites of 2,3,7,8-TCDD showed no significant effect on Ah receptor-mediated 

responses, such as body weight loss, thymus atrophy, or liver or spleen weight in male Wistar rats at 

doses as high as 5,000 μg/kg (Mason and Safe 1986a). One of the metabolites, 2-hydroxy-3,7,8-TrCDD, 

induced CYP1A1-related enzyme activities only at very high dose levels (1,000 and 5,000 μg/kg), 

whereas the other metabolite, 2-hydroxy-1,3,7,8-TCDD, lacked inducing capacity. Structure-activity 

studies of 7-substituted 2,3-CDDs, including the hydroxylated congeners, showed that the binding 

affinities of the hydroxylated congeners for the Ah receptor were significantly lower than those of the 

corresponding chlorine analogs (Denomme et al. 1985). Similar results were obtained in an additional 

study using hepatic cytosol from rat, mouse, hamster, and guinea pig (Romkes et al. 1987). 

There is some evidence that hydroxylated metabolites of CDDs interfere with the transport of T4 in 

blood by a mechanism unrelated to the Ah receptor (Lans et al. 1993, 1994). These investigators showed 

that hydroxy-CDDs with chlorine substitution adjacent to the hydroxy group (e.g., 7-hydroxy2,3,8-TrCDD, 2-hydroxy-1,3,7,8-TCDD, and 3-hydroxy-2,6,7,8-TCDD) showed similar or higher 

relative binding potency than T4 for the thyroid hormone transport protein transthyretin (TTR) in an in 

vitro assay using purified human TTR (Lans et al. 1993). In a subsequent study, they found that none of 

several hydroxylated CDDs tested inhibited T4 binding to thyroxin-binding globulin, the major 

T4-transporting plasma protein in humans, as opposed to TTR in rodents (Lans et al. 1994). This clearly 

indicated that hydroxylated CDDs may cause different effects in rodents and humans. 

The possibility exists that reactive epoxide intermediates of 2,3,7,8-TCDD that may be formed as a 

result of metabolism are involved in 2,3,7,8-TCDD-induced carcinogenicity by covalently binding to 

DNA. However, this appears unlikely since, as previously mentioned, studies using extraordinarily 

sensitive analytical methods have been unable to detect DNA adducts in rodent tissue after exposure to 

2,3,7,8-TCDD (Randerath et al. 1988; Turteltaub et al. 1990), and the fact that 2,3,7,8-TCDD is not 

mutagenic in the Salmonella/Ames test with or without an activating system (Giri 1986; Wasson et al. 

1977). 
CDDs 247 

2. HEALTH EFFECTS 

2.4.3 Animal-To-Human Extrapolations 

As discussed in the introduction to Section 2.1, there are a number of limitations in the human database; 

for most health effects, the data are inadequate to assess the potential for humans having a particular 

effect. Because the human data are incomplete, hazard and risk must be extrapolated across species. A 

large number of adverse effects have been observed in animals, and most have been observed in every 

experimental animal species tested, if the appropriate dose is administered. This is illustrated in Table 2­

8 for 8 major effects associated with CDD toxicity (acute lethality, hepatotoxicity, wasting syndrome, 

chloracne, immunotoxicity, reproductive toxicity, developmental toxicity, and cancer). With the 

exception of acute lethality in humans, positive responses have been observed in each tested species, 

when a response has been investigated. Despite the similarities in hazard response between different 

species, large species differences in sensitivity have been observed. Comparisons of species sensitivity 

demonstrate that no species is consistently sensitive or refractory for all effects and, for some effects, 

there is a small range of species sensitivity. As presented in Table 2-9, the range of LD50 values for 6 

commonly tested animal species spans several orders of magnitude. Guinea pigs have the lowest LD50 

value (0.6 μg/kg) and hamsters have the largest (1,157 μg/kg). However, if these outliers are removed, 

the range of LD50 values for mice, monkeys, rabbits, and rats is less than an order of magnitude 

(22–115 μg/kg). In contrast, the range of LOAELs for reproductive toxicity (abortions, resorptions, preand post-implantation losses) spans approximately an order of magnitude with rats (0.125 μg/kg) being 

the most sensitive and guinea pigs the least sensitive (1.5 μg/kg; NOAEL of 0.15 μg/kg). These data 

suggest that even though some effects have wide ranges of sensitivity, for most of the effects, the 

LOAELs for the majority of species cluster within an order of magnitude (Table 2-9). 

It is generally accepted that the Ah receptor plays a role in mediating many toxic responses attributed to 

exposure to CDDs (for additional information on the mechanisms of toxicity, see Section 2.4.2). For some 

responses, receptor binding appears necessary but may not be sufficient to result in downstream biological 

effects. Ah receptors have been found in most species, including humans, monkeys, rats, mice, hamsters, 

rabbits, and guinea pigs (Denison et al. 1986a; Landers and Bunce 1991). A simple way to explain 

sensitivity differences among species to 2,3,7,8-TCDD and related compounds, at least for Ah receptormediated responses, would be to assume that they are related to differences in receptor levels in target 

tissues and/or to differences in the affinity of binding of the specific CDD congeners. However, experimental data indicate that differences in such parameters cannot explain marked differences to CDD toxicity 

across species. For example, single dose LD50s range from 0.6 μg/kg in guinea pigs to 1,157 μg/kg in 


CDDs 250 

2. HEALTH EFFECTS 

hamsters, but the affinity with which 2,3,7,8-TCDD binds to the Ah receptor from guinea pigs is not 

significantly different from the affinity with which 2,3,7,8-TCDD binds to the hamster Ah receptor 

(Denison et al. 1986a). In addition, there are no significant differences in the level of the hepatic Ah 

receptor between the two species, suggesting that in addition to species differences in receptor levels and in 

their affinities for the ligand, differences in species sensitivity to 2,3,7,8-TCDD may be determined by some 

event or events occurring after the initial binding of 2,3,7,8-TCDD to the Ah receptor. These late events 

may involve a complicated interplay between genetic and environmental factors which may be key 

determinants of 2,3,7,8-TCDD biological potency and toxicity. Factors unrelated to the Ah receptor, such 

as toxicokinetic differences, may also account for some of the observed species differences (for additional 

information, see Section 2.4.2). The Ah receptor has been identified in many human tissues and human cell 

lines (Okey et al. 1994). However, considerable individual differences in the expression levels of both Ah 

receptor and Arnt mRNAs have been found in human tissues (Hayashi et al. 1994). Furthermore, based on 

findings in inbred mice, polymorphism in the Ah receptor probably exists in humans, so that a concentration of TCDD that produces a response in one individual may not do the same in another (Whitlock 

1993). This could explain why there was a wide range of serum 2,3,7,8-TCDD levels among Seveso 

residents where the occurrence of chloracne was sporadic over a generally wide range of doses (Mocarelli et 

al. 1991). 

The weight of evidence from animal species comparisons and mechanistic data indicates that caution should 

be exercised when extrapolating from animals to humans. Some theoretical models indicate a basis for 

extrapolating from animals to humans, but such models have not been validated; there is wide variation in 

the results of different models; and a great deal of uncertainty remains regarding whether valid, predictive 

extrapolations can be made. It is reasonable to assume that humans will not be the most sensitive responder 

or be refractory to all effects, and that they will have a wider range of response due to increased 

heterogeneity. Levels of exposure to CDDs that produce toxicity in experimental animals cannot be 

directly compared to levels associated with adverse health effects in humans because most epidemiologic 

studies do not provide adequate data to estimate CDD exposures in the studied populations. However, the 

CDD body-burden history can sometimes be estimated in epidemiology studies from reported serum or 

adipose concentrations and empirically based assumptions regarding whole-body elimination kinetics of 

CDDs (as discussed in the introduction to Section 2.1). Comparisons between estimated adverse body 

burdens of CDDs and related compounds (CDFs, PCBs) in experimental animals and humans have shown 

that humans and animals appear to respond to similar body burdens (DeVito et al. 1995). As presented in 

Table 2-10, the adverse effect levels identified in humans are typically within a factor of 10 of the body 

 

CDDs 252 

2. HEALTH EFFECTS 

burdens associated with similar effects in animals. The data in Table 2-10 should be interpreted cautiously 

and should not be taken to suggest that humans are more sensitive than experimental animals. The human 

body burdens were estimated from serum 2,3,7,8-TCDD levels measured many years after exposure 

termination using empirically based assumptions; small differences in one or more assumption can result in 

large differences in the estimated body burdens. For example, the body burden of 945 ng/kg in the Manz et 

al. (1991) study was estimated using a half-life of 8.5 years; if the half-life of 7.1 years were used, the 

estimated body burden would have been 1606 ng/kg. Additionally, individual serum 2,3,7,8-TCDD levels 

and length of time between exposure termination and measurement of serum 2,3,7,8-TCDD levels (latency) 

were not available for a few epidemiology studies, and mean serum levels and latency periods were used to 

estimate body burdens. The use of mean values rather than individual values and empirically based 

assumptions may have resulted in an over- or underestimation of actual body burdens. Conversely, in the 

animal studies, actual exposure levels were known and there is greater confidence in the estimated body 

burdens. An acute high-dose exposure would produce higher peak serum lipid 2,3,7,8-TCDD and target 

tissue levels than chronic exposure to lower levels, although the body burdens may be similar. Thus, it may 

be misleading to compare adverse-effect body burdens from acute studies to those identified in chronic 

studies. Another issue which needs to be considered in comparing the human and animal adverse-effect 

body burdens is that this is comparison of LOAELs not a comparison of threshold levels, and free-standing 

LOAELs may not accurately predict threshold levels. 

2.5 RELEVANCE TO PUBLIC HEALTH 

Overview 

The primary route of exposure to CDDs for the general population is the food supply. This type of 

exposure is the main contributor to the background exposure. Background exposure refers to exposure of 

the general population who are not exposed to readily identifiable point-sources of CDDs that result in 

widespread, low-level circulation of CDDs in the environment. It is generally accepted that the contribution 

of inhalation and direct contact with CDDs to the body burden of the general population is not more than a 

few percent. However, inhalation and direct contact represent major exposure routes in cases of 

occupational or accidental exposures. A background exposure level of approximately 0.7 pg 

2,3,7,8-TCDD/kg/day (assuming a 70 kg reference body weight) has been estimated for the general 

population in the United States (Travis and Hattemer-Frey 1987). If other CDD and CDF congeners are 

included, the background exposure level increases to approximately 18–192.3 pg TEQ/day (0.26–2.75 
CDDs 253 

2. HEALTH EFFECTS 

pg/kg/day using a 70 kg reference body weight) (Schecter et al. 1994b) (for additional information on TEQ, 

see the Toxic Equivalency Factor [TEF] and Toxic Equivalents [TEQ] subsections). The inclusion of 

dioxin-like PCBs further raises the estimate to 3–6 pg TEQ/kg/day (Beck et al. 1989a; WHO 1991). The 

average concentration of 2,3,7,8-TCDD in the adipose tissue of the U.S. population is 5.8 pg/g lipid (Orban 

et al. 1994). For all TEQ congeners, excluding dioxin-like PCBs, the national average was approximately 

28 pg TEQ/g lipid. In humans, the partitioning ratio of 2,3,7,8-TCDD between adipose tissue lipid and 

serum lipid is approximately 1 and remains near unity over at least a 1,000-fold concentration range over 

background levels (Patterson et al. 1988). This makes serum lipid an accurate and more practical measure 

of body burden than adipose tissue lipid. 

Data on human health effects of CDDs are derived from a variety of sources, including case reports and 

epidemiologic studies using case-control, cross-sectional, and cohort designs. While case-control and 

cohort studies have been used to investigate increases in the incidence of cancers among populations 

exposed to 2,3,7,8-TCDD, nonmalignant effects have been examined in cross-sectional medical studies. In 

many of the earliest studies, the magnitude of exposure-response relationships could not be adequately 

assessed for a number of reasons, including small sample size, poor participation, selection of inappropriate 

controls, the inability to identify confounding exposures, and short latency periods (especially important for 

assessment of cancer). A long interval between exposure and examination (up to 40 years in some cases) is 

a serious limitation when assessing noncancer responses since responses that resolve with time might not be 

detected at the time of the examination. On the other hand, health conditions that may be present at the 

time of examination may be totally unrelated to past exposure to 2,3,7,8-TCDD. An additional limitation 

was the inability to quantify exposure. However, serum or adipose tissue levels of 2,3,7,8-TCDD have 

been measured in more recent cross-sectional studies of U.S. chemical workers (Sweeney et al. 1989), 

Ranch Hand veterans (USAF 1991), and Missouri residents (Webb et al. 1989). Using a standard half-life 

equation and assuming a one-compartment model and first-order kinetics, the half-life for 2,3,7,8-TCDD in 

humans has been estimated to be 8.7 years (Michalek et al. 1996). By knowing the half-life, estimates of 

body burdens at the time of exposure can be back-calculated. These estimates, however, should be used 

with caution since little information exists regarding the metabolism of 2,3,7,8-TCDD in humans. In 

addition, there are considerable differences in the elimination half-lives for these chemicals among 

individual humans. 

For 2,3,7,8-TCDD, the majority of the effects have been reported among occupationally exposed 

individuals such as producers or users of chemicals in which 2,3,7,8-TCDD might have occurred as 
CDDs 254 

2. HEALTH EFFECTS 

impurities, and among residents of communities contaminated with 2,3,7,8-TCDD. Effects that have been 

associated with exposure to materials contaminated with 2,3,7,8-TCDD in some studies include cancer; 

dermal, hepatic and thyroid effects; effects on serum lipids; diabetes; and cardiovascular, respiratory, 

immunologic, neurologic, and reproductive effects. A number of studies have consistently found increases 

in cancer mortalities (all types combined) in the highest exposed workers with long latency periods, but the 

data on site-specific cancer are inconclusive. Among the dermal effects, chloracne is clearly a response 

associated with exposure to 2,3,7,8-TCDD and structurally related chemicals, but the threshold level of 

2,3,7,8-TCDD at which it occurs has not been established. Moreover, there seems to be a great deal of 

innate human variability in the chloracne response between individuals (see Section 2.1.2). Hepatic 

changes observed in exposed populations include hepatomegaly, increased hepatic enzyme (GGT, AST, 

ALT) levels, induced hepatic microsomal activity (measured as increased D-glucaric acid excretion), 

alterations in porphyrin metabolism, and increases in serum lipid (cholesterol, triglycerides) levels. With 

the exception of long-lasting changes in GGT (Calvert et al. 1992; USAF 1991) and in serum cholesterol 

(USAF 1991) in some exposed groups, hepatic effects were transient and appeared to have been associated 

with acute exposure to high 2,3,7,8-TCDD concentrations. Few long-term thyroid effects were found in 

Ranch Hand veterans (USAF 1991), but a recent study of nursing infants suggests that ingestion of breast 

milk containing CDD and CDF levels somewhat higher than those reported in most general population 

studies, may alter thyroid function (these data are not conclusive because the measure thyroid hormone 

levels were within the normal range) (Koopman-Esseboom et al. 1994; Pluim et al. 1992). Slightly 

increased risk of diabetes and abnormal glucose tolerance tests have been reported in populations exposed 

to high 2,3,7,8-TCDD concentrations (Sweeney et al. 1992; USAF 1991). In the former study, however, 

age and body mass index, both known risk factors for diabetes, appear to have a greater influence than 

2,3,7,8-TCDD level. Dose-related trends for deaths from cardiovascular disease and ischemic heart 

disease were observed in individuals exposed to CDDs during the BASF accident (Flesch-Janys et al. 

1995). However, other studies found no relationship between 2,3,7,8-TCDD exposure and cardiovascular 

deaths (Bertazzi et al. 1989b) or other cardiovascular effects (Hoffman et al. 1986; Wolfe et al. 1985). A 

few case reports indicate that acute exposure to high 2,3,7,8-TCDD levels can produce respiratory 

irritation, but there is no indication that exposure to 2,3,7,8-TCDD produces chronic respiratory effects. 

Although there have been some reports of alterations in some immune end points in populations exposed 

to 2,3,7,8-TCDD, there has not been a consistent pattern, and the clinical significance of the effects is not 

totally clear. The overall evidence for neurologic effects suggests that although neurologic effects are 

reported to have occurred shortly after exposure in occupationally exposed individuals, even high exposure 

to 2,3,7,8-TCDD caused no long-term sequelae (Goetz et al. 1994; Sweeney et al. 1993). More recent data 
CDDs 255 

2. HEALTH EFFECTS 

suggests that exposure to 2,3,7,8-TCDD and related chemicals in humans during the pre- and neonatal 

periods may affect neurological development (Huisman et al. 1995a), but these data need to be interpreted 

cautiously because the neurological optimality score in infants was within the normal range and CDD/CDF 

levels may have only contributed a small amount to the variance in scores. Of the many reproductive end 

points studied in populations exposed to 2,3,7,8-TCDD, the available data provide suggestive evidence of 

altered sex ratios in children of exposed parents (Basharova 1996; Dimich-Ward et al. 1996; Mocarelli et 

al. 1996) and possibly alterations in reproductive hormone levels in males (Egeland et al. 1994) are 

associated with increased serum 2,3,7,8-TCDD levels. 

Most of the toxicity studies of 2,3,7,8-TCDDs in animals have involved oral exposure, and numerous 

effects have been documented after short- and long-term exposure including lethality, and cardiovascular, 

gastrointestinal, hematological, hepatic, renal, endocrine, dermal, body weight, immunologic, 

reproductive, and developmental effects. In addition, 2,3,7,8-TCDD is a potent carcinogen in various 

species, and produces tumors in multiple sites in rodents of both sexes. However, as shown in Table 2-2, 

these effects occurred at doses several orders of magnitude higher than estimates of background exposure 

for CDDs. The most reliable and consistent sign of 2,3,7,8-TCDD toxicity in experimental animals is 

weight loss or decreased weight gain in growing rodents. Animal responses to 2,3,7,8-TCDD exposure are 

species- and strain-dependent, although almost all responses can be induced in every species and strain if 

the appropriate dose is used. The animal data suggest that the most sensitive effects of 2,3,7,8-TCDD 

exposure are immunotoxicity, and reproductive and developmental toxicity. 

In recent years considerable advances have been made regarding the mechanisms of toxicities of 

2,3,7,8-TCDD and related chemicals, as well as the pharmacokinetics of dioxins in experimental animals. 

For CDDs, toxicity and toxicokinetics cannot be dealt with separately. Based on results from research in 

these fields, it has become apparent that the comparison of responses from animals to humans (or even 

between animal species) should be done on the basis of body-burden or target-tissue dose, rather than on 

the basis of administered dose. By doing so, species-specific toxicokinetic considerations such as dosedependent distribution, the existence of tissue-specific sequestering chemical entities (i.e., CYP1A2), and 

body composition (i.e., percent fat) can be taken into account. A discussion of relationships between 

administered dose, body burden, and biological responses is presented below. 

Issues relevant to children are explicitly discussed in Section 2.6, Children’s Susceptibility, and Section 

5.6, Exposures of Children. 
 

CDDs 256 

2. HEALTH EFFECTS 

Toxic Equivalency Factors (TEFs) and Toxic Equivalents (TEQs). Humans are exposed to 

complex mixtures of CDDs and other halogenated aromatic hydrocarbons such as CDFs and PCBs which 

are found in the environment (including food). The toxicological concerns resulting from exposure to 

these mixtures, as well as the gaps in available information with which to evaluate the potential risks from 

such exposures, led the EPA Chlorinated Dibenzo-p-dioxins/Chlorinated Dibenzofurans Technical Panel 

of the Risk Assessment Forum to recommend an interim method for assisting in estimating the risk from 

exposure to these mixtures that can be used until the data gaps are filled (Barnes 1991; EPA 1989e). Since 

for many of these chemicals very limited data on toxicity exist, TEFs were developed and validated in 

studies in animals (Eadon et al. 1986; Silkworth et al. 1989a; Viluksela et al. 1998a, 1998b). 

The TEF approach involves assessment of the comparative effects of individual halogenated aromatic 

hydrocarbons congeners on various biological end points and derivation of TEFs based on the upper range 

of potency data for these effects The key assumptions unifying the diverse types of data that are 

considered in the derivation of TEFs are: that congeners exert toxicity through a common receptormediated mechanism, and that the effects of mixtures are additive (Safe 1990). The TEF approach 

compares the relative toxicity of individual congeners to that of 2,3,7,8-TCDD, which is the most 

extensively studied of the halogenated aromatic hydrocarbons that interact with the Ah receptor. The TEF 

for 2,3,7,8-TCDD is defined as unity; and TEFs for all other CDD congeners, CDFs, and dioxin-like PCBs 

are less than one, thus reflecting their lower toxic potency (see Kennedy et al. [1996] for an exception to 

this general rule). TEFs proposed earlier by EPA (1989) are presented in Table 2-11; recently revised 

values are presented in Table 2-12 (WHO 1998). The recent update also assigned a TEF of 1 to 1,2,3,7,8­

PeCDD. The toxic potency of a mixture of congeners (i.e., the TEQ) is the sum of the products of the 

TEFs for each congener and its concentration in the mixture. Thus, TEQs represent 2,3,7,8-TCDD toxic 

equivalents for mixtures of CDDs, CDFs, and/or dioxin-like PCBs. 

The TEF approach facilitates site-specific assessments that account for changes in congener composition 

due to differential environmental partitioning and transformation, as well as differences in congener profiles 

between sites and co-exposure to related halogenated aromatic hydrocarbons. The TEF approach, however, 

has several shortcomings. One problem is that very little data may be available for estimating the TEF and 

the available data are often from in vitro or single-exposure acute in vivo studies. Furthermore, there is a 

wide range in relative potency estimates derived from the literature. For example, Safe (1990) estimated 

2,3,7,8-TCDD/1,2,3,4,7,8-HxCDD potency ratios of 33/1 for rat body weight loss, 12/1 for rat thymic 

atrophy, and 8/1 for AHH induction in cultured rat liver cells. One further problem is that 


CDDs 259 

2. HEALTH EFFECTS 

differences in pharmacokinetics between two chemicals result in different estimates of the relative potency 

depending upon the exposure protocol (DeVito and Birnbaum 1995). These investigators demonstrated that 

published TEFs for 2,3,7,8-TCDD (TEF 1) and 2,3,7,8-TCDF (TEF 0.1) accurately estimated the relative 

potencies for liver EROD induction in female B6C3F1 mice after 4 weeks of treatment, but failed to do so 

after 13 weeks of treatment. The inability to estimate relative potencies after the longer treatment duration 

was attributed to the difference in half-lives between the two compounds (2 days for 2,3,7,8-TCDF and 

15 days for 2,3,7,8-TCDD). Steady-state levels of 2,3,7,8-TCDF were achieved within 4 weeks and, thus, 

EROD remained constant from 4 to 13 weeks. Steady-state levels of 2,3,7,8-TCDD were not attained 

within 4 weeks, which explained the increased hepatic EROD between 4 and 13 weeks. The results showed 

that TEFs for congeners with a short half-life may overestimate their potency and that the opposite may be 

true for congeners with a long half-life. Based on their results, DeVito and Birnbaum (1995) suggested that 

“present TEFs should be reevaluated to determine whether values have adequately incorporated 

pharmacokinetic differences between the test compound and 2,3,7,8-TCDD.” In a more recent study, the 

same group of investigators compared the relative potencies for enzyme induction of 2,3,7,8-TCDD, 

1,2,3,7,8-PeCDD, 2,3,7,8-TCDF, 1,2,3,7,8-PeCDF, 2,3,4,7,8-PeCDF, OCDF, and 2,3,7,8­

tetrabromodibenzo-p-dioxin (2,3,7,8-TBDD) in mice based on daily administered or final tissue dose 

following gavage dosing for 90 days (DeVito et al. 1997). The enzymes monitored were EROD (liver, 

lung, and skin) and ACOH (liver). After the 90-day administration period, the chemicals were assumed to 

be at or approaching steady-state conditions. Since ED50 values could not be estimated for all the 

congeners, the authors used an alternative method of comparison that fitted a function to the 2,3,7,8-TCDD 

dose-response data. The function was then used to predict the 2,3,7,8-TCDD equivalent dose of a chemical 

based on the enzymatic activity induced at a given dose of the test compound. A linear regression of the 

predicted dose of 2,3,7,8-TCDD and the actual congener dose provided the relative potency estimate. The 

results showed that, when based on administered dose, the relative potencies for the specific congeners did 

not vary substantially among tissues. However, for congeners with a much shorter half-life than 

2,3,7,8-TCDD, the relative potencies increased for all enzymes when estimated from tissue concentrations. 

For example, for 2,3,7,8-TCDF the relative potency based on administered dose varied by less than a factor 

of 2 between endpoints, ranging from 0.0076 for skin EROD to 0.014 for ACOH. The relative potency 

increased by 4- to 14-fold when based on tissue dose and varied between tissues by a factor of 4, from 

0.028 to 0.11. Because 2,3,7,8-TCDF is metabolized much faster than 2,3,7,8-TCDD, to achieve an 

equivalent tissue concentration of these chemicals, higher doses of 2,3,7,8-TCDF must be administered 

relative to 2,3,7,8-TCDD. Overall, the results confirmed their previous observations that differences in 

absorption and metabolism modulate the relative potency of this class of chemicals. DeVito et al. (1997) 
CDDs	 260 

2. HEALTH EFFECTS 

suggested that it might be useful to derive two sets of TEF values, one used for estimating intake 

equivalents and the other for estimating tissue equivalents. 

Viluksela et al. (1998a, 1998b) recently examined a wide range of endpoints in rats administered either a 

mixture of CDDs with a given TEQ or single CDD congeners at the same TEQ dose level as the mixture. 

The TEFs for the various congeners were derived from acute experiments. The dosing period was 

13 weeks. The results showed effects of similar magnitude in response to administration of the CDD 

mixture or single CDD congeners. This supported the validity of the TEF method and the notion of 

additive toxicity for the CDDs evaluated. Moreover, the concentration ratios for the various congeners in 

the liver were very similar to the ratios at which the congeners were administered. 

Neubert and coworkers (Neubert et al. 1992c) have also examined the issue of TEFs and proposed that a 

number of prerequisites need to be fulfilled in order to consider the TEF approach from a scientific point of 

view: 

•	 The actions of the congeners must be strictly additive in the dose range to be evaluated. 

•	 The organotropic manifestations in different species must be identical over the relevant dose ranges. 

•	 Dose-response curves for various toxicological end points for a given congener must run parallel. 

•	 The dose-response curves for a given toxicological end point must run parallel for the various 

congeners. 

•	 For extrapolations between species, the kinetics must be identical, or differences have to be taken into 

consideration. 

•	 With respect to a risk assessment relevant to humans, toxic or biological manifestations in the lower 

dose ranges are of special interest, and LD50 or ED50 values or effects induced by highly toxic dose are 

of minor importance. 

•	 Effects to be expected at low exposures must be identical with those observed at the high doses 

studied. 

After discussing each one of these seven points, Neubert et al. (1992c) concluded that the toxicological 

background for using the TEF approach for risk assessment must be increased considerably. A similar 

conclusion was reached by a scientific panel that examined the feasibility of developing a TEF approach 

that would be applicable to PCB mixtures (Barnes et al. 1991). In the case of PCBs, the study group 

concluded that “the application of TEF approach for PCBs would be less straightforward than it was in the 

case of chlorinated dibenzo-p-dioxins and dibenzofurans.” 

In 1992, a group of scientists met in Belgium under the auspices of the European Environmental Research 

Organization to discuss the impact of CDDs, CDFs, and PCBs on human and environmental health with 

special emphasis on application of the TEF concept. The main conclusions, relevant to the TEF concept, 
CDDs 261 

2. HEALTH EFFECTS 

were that TEFs may be useful for risk management (i.e., quantitative estimation of Ah receptor-mediated 

toxic potential) of mixtures of CDDs, CDFs and the coplanar non-ortho and mono-ortho PCBs, but that the 

TEF concept is not applicable for the various toxic responses whose mechanisms do not involve the Ah 

receptor (Ahlborg et al. 1992, 1994). 

The TEF approach in relation to cancer risk estimation has also been examined Rao and Unger (1995). 

First, the authors used the standard approach of multiplying TEF doses by the cancer slope factor for 

2,3,7,8-TCDD to estimate lifetime incremental cancer risks for a mixture of CDDs and CDFs. This method 

was compared with a modified approach in which the TEF dose was adjusted for differences in the 

probability of formation of bound receptor-ligand complexes. Briefly, using algorithms from a competitive 

binding model, the fractions of Ah receptor bound to congeners were derived. This fraction was defined as 

competitive binding ratios (CBR) in mixtures and represents the maximum likelihood estimate for the 

formation of a congener-receptor bound complex in the presence of other competing ligands. Two distinct 

risk scenarios were used for comparison: (1) human adipose tissue residue data from the national human 

adipose tissue survey (Stanley et al. 1986) were used to generate potential lifetime incremental cancer risks, 

and (2) lifetime cancer risk was characterized for a potentially exposed population ingesting contaminated 

carp. In the modified TEF approach, CBR values for individual congeners computed from the competitive 

binding algorithms were used to derive the tissue concentrations. The main findings of this analysis were 

that TEF doses calculated by using the model algorithms were lower than the combined TEF dose for all 

congeners estimated by the TEF method without considering the competitive binding. In addition, the 

combined incremental cancer risks for all congeners were generally lower when model algorithms were 

used in the dose-response analysis. Also, the standard TEF method tended to overestimate the risks of 

higher congeners with low toxicity and underestimated the risk of more toxic congeners. 

One further concern regarding the use of TEQs for risk assessment is the fact that the human diet also 

contains Ah receptor agonists, such as indole-3-carbinol and related compounds in vegetables, polynuclear 

aromatic hydrocarbons (PAHs), aromatic amines formed during cooking. These natural Ah receptor 

agonists elicit responses in humans that are consistent with a receptor-mediated pathway, but the responsespecific potencies of natural Ah receptor versus xenodioxins are unknown. Therefore, a TEF/TEQ 

approach based solely on intake of xenodioxins does not take into account the background of natural 

dioxins that may influence responses associated with persistent low-level occupation of the Ah receptor (see 

Safe 1998a, 1998b for review on this issue). 
 

C

CDDs 262 

2. HEALTH EFFECTS 

Minimal Risk Levels for 2,3,7,8-TCDD 

It is ATSDR’s policy (see Appendix B) to use health guidance values (i.e., MRLs, EMEGs) derived for 

2,3,7,8-TCDD for other dioxin-like compounds, expressed in total TEQs. 

Inhalation MRLs 

MRLs were not derived for inhalation exposure. 

Oral MRLs 

An MRL of 0.0002 (2×10-4) μg/kg/day has been derived for acute-duration oral exposure 

(14 days or less) to 2,3,7,8-TCDD. 

The acute duration oral MRL was based on a NOAEL of 0.005 μg/kg and a LOAEL of 0.01 μg/kg for 

immunological effects in female mice (Burleson et al. 1996). In this study, groups of 20 female B6C3F1 

mice were administered a single gavage dose of 0, 0.001, 0.005, 0.01, 0.05, or 0.1 μg/kg 2,3,7,8-TCDD in 

corn oil. Seven days after 2,3,7,8-TCDD exposure, the mice were infected intranasally with influenza 

A/Hong Kong/8/68 (H3N2) virus diluted at 10-48, 10-50, 10-52, or 10-54. In a separate experiment, groups of 

18 female mice received a single gavage dose of 0, 0.001, 0.01, or 0.1 μg/kg 2,3,7,8-TCDD and were 

infected 7 days later with influenza A virus at a dose not known to cause mortality (10-54 and 10-58) or were 

sham-infected. Body weight, thymus weight, and wet lung weights were measured 3, 9, or 12 days 

postinfection. Pulmonary virus titers were determined in groups of 72 mice exposed to 0, 0.001, 0.01, or 

0.01 μg/kg 2,3,7,8-TCDD and infected with influenza A virus seven days later. For the virus titer study, 

groups of mice were killed 2 hours, 1, 4, 6, 7, 8, 9, 10, and 11 days post-infection. 

Statistically significant increases in mortality were observed in the influenza A infected mice exposed to 

0.01, 0.05, or 0.1 μg/kg 2,3,7,8-TCDD. However, no between group differences in mortality were observed 

at these 2,3,7,8-TCDD dosages. Mortality in mice receiving 0.001 or 0.005 μg/kg did not significantly 

differ from the mortality in the control group. Exposure to 2,3,7,8-TCDD did not enhance the increase in 

relative lung weight normally seen in mice infected with influenza A virus. As compared to controls, no 

significant alterations in thymus weights were observed in 2,3,7,8-TCDD-exposed mice sham-infected or 

those infected with influenza A virus. 2,3,7,8-TCDD exposure did not result in a significant increase in 

viral titers in the lung, as compared to titers from the control group. The authors noted that the 
C

CDDs 263 

2. HEALTH EFFECTS 

lack of dose-response in mortality and the lack of effect on the relative lung weight, thymus weight, and 

viral titers suggest that 2,3,7,8-TCDD may be exerting an effect via an indirect mechanism such as through 

an effect on cytokines. The 0.005 μg/kg dose was considered a NOAEL for immunotoxicity. As described 

in the footnote to Table 2-2, an uncertainty factor of 30 (3 for extrapolation from animals to humans and 10 

for human variability) and modifying factor of 0.7 (to adjust for the difference in higher bioavailability of 

2,3,7,8-TCDD from an oil gavage vehicle than from food) were used to derive the MRL from the NOAEL 

value. 

An MRL of 0.00002 (2×10-5) μg/kg/day has been derived for intermediate-duration oral exposure 

(15–364 days) to 2,3,7,8-TCDD. 

The intermediate-duration oral MRL was based on a NOAEL of 0.0007 μg/kg/day for immunological 

effects in Hartley guinea pigs fed 2,3,7,8-TCDD in the diet for 90 days (DeCaprio et al. 1986). In that 

study, groups of weanling Hartley guinea pigs (10 per sex) were administered a diet that provided an 

average of 0.0001, 0.0007, 0.005, or 0.028 μg 2,3,7,8-TCDD/kg/day. This corresponds to 2, 10, 76, and 

430 ppt 2,3,7,8-TCDD in the food. A control group was fed a diet without added 2,3,7,8-TCDD. The 

recovery following treatment was studied in groups of 10 guinea pigs fed a diet containing 430 ppt 

2,3,7,8-TCDD for 11, 21, or 35 days and allowed to recover for 79, 69, or 55 additional days, respectively. 

The highest dietary level of 2,3,7,8-TCDD caused net body weight loss and mortality. Four males and four 

females died, and additional animals had to be sacrificed due to poor health. Food consumption was 

significantly reduced in the highest-dose group only. Body weight gain in the 0.0007 and 0.005 μg/kg/day 

male groups was reduced by 9 and 20%, respectively. In the corresponding female groups, body weight 

gain was reduced by 6 and 15%. Gross lesions were observed only in the highest-dose group and included 

thymic atrophy, depletion of body fat, and liver enlargement. Significant changes in organ weights included 

a decrease in absolute kidney weight and in absolute and relative thymus weight in males dosed with 

0.005 μg/kg/day, increase in relative liver weight in males and females at the 0.005 μg/kg/day level, and 

increase in relative brain weight in males at this same dose level. Organ weights from high-dose animals 

were not monitored. Administration of 2,3,7,8-TCDD did not cause any significant hematological effect 

(blood was not collected from the highest-dose group). In the 0.005 μg/kg/day groups, serum ALT was 

significantly reduced in females, and triglycerides were elevated in males. No other significant changes in 

clinical chemistries were observed. Treatment-related histological alterations were observed only in the two 

higher-dose groups and consisted of hepatocellular cytoplasmic inclusion bodies and atrophy of the thymic 

cortex. In the recovery study there was 10% mortality in the groups treated for 11 and 21 days, and 70% 

mortality in the group treated for 35 days. Surviving animals in all groups exhibited significantly 
C

CDDs 264 

2. HEALTH EFFECTS 

reduced body weight gain. The 0.0007 μg/kg/day dose represents a NOAEL for decreased thymus weight, 

and the 0.005 μg/kg/day dose is a LOAEL. As described in the footnote to Table 2-2, an uncertainty factor 

of 30 (3 for extrapolation from animals to humans and 10 for human variability) was used to derive the 

MRL from the NOAEL. 

An MRL of 0.000001 (1×10-6) μg/kg/day has been derived for chronic-duration oral exposure 

(365 days or more) to 2,3,7,8-TCDD. 

The chronic-duration oral MRL is based on a LOAEL of 0.00012 μg/kg/day for developmental toxicity in 

rhesus monkeys (Schantz et al. 1992). In this study, groups of 8 female rhesus monkeys were fed a diet 

containing 0, 5, or 25 ppt 2,3,7,8-TCDD for a total of 16.2 months (the results of the neurodevelopmental 

portion of this study were published in papers by Bowman et al. [1989a], Schantz and Bowman [1989], 

Schantz et al. [1992]). After 7 months of exposure, the monkeys were mated with unexposed males. (Only 

1 monkey in the 25 ppt group delivered a viable offspring; this offspring was not studied behaviorally.) The 

monkeys were fed the 2,3,7,8-TCDD diet throughout the mating period, gestation, and lactation. When the 

offspring (3 males and 3 females per exposure group) were 8.6 months of age, they were placed in 3 peer 

groups of 4 monkeys and allowed to play for 1.5 hours without interference. The peer groups consisted of 

two 2,3,7,8-TCDD-exposed monkeys and two control monkeys. Behavioral patterns (social interactions 

and other behaviors such as vocalization, locomotion, self-directed behavior, and environmental 

exploration) were monitored 4 days a week for 9 weeks. No overt signs of toxicity were observed in the 

mothers or offspring, and birth weights and growth were not adversely affected by 2,3,7,8-TCDD exposure. 

Significant alterations were observed in play behavior, displacement, and self-directed behavior in the 

2,3,7,8-TCDD-exposed offspring. 2,3,7,8-TCDD-exposed monkeys tended to initiate more rough-tumble 

play bouts and retreated less from play bouts than controls, were less often displaced from preferred 

positions in the playroom than the controls, and engaged in more self-directed behavior than controls. No 

other significant alterations in behavior or alterations in reflex development, visual exploration, locomotor 

activity, or fine motor control were found (Bowman et al. 1989a). In tests of cognitive function, object 

learning was significantly impaired, but no effect on spatial learning was observed (Schantz and Bowman 

1989). As described in the footnote to Table 2-2, an uncertainty factor of 90 (3 for the use of a minimal 

LOAEL, 3 for extrapolation from animals to humans, and 10 for human variability) was used to derive the 

MRL. 

It should be also noted that 10 years after termination of 2,3,7,8-TCDD exposure in the Schantz et al. 

(1992) study, Rier et al. (1993) reported a dose-related increase in the incidence and severity of 
 

CDDs 265 

2. HEALTH EFFECTS 

endometriosis in these same rhesus monkeys. Rier et al. (1993) identified a less serious LOAEL of 5 ppt 

(0.00012 μg/kg/day) for moderate endometriosis. However, monkeys appear to be more susceptible to 

endometriosis, based on a background incidence of endometriosis in monkeys of 30% (Rier et al. 1993) 

compared to a background incidence of 10% in humans (Wheeler 1992). Thus, derivation of a chronic oral 

MRL based on endometriosis would necessitate using an uncertainty factor of less than 1 (or at most, 1) to 

account for the increased sensitivity of monkeys to endometriosis as compared to humans. If the Rier et al. 

(1993) study were used to calculate an oral MRL, the LOAEL of 0.00012 μg/kg/day would be divided by 

an uncertainty factor of 100 (10 to extrapolate from a LOAEL, 10 for human variability and 1 for 

interspecies differences). This would result in a computed MRL essentially the same as the chronic oral 

MRL of 1 pg/kg/day based on developmental toxicity as described in the preceding paragraph. Moreover, 

(1) the clinical history for these rhesus monkeys during the 10-year period between the Schantz et al. (1992) 

study and examination by Rier et al. (1993) is unknown (not reported); (2) Boyd et al. (1995) did not find 

an association between exposure to CDDs, CDFs, or PCBs and endometriosis in a clinical study in women; 

and (3) the EPA (1997) concluded that “the evidence for supporting the hypothesis that CDDs and PCBs 

are causally related to human endometriosis via an endocrine-disruption mechanism is very weak.” So, 

even though there is information to indicate that endometriosis may also be a sensitive toxicological end 

point for 2,3,7,8-TCDD exposure, the developmental end point (altered social behavior) reported in the 

Schantz et al. (1992) study was determined to be the most appropriate end point for derivation of an MRL 

for chronic oral 2,3,7,8-TCDD exposure. 

Comparison of Estimated Body Burdens Associated with Effects in Experimental Animals 

and Humans. Estimated average body burdens of 2,3,7,8-TCDD in human populations in which 

various health effects of 2,3,7,8-TCDD are suspected range from 31 to 6,600 ng/kg (estimated body 

burdens at the time of exposure termination). See Table 2-1 for more information. The human body burden 

expected in populations exposed to background environmental levels of 2,3,7,8-TCDD has been estimated 

to be 1 ng TCDD/kg body weight (DeVito et al. 1995; Orban et al. 1994). This would suggest that effects 

of 2,3,7,8-TCDD in humans may occur at body burdens that are 30 to 6,600 times greater than background 

burdens for 2,3,7,8-TCDD. 

The similarities in response of humans and experimental animals to similar body burdens of CDDs and 

related chemicals (Table 2-10), along with our understanding of common mechanisms of actions of CDDs 

in humans and experimental animals lends support to both the relevance of experimental animal toxicology 

to humans and the use of experimental animal data for establishing MRLs (see Section 2.4.3 for more 
 

CDDs 266 

2. HEALTH EFFECTS 

information on the animal-to-human extrapolations). Acute, intermediate and chronic MRLs for 

2,3,7,8-TCDD were derived from experimental animal studies that identified the highest NOAELs or lowest 

LOAELs for the respective exposure-duration category (Table 2-2). The MRLs, and the NOAELs and 

LOAELs on which they are based, have been converted to their corresponding equivalent body burdens and 

are compared to background 2,3,7,8-TCDD body burdens in humans and body burdens associated with 

adverse health effects (Table 2-13). This comparison shows that the NOAELs for acute, intermediate, and 

LOAEL for chronic exposure to 2,3,7,8-TCDD in experimental animals are generally below the low end of 

the 31–6,600 ng TCDD/kg of body weight range of body burdens that is suspected to be adverse to humans 

based on the epidemiologic evidence. Body burdens that correspond to the MRLs are 2-6 times lower than 

the estimated background body burden in humans, suggesting that adverse health effects are unlikely in 

humans exposed to background levels of 2,3,7,8-TCDD. However, because of the magnitude of uncertainty 

in dose-response relationships for 2,3,7,8-TCDD, the possibility that current background exposures may be 

sufficient to contribute to a risk of adverse health effects in human populations cannot be completely 

excluded. 

Death. Information regarding mortality in humans after exposure to CDDs is limited to epidemiological 

studies in populations exposed occupationally or environmentally (Bertazzi et al. 1989b; Cook et al. 1986, 

1987b; Fingerhut et al. 1991; Ott et al. 1980, 1987; Pesatori et al. 1998; Thiess et al. 1982; Vena et al. 

1998; Wolfe et al. 1985; Zack and Suskind 1980; Zober et al. 1990). These studies did not find a 

significant increase in the overall mortality rate in populations exposed to 2,3,7,8-TCDD or other CDD 

congeners for acute or chronic durations. However, several studies did find significant increases in causespecific mortality (i.e., cancer and cardiovascular disease). These increases in cause-specific mortality are 

discussed under the specific effect. 

Several studies provided data regarding lethality following CDDs exposure in animals. Oral LD50 values 

for 2,3,7,8-TCDD were calculated in rats (NTP 1982b; Schwetz et al. 1973; Walden and Schiller 1985), 

minks (Hochstein et al. 1988), rabbits (Schwetz et al. 1973), guinea pigs (McConnell et al. 1984; Schwetz et 

al. 1973), and hamsters (Henck et al. 1981) following gavage doses in corn oil or corn oil:acetone vehicle. 

Doses that produced death were in the μg/kg range. Differences in the susceptibility to the lethality of 

2,3,7,8-TCDD were observed not only among different species, but also among different strains 

CDDs 268 

2. HEALTH EFFECTS 

within the same species, and even in the same strain of rat bred in different laboratories (Walden and 

Schiller 1985). The use of rats of different ages may have played a role in the interlaboratory differences in 

susceptibility between rats of the same strain. Toxicity results from acute- and intermediate-duration 

categories indicated that the guinea pig is the most sensitive species to 2,3,7,8-TCDD toxicity leading to 

death (DeCaprio et al. 1986; McConnell et al. 1984; Schwetz et al. 1973; Vos et al. 1973), and that the 

hamster is the most resistant (Hanberg et al. 1989; Henck et al. 1981). Experiments with mice that were 

injected with 2,3,7,8-TCDD intraperitoneally showed that the C57BL/6J mice responsive to 2,3,7,8-TCDDinduced toxicity were twice as sensitive to 2,3,7,8-TCDD-induced lethality as the less-responsive DBA/2J 

strains (Gasiewicz et al. 1983a). Increased mortality was also recorded in mice following intermediateduration dermal exposure to 2,3,7,8-TCDD (NTP 1982a). 

Toxicity data in animals indicated that similar effects occur after exposure to CDDs by oral, dermal, or 

parenteral routes. Toxicokinetic data in mice showed that 2,3,7,8-TCDD hepatic levels were similar 

following oral, intraperitoneal, and subcutaneous exposure (Nau and Bass 1981). However, recent data in 

rats showed that intratracheal administration of a 2,3,7,8-TCDD dose resulted in a relatively higher 

accumulation of 2,3,7,8-TCDD in the liver than after oral administration of the same dose (Diliberto et al. 

1996). Intraperitoneal administration of 2,3,7,8-TCDD was less toxic than oral dosing in acute-exposure 

experiments with hamsters (Olson et al. 1980a). 

Following acute oral exposure to 2,3,7,8-TCDD, death occurred within 6–42 days depending on the dose 

and species tested, indicating a delayed type of toxicity. Death was usually preceded by significant weight 

loss in all duration categories. However, weight loss did not appear to be the only cause of death. A total 

parenteral nutrition fluid given to 2,3,7,8-TCDD-exposed rats and guinea pigs protected the animals against 

wasting syndrome, but not against 2,3,7,8-TCDD-induced lethality (Gasiewicz et al. 1980; Lu et al. 1986). 

Specifically, biochemical changes indicative of severe liver damage were found in moribund rats. 

2,3,7,8-TCDD was found not only to be more toxic than its isomer 1,2,3,4-TCDD (Courtney 1976) but also 

more toxic than any other congener tested (2,7-DCDD, 1,2,3,7,8-PeCDD, 1,2,3,6,7,8-HxCDD, 

1,2,3,7,8,9-HxCDD, 1,2,3,4,6,7,8-HpCDD, and OCDD) (Courtney 1976; NCI/NTP 1980; NTP 1982b; 

Viluksela et al. 1994, 1998a). LD50 values for acute oral exposure to a mixture of 1,2,3,6,7,8-HxCDD and 

1,2,3,7,8,9-HxCDD derived for rats and mice (NCI/NTP 1980) were higher by more than 2 orders of 
 

CDDs 269 

2. HEALTH EFFECTS 

magnitude than LD50 values for 2,3,7,8-TCDD. However, 1,2,3,4,6,7,8-OCDD and 2,7-DCDD did not 

cause death in mice even at doses as high as 4,000 and 2,000 mg/kg, respectively. 

Systemic Effects. 

Respiratory Effects. No exposure-related respiratory effects were found in a group of Air Force Vietnam 

veterans exposed to 2,3,7,8-TCDD during aerial spraying studied sometime after exposure (Wolfe et al. 

1985). No respiratory effects clearly attributable to 2,3,7,8-TCDD have been found in workers potentially 

exposed (Calvert et al. 1991). In rhesus monkeys, intermediate-duration exposure to a lethal oral dose of 

2,3,7,8-TCDD caused nose bleeding (McNulty 1984), hemorrhage, and hyperplasia of the bronchial 

epithelium (Allen et al. 1977). Bronchiolar adenomatoid changes were seen in mice chronically exposed 

dermally (NTP 1982a). Furthermore, hyperplastic changes in the lungs were recorded in rats exposed orally 

to a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980). However, mostly negative 

results were obtained in other oral studies in animals regardless of duration of exposure (Holsapple et al. 

1986b; Kociba et al. 1978a; NCI/NTP 1979a; NTP 1982a, 1982b). The relevance of the animal findings to 

human health is unclear. Intense acute exposure to 2,3,7,8-TCDD can produce respiratory irritation, but the 

findings from controlled epidemiologic studies do not support an association between 2,3,7,8-TCDD 

exposure and chronic respiratory disease. It should be noted, however, that chronic bronchitis and related 

effects were observed in many Yusho and Yu-Cheng patients, who were exposed to the structurally related 

CDFs (ATSDR 1994). 

Cardiovascular Effects. While some studies have found an association between CDD exposure and 

cardiovascular disease, most studies have not found a clear association between exposure to 2,3,7,8-TCDD 

and diseases of the heart and circulatory system (Bond et al. 1983; Calvert et al. 1998; Hoffman et al. 1986; 

Moses et al. 1984; Reggiani 1980; Suskind and Hertzberg 1984; Wolfe et al. 1985). However, human 

studies have suffered from limitations such as examination of the cohorts after exposure has ended, thus 

allowing for tissue repair to occur; lack of good exposure data; and inability to examine the relationship 

between serum 2,3,7,8-TCDD levels and cardiovascular disease in most studies. In the Ranch Hand study 

(USAF 1991), a weak association was found between decreased mean diastolic blood pressure, cardiac 

arrhythmias, and decreases in peripheral pulses and exposure to 2,3,7,8-TCDD. Bertazzi et al. (1989b) and 

Pesatori et al. (1998) found increases in deaths from ischemic heart disease and cardiovascular disease in 

men and chronic rheumatic heart disease in women in the 10-year period following the Seveso accident. 

However, the authors attributed these findings to post-accident stress rather than the 2,3,7,8-TCDD 
CDDs 270 

2. HEALTH EFFECTS 

exposure. In workers exposed occupationally to 2,3,7,8-TCDD and other CDD congeners at the Boehringer 

Hamburg plant, a statistically significant trend for increased risk of cardiovascular disease and ischemic 

heart disease mortalities with increasing serum lipid levels of 2,3,7,8-TCDD or TEQ (CDDs and CDFs) was 

found (Flesch-Janys et al. 1995). An international study of more than 20,000 workers followed from 1939 

to 1992 found an increased risk for death from cardiovascular disease, especially ischemic heart disease, 

among exposed workers, but the authors did not rule out the influence of risk factors such as cigarette 

smoking, high fat diet, obesity, physical inactivity, and serum lipids (Vena et al. 1998). 

Experiments in animals demonstrated that exposure to relatively high doses of 2,3,7,8-TCDD can cause 

various pathophysiological effects. Acute oral exposure of rats increased the basal tension of the left 

cardiac atria (Kelling et al. 1987) or decreased the basal rate for spontaneous beating, depending on the dose 

used (Hermansky et al. 1988; Kelling et al. 1987). Reduced blood pressure and increased myocardial 

peroxidase activity were also recorded. All of the effects on the heart in these two studies were attributed to 

a hypothyroid condition caused by near-lethal doses (Hermansky et al. 1988). Other studies have suggested 

a direct effect of 2,3,7,8-TCDD on cardiac muscle, as for example an intraperitoneal injection of 

2,3,7,8-TCDD to guinea pigs reduced the contractive responsiveness of isolated myocardium (Brewster et 

al. 1987; Canga et al. 1988). 

Reports of histological findings are few. Myocardial degenerative changes were reported in rats after 

chronic oral exposure to a lethal dose of 2,3,7,8-TCDD (Kociba et al. 1978a), and hemorrhages were 

reported in monkeys after intermediate-duration dietary exposure to near-lethal doses (Allen et al. 1977). 

However, most studies did not find any histopathological changes in rats and mice following chronic oral 

exposure to 2,3,7,8-TCDD (NTP 1982b), 2,7-DCDD (NCI/NTP 1979a), or a mixture of 1,2,3,6,7,8-HxCDD 

and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980). Similarly, no changes were reported after chronic dermal 

exposure of mice to 2,3,7,8-TCDD (NTP 1982a). There is no conclusive evidence that the cardiovascular 

system is a target for 2,3,7,8-TCDD toxicity. 

Gastrointestinal Effects. Limited information is available regarding gastrointestinal effects of 

2,3,7,8-TCDD in humans. Earlier studies of individuals with exposure to substances contaminated with 

2,3,7,8-TCDD found significant elevations in self-reported ulcers (Bond et al. 1983; Suskind and Hertzberg 

1984), but a study of Vietnam veterans (USAF 1991) failed to find such effect. A more recent crosssectional medical study of workers employed more than 15 years earlier in the production of 
CDDs 271 

2. HEALTH EFFECTS 

2,3,7,8-TCDD-contaminated chemicals found no association between 2,3,7,8-TCDD exposure (body 

burden) and gastrointestinal disease (Calvert et al. 1992). 

Only a few of the numerous animal studies found any effects. Gastrointestinal ulcerations were reported in 

minks after an acute oral exposure to a lethal dose of 2,3,7,8-TCDD (Hochstein et al. 1988), and 

hemorrhages were reported in rats following chronic exposure (Van Miller et al. 1977). Ileitis and 

peritonitis were observed in hamsters receiving a single lethal dose of 2,3,7,8-TCDD (Olson et al. 1980a). 

A trophic effect on the antral mucosa was found in 2,3,7,8-TCDD treated rats, in contrast to atrophy found 

in pair-fed control animals (Theobald et al. 1991). Although these authors attempted to relate the 

mechanism of action to hormonal effects, a definitive mechanism was not established. Changes progressing 

from epithelial hyperplasia and metaplasia of gastric mucosa to stomach ulcerations were observed in 

rhesus monkeys with prolonged oral exposure (Allen et al. 1977; McConnell et al. 1978a; McNulty 1984). 

These data indicate that primates are particularly sensitive to 2,3,7,8-TCDD-induced gastrointestinal 

toxicity; however, the effects are seen at doses which caused severe toxicity at multiple sites. In contrast, 

most studies in rodents did not find any gastrointestinal effects after oral or dermal exposure to 

2,3,7,8-TCDD (Christian et al. 1986a; Henck et al. 1981; Kociba et al. 1978a; NTP 1982a, 1982b), or after 

oral exposure to a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980) and 2,7-DCDD 

(NCI/NTP 1979a). The available information suggests that the gastrointestinal tract is not a target for 

2,3,7,8-TCDD toxicity in humans. 

Hematological Effects. Limited human data were located regarding hematological effects following 

exposure to CDDs. Increases in leukocyte and platelet counts were reported in Vietnam veterans involved 

in Operation Ranch Hand (USAF 1991), which suggested the presence of a low-level, chronic inflammatory 

response related to higher levels of 2,3,7,8-TCDD exposure. Increased prevalence of high white blood cell 

counts was found in a population exposed to 2,3,7,8-TCDD in the environment, but the increase was not of 

clinical importance (Hoffman et al. 1986), and other epidemiological studies reported negative results 

(Stehr et al. 1986; Wolfe et al. 1985). 

Increased packed-cell volume was found in guinea pigs following a single intraperitoneal injection of 

2,3,7,8-TCDD; however, this was considered to be secondary to progressive dehydration of exposed 

animals with decreased water consumption (Gasiewicz and Neal 1979). Several studies in animals reported 

hematological effects after intermediate-duration exposure to 2,3,7,8-TCDD. Among the effects observed 

were reduced leukocytes in guinea pigs at 0.001 μg/kg/day (Vos et al. 1973), thrombocytopenia and 
CDDs 272 

2. HEALTH EFFECTS 

hemoconcentration in rats at 0.8–1 μg/kg/day (Viluksela et al. 1994; Zinkl et al. 1973), and anemia and 

bone marrow hypoplasia in rhesus monkeys at 0.1 μg/kg/day (McNulty 1984). Reduced erythrocytes was 

reported in rats chronically exposed to 0.1 μg 2,3,7,8-TCDD/kg/day in the feed (Kociba et al. 1978a). 

Splenic changes include reduced germinal centers after acute (Christian et al. 1986a) and splenic atrophy 

after chronic (Van Miller et al. 1977) oral exposures in rats. Furthermore, splenic hyperplasia was found in 

rats orally exposed to a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD for an intermediate duration 

(NCI/NTP 1980). Whether or not the splenic changes were secondary to hematopoietic effects is unclear. 

No hematological effects were found after acute oral exposure in minks (Hochstein et al. 1988) and mice 

(Holsapple et al. 1986a), intermediate oral exposure in guinea pigs (DeCaprio et al. 1986), chronic oral 

exposure in rats and mice (NTP 1982a; Oughton et al. 1995), chronic dermal exposure in mice (NTP 

1982a), and acute intraperitoneal exposure in rats (Mason and Safe 1986a). Similarly, no effects were 

reported in rodents exposed chronically to 2,7-DCDD (NCI/NTP 1979a) or a mixture of 1,2,3,6,7,8­

HxCDD and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980) by the oral route, but thrombocytopenia was reported in 

male Sprague-Dawley rats exposed for 13 weeks to 1,2,3,4,6,7,8-HpCDD (Viluksela et al. 1994). 

Decreased hematocrit and reduced platelet counts were reported in rats administered 1,2,3,7,8-PeCDD or 

1,2,3,4,7,8-HxCDD for 13 weeks at dose levels that caused lethality (Viluksela et al. 1998a). 

No clear picture regarding hematologic effects of 2,3,7,8-TCDD emerges from the studies in animals. From 

the limited data, it appears, however, that mice are less sensitive than other species. The relevance of the 

findings in animals to human health is difficult to ascertain. 

Musculoskeletal Effects. No relevant information was located regarding musculoskeletal effects in 

humans exposed to CDDs. However, evidence from case reports in the Yu-Cheng incident (which involved 

oral exposure to the structurally related CDFs and PCBs) indicate that musculoskeletal effects may occur 

after oral exposure to CDDs. Guo et al. (1994) reported that Yu-Cheng children were smaller and had less 

total lean mass and soft-tissue mass compared to matched control subjects. Hemorrhages in the 

musculoskeletal system of monkeys were observed following an oral intermediate-duration exposure to 

2,3,7,8-TCDD (Allen et al. 1977). However, monkeys in this experiment were in terminal stages, and 

hemorrhages were found in several other systems. There is no evidence that would indicate that the 

musculoskeletal system is a target for 2,3,7,8-TCDD toxicity in humans or in animals. 
CDDs 273 

2. HEALTH EFFECTS 

Hepatic Effects. Exposure to 2,3,7,8-TCDD induces liver microsomal enzymes in both humans and 

animals, regardless of the route or duration of exposure. Increased urinary δ-glucaric acid (UGA) excretion, 

an indirect index of enzyme induction, was found in children with chloracne living in the Seveso area 

following the 1976 industrial accident (Ideo et al. 1982). Biochemical changes (increased cholesterol and 

bilirubin levels, induced GGT and ALT activities) indicated liver effects in exposed humans (Hoffman et al. 

1986; Mocarelli et al. 1986). Biochemical changes indicative of a subclinical effect on lipid metabolism 

were found in Vietnam veterans involved in Operation Ranch Hand (USAF 1991). Biochemical 

examinations found disorders in the metabolism of porphyrins, lipids, carbohydrates, and plasma proteins in 

workers exposed to 2,3,7,8-TCDD during the manufacture of herbicides (Jirasek et al. 1976; PazderovaVejlupkova et al. 1981). In addition, histopathological changes (steatosis, fibrosis) were also documented. 

A more recent and better-designed study of workers employed at 2 chemical plants in the manufacture of 

sodium trichlorophenol and its more than 15 years earlier derivatives found no evidence of an elevated risk 

for long-term clinical hepatic disease (Calvert et al. 1992). Exposure was assessed by measuring lipidadjusted serum 2,3,7,8-TCDD levels, and exposed workers had serum 2,3,7,8-TCDD levels significantly 

higher than unexposed controls. The negative findings, however, are not necessarily inconsistent with 

results from earlier studies, but suggest that hepatic effects observed in humans immediately after exposure 

probably resolve with time. A follow-up study of the same cohort found a positive association between 

serum 2,3,7,8-TCDD levels and the concentration of triglycerides and a negative correlation with HDL 

cholesterol; these associations were small when compared with the influence of many other factors (Calvert 

et al. 1996). 

Studies in animals have shown that exposure to 2,3,7,8-TCDD can induce hepatotoxicity in several species 

administered the chemical by various exposure routes for several exposure durations. The severity of the 

lesion is dependent not only on the species, but also on the strain. In general, it appears that rats exhibit 

more signs of hepatotoxicity than guinea pigs and hamsters. Histological alterations of the liver are 

common findings observed in animals exposed to 2,3,7,8-TCDD. These effects have been reported after 

acute exposure in rats (Christian et al. 1986; Hermansky et al. 1988), mice (Greig 1984; Greig et al. 1987; 

Kelling et al. 1985), and guinea pigs (Turner and Collins 1983); after intermediate-duration exposure in rats 

(NTP 1982b; Van Miller et al. 1977), mice (Thigpen et al. 1975), guinea pigs (DeCaprio et al. 1986), and 

monkeys (Allen et al. 1977; McNulty 1984); and after chronic exposure in rats (Kociba et al. 1978a). 

Hepatic lesions in rats are characterized by degenerative and necrotic lesions with the appearance of 

mononuclear cell infiltration, multinucleated giant hepatocytes, and increased number of mitotic figures and 

intracytoplasmic lipid droplets. Markers of hepatic damage such as serum ALT and AST activities usually 
CDDs 274 

2. HEALTH EFFECTS 

increase in animals that exhibit altered liver histology (Greig 1984; Rosenthal et al. 1989; Smith et al. 

1981). DBA/2J mice developed hepatic necrosis and inflammation without fatty changes after acute 

intraperitoneal exposure to 2,3,7,8-TCDD (Shen et al. 1991). Only slight lipid accumulation was found 

after exposure to a high dose (600 μg/kg). In contrast, severe fatty changes were observed in C57BL/6J 

mice, indicating that the steatitic effect may depend on the Ah locus. Histological lesions may be severe 

enough to be a contributing factor in death. Dose-related increases in intracellular and paracellular 

permeability of the biliary tree was observed in rats administered ip doses of 2,3,7,8-TCDD (Davidson and 

Fujimoto 1987). 

2,3,7,8-TCDD has been found to be porphyrogenic in both rats and mice (Cantoni et al. 1981; Goldstein et 

al. 1977, 1982; Jones and Sweeney 1977, 1980). The mechanism of induction of porphyria is not known. 

2,3,7,8-TCDD is a potent inducer of the initial and rate-limiting enzyme involved in heme synthesis, ALAsynthetase, but no increased activity was seen in mice which exhibited porphyria after treatment with 

2,3,7,8-TCDD for 11 weeks (Jones and Sweeney 1980). A more likely explanation is that the primary 

event in 2,3,7,8-TCDD-induced porphyria is inhibition of hepatic porphyrinogen decarboxylase (Jones and 

Sweeney 1980). Crossbreeding experiments have shown that porphyrinuria was inherited together with 

AHH inducibility (Jones and Sweeney 1980), indicating that the Ah locus is involved in the porphyrogenic 

response to 2,3,7,8-TCDD. 

Enzyme induction is one of the most sensitive responses to 2,3,7,8-TCDD exposure, and has been one of 

the most extensively studied biochemical responses produced by 2,3,7,8-TCDD. The MFO system is the 

most thoroughly investigated, and AHH and EROD (CYP1A1 markers) and acetanilide-4-hydroxylase 

(ACOH) (CYP1A2 marker) are the most frequently assayed enzyme activities. The lowest single oral dose 

of 2,3,7,8-TCDD shown to induce AHH activity in rats was 0.002 μg/kg (Kitchens and Woods 1979). 

Similarly, the induction of EROD was observed in the liver in Wistar rats (Abraham et al. 1988) and 

C57BL/6 mice (Harris et al. 1990) following subcutaneous and intraperitoneal injection, respectively. In 

female B6C3F1 mice, administration of a single oral dose of $0.1 μg 2,3,7,8-TCDD/kg significantly 

increased liver, lung, and skin EROD activities and liver acetanilide-4-hydroxylase activity (CYP1A2 

marker) (Diliberto et al. 1995). In the three tissues examined, induction of EROD was dose-dependent. 

Also in B6C3F1 female mice, repeated oral administration of doses as low as 1.5 ng 2,3,7,8-TCDD/kg day 

significantly increased liver, lung, and skin EROD activities and liver acetanilide-4-hydroxylase activity 

(DeVito et al. 1994). In both studies (DeVito et al. 1994; Diliberto et al. 1995), liver, lung, and skin 

exhibited different sensitivities for enzyme induction. In male C57BL/6J and DBA/2J mice, the ED50 
CDDs 275 

2. HEALTH EFFECTS 

values for induction of hepatic EROD after a single dose of 2,3,7,8-TCDD were 1.1 and 16 μg/kg, 

respectively (Weber et al. 1995). In an intermediate-duration dietary study in Sprague-Dawley rats, doses 

as low as 0.014 μg/kg/day induced both EROD and acetanilide-4-hydroxylase (Van Birgelen et al. 1995). 

Enzyme induction is a reversible process dependent on the dose and the dosing regime (Fan and Rozman 

1995; Li and Rozman 1995). In male C57BL/6N mice, Pegram et al. (1995) showed that induction of 

acetanilide-4-hydroxylase by 2,3,7,8-TCDD was age-dependent, as it was significantly greater in old than 

in young mice. 

In addition to altering the activities of enzymes from the MFO system in the liver, 2,3,7,8-TCDD also alters 

the activities of some key liver enzymes of the intermediary metabolism. These effects are intimately 

related with the wasting syndrome as discussed below (see Body Weight Effects). For example, 

2,3,7,8-TCDD decreased the activities of hepatic PEPCK and G-6-Pase (key enzymes of gluconeogenesis) 

in mice and rats (Fan and Rozman 1995; Li and Rozman 1995; Viluksela et al. 1994; Weber et al. 1995) 

and also reduced the activity of TdO (key enzyme of tryptophan metabolism) in rats (Li and Rozman 1995; 

Viluksela et al. 1994), but not in mice (Weber et al. 1995). 

Vitamin A (retinol) is essential for normal growth and cell differentiation, particularly for epithelial cells. 

2,3,7,8-TCDD has been shown to decrease the storage of vitamin A in rodents. Decreased ability to store 

vitamin A (retinol) was found in rats and guinea pigs; however, partial recovery of the retinol content by 

week 16 postexposure was reported only in rats. A single oral dose of 2,3,7,8-TCDD caused a 70% 

reduction in the liver storage of retinol in rats when measured 2 months postexposure (Thunberg et al. 

1979). The reduction was dose-related within the dose range studied (0.1–10 μg/kg) (Thunberg et al. 

1980). Reduction of hepatic retinol by 2,3,7,8-TCDD was greater (87%) in younger rats with lower initial 

weights (Thunberg et al. 1984) than in more mature rats (60%) (Thunberg et al. 1979, 1980). In a 13-week 

dietary study in female Sprague-Dawley rats, dose of $0.014 μg/kg/day produced a dose-dependent 

reduction in hepatic retinol (Van Birgelen et al. 1995). In addition to reducing hepatic retinol storage, 

2,3,7,8-TCDD exposure induced the activity of UDPGT and AHH in the rat, results obtained with other 

compounds (polychlorinated phenols, methylcholanthrene, phenobarbital) suggest that the effect is not 

mediated by the Ah receptor (Thunberg et al. 1984). Experiments with Sprague-Dawley rats showed that 

pretreatment with 2,3,7,8-TCDD influenced not only the storage, but also urinary and fecal excretion of a 

subsequent dose of radioactively labeled retinyl acetate. Retinol-content was also altered in various tissues; 

liver, intestine, and epididymis content decreased by 39–53, 19–67, and 18–44%, respectively, while renal 

content increased 3–30 times (Håkansson et al. 1989b). The 2,3,7,8-TCDD pretreated rats, though not 
CDDs 276 

2. HEALTH EFFECTS 

retinol-deficient, used a subsequently administered dose of retinyl acetate in a manner similar to retinoldeficient animals. 

Mild-to-moderate hepatic effects of 2,3,7,8-TCDD exposure were also found after intermediate-duration 

dermal exposure in mice (Hebert et al. 1990; NTP 1982a), after acute intratracheal instillation in rats 

(Nessel et al. 1990, 1992), after acute intraperitoneal exposure in rats (DiBartolomeis et al. 1986; Mason 

and Safe 1986a), hamsters (Olson et al. 1980a), and guinea pigs (Gasiewicz and Neal 1979; Holcomb et al. 

1988; Lu et al. 1986), and after acute subcutaneous exposure in mice (Courtney 1976). Some effects 

(increased plasma albumin, total protein) were considered to be secondary to progressive dehydration of 

exposed animals with decreased water consumption (Gasiewicz and Neal 1979). 

Hepatotoxicity was observed in rats and mice chronically exposed by gavage to a mixture of 1,2,3,6,7,8­

HxCDD and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980). Liver effects were also found in rats exposed by diet to 

HpCDD (Viluksela et al. 1994), 1,2,3,7,8-PeCDD and 1,2,3,4,7,8-HxCDD (Viluksela et al. 1998b), and 

OCDD for subchronic durations (Birnbaum et al. 1989a; Couture et al. 1988), and to 2,7-DCDD for a 

chronic duration (NCI/NTP 1979a). 

It is clear that the liver is a target for 2,3,7,8-TCDD toxicity in animals. In humans, hepatic alterations have 

been observed sometimes following exposure to high 2,3,7,8-TCDD levels. In general, the effects are mild 

and transient, which might explain the negative findings of Calvert et al. (1992). 

Renal Effects. The overall evidence from studies of populations exposed to high concentrations of 

2,3,7,8-TCDD suggests that the kidney is not a target for 2,3,7,8-TCDD toxicity in humans (Stehr et al. 

1986; USAF 1991; Wolfe et al. 1985). 

Likewise, the kidney is not a target organ in adult animals. No effects were found in mice exposed acutely 

(Holsapple et al. 1986a; Weber et al. 1995) and in rats exposed chronically (Kociba et al. 1978a; NTP 

1982b) to 2,3,7,8-TCDD by the oral route or in mice exposed by the dermal route (NTP 1982a). Similarly, 

no changes were found in rodents exposed chronically to 2,7-DCDD (NCI/NTP 1979a) or a mixture of 

1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980) by the oral route. 

Renal effects which were seen at near-lethal doses and considered secondary to frank toxicity included pale 

kidneys in minks after acute oral exposure to 2,3,7,8-TCDD (Hochstein et al. 1988), and enlarged 
CDDs 277 

2. HEALTH EFFECTS 

convoluted tubules and Bowman's spaces together with epithelial hyperplasia in rats (Christian et al. 1986) 

and monkeys (McConnell et al. 1978a) orally exposed for acute durations and in monkeys orally exposed 

for intermediate durations (Allen et al. 1977). An increased incidence of renal inflammatory changes 

recorded in mice after chronic oral exposure to 2,3,7,8-TCDD (NTP 1982b) was not a primary effect. 

Decreased glomerular filtration rate (Anaizi and Cohen 1978; Pegg et al. 1976) and increased tubular 

filtration rate (Anaizi and Cohen 1978) were reported in rats treated with a single intraperitoneal dose of 

2,3,7,8-TCDD. The authors concluded that observed renal effects were probably secondary to the general 

toxic reaction to 2,3,7,8-TCDD (Pegg et al. 1976). However, renal effects (mainly hydronephrosis) were 

found in pups of 2,3,7,8-TCDD-exposed pregnant rodents (Abbott et al. 1987a, 1987b; Courtney 1976; 

Schwetz et al. 1973) indicating special ability of CDDs to induce effects in the developing kidneys. 

Endocrine Effects. No biochemical evidence of thyroid dysfunction, as evaluated by serum levels of T4, 

triiodothyronine, and TSH, were reported in a group of 18 workers examined 17 years after an industrial 

accident during the manufacture of 2,4,5-T (Jennings et al. 1988). The small sample size, the fact that no 

measure of exposure was provided, and the long period of time between exposure and examination preclude 

any conclusion regarding possible effects of 2,3,7,8-TCDD. Zober et al. (1994) found a significant increase 

in the incidence of thyroid disease (no further details provided) 35 years after the BASF accident. An 

increased incidence of diabetes and subclinical decreases in thyroid function were found in Vietnam 

veterans who participated in operation Ranch Hand (USAF 1991). 

A strong positive association was found between glucose intolerance or increased risk of diabetes and 

2,3,7,8-TCDD serum levels (USAF 1991). The diabetes finding remained significant even after adjusting 

for body fat. Furthermore, subclinical effects in thyroid function (significant decrease in mean T3 % uptake 

and increases in mean TSH) were reported for Operation Ranch Hand veterans with high 2,3,7,8-TCDD 

serum levels (USAF 1991). However, the magnitude of the differences was not considered physiologically 

significant. Diabetes and glucose intolerance were also found in workers exposed occupationally 

(Pazderova-Vejlupkova et al. 1981; Sweeney et al. 1992). However, in the Sweeney et al. (1992) study, age 

and body mass index, both known risk factors for diabetes, appear to have a greater influence on the 

increase in both the risk of diabetes and elevated fasting serum glucose levels than 2,3,7,8-TCDD level. A 

follow-up study of Operation Ranch Hand veterans confirmed earlier findings of glucose abnormalities and 

increased risk of diabetes mellitus in exposed subjects (Henriksen et al. 1997). Furthermore, a follow-up of 

Seveso residents found a significant increase in deaths from diabetes among women from zone B (Pesatori 

et al. 1998). 
CDDs 278 

2. HEALTH EFFECTS 

The evidence available from epidemiological studies suggests that exposure to high concentrations of CDDs 

may induce long-term alterations in glucose metabolism and subtle alterations (of unknown clinical 

relevance) in thyroid function. 

Numerous studies in rodents have reported alterations in thyroid status after exposure to 2,3,7,8-TCDD. 

End points commonly examined included serum levels of T4, T3, TSH, and activity of hepatic microsomal 

UDPGT, an enzyme which increases glucuronation of T4 and clearance. The effects on T4 levels are dosedependent and also appear to be species-dependent. For example, serum T4 was decreased in rats after 

acute (Bastomsky 1977; Fan and Rozman 1995; Gorski and Rozman 1987; Henry and Gasiewicz 1987; 

Hermansky et al. 1988; Potter et al. 1986) and intermediate exposure (Li and Rozman 1995; Sewall et al. 

1995; Van Birgelen et al. 1995; Viluksela et al. 1994). In contrast, the response of serum T3 levels ranged 

from increased (Bastomsky 1977; Potter et al. 1986) to no change or inconsistent change (Fan and Rozman 

1995; Henry and Gasiewicz 1987; Potter et al. 1983; Sewall et al. 1995) to decreased (Pazdernik and 

Rozman 1985). In hamsters, a species less susceptible to 2,3,7,8-TCDD toxicity than rats, T4 serum levels 

were increased by 2,3,7,8-TCDD even though hepatic microsomal UDPGT activity was significantly 

increased, suggesting that mechanisms other than induction of this enzyme must account for the speciesspecific alterations in T4 (Henry and Gasiewicz 1987). A further species-specific response was noted by 

Weber et al. (1995) who reported that in C57BL/6J and DBA/2J mice, both T4 and T3 levels were 

decreased in a parallel fashion as a result of a single dose of 2,3,7,8-TCDD. According to Weber et al. 

(1995), the decrease in T3 in mice would reduce the de novo synthesis of fatty acids, thus improving the 

balance of metabolic energy which might explain, at least in part, the reduced susceptibility of mice to 

2,3,7,8-TCDD toxicity compared to rats. It is interesting to note that Long Evans rats exhibited a doserelated increase in T4 and T3 90 days after single doses of 2,3,7,8-TCDD that decreased T4, but did not 

significantly alter T3 4 days after dosing (Fan and Rozman 1995). The significance of this finding was not 

entirely clear, but according to the authors it indicated that sustained effects of 2,3,7,8-TCDD on thyroid 

homeostasis trigger adaptive responses which persist even after most of the 2,3,7,8-TCDD has been cleared. 

Decreased levels of serum T4 have also been reported in rats administered other CDD congeners such as 

1,2,3,7,8-PeCDD and 1,2,3,4,7,8-HxCDD in intermediate exposure duration studies (Viluksela et al. 

1998b). 

The reduction in circulating T4 levels observed in rats appears, in part, to be due to the increased activity of 

UDPGT (Bastomsky 1977; Sewall et al. 1995), but other possibilities have also been discussed. McKinney 

et al. (1985b) proposed that T4 and T3 might be endogenous ligands for the Ah receptor, and 
CDDs 279 

2. HEALTH EFFECTS 

that 2,3,7,8-TCDD might be an agonist for the T4 receptor. Although some evidence was presented in 

support for this hypothesis, the weight of evidence, summarized by Goldstein and Safe (1989), argues 

against 2,3,7,8-TCDD being a thyroid agonist. Lans et al. (1993, 1994) have explored the possibility that 

hydroxylated 2,3,7,8-TCDD metabolites competitively interact with plasma thyroid hormone transport 

proteins, thus facilitating clearance and excretion of T4. They tested various hydroxy-CDDs in an in vitro 

competitive-binding assay using purified human TTR and found that those with chlorine substitution 

adjacent to the hydroxy group (7-OH-2,3,8-TrCDD and 2-OH-1,3,7,8-TCDD) showed similar or higher 

relative-binding potency than T4 (Lans et al. 1993). 8-OH-2,3-DCCD, which did not contain chlorine 

substitution adjacent to the OH group, did not displace T4. In a subsequent study, Lans et al. (1994) studied 

the displacement of T4 from globulin, the major T4-binding protein in humans by hydroxy CDD 

metabolites (in contrast to TTR in rodents). The results showed than none of the tested hydroxylated CDD 

metabolites inhibited binding of T4 by T4-binding globulin and suggested that hydroxylated CDD 

metabolites can cause different effects in rodents and humans. 

The overall evidence suggests that in rodents, thyroid hormones modify 2,3,7,8-TCDD toxicity, but a 

reduction in T4 (at least in rats) does not mediate toxicity. 

Administration of 2,3,7,8-TCDD to rodents was also shown to reduce blood corticosterone levels (Balk and 

Piper 1984; DiBartolomeis et al. 1987; Mebus and Piper 1986). This effect has been attributed to decreased 

corticosterone synthesis by decreasing cholesterol side-chain cleavage in the adrenal gland. More recent 

studies suggested that 2,3,7,8-TCDD may interfere with secretion or synthesis of appropriate, bioactive 

ACTH from the anterior pituitary gland, which could compromise adrenal steroidogenenesis (Bestervelt et 

al. 1993). 

Administration of 2,3,7,8-TCDD to animals results in a wide range of endocrine responses which are not 

only species-dependent, but also exhibit variability within species. Endocrine effects observed in humans 

have not been limited to thyroid effects and diabetes; alterations in levels of reproductive hormones, as 

summarized in the sections on reproductive effects have also been observed. The wide array of endocrine 

effects induced by CDDs and structurally-related chemicals has triggered increased interest within the 

scientific community and the term “endocrine disruptors” is currently being used to describe some members 

of this class of chemicals. The available information suggests that CDDs may cause adverse endocrine 

effects in humans. 
CDDs 280 

2. HEALTH EFFECTS 

Dermal Effects. Chloracne is the most easily recognized effect of exposure to 2,3,7,8-TCDD and 

structurally related chlorinated organic chemicals. Chloracne is a high-dose response in animals and 

humans; and its presence in humans indicates exposure to CDDs and other chlorinated organic compounds, 

but its absence does not preclude such exposure. Furthermore, the variability of the response in more 

highly exposed individuals suggests that susceptibility varies among individuals. Chloracne can first occur 

on the face, particularly under the eyes and behind the ears. With increasing exposure, the rest of the face 

and neck, upper arms, chest, back, abdomen, outer thighs, and genitalia may be affected. When severe, 

chloracne can cover the entire body. Clinically, changes vary from an eruption of comedones to occurrence 

of papules and pustules. Histologically, the lesions consist of keratinous cysts caused by squamous 

metaplasia of sebaceous glands. The acute stage is followed by vermiculite skin atrophy. The incidence of 

other dermal effects, including hyperpigmentation and hirsutism, correlates with the intensity of chloracne 

(Poland et al. 1971). Chloracne has been reported to have occurred in at least a small number of workers in 

all accidents at TCP-production facilities (Jansing and Korff 1994; May 1973; Schecter et al. 1993; Suskind 

1985; Zober et al. 1990); among subjects involved in production of 2,3,7,8-TCDD-contaminated products 

(Bond et al. 1989a; Moses and Prioleau 1985; Pazderova-Vejlupkova et al. 1981; Poland et al. 1971; 

Suskind and Hertzberg 1984); in laboratory workers exposed to 2,3,7,8-TCDD (Oliver 1975); and among a 

small percentage of Seveso residents (Assennato et al. 1989; Caramaschi et al. 1981; Mocarelli et al. 1986; 

Reggiani 1980). Chloracne, however, was not observed among Missouri residents (Hoffman et al. 1986; 

Webb et al. 1989) examined 10 years after exposure or among Ranch Hand veterans (Burton et al. 1998; 

USAF 1991). 

The dermal changes induced by 2,3,7,8-TCDD may appear as soon as 2 days after exposure (Ott et al. 1993; 

Zober et al. 1990) or within months (Caramaschi et al. 1981; Reggiani 1980). The lesions may heal within 

a few months after cessation of exposure (Assennato et al. 1989) despite high serum 2,3,7,8-TCDD levels 

(Mocarelli et al. 1991) or persist for over 15 years, depending upon severity (Crow 1978; Jansing and Korff 

1994; Moses and Prioleau 1985; Schecter et al. 1993; Suskind and Hetzberg 1984). Children exposed to 

2,3,7,8-TCDD appear to be more sensitive than adults, and individuals similarly exposed have variable 

susceptibility to chloracne (Mocarelli et al. 1991). Data from analyses of cases among chemical workers 

suggested that the risk for developing chloracne was highest among workers who were exposed at younger 

ages, among those who had been exposed for the longest periods, and among workers whose jobs rated at 

the highest intensity of exposure (Ott et al. 1987). The variability in the chloracneic response can be 

illustrated the following evidence from the Seveso incident: no chloracne was observed in subjects with 

initial serum lipid 2,3,7,8-TCDD levels of <800 ppt, chloracne was present at serum lipid levels of 
CDDs 281 

2. HEALTH EFFECTS 

>12,000 ppt; and between 800 and 12,000 ppt the occurrence of chloracne was sporadic (Mocarelli et al. 

1991); this suggested that, in this population, 8,000 to 10,000 ppt 2,3,7,8-TCDD in blood was necessary for 

expression of chloracne. German workers involved in TCP production who had chloracne had estimated 

adipose levels $200 ppt 2,3,7,8-TCDD and $2,000 ppt HxCDD at the time of diagnosis (Beck et al. 1989a). 

In another study of German workers, 80% of the severe chloracne cases had 2,3,7,8-TCDD levels of 

$250 ppt; however, 26% of the workers without chloracne also had 2,3,7,8-TCDD levels of $250 ppt (Ott 

et al. 1993). Schecter et al. (1993) provided the first reported incidence of chloracne in females with 

elevated dioxin blood levels from occupational exposure. Their observation that one worker diagnosed 

with chloracne in their study had the lowest 2,3,7,8-TCDD blood concentration, whereas two workers with 

the higher levels did not display chloracne, confirmed the view that chloracne indicates exposure to dioxin, 

but its absence does not preclude such exposure. 

No studies were located regarding dermal effects in humans exposed specifically to CDDs by the oral route. 

Evidence from human case reports in the Yusho/Yu-Cheng incidents (which involved exposure to CDFs, 

PCBs, and CDDs) and from animal studies, however, indicates that dermal effects could occur after 

exposure by the oral route (ATSDR 1994). 

Oral studies of 2,3,7,8-TCDD showed the development of rough hair in hamsters (Henck et al. 1981) and 

skin thickening in A2G-hr/+ mice (Greig 1984) after acute exposure. Chronic oral exposure to 

2,3,7,8-TCDD caused dermatitis in B6C3F1 mice (Della Porta et al. 1987) and amyloidosis in Swiss mice 

(Toth et al. 1979). Rhesus monkeys proved to be very sensitive to 2,3,7,8-TCDD-induced dermal effects. 

The changes consisted of swollen eyelids, nail loss, facial alopecia, and acneform lesions after both acute­

(McConnell et al. 1978a) and intermediate-duration oral exposures (Allen et al. 1977; McNulty 1984). 

Dermal exposure to 2,3,7,8-TCDD induced hyperkeratosis and epidermal hyperplasia in hairless HRS/J 

mice after acute- (Puhvel and Sakamoto 1988) and intermediate-duration (Puhvel et al. 1982) exposures. 

While acneiform lesions were reported in CD-1 mice after intermediate-duration dermal exposure to 

2,3,7,8-TCDD (Berry et al. 1978, 1979), no effects were found in Swiss Webster mice chronically exposed 

to lower levels (NTP 1982a). 

The data available suggest that 2,3,7,8-TCDD is a dermal toxicant both in humans and animals. 

Erythematous skin rashes and chloracne are considered one of the hallmarks of 2,3,7,8-TCDD toxicity, 

although it can be caused also by exposure to other polyhalogenated aromatic compounds. It is also worth 
CDDs 282 

2. HEALTH EFFECTS 

mentioning that although in humans chloracne indicates exposure to chlorinated or halogenated aromatics, 

lack of chloracne does not indicate that exposure has not occurred. 

Ocular Effects. The incidence of eye irritation correlated with the intensity of chloracne in a study of 

workers employed in a 2,4,5-T factory (Poland et al. 1971), but the role of 2,3,7,8-TCDD, if any, cannot be 

established. No studies were located regarding ocular effects in humans exposed specifically to CDDs by 

the oral route. Evidence from the human case reports in the Yusho/Yu-Cheng incidents (which involved 

exposure to CDFs, PCBs, and less so CDDs) and from animal studies, however, indicates that ocular effects 

could occur after exposure by the oral route (ATSDR 1994). Ocular effects observed in the Yusho and YuCheng victims included hypersecretion of the Meibomian glands, abnormal pigmentation of the 

conjunctiva, and swelling of the eyelids (Hsu et al. 1994; Masuda 1994). 

Topical application of 2,7-DCDD, with no toxic 2,3,7,8-chlorine pattern, 2,3,7,8-TCDD, and mixed 

HxCDD, or OCDD into the conjunctival sac of rabbits caused transient pain and conjunctival inflammation 

(Schwetz et al. 1973). 

Based on adverse ocular effects observed in humans and animals exposed to chemicals structurally-related 

to CDDs and animals (monkeys) exposed to 2,3,7,8-TCDD itself, it is reasonable to assume that CDDs will 

cause similar effects under similar exposure conditions. 

Body Weight Effects. A transient weight loss was reported in a small number of subjects exposed to 

2,3,7,8-TCDD in the workplace (Jirasek et al. 1976; Oliver 1975). However, due to the lack of data from 

controlled studies, the role of 2,3,7,8-TCDD, if any, is difficult to ascertain. Although weight loss has not 

been well documented in humans following exposure to 2,3,7,8-TCDD, numerous animal studies provide 

evidence that exposure to CDDs causes the wasting syndrome. Acute oral exposure to 2,3,7,8-TCDD 

induced weight loss in rats (Christian et al. 1986a; Moore et al. 1985; Roth et al. 1988; Seefeld and Peterson 

1984; Seefeld et al. 1984a, 1984b; Walden and Schiller 1985), mice (Hanberg et al. 1989; Kelling et al. 

1985; Smith et al. 1976; Weber et al. 1995), guinea pigs (Hanberg et al. 1989; Umbreit et al. 1985), 

hamsters (Hanberg et al. 1989), and monkeys (McConnell et al. 1978a). Similarly, body weight changes 

were found after intermediate-duration oral exposure to 2,3,7,8-TCDD in rats (Diliberto et al. 1996; NTP 

1982b; Van Birgelen et al. 1995; Viluksela et al. 1994; Vos et al. 1973), guinea pigs (DeCaprio et al. 1986; 

Vos et al. 1973), mice (Thigpen et al. 1975; Vos et al. 1973), and monkeys (McNulty 1984), and after 

chronic exposure in rats (Kociba et al. 1978a; NTP 1982b; Van Miller et al. 1977) and mice (Della 
CDDs 283 

2. HEALTH EFFECTS 

Porta et al. 1987). In addition, milder changes, represented usually by decreases in body weight gain, were 

seen after oral exposure to other congeners (NCI/NTP 1979a, 1980; Schwetz et al. 1973). The wasting 

syndrome does not appear to be a route-specific effect since body weight changes were observed in mice 

exposed to 2,3,7,8-TCDD dermally (NTP 1982a; Puhvel et al. 1982), and numerous studies reported body 

weight changes in animals that were injected with 2,3,7,8-TCDD subcutaneously or intraperitoneally 

(Canga et al. 1988; Chahoud et al. 1989; Della Porta et al. 1987; Gorski et al. 1988b; Holcomb et al. 1988; 

Kelling et al. 1985; Lu et al. 1986; McConkey and Orrenius 1989; Pohjanvirta et al. 1989). 

The mechanism of the wasting syndrome has been extensively investigated. Results of studies in C57BL/6 

mice, guinea pigs, and Fischer 344 rats showed that 2,3,7,8-TCDD exposure induces appetite suppression 

resulting in loss of adipose and lean tissue, and eventually death (Kelling et al. 1985). However, by using 

pair-fed animals as controls, it was clear that body weight loss alone was not the cause of death. This is 

also supported by the fact that the weight loss, but not the lethality of 2,3,7,8-TCDD, can be prevented by 

parenteral feeding of rats and guinea pigs (Gasiewicz et al. 1980; Lu et al. 1986). Seefeld and Peterson 

(1984) showed that in rats, fecal energy loss as a percentage of daily feed energy uptake was not 

significantly altered by treatment with 2,3,7,8-TCDD. Furthermore, the percentage of feed energy absorbed 

by the gastrointestinal tract was not changed by 2,3,7,8-TCDD, which ruled out the possibility of a 

2,3,7,8-TCDD-induced gross malabsorption syndrome. The same group of investigators (Seefeld et al. 

1984a, 1984b) showed that 2,3,7,8-TCDD does not impair the animals' capacity to feed since rats that lost 

weight prior to treatment with 2,3,7,8-TCDD ate and gained weight after treatment with 2,3,7,8-TCDD. 

Based on their results, Seefeld and coworkers (Seefeld and Peterson 1984; Seefeld et al. 1984a, 1984b) 

proposed that 2,3,7,8-TCDD lowers a “set point” for regulated body weight, and hypophagia serves as a 

secondary response to reduce the animal’s body weight to the lower regulation level determined by the dose 

of 2,3,7,8-TCDD administered. The ability of an animal to recover from the 2,3,7,8-TCDD-induced 

hypophagia may be species- and/or strain-specific (Tuomisto and Pohjanvirta 1991). Within 1-2 weeks 

after a single dose of 2,3,7,8-TCDD, feed intake increased in Hans/Wistar rats but not in Long Evans rats; 

the Long Evans rats died by week 3. Although Hans/Wistar and Long Evans rats have similar Ah receptor 

binding and cytochrome P-450 induction properties, the wide differences in sensitivity to 2,3,7,8-TCDD 

suggest that other mechanisms may be involved in the wasting syndrome. 

Numerous studies have examined the possibility that the wasting syndrome results from 2,3,7,8-TCDDinduced alterations in intermediate metabolism. For example, it has been shown that in male SpragueDawley rats, lethal doses of 2,3,7,8-TCDD severely alter glucose homeostasis (Gorski and Rozman 1987; 
CDDs 284 

2. HEALTH EFFECTS 

Gorski et al. 1990; Potter et al. 1983). Hypoglycemia was not the result of hyperinsulinemia since insulin 

levels were also depressed with 2,3,7,8-TCDD treatment (Potter et al. 1983). Also, hypophagia did not 

account for hypoglycemia since pair-fed rats also exhibited hypoglycemia (Potter et al. 1983). Further 

studies showed that decreased gluconeogenesis was the result of significantly reduced activity of hepatic 

PEPCK, the rate-determining enzyme in the pathway (Weber et al. 1991a). Other gluconeogenic enzymes 

such as glucose-6-phosphatase and pyruvate carboxylase were also decreased by treatment with 

2,3,7,8-TCDD, but pyruvate kinase, a glycolytic enzyme, was not affected (Weber et al. 1991a). It was also 

shown that changes in gluconeogenic enzyme activities preceded hormonal changes (insulin, 

corticosterone) by at least 2 days (Weber et al. 1991b), which led the authors to suggest that 2,3,7,8-TCDDinduced changes in hormonal homeostasis are adaptive responses of the organism to stimulate 

gluconeogenesis. Reduced liver PEPCK activity as a result of 2,3,7,8-TCDD treatment has also been 

observed in Long Evans rats (Fan and Rozman 1995) and in C57BL/6J and DBA/2J mice (Weber et al. 

1995). 

Some investigators suggested that the wasting syndrome may be linked to 2,3,7,8-TCDD-induced effects on 

the thyroid (Rozman 1984; Rozman et al. 1985). In thyroidectomized rats, the weight loss after 

2,3,7,8-TCDD exposure was slow, suggesting that the lack of thyroid hormone reduced the rate of stored fat 

utilization (Rozman et al. 1985). Thyroidectomy protected rats from immunotoxicity induced by an 

intraperitoneal dose of 2,3,7,8-TCDD (Pazdernik and Rozman 1985). Replacement therapy with T4 

partially reversed the effects of thyroidectomy on T4 and triiodothyronine serum levels, body weight, and 

immune function. The authors suggested that 2,3,7,8-TCDD-induced hypothyroidism may be a protective 

mechanism against 2,3,7,8-TCDD-induced wasting syndrome and lethality. Thyroid hormones regulate fat 

mobilization and use of fatty acids in adipose tissue and influence norepinephrine-mediated nonshivering 

thermogenesis that is also linked to brown adipose tissue. It was also suggested that the effect of 

2,3,7,8-TCDD on the thyroid causes activation of thyrotropin-releasing hormone, which results in anorexia 

(Aust 1984). Anorexia and 2,3,7,8-TCDD-induced retinol depletion would then lead to the body weight 

loss. 

Based on the findings that 2,3,7,8-TCDD administered into the lateral cerebral ventricles does not cause 

death or decreased feed intake in rats (Stahl and Rozman 1990), Rozman et al. (1991) examined the 

possibility that 2,3,7,8-TCDD suppresses appetite via peripheral mechanisms acting on the central nervous 

system. The results of experiments of transfusion of blood from rats with 2,3,7,8-TCDD-induced appetite 

suppression and normal satiated rats suggested that 2,3,7,8-TCDD-treated rats are not satiated, rather than 
CDDs 285 

2. HEALTH EFFECTS 

that they are not hungry. In a second experimental series, the possible role of norepinephrine, dopamine, 

and serotonin as central mediators of appetite suppression induced by 2,3,7,8-TCDD was investigated. No 

changes were found in epinephrine and dopamine in the hypothalamus or in dopamine and its metabolites in 

the striatum. However, tryptophan (a precursor of serotonin) levels in plasma and brain were increased and 

this was paralleled by increases in brain serotonin and 5-hydroxyindolacetic acid (the major serotonin 

metabolite) (Rozman et al. 1991). Based on the results of these experiments, Rozman et al. (1991) 

proposed that decreased PEPCK activity decreases gluconeogenesis and leads to increased plasma 

concentrations of glycogenic amino acids, such as tryptophan. Increased tryptophan leads to increase in 

serotonin release in the brain and to appetite suppression. It was subsequently shown, however, that lethal 

doses of 2,3,7,8-TCDD reduces the activity of TdO, the key enzyme of the major tryptophan degradation 

pathway (Weber et al. 1992c, 1994). Whether due to reduction in TdO activity, reduced glyconeogenesis, 

or both, Weber et al. (1994) proposed that an initial increase in tryptophan levels result in some initial feed 

refusal, which in turn initiates the wasting of body mass and increases the supply of tryptophan with which 

the animals cannot deal. A vicious cycle develops which results in strongly elevated tryptophan levels and 

increased serotonin turnover, which acts as an appetite suppressant. 

Alternative explanations for the increased levels of plasma-free tryptophan in 2,3,7,8-TCDD-treated rats 

have been offered. Four possibilities were discussed by Unkila et al. (1994): 2,3,7,8-TCDD may reduce 

the binding capacity of the blood, i.e., may decrease plasma albumin levels; 2,3,7,8-TCDD may stimulate 

the production of some competing factor in the blood (e.g., nonesterified fatty acids or bilirubin) which are 

also bound to albumin; 2,3,7,8-TCDD might affect the binding properties of the albumin molecule; and 

2,3,7,8-TCDD might inhibit tryptophan catabolism. Of the four factors examined that might affect the 

binding of tryptophan to albumin, Unkila et al. (1994) indicated that the most important is probably plasma 

bilirubins and suggested that disturbances in liver function may be involved in the changes in tryptophan 

metabolism. 

The wasting syndrome is a characteristic effect of exposure to 2,3,7,8-TCDD in animals and, in its most 

severe form, is usually associated with lethality, particularly in rodents. The fact that the wasting syndrome 

has not been demonstrated in humans does not necessarily indicate that humans are insensitive to this effect 

of dioxins, but may indicate that human exposure has not approached acutely high enough levels. 
 

CDDs 286 

2. HEALTH EFFECTS 

Immunological and Lymphoreticular Effects. No consistent exposure-related effects on the 

immune system have been observed in human populations exposed to above-background levels of 

2,3,7,8-TCDD (Ernst et al. 1998; Jansing and Korff 1994; Jennings et al. 1988; Jung et al. 1998; Mocarelli 

et al. 1986; Neubert et al. 1993, 1995; Svensson et al. 1994; Tonn et al. 1996; USAF 1991; Webb et al. 

1989; Zober et al. 1994). The immunological effects of 2,3,7,8-TCDD were recently reviewed by Kerkvliet 

(1995) who identified a number of factors on which the results of immunological assessments may be 

dependent. One of these factors, perhaps the most notable, is the inherent difficulty in assessing subclinical 

immunological effects in an outbred human population. In addition, the wide range of normal responses of 

most immunological assays diminishes the sensitivity to detect small changes. Another factor to consider is 

that assays used to assess immune function in humans exposed to 2,3,7,8-TCDD and related chemicals have 

been based for the most part on what was clinically feasible rather than on assays proven to be sensitive in 

animal studies (i.e., the antibody response to SRBC). Therefore, the lack of consistent or significant 

immunotoxic effects in humans exposed to 2,3,7,8-TCDD may be a function of both the type of assay and 

the immune status of the population studied. Furthermore, often the cohort exposure is not validated and 

the immune status has been examined long after exposure allowing for recovery from any immunotoxic 

effect that may have occurred shortly after exposure. 

A potentially useful approach to studying the sensitivity of the human immune system to 2,3,7,8-TCDD 

has been to examine the direct in vitro effects of 2,3,7,8-TCDD on human cell cultures. For example, 

Cook et al. (1987a) observed concentration dependent immunosuppressive responses of cultured human 

thymic epithelial cell and thymocytes exposed to 2,3,7,8-TCDD. The proliferative response of human 

lymphocytes in vitro to stimulation with mitogens is extremely sensitive to 2,3,7,8-TCDD. Concentrations 

of 2,3,7,8-TCDD as low as 10-12 to 10-14 M reduced the percentage of CD20+

 B cells and CD4+CDw29+ 

T cells (Neubert et al. 1991). However, these results could not be corroborated in a similar study by Lang 

et al. (1994) who used 2,3,7,8-TCDD concentrations ranging from 10-7 to 10-11 M. In another study, 

proliferation of human tonsillar lymphocytes (HTLs) cultured in vitro was inhibited by 3×10-8 M 

2,3,7,8-TCDD, but pokeweed mitogen (PWM) induced proliferation was not affected by 2,3,7,8-TCDD 

concentrations ranging from 3×10-8 to 10-10 M (Wood et al. 1992). However, when low density ß cells 

from HTLs were purified and cultured in vitro in the same laboratory and stimulated with 

lipopolysaccharide and TRF (T-cell replacing factor), 3×10-8 to 10-10 M 2,3,7,8-TCDD suppressed the IgG 

secretion in a dose related manner. HTLs possess the Ah receptor as indicated by the induction of 

7-ethoxycoumarin-o-deethylase (EROD) in a dose-related manner at the above doses when the HTLs are 

stimulated with phytohemagglutenin (PHA) or PWM (Wood et al. 1993). A promising animal model for 
CDDs 287 

2. HEALTH EFFECTS 

assessing the potential immunotoxicity of CDDs in humans is the SCID mice, which can be engrafted with 

human fetal thymus and liver tissue fragments under the kidney capsule. Using the SCID mice model, it 

was shown that human thymus cells are as sensitive to 2,3,7,8-TCDD as the thymus of Wistar rats (De 

Heer et al. 1995). 

The immune system appears to be one of the most sensitive targets for CDDs in animals. However, it is 

difficult to make interspecies or congener comparisons due to interlaboratory variability of functional tests 

and use of different end points in various studies. Thymic atrophy was observed in rats, mice, guinea pigs, 

(De Heer et al. 1994a; Hanberg et al. 1989), hamsters (Hanberg et al. 1989; Olson et al. 1980a), and 

monkeys (McConnell et al. 1978a) after acute exposure; in guinea pigs (Vos et al. 1973) after 

intermediate-duration exposure; and in rats (Kociba et al. 1978a) after chronic oral exposure to 

2,3,7,8-TCDD. Furthermore, lymph node atrophy (Allen et al. 1977) and bone marrow degeneration 

(Hong et al. 1989) were reported in monkeys after intermediate- and chronic-duration exposure to 

2,3,7,8-TCDD, respectively. In support of these data, thymic atrophy was also induced by a single 

intraperitoneal injection of 2,3,7,8-TCDD in Sprague-Dawley rats (Gorski et al. 1988b), Syrian hamsters 

(Olson et al. 1980a), and C57BL/6J mice (Poland and Glover 1980). Only increased thymus/body weight 

ratio was found in HRS/J mice exposed to 2,3,7,8-TCDD dermally (Hebert et al. 1990). Effects on 

peripheral lymphocytes following acute subcutaneous and in vitro exposure, particularly changes in 

percentages of lymphocyte subpopulations, suggest that marmoset monkeys may be particularly sensitive 

to immunologic effects of 2,3,7,8-TCDD (Neubert et al. 1990a, 1991). In general, relatively high doses 

cause lymphoid depletion, lower doses cause thymic cellular depletion in young animals, and much lower 

doses affect specific immune receptor functions. 

Administration of total parenteral nutrition did not protect rats from thymic atrophy with decreased 

numbers of cortical lymphocytes that developed after acute intraperitoneal 2,3,7,8-TCDD exposure 

(Gasiewicz et al. 1980). 2,3,7,8-TCDD-induced thymic atrophy in BALB/CJ and DBA/2J mice correlated 

with a reduction in thymic and bone marrow terminal deoxynucleotidyl transferase synthesis (Fine et al. 

1990b). The prothymocyte activity was severely damaged by 2,3,7,8-TCDD exposure. The authors 

concluded that 2,3,7,8-TCDD produces atrophy by damaging the capability of bone marrow prethymic 

stem cells to seed the thymus. In addition to bone marrow effects, 2,3,7,8-TCDD may also inhibit normal 

thymocyte maturational processes. When B6C3F1 mice were exposed in utero by dosing the dam at 

3 μg/kg/day between Gd 6–14 and foster-nursing the offspring with unexposed females, thymic atrophy 

was seen at Gd 18 or on postnatal day 6, but the thymic effects were no longer seen by day 14. In utero 
CDDs 288 

2. HEALTH EFFECTS 

effects on the thymus were at much lower doses than effects in animals exposed postnatally. Rodents are 

born with an immature immune system which develops in the first few days after birth. These mice were 

tested for immune function at 7–8 weeks of age; the cytotoxic T-lymphocyte was still suppressed, but the 

mitogen response to SRBC was not suppressed (Holladay et al. 1991). Humans, in contrast to rodents, 

have a more mature immune system at birth. The role of the thymus is important in prenatal and perinatal 

development of the immune system, but its role in adult life has not been established. Concentrations of 

140 fg 2,3,7,8-TCDD/mg in the thymus of mice at Gd 18 were associated with thymic atrophy (Fine et al. 

1990a). 

In addition to causing lymphoid organ weight changes, 2,3,7,8-TCDD has been shown to cause functional 

alterations in the immune response (Vecchi et al. 1980a). Studies have shown that suppressed antibody 

response (Holsapple et al. 1986a; Vecchi et al. 1980a, 1983b) decreased host resistance to Streptococcus 

pneumoniae or influenza A virus (Burleson et al. 1996; White et al. 1986), and suppressed serum 

complement activity (White et al. 1986) occur in B6C3F1 mice after single or repeated oral dose(s) of 

2,3,7,8-TCDD. Immunological effects occurred at the lowest LOAEL in acute- and intermediate-duration 

exposure studies and indicated that the immunological system is very sensitive to 2,3,7,8-TCDD-induced 

toxicity. The dose of 0.01 μg/kg for impaired resistance was the lowest LOAEL for acute oral exposure 

(Burleson et al. 1996). The NOAEL of 0.005 μg/kg identified in this study was used to derive an acute 

oral MRL of 2×10-4 μg/kg/day. The 0.0007 μg/kg/day dose for reduced thymus weight for intermediateduration oral exposure (DeCaprio et al. 1986) was used to derive an intermediate-duration oral MRL for 

2,3,7,8-TCDD of 2×10-5 μg/kg/day. Immunosuppression as evidenced by increased mortality when 

challenged with bacteria was demonstrated in C57BL/6J mice after administration of a dose of 1 μg 

2,3,7,8-TCDD/kg/week over a period of 4 weeks (Thigpen et al. 1975); this occurred without any other 

apparent signs of toxicity. In addition, an intermediate-duration exposure to 2,3,7,8-TCDD induced 

decreased cell-mediated (mice and guinea pigs) and humoral (guinea pigs) immunity (Vos et al. 1973). 

The results indicated that guinea pigs are the most sensitive species tested. 2,3,7,8-TCDD-induced 

suppression of humoral immunity was also reported in animals exposed parenterally. 

In general, the route of exposure does not affect the immune response. Several tests of immunotoxicity 

dosed the animals by parenteral routes (intravenous, subcutaneous, or intraperitoneal). Acute 

intraperitoneal 2,3,7,8-TCDD exposure inhibited the primary and secondary humoral response to 

T-dependent (SRBC) and T-independent (pneumococcal polysaccharide) antigens in C57BL/6 mice 

(Vecchi et al. 1980b, 1983b); doses were comparable to those causing effects by the oral route. A dose­
CDDs 289 

2. HEALTH EFFECTS 

related suppression of IgM and IgG antibody-forming cells was induced by exposure to a single 

intraperitoneal injection of 2,3,7,8-TCDD in B6C3F1 mice (House et al. 1990). Furthermore, doses as low 

as 0.1 μg/kg decreased survival after influenza virus infection, and exposure at 10 μg/kg suppressed 

production of antibody to viral hemagglutinin. Cytolytic and NK-cell-mediated immunity was impaired in 

C57BL/6 mice after a single intraperitoneal injection of 2,3,7,8-TCDD due to the decreased number of 

peritoneal macrophages and splenocytes (Mantovani et al. 1980). However, the immune function per unit 

was not damaged. In addition, an in vitro study demonstrated that 2,3,7,8-TCDD induces tumor necrosis 

factor in human keratinocytes that could affect tumor promotion and affect immune parameters (Choi et al. 

1991). 

The role of the Ah receptor in the immune responses to CDDs has been examined in several studies. A 

correlation between the AHH inducibility and suppression of humoral immunity caused by 2,3,7,8-TCDD 

injection was observed in several strains of mice (Vecchi et al. 1983a). Similarly, when three strains of Ah 

responsive mice (C57BL/6nQdj, BALB/cCrj, C3H/HeNQdj) were compared with nonresponsive 

(AKR/JSea, DBA/2JCrj, DDD:Qdj) strains of mice, decreased thymus weight was found only in the 

responsive animals (Nagayama et al. 1989). The C57 strain also had decreased lymphocyte counts. 

Results of an in vitro experiment supported these observations (Dencker et al. 1985). Thymus cultures 

from Ah locus responsive C57BL/6 mice were very sensitive to the toxicity of 2,3,7,8-TCDD compared 

with thymus cultures from the nonresponsive DBA/2J mice. 2,3,7,8-TCDD exposure of cultures of thymic 

epithelial cells from responsive C57BL/6 mice indicated that 2,3,7,8-TCDD alters the maturation of 

thymocytes (Greenlee et al. 1985). It was further demonstrated that 2,3,7,8-TCDD toxicity in human 

thymic epithelial cells was mediated by a protein receptor (Cook and Greenlee 1989). Similarly, in vitro 

studies with lymphocytes, spleen cells, and bone marrow cells from 2,3,7,8-TCDD-pretreated mice 

indicated that 2,3,7,8-TCDD acts by an Ah locus-dependent mechanism to obstruct the formation of 

cytotoxic T cell generation from their precursors (Dooley et al. 1990; Holladay et al. 1991; Nagarkatti et 

al. 1984). A brief summary of the possible mechanisms of 2,3,7,8-TCDD immunotoxicity can be found in 

Section 2.4.2. 

Oral experiments with other congeners reported suppressed antibody response in B6C3F1 mice after acute 

exposure to 2,7-DCDD, a non 2,3,7,8-chlorine substituted CDD (Holsapple et al. 1986b) and splenic 

hyperplasia in rats after intermediate-duration exposure to a mixture of 1,2,3,6,7,8-HxCDD and 

1,2,3,7,8,9-HxCDD (NCI/NTP 1980). 
 

CDDs 290 

2. HEALTH EFFECTS 

Although human exposure studies to date found no conclusive evidence of immunotoxicity, the animal 

data show that the immune system is a target for CDD toxicity in many species. However, a defined 

2,3,7,8-TCDD-induced immune deficiency syndrome has not emerged largely because in animals, the 

immune response depends on the species studied, the dose of 2,3,7,8-TCDD, and the antigen and exposure 

protocol studied. 

Neurological Effects. Some psychological effects were reported in Vietnam veterans potentially 

exposed to 2,3,7,8-TCDD-contaminated herbicides. These included depression in Air Force and ground 

troop veterans (Levy 1988; Wolfe et al. 1985) and hypochondria and hysteria in Air Force veterans (Wolfe 

et al. 1985). In contrast, a more recent study did not find any association between 2,3,7,8-TCDD exposure 

and neurological or psychological diseases in Air Force personnel (USAF 1991). These psychological 

effects could be due to a number of stress-related factors in the veterans. Recently, a group of 16 scientific 

experts from the National Academy of Sciences’ Institute of Medicine, who evaluated the strength of 

evidence for human health effects among veterans exposed to herbicides used in Vietnam, found no strong 

evidence establishing an association between herbicide use in Vietnam and clinical neurologic disorders 

(Goetz et al. 1994). However, psychological changes were reported in relatively small cohorts of exposed 

individuals (Oliver 1975; Pazderova-Vejlupkova et al. 1981; Peper et al. 1993). Subclinical peripheral 

neuropathy, encephalopathy, and sensory impairment were reported in workers exposed to higher levels of 

2,3,7,8-TCDD (Goldman 1973; Jirasek et al. 1976; Pazderova-Vejlupkova et al. 1981) and in the general 

population exposed to 2,3,7,8-TCDD after an industrial accident (Barbieri et al. 1988; Filippini et al. 1981; 

Pocchiari et al. 1979). Decreased nerve conduction velocity was observed in phenoxy herbicide 

production workers (Singer et al. 1982). In contrast, exposure to 2,3,7,8-TCDD (confirmed by elevated 

serum levels) was not related to chronic peripheral neuropathy in a group of workers exposed 15–37 years 

earlier compared to referent controls (Sweeney et al. 1993). These authors suggest that the finding of 

peripheral neuropathy in the earlier studies indicate that this condition may occur shortly after exposure 

and resolve over time. 

Data regarding neurological or neurophysiological effects following exposure to CDDs in animals are 

limited. Decreased motor activity was seen in rats at 2,3,7,8-TCDD dose levels of #5 μg/kg (Giavini et al. 

1983; Seefeld et al. 1984a). Time-dependent increases in tryptophan (amino acid precursor of the 

neurotransmitter serotonin) levels in plasma and brain (hypothalamus, striatum) correlated with elevations 

in brain serotonin and 5-hydroxyindoleacetic acid levels in rats after a single intraperitoneal injection of 50 

or 120 μg 2,3,7,8-TCDD/kg (Rozman et al. 1991; Tuomisto et al. 1990). Furthermore, slight changes 
 

CDDs 291 

2. HEALTH EFFECTS 

were observed in levels of noradrenaline, dopamine, dihydroxyphenylacetic acid (DOPAC), homovanillic 

acid (HVA), 5-hydroxytryptamine (5-HT), and 5-hydroxyindoleacetic acid (5-HIAA) (Tuomisto et al. 

1990). Results from a more recent study in rats showed that 2,3,7,8-TCDD increases neuronal serotonin 

turnover in TCDD-susceptible Long Evans rats, but not in TCDD-resistant Han/Wistar rats or in foodrestricted Long Evans rats (Unkila et al. 1994). By using much lower intraperitoneal doses of 

2,3,7,8-TCDD (2.2–8.8 μg/kg) in adult male Han/Wistar rats, thus avoiding manifestation of the wasting 

syndrome, Grahmann et al. (1993) and Grehl et al. (1993) observed electrophysiological (decreased 

conduction velocity) and histological signs (nerve degeneration) of peripheral neuropathy several months 

after a single injection of 2,3,7,8-TCDD. The possible mechanism of 2,3,7,8-TCDD neurotoxicity was 

not discussed. The existing information suggests that 2,3,7,8-TCDD causes minor alterations in brain 

neurotransmitter systems. 

The overall evidence suggests that adverse neurological effects may occur in subjects exposed to relatively 

high levels of dioxins, or at least to levels that cause frank dermal effects. The neurological effects, 

however, may be transient and therefore, difficult to diagnose if examination is conducted several years 

after exposure. The nervous system in adults does not seem to be a particularly sensitive target for CDDs 

toxicity, but CDDs may represent a neurological hazard to the developing organism by, for example, 

altering hormone levels at critical times during the maturation of the central nervous system. 

Reproductive Effects. The weaknesses of the epidemiology studies examining reproductive end points 

limits drawing conclusions regarding the reproductive toxicity of 2,3,7,8-TCDD in humans. Some common 

weaknesses include lack of exposure data (many of the studies rely on self-reported 2,3,7,8-TCDD 

exposure; CDC (1987) found that 2,3,7,8-TCDD blood levels of Vietnam veterans reporting direct or 

indirect exposure to Agent Orange were not significantly different from levels in non-Vietnam veterans), 

concomitant exposure to other chemicals, and lack of data on 2,3,7,8-TCDD levels at the time of 

conception. Several studies looked for an association between 2,3,7,8-TCDD exposure and an increased 

risk of spontaneous abortions, most did not find any statistically significant alterations following paternal 

exposure to 2,3,7,8-TCDD (Aschengrau and Monson 1989; Smith et al. 1982; Wolfe et al. 1995). An 

increased incidence of spontaneous abortions, was observed in women living near an herbicide 

manufacturing facility (Forsberg and Nordstrom 1985). However, this study has been criticized for its 

small sample size, inadequate discussion of sample selection, and concomitant exposure to other chemicals, 

including dibenzofurans (Sweeney 1994). In Vietnamese residents living in areas sprayed with Agent 

Orange, an increased incidence of hydatiform moles was observed (Phuong et al. 1989a). A later case­
CDDs 292 

2. HEALTH EFFECTS 

control study by Ha et al. (1996) did not confirm the results of the Phuong et al. (1989a) study. In the 

71⁄2-year period after the Seveso accident, the number of female children born to parents living in area A 

was significantly higher than the number of male children (48 versus 26) (Mocarelli et al. 1996). An 

increased ratio of female to male children was also reported in workers of a 2,4,5-T production facility in 

Ufa, Russia (Basharova 1996) and in men exposed to chlorophenate wood preservatives contaminated with 

CDD (Dimich-Ward et al. 1996; James 1997). No alterations were found in the Missouri cohort of women 

living in 2,3,7,8-TCDD-contaminated areas (Stockbauer et al. 1988). Although several studies provide 

suggestive evidence of a relationship between CDD exposure and alterations in the sex ratio, the data are 

inadequate to establish a causal relationship. Additionally, it is not known how 2,3,7,8-TCDD affects the 

sex ratio. It has been postulated that the effect may be due to an alteration in hormonal balance or a 

disproportional number of miscarriages of male fetuses. 

Data on 2,3,7,8-TCDD-induced alterations in gonads and reproductive endocrine function in humans are 

limited to effects observed in males. Decreased testicular size without any hormonal changes was found in 

Air Force Vietnam veterans exposed to 2,3,7,8-TCDD during Operation Ranch Hand (USAF 1991). This 

finding (decreased testicular size) was not confirmed when a more sensitive measurement device (ultrasound) was used (Henriksen et al. 1996). Wolfe et al. (1985) found no alterations in sperm count or 

morphology in veterans involved in Operation Ranch Hand. Henriksen et al. (1996) assessed the possible 

relationship between 2,3,7,8-TCDD exposure and alterations in testosterone levels, FSH, LH, testicular 

abnormalities, sperm abnormalities, and sperm counts in the Operation Ranch Hand cohort (reproductive 

parameters were assessed in 1982, 1987, and 1992) and found no consistent, statistically significant 

alterations. Increases in FSH and LH levels and decreases in testosterone levels were observed in males 

working in 2,4,5-trichlorophenol manufacturing facilities (NIOSH cohort); however the magnitude of the 

changes in hormone levels was small (Egeland et al. 1994). The study authors note that increases in LH 

levels and decreases in testosterone levels were not found in the same men, suggesting that 2,3,7,8-TCDD 

may result in subtle alterations rather than primary gonadal failure. 

A number of reproductive effects, including decreased fertility, damage to the gonads, and alterations in 

hormone levels, have been observed in male and female animals orally exposed to 2,3,7,8-TCDD. In male 

rats, a dose- and time-dependent reduction of serum testosterone and dihydrotestosterone levels was 

observed after acute oral exposure to 2,3,7,8-TCDD (Mebus et al. 1987; Moore et al. 1985, 1991). 

Furthermore, male rats had decreased weight of seminal vesicles following oral exposure to 2,3,7,8-TCDD 

(Al-Bayati et al. 1988; Moore et al. 1985) and reduced spermatogenesis after oral and subcutaneous 
CDDs 293 

2. HEALTH EFFECTS 

exposure (Al-Bayati et al. 1988; Chahoud et al. 1989; Van Miller et al. 1977). Biochemical changes in rat 

testes included dose- and time-dependent decreases in 17-hydroxylase activity and 20-lyase activity and 

reduced microsomal cytochrome P-450 (Mebus et al. 1987). Decreases in testicular superoxidase dismutase 

and glutathione peroxidase activities, and increases in protein kinase C activity and lipid peroxidation were 

also found in 2,3,7,8-TCDD-exposed rats (Al-Bayati et al. 1988). On the basis of the above data, it was 

postulated that the androgen deficiency is due to decreased androgen synthesis. It was further suggested 

that the morphological changes in rat testes may be due to changes in lipid peroxidation. 

Pre- and/or postimplantation losses have been observed in rats (Giavini et al. 1983; Sparschu et al. 1971a), 

mice (Neubert and Dillman 1972; Smith et al. 1976), and rabbits (Giavini et al. 1982) following acute oral 

exposure to 2,3,7,8-TCDD. A single intraperitoneal injection of 2,3,7,8-TCDD (100 μg/kg) given between 

Gd 2–6 caused a high incidence of resorptions in C57BL/6J mice (Pratt et al. 1984). Similarly, increased 

resorptions were reported in rats exposed to mixed HxCDD during gestation, but not in those exposed to 

2,7-DCDD or OCDD (Schwetz et al. 1973). In addition, abortions were observed in monkeys exposed to 

2,3,7,8-TCDD for 3 weeks by gavage (McNulty 1984), and reduced reproduction was observed in those 

exposed chronically in the feed (Bowman et al. 1989b; Hong et al. 1989; Schantz et al. 1992). Finally, 

significantly decreased fertility in F1 and F2 generations was reported in a 3-generation reproductive study 

in rats exposed to 2,3,7,8-TCDD (Murray et al. 1979). 

Investigations into the mechanism of CDD-induced decreased fertility revealed blocked estrous cycle in 

female mice exposed orally to 2,3,7,8-TCDD for an intermediate duration (Umbreit et al. 1987) and dosedependent decreases in uterine and hepatic cytosolic, and nuclear estrogen and progesterone receptor levels 

in rats after intraperitoneal 2,3,7,8-TCDD injection (Romkes and Safe 1988). Furthermore, 2,3,7,8-TCDD 

antagonized the estradiol-mediated increases in these levels. In addition, a dose-related reduction of uterine 

peroxidase activity and decreased uterine wet weight were seen after a single 2,3,7,8-TCDD injection in rats 

(Astroff and Safe 1990). 2,3,7,8-TCDD application also antagonized the treatment with estradiol. The 

authors concluded that 2,3,7,8-TCDD antagonized the estrogen-induced uterine response and that the Ah 

receptor was involved in mediating the reaction. Other authors suggest that the anti-estrogen effect is 

mediated by 2,3,7,8-TCDD-induced metabolism of estrogens (Gierthy et al. 1987). 

In non-pregnant female rats, decreases in ovarian weight, estrous cyclicity, ovulation rate, and the number 

of ova released were observed following a single dose of 2,3,7,8-TCDD (Li et al. 1995a, 1995b). Increases 

in LH and follicle stimulating hormone levels were also observed. The mechanisms involved in these 
 

CDDs 294 

2. HEALTH EFFECTS 

effects are thought to involve direct effects on the ovaries and effects on the hypothalamus/pituitary axis. 

The normal preovulatory surge of LH was not observed in the 2,3,7,8-TCDD-exposed rats, suggesting that 

2,3,7,8-TCDD inhibited the positive feedback action of 17β-estradiol at the hypothalamic-pituitary axis (Li 

et al. 1995a). In hypophysectomized rats, 2,3,7,8-TCDD exposure resulted in a reduction of ovulation; Li et 

al. (1995a) suggests that this may be the result of a direct effect on the ovary, although the mechanism has 

not been elucidated. 

Endometriosis has been observed in monkeys chronically exposed to 2,3,7,8-TCDD in the diet (Rier et al. 

1993). A possible association between 2,3,7,8-TCDD and endometriosis is supported by rat and mouse 

studies using surgically induced models of endometriosis (Cummings et al. 1996; Johnson et al. 1997). In 

contrast, Foster et al. (1997) found that 2,3,7,8-TCDD exposure diminished endometrial tissue growth in 

mice. These studies used different models of surgically induced endometriosis and highlight the complexity 

of the disease. In the Cummings et al. (1996) and Johnson et al. (1997) studies, the animals were exposed 

to 2,3,7,8-TCDD prior to the development of endometriosis, and immune suppression probably facilitated 

the growth of endometrial tissue. In the Foster et al. (1997) model, 2,3,7,8-TCDD was administered after 

endometriosis development and 2,3,7,8-TCDD, via its anti-estrogenic effects, inhibited tissue growth. The 

relationship between CDD exposure and endometriosis in humans has not been adequately studied. In 

humans, the etiology of endometriosis likely involves a complex interplay between a number of diverse 

physiological factors including altered cell-mediated immunity and increased levels of growth hormone. 

Although the human data regarding reproductive effects are inconsistent, a number of reproductive effects 

have been observed in animals, including deceased fertility, altered hormone levels, and gonad damage in 

males and females. The similarity between some of the effects observed in humans and animals suggest 

that reproductive effects may also occur in humans. 

Developmental Effects. The developmental toxicity of 2,3,7,8-TCDD has been investigated in 

several human populations, with conflicting results. Most studies did not find increases in the number of 

birth defects in the children of men exposed to 2,3,7,8-TCDD in a chlorophenols manufacturing facility 

(Townsend et al. 1982) or during the Vietnam war (Aschengrau and Monson 1990; Erickson et al. 1984; 

Wolfe et al. 1995); or the children of parents living in Seveso, Italy (Bisanti et al. 1980; Mastroiacovo et 

al. 1988). Some studies did find increases in the incidence of specific defects (e.g., talipes, ventricular 

septal defect) in the infants of exposed fathers or mothers and fathers (Aschengrau and Monson 1990; 

Erickson et al. 1984; Hanify et al. 1981; Wolfe et al. 1995), but there was little consistency regarding the 
CDDs 295 

2. HEALTH EFFECTS 

type of defect or the target organ/system. The lack of exposure data, small sample sizes, and the lack of 

reliable data for birth defect rates prior to 2,3,7,8-TCDD exposure precludes drawing conclusions from 

these human studies. A section below summarizes information on health effects in humans associated with 

exposure to CDDs in utero and/or via breast milk. 

Developmental toxicity has been observed in rats, mice, rabbits, hamsters, and monkeys exposed to 

2,3,7,8-TCDD and other CDD congeners. Perinatal exposure to 2,3,7,8-TCDD results in structural 

malformations, functional alterations, decreased growth, and fetal/newborn mortality. Many of the effects 

occurred at 2,3,7,8-TCDD doses which were not maternally toxic. Acute oral exposure to 2,3,7,8-TCDD 

during gestation caused an increased incidence of cleft palate and skeletal anomalies in offspring of rats 

(Giaviani et al. 1983; Huuskonen et al. 1994), mice (Abbott and Birnbaum 1989a; Courtney 1976; 

Dasenbrock et al. 1992; Neubert and Dillman 1972; Smith et al. 1976; Weber et al. 1985), and rabbits 

(Giavini et al. 1983). These effects were also observed in fetuses of mice that received subcutaneous 

injections of 2,3,7,8-TCDD during gestation (Courtney 1976; Poland and Glover 1980). The 

2,3,7,8-TCDD-induced cleft palate is caused by the failure of the opposing palatal shelves to fuse (Pratt et 

al. 1984); 2,3,7,8-TCDD does not alter the size of the palatal shelves or interfere with the opposing shelves 

coming into contact. Under normal conditions, there is a cessation of medial cell proliferation, a 

degeneration of peridermal medial cells, and a transformation of basal cells to mesenchymal cells as the 

opposing palatal shelves come into contact and fuse (Abbott and Birnbaum 1989b). 2,3,7,8-TCDD 

exposure alters medial cell proliferation and differentiation resulting in the formation of stratified 

squamous epithelium. Abbott and Birnbaum (1990a) suggest that the altered proliferation and 

differentiation of the medial cells is due to 2,3,7,8-TCDD-induced reductions of several growth factors 

(EGF, TGF-α, and TGF-β1) and increases in EGF receptor expression. EGF and TGF-α (which both bind 

to the EGF receptor) stimulate epithelial proliferation and differentiation and TGF-β1 inhibits epithelial 

proliferation. The increased levels of EGF receptor appear to compensate for the decreased EGF and TGFα levels resulting in continued proliferation. Abbott et al. (1994a, 1994b) suggest that the altered 

expression of growth factors may be mediated by the Ah receptor. Exposure to 2,3,7,8-TCDD resulted in 

a dose-dependent downregulation of the Ah receptor throughout the palate; this probably occurs at the 

transcriptional level as decreases in mRNA were also observed (Abbott et al. 1994b). There is no evidence 

for direct Ah regulation of growth factors; rather, transcriptional regulation of related genetic activity may 

indirectly influence growth factor expression. Data which support an association between Ah receptor and 

cleft palate include a correlation between 2,3,7,8-TCDD binding to the Ah receptor and altered growth 

factor expression (Abbott et al. 1994b); finding of 2,3,7,8-TCDD-induced altered Ah receptor expression 
 

CDDs 296 

2. HEALTH EFFECTS 

and altered growth factor expression at doses which do not induce cleft palate (Abbott et al. 1994b); and 

the inability of 2,3,7,8-TCDD to induce cleft palate in strains of mice which have low affinity for Ah 

receptors (Pratt et al. 1984; Silkworth et al. 1989b). 

Kidney malformations, particularly hydronephrosis, were observed in the offspring of rats (Giavini et al. 

1983; Huuskonen et al. 1994), mice (Abbott et al. 1987a, 1987b; Courtney 1976; Moore et al. 1973; 

Silkworth et al. 1989b), and hamsters (Gray et al. 1995) orally exposed to 2,3,7,8-TCDD during gestation. 

Kidney defects were also observed in mouse offspring following in utero subcutaneous exposure to 

2,3,7,8-TCDD (Courtney 1976) and in mice postnatally exposed to 2,3,7,8-TCDD via contaminated 

mothers' milk (Couture-Haws et al. 1991b). The hydronephrosis observed in these offspring is the result 

of occlusion of the ureter and subsequent accumulation of urine in the kidney (Abbott et al. 1987a). 

Prenatal exposure to 2,3,7,8-TCDD results in hyperplasia of the epithelium in the ureter, obstruction of the 

ureteric lumen, and a restriction of the flow of urine. Abbott and Birnbaum (1990b) found that 

2,3,7,8-TCDD interfered with the normal decline in EGF receptors in the ureteric epithelia, resulting in 

excessive proliferation. In the bladder, 2,3,7,8-TCDD exposure also resulted in an increase in the 

epithelial thickness and continued expression of EGF receptors. 2,3,7,8-TCDD also appears to directly 

damage the kidney. Under normal conditions, there is an increase in laminin and type IV collagen levels 

and a thickening of the lamina densa of the glomerular basement membrane, which is important in 

establishing the filtration barrier. Following exposure to 2,3,7,8-TCDD, there is a decreased expression of 

laminin and type IV collagen and a diminished thickening of the lamina densa (Abbott et al. 1987b). This 

immature filtration barrier is likely to result in proteinuria and may result in increased urine volume. 

A number of recently published studies have shown that the developing reproductive system is very 

sensitive to the toxicity of 2,3,7,8-TCDD. In female rats, exposure to 2,3,7,8-TCDD on Gd 8 caused 

functional reproductive toxicity (accelerated onset of constant estrus, shortened reproductive lifespan, 

reduced ovarian weight, and cystic hyperplasia of the endometrium) (Gray and Ostby 1995). Although 

there were no effects on fertility or estrous cyclicity when 2,3,7,8-TCDD exposure occurred after 

organogenesis (exposure on Gd 15) (Gray and Ostby 1995), external urogenital malformations (clefting, 

hypospadias, vaginal thread, and delayed vaginal opening) were observed (Flaws et al. 1997; Gray and 

Ostby 1995; Gray et al. 1997a; Heimler et al. 1998). These malformations to external genitalia are likely 

to interfere with mating (Gray and Ostby 1995). The authors note that the effects on the external genitalia 

are similar to effects observed in animals exposed to potent estrogen-like chemicals (e.g., DES, estradiol), 

although it likely that these effects occur by a different mechanism. 
CDDs 297 

2. HEALTH EFFECTS 

In male rats, perinatal exposure to 2,3,7,8-TCDD resulted in alterations in androgen status (decreased 

plasma testosterone levels, delay in testes descent, delay in external signs of puberty, and decreased ventral 

prostate and seminal vesicle weights), testes and cauda epididymis weights, and spermatogenesis 

(decreased daily sperm production, amount of mature sperm in cauda epididymis, and amount of sperm 

ejaculated), and in demasculinization and partial feminization of sexual behavior following exposure on 

Gd 15 (Bjerke and Peterson 1994; Bjerke et al. 1994a, 1994b; Gray et al. 1995; 1997b; Mably et al. 

1992a, 1992b, 1992c; Sommer et al. 1996). In most of these studies, the experimental protocol involved 

gavaging the dams with a single dose of 2,3,7,8-TCDD on Gd 8 (Gray et al. 1995) or 15 (Bjerke and 

Peterson 1994; Bjerke et al. 1994a, 1994b; Gray et al. 1995; Mably et al. 1992a, 1992b, 1992c) and 

assessing a number of indices of reproductive development and function in newborn, juvenile, prepubescent, post-pubescent, and mature male rats. Because 2,3,7,8-TCDD is lipophilic and has a relatively 

long half-life, a single dose on Gd 15 will result in transplacental exposure from Gd 15 to birth and 

exposure via contaminated milk. Bjerke and Peterson (1994) compared the reproductive effects of 

2,3,7,8-TCDD in rats exposed in utero to the effects observed in rats exposed to 2,3,7,8-TCDD only 

during lactation. Both in utero and lactational exposure resulted in decreased plasma testosterone level, 

decreased seminal vesicle and ventral prostate growth, and decreased epididymal sperm reserves. 

Exposure in utero only also resulted in decreased daily sperm production and delayed puberty; and 

exposure by lactation only resulted in partial feminization of sexual behavior. These data suggest that the 

timing of the 2,3,7,8-TCDD exposure is important. The mechanism by which 2,3,7,8-TCDD disrupts the 

development of the reproductive system and whether all of the reproductive effects have similar 

mechanisms is not known. Early investigators of the effects of 2,3,7,8-TCDD on sexual behavior 

suggested that perinatal exposure to 2,3,7,8-TCDD resulted in impaired sexual differentiation of the central 

nervous system (Mably et al. 1992b). The results of the Bjerke et al. (1994b) study suggest that the 

2,3,7,8-TCDD-induced alterations in sexual behavior were not due to 2,3,7,8-TCDD acting as an estrogen 

antagonist or altering ER capacities of hypothalamic nuclei. The volume of the sexually dimorphic 

nucleus in the preoptic area of the hypothalamus (SDN-POA), which is dependent upon testosteronederived estradiol in the brain during perinatal development, was not altered in 2,3,7,8-TCDD-exposed rats. 

Additionally, the sexual differentiation of ER concentration in brain nuclei which exhibit sexual 

dimorphism (ventromedial nuclei, medial preoptic nuclei, bed nucleus of the stria terminalis, 

periventricular preoptic area nucleus, cortical and medial amygdala, and arcuate nucleus) were not affected 

by 2,3,7,8-TCDD. Thus, 2,3,7,8-TCDD effects did not parallel those of either estrogen or androgen 

antagonists. Gray et al. (1995) also support the theory that 2,3,7,8-TCDD does not interfere with 

testosterone- and estrogen-dependent central nervous system sexual differentiation. In their study, no 

alterations in mounting behavior were observed in male hamsters perinatally exposed to 2,3,7,8-TCDD (in 
CDDs 298 

2. HEALTH EFFECTS 

hamsters, masculinization of the central nervous system requires perinatal exposure to testosterone). 

Bjerke et al. (1994b) proposed that 2,3,7,8-TCDD may affect other systems, such as brain amine content or 

growth factor expression of function, which would indirectly impact sexual differentiation. Similarly, 

Gray et al. (1995) suggested that 2,3,7,8-TCDD-induced alterations in the growth factors and receptors 

involved in urogenital system cell differentiation and proliferation may result in alterations in 

morphological sexual differentiation. Bjerke et al. (1994a) also found that the 2,3,7,8-TCDD-induced 

inhibition of ventral prostate weight and protein content imprinting was not due to perinatal reductions in 

plasma androgen levels because no effect on imprinting of the seminal vesicle, penis, or pituitary were 

observed in the 2,3,7,8-TCDD-exposed rats. Using a treatment regime that consisted of administration of 

a loading subcutaneous dose of 2,3,7,8-TCDD to female rats prior to mating, followed by weekly 

maintenance subcutaneous doses during mating, pregnancy, and lactation, Faqi et al. (1998) reported that 

sperm parameters were the most susceptible end points in male offspring examined at puberty (70 days 

old) and adulthood (170 days old). Based on pharmacokinetic considerations, the authors estimated that 

the lowest effective dose was <0.8 ng/kg/day. The sperm parameters that were altered were sperm number 

from cauda epididymis, daily sperm production, sperm transit rate, and percent abnormal sperm (more so 

in adults than in pubertal rats). No significant and/or consistent effects were observed on litter size, sex 

ratio, body weights, developmental landmarks, weight of sex organs, and sexual behavior. Testosterone 

levels were significantly reduced at age 170 days but not at age 70 days. In spite of sperm alterations, all 

exposed males exhibited normal reproductive performance and successfully impregnated untreated female 

to produce viable fetuses. 

Recent studies have also focused on the role of the Ah receptor in the 2,3,7,8-TCDD-induced alterations in 

the development of the male reproductive system. Roman et al. (1998a) recently demonstrated the presence 

of both the Ah receptor and the receptor nuclear translocator (Arnt) in the testis, epididymis, vas deferens, 

ventral and dorsolateral prostate, and seminal vesicles from adult Holtzman rats. Arnt was localized in all 

organs in a variety of cell types; subcellular localization varied across organs and cell types within these 

organs. Unfortunately, technical difficulties precluded the evaluation of the Ah receptor distribution in the 

various organs. The authors also showed that a single oral dose of 25 μg 2,3,7,8-TCDD/kg produced 

significant induction of CYP1A1 in the ventral and dorsolateral prostate. CYP1A1 expression was 

localized in the epithelial cells of the ventral and lateral lobes of the prostate. Less CYP1A1 induction was 

seen in selected epithelial cells from other tissues, and no induction was detected in the testis. Also, 

2,3,7,8-TCDD had no effect on Arnt protein expression, but Ah receptor expression was significantly 

reduced in all organs examined. In another study from this series, Roman and Peterson (1998) found that, 
CDDs 299 

2. HEALTH EFFECTS 

relative to controls, in utero exposure to 2,3,7,8-TCDD (1 μg/kg) transiently decreased the amount of 

several prostate-specific androgen-regulated mRNAs, all of which are markers of a differentiated ductal 

epithelium. This was in contrast with observations in adults, in which 2,3,7,8-TCDD induced CYP1A1 

mRNA without altering the amount of prostate-specific, androgen-regulated mRNAs. These results 

suggested that the developing prostate can directly respond to in utero and lactational exposure to 

2,3,7,8-TCDD, and that this exposure not only impairs prostate growth but also delays prostate luminal 

epithelial cell differentiation. In yet an additional study from this series, Roman et al. (1998b) reported that 

in the most severely affected animals, 2,3,7,8-TCDD produced alterations in the histological arrangement of 

epithelial and stromal cells and in the spatial distribution of androgen receptor expression. 

Other developmental effects that have been observed in animals include immunotoxicity (thymic atrophy, 

immunosuppression, and alterations in thymocyte phenotypes) (Fine et al. 1989; Gehrs et al. 1997a, 1997b; 

Håkansson et al. 1987; Huuskonen et al. 1994; Luster et al. 1980; Madsen and Larsen 1989; Thomas and 

Hinsdill 1979), decreased fetal and newborn body weight (Abbott et al. 1992; Bjerke et al. 1994a; Bjerke 

and Peterson 1994), fetal/newborn mortality or decreased survival (Bjerke et al. 1994a; Bjerke and Peterson 

1994; Huuskonen et al. 1994; McNulty 1984; Murray et al. 1979; Nau et al. 1986), and altered social 

behavior (Schantz et al. 1992). 

Developmental toxicity has also been observed in animals exposed to other CDDs. These effects include 

heart defects in rats exposed to 2,7-DCDD (Schwetz et al. 1973); decreased thymic weight in rats exposed 

to 1,2,3,7,8-PCDD (Madsen and Larsen 1989); subcutaneous edema, decreased fetal growth, delayed 

ossification, dilated renal pelvis, and cleft palate in rats exposed to HxCDD (Schwetz et al. 1973); and 

subcutaneous edema in rats exposed to OCDD (Schwetz et al. 1973). 

The animal database provides strong evidence that developmental toxicity is a sensitive end point following 

2,3,7,8-TCDD exposure. Structural malformations, functional alterations (including impaired development 

of reproductive system), decreased growth, and fetal/newborn mortality have been observed in several 

animal species. Limited human data on the developmental toxicity of CDDs is available. Most of these 

studies examined the occurrence of birth defects in children of males exposed to 2,3,7,8-TCDD. 

Deficiencies in the human data preclude drawing firm conclusion on the potential of 2,3,7,8-TCDD to 

induce developmental effects in humans. However, the animal data suggest that 2,3,7,8-TCDD is a likely 

human developmental toxicant. 
CDDs 300 

2. HEALTH EFFECTS 

Health Effects Associated with Exposure to CDDs in Breast Milk. The developing organism is very 

susceptible to the toxicity of CDDs, in particular 2,3,7,8-TCDD. Prenatal or perinatal exposure has resulted 

in structural malformations (e.g., cleft palate, hydronephrosis), functional alterations (e.g., damage to the 

immune system, impaired development of the reproductive system), decreased growth, and fetal/newborn 

mortality in several animal species. Additionally, several animal studies (summarized in Table 2-14) 

provide evidence that lactation-only exposure to 2,3,7,8-TCDD can adversely affect the developing animal. 

Impaired development of the reproductive system (Bjerke and Peterson 1994), increased incidence of 

hydronephrosis (Couture-Haws et al. 1991a, 1991b), decreased weight gain, thymic atrophy (Faith and 

Moore 1977; Håkansson et al. 1987), and suppression of cell-mediated immunity (Faith and Moore 1977) 

have been observed in rats and mice exposed to 2,3,7,8-TCDD during lactation but not during gestation. 

The authors of these studies noted that most of the effects observed following lactation-only exposure were 

similar to those observed in animals exposed to 2,3,7,8-TCDD during gestation. 

Because CDDs are efficiently absorbed following ingestion of breast milk (approximately 95% absorption 

efficiency for most congeners, see Section 2.3.4.4 for more information) and animal studies have found that 

lactation-only exposure can result in developmental effects, there is concern that breast-fed infants of 

women with high background levels of CDDs may be at risk. Ayotte et al. (1996) predicted that exposure 

to CDDs and related chemicals from breast milk will strongly influence the body burden of these chemicals 

during childhood and adolescence. Several human studies have examined the possible association between 

background CDD and CDF levels from in utero exposure and exposure from breast milk, and adverse 

health effects in infants. These studies (summarized in Table 2-15) found alterations in the levels of some 

markers of liver function (plasma ALT and AST) (Pluim et al. 1994a), thyroid function (thyroxine, thyroid 

stimulating hormone) (Koopman-Esseboom et al. 1994; Pluim et al. 1993b), and immune function (T-cell 

markers [TcRγδ+

, CD3+CD8+, and TcRαβ+] and monocyte) (Weisglas-Kuperus et al. 1995), and the 

neurological optimality score in infants (Huisman et al. 1995a) which significantly correlated with CDD 

and CDF TEQ levels in breast milk. In follow-up studies and studies of older infants or children, no 

relationship between high levels of CDDs, CDFs, and PCBs in breast milk and neurological development, 

neurological optimality score, and/or reflexes was found at ages 6 (Pluim et al. 1996), 18 (Huisman et al. 

1995b), or 31 months (Ilsen et al. 1996). Although the Ilsen et al. (1996) study of 31-month-old children 

did not find any alterations in overall neurological optimality or suboptimality scores, significant 

alterations, indicative of enhanced neuromuscular maturation and higher reflexes, were found in some tests 

(results were still within the normal range). Hypomineralization of teeth was found in 6- to 7-year-old 

children who 



 

CDDs 304 

2. HEALTH EFFECTS 

received higher than background levels of CDDs and CDFs in breast milk (Alaluusua et al. 1996). Other 

studies have not found a relationship between higher background levels of CDDs, CDFs, and PCBs in 

breast milk and adverse health effects (summarized in Table 2-15); decarboxylated prothrombin and 

vitamin K levels in 11-week-old infants (Pluim et al. 1994b); birth weight (Pluim et al. 1996; Vartiainen et 

al. 1998), head circumference, and body weight at 1, 10, and 26 weeks of age (Pluim et al. 1996); or liver 

size (Pluim et al. 1996) were not adversely affected. Although significant correlations were found, the data 

should be interpreted cautiously because the levels of these markers were within the normal range and the 

correlation coefficients were low suggesting that only a small amount of the variance in the marker 

concentrations can be attributed to CDD and CDF levels. 

The animal data suggest that lactation-only exposure to relatively high concentrations of CDDs can result in 

serious health effects. However, the human data show that the risk of CDD-induced health effects in infants 

exposed to background levels of CDDs and CDFs in breastmilk is small and this risk, in most cases, does 

not outweigh the benefits of breast-feeding. 

Genotoxic Effects. In vivo genotoxicity studies are summarized in Table 2-16. Human studies have 

been conducted on populations exposed to 2,3,7,8-TCDD. An increased incidence of chromosomal 

aberrations was found in the fetal tissues, but not in the maternal tissues, following induced abortions in a 

group of women exposed to 2,3,7,8-TCDD in the Seveso accident (Tenchini et al. 1983). However, cases 

treated for chloracne in the area did not have an elevated frequency of chromosomal aberrations (Reggiani 

1980). Results of a higher incidence of chromosomal aberrations were inconsistent in groups of Vietnam 

veterans (Kaye et al. 1985) or no cytogenetic changes were reported (Mulcahy et al. 1980). Fewer birth 

defects due to chromosomal abnormalities in children of Vietnam veterans were reported in another study 

(Erickson et al. 1984). Human studies cited above were limited by several factors. Generally, the levels of 

exposure to 2,3,7,8-TCDD were not known and coexposure to other potentially active compounds occurred 

in all studies. In the case of Vietnam veterans, a long postexposure period passed before the cytogenetic 

analysis was done. Furthermore, most of the studies used groups that were too small (less than 20 

individuals) to have the statistical power to detect any changes. 

CDDs 306 

2. HEALTH EFFECTS 

In a study in which current 2,3,7,8-TCDD blood levels of previously exposed workers were approximately 

25 times higher than in referents, there was no evidence of increased incidence of chromosomal aberrations 

or sister chromatid exchanges (Zober et al. 1993). 

Animal studies on the genotoxicity of CDDs are inconclusive. When Osborne-Mendel rats were given 

2,3,7,8-TCDD (0.25, 0.5, 1, 2, or 4 μg/kg) by gavage twice a week for 13 weeks, increased incidence of 

chromosomal aberrations was observed in the highest-exposure group (Green et al. 1977). Increased 

incidences of gaps and chromatid aberrations were observed in bone marrow cells of CD-1 mice following 

an intraperitoneal injection of 10 μg/kg 2,3,7,8-TCDD (Loprieno et al. 1982). Positive results were 

obtained at 96 hours, but not 24 hours, posttreatment. In contrast, no induction of structural chromosomal 

changes was found in CD-COBS rats orally exposed to 1.0, 0.1, or 0.01 μg/kg 2,3,7,8-TCDD once a week 

for 45 weeks (Loprieno et al. 1982). In addition, no differences in the frequency of sister chromatid 

exchanges and chromosomal aberrations in peripheral lymphocytes were observed in a group of rhesus 

monkeys receiving 0.001 μg/kg 2,3,7,8-TCDD in the feed for 4 years and their matching controls (Lim et 

al. 1987). Furthermore, no induction of chromosomal aberrations or sister chromatid exchanges, or 

increases in the frequency of micronuclei, were found in bone marrow cells of C57BL/6J (with high-affinity 

2,3,7,8-TCDD receptor) or DBA/2J mice (with low-affinity 2,3,7,8-TCDD receptor) following a single 

intraperitoneal injection of 2,3,7,8-TCDD at doses of 50, 100, or 150 μg/kg (Meyne et al. 1985). The 

samples were examined within 8–48 hours. The negative results may, however, have been due to the timedependent detectability of chromosomal changes after CDD exposure reported earlier (Loprieno et al. 

1982). 

In addition to studies dealing with structural chromosomal changes, effects on DNA were also investigated. 

Oral exposure to 1 μg/kg/week of 2,3,7,8-TCDD or 1,2,3,7,8-PCDD for up to 6 months did not increase the 

formation of DNA adducts in Sprague-Dawley rats (Randerath et al. 1989). A single oral dose of 

2,3,7,8-TCDD (25–100 μg/kg) caused time-dependent increases in the induction of DNA single-strand 

breaks (and lipid peroxidation) in hepatic cells of Sprague-Dawley rats terminated within 3–14 days after 

the treatment (Wahba et al. 1989). 

Negative results were obtained in reproductive tests including a dominant-lethal test following 7 daily oral 

doses of 2,3,7,8-TCDD (4, 8, or 12 μg/kg/day) to male Wistar rats (Khera and Ruddick 1973) and a sexlinked recessive-lethal test with 2,3,7,8-TCDD in Drosophila melanogaster (Zimmering et al. 1985). 
 

CDDs 307 

2. HEALTH EFFECTS 

In vitro genotoxicity studies are summarized in Table 2-17. Eukaryotic cell systems were used for detecting 

the effects of 2,3,7,8-TCDD exposure on DNA. Exposure to 2,3,7,8-TCDD did not stimulate the 

unscheduled DNA synthesis in cultural human cells (Loprieno et al. 1982), but inhibited DNA, ribonucleic 

acid (RNA), and protein synthesis in mouse lymphocytes (Luster et al. 1979); caused gene mutations in 

mouse lymphoma cells (Rogers et al. 1982); and induced sister chromatid exchanges in Chinese hamster 

cells (Toth et al. 1984). 

Several researchers used the Ames test with Salmonella typhimurium to assess the mutagenicity of 

2,3,7,8-TCDD in prokaryotic organisms. Predominantly negative results were obtained with tester strains 

G46, TA 1530, TA 1535, TA 100, TA 1950, and TA 1975, revealing base pair substitutions; and with 

strains TA 1531, TA 1532, TA 1534, TA 1538, TA 98, and TA 1978, revealing frame shift mutations 

(Geiger and Neal 1981; Gilbert et al. 1980; Mortelmans et al. 1984; Toth et al. 1984). However, some of 

the studies were limited by using 2,3,7,8-TCDD concentrations in excess of its solubility in water. Only 

two early studies reported positive results (Hussain et al. 1972; Seiler 1973). However, the results were 

limited by failure to demonstrate a dose-response relationship and by low bacterial survival rates. In 

addition, 2,3,7,8-TCDD exposure induced reverse mutations in Escherichia coli (Hussain et al. 1972) and in 

Saccharomyces cerevisiae (Bronzetti et al. 1983). The conflicting data obtained in the above studies may 

result from technical difficulties in testing 2,3,7,8-TCDD rather than from a lack of biological activity. 

Testing difficulties arise from an extreme insolubility of this compound and a high toxicity observed in 

some test systems, which would be anticipated to result in a very narrow window for effective genotoxic 

doses. 

Considering the inconclusive results reported above and the severe limitations of some studies, there is no 

strong evidence for 2,3,7,8-TCDD genotoxicity. The information regarding the mutagenic potential of 

other CDDs is even more limited. 

Cancer. Numerous epidemiological studies investigated the effects of 2,3,7,8-TCDD exposure on the 

development of cancer. A number of large-scale retrospective cohort mortality studies (Becher et al. 1996; 

Fingerhut et al. 1991; Hooiveld et al. 1998; Kogevinas et al. 1993, 1997; Manz et al. 1991; Ott and Zober 

1996; Zober et al. 1990) have found significant increases in cancer mortalities (all types of cancers 

combined). These increases were typically found in the highest exposed workers and in workers with the 

longest latency periods. In general, the SMRs were low (less than 1.5); however, the high degree of 

consistency between studies suggests that the increases in mortalities were not due to chance. The possible 


CDDs 310 

2. HEALTH EFFECTS 

risks of several specific types of cancer have also been found, but the data are somewhat inconsistent. The 

site-specific cancers with elevated possible risks include soft-tissue sarcoma (Eriksson et al. 1981, 1990; 

Fingerhut et al. 1991; Hardell and Eriksson 1988; Hardell and Sandrom 1979; Hardell et al. 1995; 

Kogevinas et al. 1995, 1997; Saracci et al. 1991; Smith et al. 1984a), non-Hodgkin's lymphoma or 

malignant lymphoma (Becher et al. 1996; Cantor 1982; Hardell et al. 1981; Kogevinas et al. 1995), 

respiratory tract cancer (Fingerhut et al. 1991; Kogevinas et al. 1997; Manz et al. 1991; Zober et al. 1990), 

and gastrointestinal organ cancers (Axelson et al. 1980; Thiess et al. 1982). Furthermore, an increased risk 

of benign systemic neoplasms was reported in Vietnam Air Force veterans involved in Operation Ranch 

Hand (USAF 1991). There is some uncertainty regarding the interpretation of the epidemiology study 

results. In most studies, the cohort was also exposed to chemicals other than 2,3,7,8-TCDD and exact 

levels of exposure were not known. Furthermore, in some studies the exposure data were based solely on 

questionnaires and some recall bias could have been present. Other studies suffered from examining small 

cohorts or investigating the effects after a short latency period. The long latency period is important for 

detecting increases in soft-tissue sarcomas, presumably a major cancer outcome of CDD exposure in 

humans. 

Several studies provided evidence of CDD-related carcinogenicity in animals. In general, the effects were 

dependent on the congener, species, sex, and route of administration, and were seen at doses that were close 

to doses that are toxic in the same animal species. Intermediate- and chronic-duration oral exposure to 

2,3,7,8-TCDD induced multiple-site carcinomas and/or sarcomas in rats (Kociba et al. 1978a; NTP 1982b) 

and mice (Della Porta et al. 1987; NTP 1982a, 1982b; Toth et al. 1979). Similarly, chronic oral exposure to 

a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD induced carcinomas in mice and rats (NCI/NTP 

1980), and exposure to 2,7-DCDD caused carcinomas and sarcomas in mice (NCI/NTP 1979a). However, 

no cancer effects were found following chronic exposure to 2,7-DCDD in rats (NCI/NTP 1979a). 

Furthermore, squamous cell carcinoma developed in hamsters (Rao et al. 1988) following intermediateduration intraperitoneal exposures. 

Short-term dermal studies with 2,3,7,8-TCDD had controversial results. Some studies reported its 

inhibitory effects on the development of skin tumors in mice otherwise initiated by 13-dimethylbenz­

(o)anthracene (Berry et al. 1978, 1979). Others cited its ability to promote tumors initiated by N-methylN-nitro-N-nitrosoguanidine (Hebert et al. 1990; Poland et al. 1982). Further, intraperitoneal injection of 

2,3,7,8-TCDD given 2 days prior to or concurrently with methylcholanthrene did not affect methylcholanthrene-induced carcinogenicity in C57BL/6 mice (Kouri et al. 1978); in contrast, 2,3,7,8-TCDD 
CDDs 311 

2. HEALTH EFFECTS 

pretreatment (intraperitoneal or subcutaneous) of DBA/2 mice slightly increased the carcinogenic index. In 

support of these data, promotion of GGT-positive hepatic foci and/or development of tumors was observed 

after initiation with nitrosodiethylamine in rats (Flodstrom and Ahlborg 1989; Flodstrom et al. 1991; Pitot 

et al. 1980) that were injected with 2,3,7,8-TCDD for an intermediate duration. A recent study of the 

promoting activity of 2,3,7,8-TCDD in the liver from female Sprague-Dawley rats showed that increased 

tissue burden of 2,3,7,8-TCDD, which correlated with increased CYP1A1 activity, did not necessarily lead 

to increased cell proliferation (Walker et al. 1998). Experimentally, cell proliferation was increased after 30 

or more weeks of treatment, but not after only 14 weeks of treatment, whereas both tissue burden and 

CYP1A1 activity exhibited similar significant increases at both time points. Walker et al. (1998) noted that 

a dose metric such as the area under the curve, which measures total dose over time, did not correspond to 

either 2,3,7,8-TCDD-induced changes in cell proliferation or changes in CYP1A1 expression. This, 

according to the authors, suggested that for a number of 2,3,7,8-TCDD-induced responses, particularly 

those involving integrated signal transduction pathways such as altered cell/tissue growth and 

differentiation, dose metrics that incorporate not only magnitude of exposure, but also duration of exposure 

and temporal windows of sensitivity for the response, may be more appropriate. 

The available data provided sufficient evidence that 2,3,7,8-TCDD is a carcinogen in animals and its action 

is not solely dependent upon initiation by other substances. This is in conflict with the inconclusive genotoxicity data. Significant binding of radioactivity derived from labeled 2,3,7,8-TCDD to liver proteins was 

observed in several studies. However, covalent binding to hepatic DNA was four times less than the levels 

of binding with other carcinogens (Poland and Glover 1979). This indicates that the typical mutation 

mechanism model (covalent binding/DNA alteration) may not be applicable in the case of CDDs. In 

addition, there is an evidence that 2,3,7,8-TCDD acts as a tumor promoter (Hebert et al. 1990; Poland et al. 

1982), which is consistent with the increases in multiple-site tumors observed in exposed humans and 

animals. 

2,3,7,8-TCDD is an atypical chemical because of its accumulation and long persistence in the body. 

Several studies demonstrated that 2,3,7,8-TCDD affects the adrenals, thymus (DiBartolomeis et al. 1987; 

Gorski et al. 1988b; Greenlee et al. 1985; Hochstein et al. 1988), and thyroid (Henry and Gasiewicz 1987; 

Hermansky et al. 1988; Hong et al. 1987; Lu et al. 1986; Rozman et al. 1985) and also alters the estradiol 

(Umbreit et al. 1987), testosterone, and dihydrotestosterone (Mebus et al. 1987; Moore et al. 1985) levels in 

the organism. A study with intact and ovariectomized rats indicated that ovarian estrogens are involved in 

2,3,7,8-TCDD induced hepatocarcinogenesis (Lucier et al. 1991). Assuming that there is a relationship 
 

CDDs 312 

2. HEALTH EFFECTS 

between the 2,3,7,8-TCDD Ah receptor protein function and the steroid and thyroid receptor protein 

functions, 2,3,7,8-TCDD would interact with various hormone receptors (Holder and Menzel 1989). It has 

been proposed that 2,3,7,8-TCDD is a hormonal carcinogen causing effects in targeted organs and in 

secondary targets through hormonal imbalance. Furthermore, 2,3,7,8-TCDD may also promote the metabolism of procarcinogens to active intermediates by the induction of metabolizing enzymes as demonstrated 

in vitro in cultured human lymphocytes (Jaiswal et al. 1985; Kouri et al. 1974, 1978). The induction of 

these enzymes appears to be the subject of genetic polymorphism such that individuals who are highly 

inducible may be at high risk for the development of tumors. 

Taken together, the results of the epidemiology studies and the animal studies suggest that 2,3,7,8-TCDD 

may be a human carcinogen. This is consistent with conclusions of several regulatory agencies. NTP 

(1998) considers 2,3,7,8-TCDD to be a substance that may reasonably be anticipated to be a carcinogen 

(limited evidence in humans, sufficient evidence in animals); NTP is currently considering a reclassification 

of 2,3,7,8-TCDD and the decision is pending. IARC (1997) has recently classified 2,3,7,8-TCDD in Group 

1 based on limited evidence of carcinogenicity in humans and sufficient evidence in animals. EPA had 

classified 2,3,7,8-TCDD as a Group B2 carcinogen when considered alone and a Group B1 carcinogen 

when considered in association with phenoxyherbicides and/or chlorophenols (EPA 1985d, 1989d, 1991a). 

The Group B2 classification indicates that although evidence in humans is inadequate, the evidence in 

animals is sufficient to consider 2,3,7,8-TCDD a probable human carcinogen. The Group B1 classification 

indicates that there are not only sufficient animal data but also limited human data to support the 

consideration that 2,3,7,8-TCDD, in association with phenoxyherbicides and/or chlorophenols, is a probable 

human carcinogen. Moreover, in a proposed rule to add “Dioxin and Dioxin-Like Compounds” to the list 

of chemicals subject to release reporting requirements, EPA reiterated that, “Based on the EPA weight of 

evidence classification criteria, there is sufficient evidence to conclude that 2,3,7,8-TCDD is a probable 

human carcinogen” (EPA 1997c). 

2.6 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 due to maternal exposure during gestation and lactation. Relevant animal and in vitro 

models are also discussed. 
CDDs 313 

2. 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 5.6, Exposures of Children. 

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 pre-natal and post-natal 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 and 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 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 the newborn who has a low glomerular filtration 

rate and has 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 lifetime in which to express damage from chemicals; this 

potential is particularly relevant to cancer. 

Certain characteristics of the developing human may increase exposure or susceptibility while others may 

decrease susceptibility to the same chemical. For example, the fact that infants breathe more air per 
CDDs 314 

2. HEALTH EFFECTS 

kilogram of body weight than adults may be somewhat counterbalanced by their alveoli being less 

developed, so there is a disproportionately smaller surface area for absorption (NRC 1993). 

There is a limited amount of information available on the toxicity of CDDs in children. Most of the 

available data come from a series of studies on children living in Seveso during the accidental release of 

airborne trichlorophenol contaminated with 2,3,7,8-TCDD. Shortly after the accident, early irritative 

dermal lesions (this effect may not have been related to 2,3,7,8-TCDD exposure) and chloracne were 

observed in a number of children. Erythema and edema, the main clinical features of the early irritative 

lesions, were only observed in children and young adults (less than 20 years old) (Caputo et al. 1988). 

Chloracne was observed in 187 individuals, 88% of them were children aged 0 to 14 years (Bisanti et al. 

1980). Based on serum 2,3,7,8-TCDD levels measured in 30 Seveso residents with and without chloracne, 

Mocarelli et al. (1991) suggested that children may develop chloracne at lower 2,3,7,8-TCDD body burdens 

than adults following acute exposure to 2,3,7,8-TCDD. Other effects observed in the exposed children 

include a significant increase in the number of children with chloracne having clinical and 

electrophysiological signs of peripheral nervous system involvement (assessed 6 years after the accident) 

(Barbieri et al. 1988) and slight transient increases in serum γ-glutamyltransferase and alanine 

aminotransferase levels in boys aged 6-10 years (Mocarelli et al. 1986). Although the serum enzyme levels 

were higher than in non-exposed children, the values were within the normal range and were elevated 1, 2, 

and 3 years after the accident, but not after 4 or 5 years. Increased risks of Hodgkin’s lymphoma, myeloid 

leukemia, and thyroid cancer were also reported among children who were 0–19 years old at the time of the 

Seveso accident (Pesatori et al. 1993). However, the differences in relative risks (RRs) for these cancer 

types between the Seveso residents and the control population did not reach statistical significance. Similar 

results were found in a 15-year follow-up study of this cohort (Bertazzi et al. 1997). 

A wide variety of effects have been observed in adults exposed to 2,3,7,8-TCDD at work or following an 

accidental release of 2,3,7,8-TCDD into the environment. The primary targets appear to be the skin, liver, 

thyroid, and cardiovascular, endocrine, and immune systems; an increased cancer risk has also been 

observed. In the absence of data to the contrary, it is likely that these organs/systems will also be sensitive 

targets in children. 

A number of human studies have investigated the potential of 2,3,7,8-TCDD to induce developmental 

effects. No significant increases in the incidence of birth defects have been observed in the children of 

parents living in Seveso at the time of the accident or during the next 6-year period (Bisanti et al. 1980; 
CDDs 315 

2. HEALTH EFFECTS 

Mastroiacovo et al. 1988) or in the children of men involved in the manufacture of chlorophenols 

(Townsend et al. 1982). In contrast, other studies have found increases in specific types of defects, 

although the total number of defects was not significantly altered. It is difficult to interpret these data 

because there is little consistency regarding the type of defect or the target organ/system. For example, a 

significant association between nervous system defects and paternal serum 2,3,7,8-TCDD levels was 

observed in the Ranch Hand cohort (Wolfe et al. 1995) and facial clefts were observed in Arkansas 

residents exposed to sprayed 2,4,5-T. The lack of exposure data, small sample sizes, and the lack of reliable 

data for birth defect rates prior to 2,3,7,8-TCDD exposure limits the power of the human studies to 

determine if an association between 2,3,7,8-TCDD exposure and developmental toxicity exists in humans. 

The toxicity of 2,3,7,8-TCDD has been extensively examined in animal oral toxicity studies, and effects 

have been observed in most organs/systems. The animal studies clearly demonstrate that the developing 

organism is very sensitive to the toxicity of 2,3,7,8-TCDD. The types of effects observed in the offspring 

of animals exposed to 2,3,7,8-TCDD include fetal/newborn mortality, decreased growth, structural 

malformations, kidney anomalies, immunotoxicity, thymic atrophy impaired development of the reproductive system, and neurodevelopmental effects. The LOAELs for developmental effects are among the 

lowest identified in animals, and the chronic oral MRL is based on a developmental effect. The most 

sensitive developmental effects are impaired development of the reproductive system and neurobehavioral 

effects. In utero exposure to 2,3,7,8-TCDD adversely affects the development of the reproductive system in 

male and female offspring; studies have shown alterations in androgen levels, secondary sex organs, 

spermatogenesis, fertility, and sexual behaviors (Bjerke and Peterson 1994; Bjerke et al. 1994a, 1994b; 

Chaffin et al. 1996, 1997; Flaws et al. 1997; Gray and Ostby 1995; Gray et al. 1995, 1997a, 1997b; Heimler 

et al. 1998; Mably et al. 1992a, 1992b, 1992c). Gray and Ostby (1995) found decreased fertility in the 

female offspring exposed on Gd 8; no effects on fertility have been observed in female offspring exposed 

on Gd 15 (Gray et al. 1997a) or in male offspring (Gray et al. 1995; Mably et al. 1992c). Schantz and 

Bowman (Bowman et al. 1989b; Schantz et al. 1986, 1992; Schantz and Bowman 1989) found neurobehavioral alterations in the offspring of monkeys chronically exposed to dietary 2,3,7,8-TCDD (7 months 

prior to mating and during mating and lactation). Altered peer group behavior, cognitive deficits, and 

prolonged maternal care were observed. There are some data to suggest that other CDDs (2,7-DCDD, 

1,2,3,7,8-PCDD, OCDD, and HxCDD) are also toxic to the developing organism (Madse and Larsen 1989; 

Schwetz et al. 1973). The observed developmental effects appear to be similar to those observed following 

in utero exposure to 2,3,7,8-TCDD. More details about these studies can be found in Sections 2.2.2.6, 

Developmental Effects, and 2.4.2, Mechanisms of Toxicity. 
CDDs 316 

2. HEALTH EFFECTS 

There is a limited amount of data on the toxicokinetic properties of CDDs in children or immature animals. 

A toxicokinetic model was constructed that accurately predicted the lifetime concentrations of 

2,3,7,8-TCDD in adipose tissue, blood, liver, and feces at different ages (Kreuzer et al. 1997). In formulafed infants, the model predicted that 2,3,7,8-TCDD lipid levels would decrease during the first year and 

subsequently increase, reaching a maximum at 16 years of age. In contrast, the model predicted an initial 

increase in 2,3,7,8-TCDD lipid levels in exclusively breast-fed infants followed by a 3-year decrease after 

weaning and merging at about 7 years with concentrations in formula-fed individuals. The half-life of 

nonmetabolic elimination (unchanged 2,3,7,8-TCDD) was calculated to be 0.42 years in newborns and 

9.5 years in 40-year-old adults. The half-life of the fraction metabolized by the liver ranged from 1.5 years 

for newborns to approximately 10 years for a 40-year-old individual. The three times greater elimination 

half-life for the metabolized fraction relative to the nonmetabolized fraction in infants suggests that 

metabolic elimination does not play a major role in the elimination of 2,3,7,8-TCDD in infants. 

2,3,7,8-TCDD accumulates preferentially in liver and adipose tissue. Accumulation in the liver is due to 

sequestration by the microsomal binding protein, CYP1A2. To the extent that this protein is developmentally regulated (Leeder and Kerns 1997), infants (<4 months old) might accumulate relatively less 

2,3,7,8-TCDD in their livers than adults. Little is known about the metabolism of 2,3,7,8-TCDD in humans 

and it is unknown whether the metabolism of 2,3,7,8-TCDD or other CDDs differs between adults and 

children. In animals, phase II enzymes play an important role in the biotransformation and elimination of 

2,3,7,8-TCDD. If this were the case in humans, it would be expected that very young infants would 

metabolize and eliminate 2,3,7,8-TCDD slower than adults since glucuronosyltransferase activity achieves 

adult levels by 6–18 months of age (Leeder and Kearns 1997). 

CDDs are transferred from mother to offspring through the placenta and breast milk. Although there are 

human data indicating placental transfer of 2,3,7,8-TCDD (Kreuzer et al. 1997; Schecter et al. 1996b), 

quantitative data are not available. A study in mice administered a single dose of 2,3,7,8-TCDD on Gd 12 

showed that the rate of accumulation of 2,3,7,8-TCDD in placental tissue reached a maximum in about 

3 hours (Abbott et al. 1996); after 24 hours, 0.27% of the maternal dose was detected in the placenta. This 

issue is discussed in more detail in Section 2.3.4.4, Transfer of CDDs Through the Placenta and Breast 

Milk. 

CDDs are lipophilic compounds that can concentrate in maternal milk and be transferred to the nursing 

infant. Numerous studies have examined the transfer of 2,3,7,8-TCDD and related chemicals to infants via 

breast milk and for the most part, the results showed that infants may absorb up to 95% of the administered 
CDDs 317 

2. HEALTH EFFECTS 

dose (Abraham et al. 1994, 1996; Dahl et al. 1995; McLachlan 1993; Pluim et al. 1993b). This percentage 

is similar to the percent of 2,3,7,8-TCDD absorbed (>87%) by an adult volunteer after ingestion of a single 

oral dose of 2,3,7,8-TCDD (Poiger and Schlatter 1986). As stated previously, it has also been shown that 

breast-fed infants have a larger 2,3,7,8-TCDD burden during the first year of life compared to formula-fed 

infants (Kreuzer et al. 1997). However, this initial higher burden does not translate into a higher lifetime 

burden. A number of human studies have examined breast-fed infants of mothers with high background 

levels of CDDs. These studies have found alterations in some markers of liver, thyroid, and immune 

function and neurodevelopment (neurological optimality score) (Huisman et al. 1995a; Koopman-Esseboom 

et al. 1994; Pluim et al. 1993b, 1994a; Weisglas-Kuperus et al. 1995); however, all of the markers were 

within the normal range. The impaired neurological optimality score that was observed in newborns was 

not significantly altered in children aged 6, 18, or 31 months (Ilsen et al. 1996; Huisman et al. 1995b; Pluim 

et al. 1996). 

Subsequent sections of this chapter (Sections 2.7, 2.8, and 2.10) discuss the available information on 

biomarkers, interactions, and methods for reducing toxic effects. Most of the available information is from 

adults and mature animals; no child-specific information was identified, with the possible exception of 

biomarker data. However, there are some data to suggest that interactions with PCBs and CDFs may 

influence the developmental toxicity of 2,3,7,8-TCDD. Data from children living in Seveso suggest that 

serum 2,3,7,8-TCDD levels are reflective of exposure levels and are a sensitive indicator of past exposure. 

Likewise, it is likely that the available information in adults on interactions and methods for reducing toxic 

effects will also be applicable to children. 

As discussed previously, children appear to be unusually susceptible to the dermal toxicity of 

2,3,7,8-TCDD. The data are inadequate to assess whether they will also be more sensitive to other CDD 

effects. Additionally, the available animal data suggest that the developing fetus is very sensitive to 

2,3,7,8-TCDD-induced toxicity. 2,3,7,8-TCDD appears to interfere with the development of the 

reproductive, immune, and nervous systems; the mechanisms of action for these toxic effects have not been 

elucidated. 
CDDs 318 

2. HEALTH EFFECTS 

2.7 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 (NRC 1989). The 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 CDDs are discussed in Section 2.7.1. 

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 CDDs are 

discussed in Section 2.7.2. 

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 
CDDs 319 

2. HEALTH EFFECTS 

biologically effective dose, or a target tissue response. If biomarkers of susceptibility exist, they are 

discussed in Section 2.9, Populations That Are Unusually Susceptible. 

2.7.1 Biomarkers Used to Identify or Quantify Exposure to CDDs 

Methods for measuring CDDs in biological fluids and tissues are available. Adipose tissue and liver are the 

primary storage site for CDDs and tissue samples have been analyzed in several studies. It was 

demonstrated that the relative (lipid-based) levels of 2,3,7,8-TCDD are similar in hepatic and adipose 

tissues (Leung et al. 1990a) and between adipose tissue and serum (Patterson et al. 1988; Schecter et al. 

1990) from the same patients. However, this was not the case for more highly chlorinated dioxins; for 

example, for OCDD there is a 2:1 ratio between serum and adipose tissue lipid fractions (Schecter et al. 

1990) and a 12:1 ratio between liver and adipose tissue levels (Thoma et al. 1990). However, the important 

TEQ variable was close to 1:1 ratio. 

In the general population, adipose tissue levels of 2,3,7,8-TCDD ranged from non-detectable to 20.2 ppt in 

128 Kansas City, St. Louis, and Springfield, Missouri, residents with no known special exposure to CDDs 

(Andrews et al. 1989). Similarly, 2,3,7,8-TCDD levels in adipose tissues were between 5 and 10 ppt in a 

sample of the general population in Canada, while OCDD levels ranged from 600 to 800 ppt in the cohort 

(Ryan et al. 1985a). Increased environmental exposure to CDDs was reflected by increased levels in 

adipose tissues. Residents of two California households who had eaten dioxin-contaminated beef and eggs 

had significantly elevated serum levels of 2,3,7,8-TCDD, PCDD, and HxCDD, compared with rural 

Californians who did not eat contaminated beef and eggs (Goldman et al. 1989). 2,3,7,8-TCDD serum lipid 

levels ranged from 2.8 to 750 ppt in individuals with possible recreational, residential, or occupational 

exposure in Missouri (Patterson et al. 1986a). Several other studies reported increased concentrations of 

2,3,7,8-TCDD in adipose tissues (Beck et al. 1989c; Fingerhut et al. 1989; Patterson et al. 1989b) or serum 

lipid (Fingerhut et al. 1989; Patterson et al. 1989b) of occupationally exposed workers. The highest 

2,3,7,8-TCDD levels reported ranged between 42 and 750 ppt in adipose tissue lipid and between 61 and 

1,090 ppt in the serum lipid of Missouri chemical workers (Patterson et al. 1989b). Based on a mean of 

208 ppt 2,3,7,8-TCDD in the serum lipid of chemical workers measured at least 17 years postexposure, and 

a mean biological half-life level of 7 years, it was estimated that the serum lipid level shortly after exposure 

would have been approximately 2,313 ppt (range, 6.4–14,673) (Fingerhut et al. 1989). The major 

disadvantage of these studies is a lack of information regarding actual 2,3,7,8-TCDD exposure. Studies in a 

population exposed to 2,3,7,8-TCDD in the Seveso industrial accident found serum lipid 
CDDs 320 

2. HEALTH EFFECTS 

2,3,7,8-TCDD levels in exposed individuals as high as 56,000 ppt in the highly contaminated zone where 

soil samples ranged from 956 to 1,185 μg 2,3,7,8-TCDD/m2

 (Mocarelli et al. 1991); these serum samples 

were collected about a month after exposure and were recently analyzed. Chloracne did not always develop 

in the individuals with highly elevated serum 2,3,7,8-TCDD levels, although it did occur in all who had 

levels above 12,000 ppt. In the most recent National Human Adipose Tissue Survey (NHATS), conducted 

in fiscal year 1987, it was found that the average concentration of 2,3,7,8-TCDD in the adipose tissue of the 

U.S. population was 5.38 ppt, increasing from 1.98 ppt in children under 14 years of age to 9.4 ppt in adults 

over 45 (Orban et al. 1994). The average concentration of 2,3,7,8-TCDD was found to increase at an 

average rate of 0.83 ppt per decade for individuals under age 31, and at an average rate of 1.52 ppt per 

decade for the older population. The study also found no statistical evidence of differences in the average 

levels for populations representing different sexes and racial groups nationwide. Furthermore, a 

comparison of mean concentrations of 2,3,7,8-TCDD and OCDD between the 1982 and 1987 NHATS 

revealed no statistically significant differences between the two surveys. For OCDD the values were 768 

and 724 ppt in the 1982 and 1987 surveys, respectively (Orban et al. 1994). 

Similarly, no exact exposure data were available in Vietnam veteran studies. In general, tissue samples for 

analyses were taken several (approximately 10–20) years after exposure, which represents another 

limitation of these studies. Increased adipose 2,3,7,8-TCDD levels (up to 99 ppt) were recorded in Vietnam 

War veterans involved in Operation Ranch Hand (Gross et al. 1984; Schecter et al. 1989a). In a small 

group of potentially exposed Vietnam veterans, adipose tissue 2,3,7,8-TCDD levels ranged from nondetectable to 11 ppt with a mean of 5.8 ppt (Schecter et al. 1989a). Schecter et al. (1989a) noted that the 

veterans with adipose tissue levels of $8 ppt were considered to have slightly elevated values or values 

within the normal range. In another study of 646 ground troop veterans, only two individuals had serum 

2,3,7,8-TCDD levels above 20 ppt in the lipid fraction (CDC 1988). In the rest of the cohort, the median 

2,3,7,8-TCDD levels ranged from 3.2 to 4.3 ppt and did not differ significantly from the levels found in the 

control group of non-Vietnam veterans (CDC 1988; MMWR 1987). It was concluded that those who did 

not handle or spray herbicides were not highly exposed to 2,3,7,8-TCDD (CDC 1988). With regard to the 

long time period between exposure and serum analysis, the authors argued that, assuming first-order 

kinetics and 2,3,7,8-TCDD's half-life of 7 years in humans, the study had enough statistical power to detect 

differences between the exposed and control groups (CDC 1988). Elevated CDD levels were also measured 

in some patients treated in a hospital in South Vietnam (Phiet 1989; Phuong et al. 1989b). However, these 

reports involved too few patients to give any conclusive results. In a more recent expanded half-life study 

of 337 Vietnam veterans, a median observed half-life of 11.5 years was calculated for 
CDDs 321 

2. HEALTH EFFECTS 

2,3,7,8-TCDD (Wolfe et al. 1994). The nonparametric 95% CI was 10–14.1 years. A review of CDD 

levels in human tissues from various populations can be found in Schecter et al. (1994c). 

CDDs have also been detected in breast milk of women exposed to high levels of CDDs and in women 

presumably exposed to background levels. High 2,3,7,8-TCDD levels (mean of 484 pg/g milk fat; 18 pg/g 

whole milk) were found in the milk of mothers from South Vietnam in 1970; the levels dropped to a mean 

of 12 pg/g milk fat (0.47 pg/g milk) by 1985 (Schecter et al. 1987a) and 7.5 pg/g milk fat in samples 

collected between 1984 and 1992 (Schecter et al. 1995). Mean 2,3,7,8-TCDD levels in breast milk samples 

(collected in 1984) in mothers from South Vietnam, North Vietnam, and the United States were 0.68, not 

detectable, and 0.19 pg/g whole milk (Schecter and Gasiewicz 1987b). The total CDD and CDF levels 

(expressed as TEQs) in these samples were 1.11, 0.065, and 1.04 pg/g milk. Results from the analysis of 

526 individual milk samples from the German general population revealed a mean 2,3,7,8-TCDD 

concentration of 3.2 pg/g milk fat (Fürst et al. 1994). The analysis also showed the presence of only 

2,3,7,8-chlorine-substituted congeners. OCDD was the most concentrated congener with a mean level of 

208 pg/g milk fat. In general, the levels in milk decreased with decreasing degree of chlorination from 

octa- to tetra-CDD. The total TEQs, including CDFs, was 29.3 pg/g milk fat. Fürst et al. (1994) estimated 

that the average daily intake of 2,3,7,8-TCDD via human milk for an infant weighing 5 kg is 15.4 

pg/kg/day, and the mean total dioxin equivalents amounted to 140.6 pg/kg/day. Both parity and the length 

of time the woman has been lactating influence the CDD concentration in breast milk. 

A reverse transcriptase polymerase chain reaction (RT-PCR) method was developed to quantitate CYP1A1 

mRNA levels on total RNA extracts from mitogen-stimulated human blood lymphocytes cultured in the 

presence or absence of 10 nM 2,3,7,8-TCDD (Van den Heuvel et al. 1993). Although CYP1A1 gene 

expression can be monitored by measuring EROD activity (CYP1A1) or mRNA expression (using 

conventional RNA hybridization), RT-PCR is a much more sensitive approach. The average CYP1A1 

mRNA levels in the cultured, 2,3,7,8-TCDD-treated cells was approximately 21 times higher than that in 

the non-2,3,7,8-TCDD-treated cells. In uncultured, nonstimulated lymphocytes from volunteers, CYP1A1 

mRNA could be reproducibly measured at levels that were 10–40-fold lower than in mitogen-stimulated 

lymphocytes. In comparison, EROD activity measured in uncultured, nonstimulated lymphocytes was 

indistinguishable from measurements on reagents controls, which proved the high sensitivity of the RTPCR approach. In a group of 6 smokers, the average level of CYP1A1 message was approximately 2 times 

higher than in a group of 6 nonsmokers, although there was great variability in the group of smokers. 

Based on these preliminary results, Van den Heuvel et al. (1993) suggested that CYP1A1 gene expression 
CDDs 322 

2. HEALTH EFFECTS 

in peripheral blood lymphocytes may be used as a human exposure marker for 2,3,7,8-TCDD and related 

compounds. 

For more information on biomarkers for renal and hepatic effects of chemicals, see the ATSDR/CDC 

Subcommittee Report on Biological Indicators of Organ Damage (1990), and for information on biomarkers 

for neurological effects, see OTA (1990). 

2.7.2 Biomarkers Used to Characterize Effects Caused by CDDs 

Chloracne is one effect that is clearly associated with exposure to high levels of CDDs and other 

halogenated organic chemicals, and has been observed in some individuals who were exposed 

occupationally or in the environment to increased levels of 2,3,7,8-TCDD or chemicals contaminated with 

2,3,7,8-TCDD. However, while the presence of chloracne indicates exposure to CDDs or other halogenated 

organic compounds, its absence does not preclude such exposure. For example, in a cohort from the Seveso 

incident, no chloracne was observed below an initial serum lipid 2,3,7,8-TCDD level of 800 ppt (body 

burden of 2.5 μg/kg, assuming 22% body fat and 70 kg body weight); above 12,000 ppt (body burden of 

38 μg/kg) chloracne was always observed; and between 800 and 12,000 ppt the occurrence of chloracne 

was sporadic (Mocarelli et al. 1991). In the Yu-Cheng population, chloracne was associated with a body 

burden in 2,3,7,8-TCDD equivalents of 2–3 μg/kg body weight, or about 140–210 μg for a 70-kg adult 

(Ryan et al. 1990). 

Biochemical changes (raised serum hepatic enzyme levels, disorders of lipid and carbohydrate metabolism, 

unbalanced porphyrin metabolism) and/or an enlarged liver can indicate effects induced by 2,3,7,8-TCDD 

exposure, but these effects are not specific for this or other compounds. Light and electron microscope 

changes in the liver (e.g., lipid droplets in parenchymal cells, increased endoplasmic reticulum, enlarged 

and pleomorphic mitochondria) are also sensitive but nonspecific biomarkers for exposure to CDDs 

(Schecter et al. 1985b). When biochemical changes in the placenta of women exposed in the Yu-Cheng 

incident were evaluated for use as possible biomarkers, the EGF receptor autophosphorylation effect was 

found to be associated with decreased birth weight in the neonates (Lucier et al. 1986). The authors 

suggested using this response as a biomarker of effect for all toxic chlorinated aromatic compounds. 
CDDs 323 

2. HEALTH EFFECTS 

2.8 INTERACTIONS WITH OTHER CHEMICALS 

Several studies were located regarding interactions that affect the toxicity of CDDs. Probably the most 

important interactions that have an impact on human health are those involving CDDs, CDFs, and PCBs. It 

has been recognized that chloroaromatics cause a complex of similar effects that vary in severity depending 

on the number of chlorine atoms, positional substitution, and species susceptibility. Sufficient information 

is available for assessment of risk associated with exposure to 2,3,7,8-TCDD. However, exposure to a 

mixture of chloroaromatics is common in the general environment. The assessment of health risk resulting 

from exposure to chemical mixtures of chloroaromatics was enabled by the development of TEFs 

(2,3,7,8-TCDD equivalence factors) that relate the relative toxic potency for CDDs and CDFs to that of 

2,3,7,8-TCDD (EPA 1989). It was assumed based on previous literature data (Eadon et al. 1986) and in 

animal dosing studies (Van den Berg et al. 1989), that CDDs and CDFs have an additive effect in the 

organism when weighted for relative toxicity compared to 2,3,7,8-TCDD (for further information see 

Sections 2.4 and 2.5). The assumption of additivity was later supported by experimental data. The concept 

of TEFs was used, for example, to assess the potential toxicity of background levels of CDFs and CDDs in 

general populations based on body burdens of indicator CDDs that were associated with chloracne and 

other effects in the Yusho and Yu-Cheng rice oil poisoning incidents (Ryan et al. 1990). 

However, some recent studies further investigated the interactions of various chloroaromatics and indicated 

that the interactions may be more complicated. In vitro studies compared relative toxicity of various 

chloroaromatics in human cell lines monitoring enzyme induction and binding to the Ah receptor that 

mediates the induced responses (Nagayama et al. 1985; Safe 1987). In vivo studies concentrated on 

monitoring of enzyme induction, inhibition of body weight gain and immunotoxic and teratogenic effects. 

Coexposure of Long Evans rats to 6-methyl-1,3,8-trichlorodibenzofuran (MCDF) and 2,3,7,8-TCDD 

induced a partial inhibition of the monooxygenase enzyme-induction response caused by 2,3,7,8-TCDD 

treatment alone (Harris et al. 1989b). Although MCDF did not decrease the levels of occupied nuclear 

2,3,7,8-TCDD Ah receptors, it inhibited the effects of 2,3,7,8-TCDD on the cytosolic Ah receptor (Harris et 

al. 1989b). 

Other studies further indicated that PCBs may antagonize Ah receptor-mediated responses to 

2,3,7,8-TCDD. In a recent review, Van den Berg et al. (1994) suggested that toxicokinetic factors 

contribute to the observed nonadditive toxicological and biological effects. Co-treatment of C57BL/6 mice 

with various commercial Aroclors (PCB mixtures) and 2,3,7,8-TCDD resulted in antagonizing the 
CDDs 324 

2. HEALTH EFFECTS 

2,3,7,8-TCDD-mediated inhibition of the splenic plaque-forming cell response (Bannister et al. 1987; Davis 

and Safe 1989). Similarly, significant antagonism of 2,3,7,8-TCDD and Aroclor 1254 was observed in the 

induction of cytochrome P-450-dependent monooxygenases in C57BL/6J mice (Bannister et al. 1987). The 

effects were dependent on the dose of both 2,3,7,8-TCDD and Aroclor 1254 and on their respective ratios. 

The ratios of Aroclor 1254/2,3,7,8-TCDD that induced antagonist reactions were comparable to the ratios of 

PCBs/CDDs found in human tissues and environmental samples. The authors speculated that less-toxic 

chlorinated compounds may have a protective effect against the more-toxic compounds in the environment. 

However, by comparing the immune sensitivities of both Ah responsive and Ah less-responsive mouse 

strains, it was demonstrated that a complex mixture of contaminants taken from the Love Canal site was 

immunosuppressive and that this effect was primarily due to the 2,3,7,8-TCDD component of the mixture, 

although 2,3,7,8-TCDD was a very minor component, and there was little interaction with the other 

hydrocarbon components of the mixture (Silkworth et al. 1989a). 

Experimental studies have shown that interactions of 2,3,7,8-TCDD and CDFs or PCBs resulted in fetotoxic 

and teratogenic effects in the offspring of exposed animals. Exposure of pregnant mice to 2,3,7,8-TCDF 

resulted in cleft palates and hydronephrosis in the offspring (Hassoun et al. 1984). The results obtained in 

different strains of mice indicated an association with the Ah locus. Comparable results were obtained 

previously in mice exposed to 2,3,7,8-TCDD (Abbott and Birnbaum 1989a; Abbott et al. 1987a, 1987b; 

Courtney 1976). When C57BL/6N mice were treated orally with 2,3,7,8-TCDD and 2,3,7,8-TCDF on 

gestational day (Gd) 10, hydronephrosis and cleft palates were observed in the offspring (Weber et al. 

1985). The effects of both chemicals were additive. Similarly, an increased incidence (10-fold) of cleft 

palates was observed in offspring of C57BL/6N mice after a combined treatment with 2,3,7,8-TCDD and 

2,3,4,5,3',4'-hexachlorobiphenyl during gestation, as compared with those treated with 2,3,7,8-TCDD alone 

(cleft palate was not observed when 2,3,4,5,3',4'-hexachlorobiphenyl was administered alone) (Birnbaum et 

al. 1985). In contrast, no potentiation of CDD-mediated effect was found with 2,4,5,2',4',5'-hexachlorobiphenyl. Furthermore, co-treatment of pregnant C57BL/6J mice with Aroclor 1254 and 2,3,7,8-TCDD 

resulted in a sharp decrease in the incidence of cleft palate per litter (8.2%) compared with those treated 

with 2,3,7,8-TCDD alone (62%) (Haake et al. 1987). 

Similarly, 2,3,7,8-TCDD-induced fetotoxicity and teratogenicity were altered by co-exposure to other 

chemicals. A synergistic effect on the induction of cleft palates was observed in offspring of C57BL/6N 

mice treated orally with 2,3,7,8-TCDD and retinoic acid on Gd 10 or 12 (Abbott and Birnbaum 1989b; 

Birnbaum et al. 1989b). However, the co-administration of retinoic acid did not influence the incidence of 
CDDs 325 

2. HEALTH EFFECTS 

2,3,7,8-TCDD-induced hydronephrosis, nor did 2,3,7,8-TCDD affect the incidence or severity of limb-bud 

defects induced by retinoic acid (Birnbaum et al. 1989b). A synergistic effect was observed when 

2,3,7,8-TCDD (orally) and hydrocortisone (subcutaneously) were administered to C57BL/6N mice on 

Gd 10–13 (Birnbaum et al. 1986). The incidence of cleft palate in the offspring increased to 100% 

following the combined treatment. Pretreatment of pregnant NMRI mice with benzo(a)pyrene 

subcutaneously 5 hours prior to an intraperitoneal injection of 2,3,7,8-TCDD caused an increase in CDDinduced lethality but did not alter the rate of cleft palate formation (Hassoun 1987). Offspring of male 

mice, treated with chlorinated phenoxy acids and 2,3,7,8-TCDD in their feed for 8 weeks before the mating, 

did not differ in their development or survival from offspring in the control group (Lamb and Moore 1981). 

Results in B6C3F1 mice indicated that α-naphthoflavone antagonizes 2,3,7,8-TCDD in induction of 

splenocyte EROD activity (Blank et al. 1987). It was further suggested that α-naphthoflavone impedes 

2,3,7,8-TCDD suppression of B lymphocyte differentiation by competing for binding to the Ah receptor. 

The mechanism of interaction of these chemicals was studied in vitro using rat hepatic cytosol or mouse 

hepatoma cells (Gasiewicz and Rucci 1991). The results indicated that α-naphthoflavone acts as a 

2,3,7,8-TCDD antagonist by binding to the Ah receptor and forcing on it a conformation that cannot 

identify the DNA recognition sequence contained in the dioxin-responsive enhancer element of the 

CYP1A1 gene. In contrast, in vitro experiments showed that co-exposure of a thymus organ culture with 

the weakly toxic β-naphthoflavone and 2,3,7,8-TCDD results in a significant increase in the lymphoid 

inhibitory effect mediated by 2,3,7,8-TCDD (Hassoun 1987). 

Hexachlorobenzene acted like a weak Ah receptor agonist and caused an up to 40% decrease in specific 

hepatic cytosol binding of 2,3,7,8-TCDD in rat cells (Hahn et al. 1989b). Similarly, 2,3,7,8-TCDD-induced 

myelotoxicity and enzyme induction was antagonized by 1-amino-3,7,8-trichlorodibenzo-p-dioxin in 

B6C3F1 mice presumably by competitive binding to the cytosolic Ah receptor (Luster et al. 1986). 

Comparable effects were observed in vitro in murine bone-marrow-cells cultures. Treatment of Fischer 344 

rats orally with di(2-ethylhexyl)phthalate (DEHP) before or after oral administration of 2,3,7,8-TCDD 

reduced the hyperlipidemia induced by the latter compound (Tomaszewski et al. 1988). Furthermore, 

DEHP pretreatment followed by daily doses of this hypolipidemic substance was partially protective against 

2,3,7,8-TCDD-induced mortality, wasting, and liver fatty changes. 
CDDs 326 

2. HEALTH EFFECTS 

The addition of activated charcoal or dehydrocholic acid to the feed, protected animals (C57BL/6J mice, 

CD-COBS rats, and guinea pigs) from increased mortality caused by a single lethal dose of 2,3,7,8-TCDD 

(Manara et al. 1984). In the case of the former agent, the effect was probably due to the general high 

binding ability of superactivated charcoal; since no other antidote is known, its use for therapeutic purposes 

was recommended. Protective effects of ascorbic acid (administered orally) and butylated hydroxyanisole 

(BHA) (administered orally) against 2,3,7,8-TCDD given by gavage were investigated in Sprague-Dawley 

rats (Hassan et al. 1987). BHA administration partially protected rats from losses in organ weights and 

2,3,7,8-TCDD-induced lipid peroxidation and inhibition of glutathione peroxidase activity. In contrast, 

ascorbic acid had no protective effects. 

Data regarding interactions affecting the toxicity or toxicokinetics of other chemicals by 2,3,7,8-TCDD 

were limited. Dermal pretreatment with 2,3,7,8-TCDD inhibited the induction of skin tumors by 

subsequently applied benzo(a)pyrene or dimethylbenz(a)anthracene in Sencar mice (Cohen et al. 1979). It 

was proposed that 2,3,7,8-TCDD caused qualitative alteration of hydrogen binding to DNA. In addition, 

2,3,7,8-TCDD may also promote the metabolism of procarcinogens (e.g., 3-methylcholanthrene) to active 

metabolites by the induction of metabolizing enzymes (Kouri et al. 1974, 1978). 

2.9 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE 

A susceptible population will exhibit a different or enhanced response to CDDs than will most persons 

exposed to the same level of CDDs 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 may 

result in reduced detoxification or excretion of CDDs, or compromised function of target organs affected by 

CDDs. Populations who are at greater risk due to their unusually high exposure to CDDs are discussed in 

Section 5.6, Populations With Potentially High Exposure. 

No data were located regarding unusually susceptible subpopulation in humans. Animal data showed 

developmental effects of 2,3,7,8-TCDD in fetuses and newborns exposed in utero and via breast-feeding, 

respectively (Abbott and Birnbaum 1989b; Giavini et al. 1982, 1983; Håkansson et al. 1987; Weber et al. 

1985) (see Section 2.2.2.6). The experimental data suggest that the prenatal and postnatal population may 

be more sensitive to 2,3,7,8-TCDD-induced effects; however, the levels of exposure necessary to induce 

such effects are not known. Data in mice indicated that strain differences in sensitivity to 2,3,7,8-TCDD 

toxicity exist and are associated with the Ah receptor (Poland and Glover 1980). It has been shown that 
CDDs 327 

2. HEALTH EFFECTS 

the Ah receptor exists in human lymphoid tissue and that its concentration is variable between persons 

(Hayashi et al. 1994; Lorenzen and Okey 1991). As noted in Section 2.4, 2,3,7,8-TCDD may promote the 

metabolism of procarcinogens (e.g., contained in cigarette smoke) to active intermediates by the induction 

of metabolizing enzymes. The induction of these enzymes in humans appears to be subject to genetic 

polymorphism so that individuals who have an Ah receptor with high affinity for 2,3,7,8-TCDD and related 

chemicals may be at the highest risk for the development of lung tumors (Antilla et al. 1991; Bartsch et al. 

1990; Kawajiri et al. 1990; McLemore et al. 1990; Uematsu et al. 1991). 

2.10 METHODS FOR REDUCING TOXIC EFFECTS 

This section will describe clinical practice and research concerning methods for reducing toxic effects of 

exposure to CDDs. 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 CDDs. When specific 

exposures have occurred, poison control centers and medical toxicologists should be consulted for medical 

advice. 

No texts were found that provided specific information about treatment following exposures to CDDs. 

2.10.1 Reducing Peak Absorption Following Exposure 

No specific information was located regarding the reduction of peak absorption of CDDs by the oral and 

inhalation routes of exposure in humans. Weber et al. (1992d) examined the effect of four decontamination 

protocols on the residency time of 2,3,7,8-TCDD in intact or damaged human post-mortem skin in vitro. 

Damage was simulated by stripping of the stratum corneum. 2,3,7,8-TCDD was applied to the skin for 

100 minutes and one of the following protocols was performed: the sample was wiped with dry, adsorbent 

material (cotton balls); a 10-minute topical treatment with mineral oil was followed by dry wiping with 

cotton balls; a 10-minute topical treatment with mineral oil was followed by wiping with acetone-soaked 

cotton balls; and the sample was washed with water and soap. In intact skin, mineral oil treatment and 

acetone wipes reduced by about two-fold the amount of 2,3,7,8-TCDD in the stratum corneum. Mineral oil 

plus dry wipes reduced the amount of 2,3,7,8-TCDD in the stratum corneum by about 33%, whereas dry 

wiping alone was ineffective. However, all protocols were equally effective in reducing the amount of 

2,3,7,8-TCDD in the epidermis and upper dermis by factors of up to 10. In damaged skin, by dry wiping 

with adsorbent material 2,3,7,8-TCDD was rubbed into the skin, leading to increased concentrations in the 
CDDs 328 

2. HEALTH EFFECTS 

various layers of the skin. In contrast, mineral oil treatment followed either by dry wipes or by acetone 

wipes, and washing with water and soap decontaminated the damaged skin quite effectively. The authors 

(Weber et al. 1992d) noted that the effect of decontamination was most pronounced at a skin depth between 

160 and 500 μm, which is the depth at which vascularization begins and, therefore, the protocols discussed 

should be particularly effective in reducing systemic absorption of 2,3,7,8-TCDD from the skin. 

2.10.2 Reducing Body Burden 

Limited information was located regarding reducing body burden following exposure to CDDs in humans. 

A recent study examined the influence of short-term dietary measures on CDD and CDF concentrations in 

human milk (Pluim et al. 1994c). The authors hypothesized that mobilization of fatty acids from adipose 

tissue cause the concomitant release of CDDs and CDFs, which will then be eliminated in the breast milk. 

Two diets were tested for their ability to reduce the concentration of CDDs and CDFs in human milk: a 

low-fat/high carbohydrate/low CDD and CDF diet (16 women), and a high-fat/low carbohydrate/low CDD 

and CDF diet (18 women). The authors also analyzed the fatty acid pattern of the milk to determine 

whether the dietary changes were sufficient to change the milk-fat composition. The test diets were 

followed for 5 consecutive days in the fourth week after delivery. Body weights were not affected by the 

experimental diets. The results showed that the fat content of the milk did not decrease in either group 

during the test diet. Furthermore, there was no significant change in CDD and CDF concentration in milk 

fat after treatment with the experimental diets. However, the percentage of medium-chain fatty acids 

(MCFA) changed significantly. In the low-fat/high carbohydrate diet group, the percentage of MCFA 

increased while the percentage of C18:1ω9 fatty acids (fatty acid with 18 straight-chain carbon atoms, 

1-methylene-interrupted double bond and 9 carbon atoms from the terminal methyl group to the nearest 

double bond) decreased. In the high-fat/low carbohydrate diet group the changes were in the opposite 

direction. According to the authors, the results would indicate that the concentration of CDDs and CDFs in 

milk fat may be independent of the source of the fatty acids. Alternatively, they indicate that the dieting 

period may have been too short or the dietary changes in fat and carbohydrate intake may have not been 

large enough. 

Using a fugacity-based PBPK model to evaluate elimination of 2,3,7,8-TCDD from humans, assuming a 

background 2,3,7,8-TCDD intake of 50 pg/day, Kissel and Robarge (1988) estimated that daily 

consumption of 10 g of a nonabsorbable oil would reduce the steady-state adipose tissue concentration of 

2,3,7,8-TCDD from 7.7 ppt to 3 ppt. At an adipose tissue level of 50 ppt, the apparent half-life of 
CDDs 329 

2. HEALTH EFFECTS 

2,3,7,8-TCDD would be decreased by consumption of the nonabsorbable oil from 5.2. to 2.1 years. 

However, this is a theoretical assumption based on the PBPK model. 

Zober and Päpke (1993) reported that serum lipid 2,3,7,8-TCDD levels increased from 17 ppt to 32 ppt in a 

patient who lost >7 kg of weight in a 5-month period. The serum lipid concentrations of 

1,2,3,6,7,8-HxCDD, 1,2,3,4,5,7,8-HpCDD, and OCDD also increased during this period. Fasting appeared 

to relieve signs and symptoms of intoxication in a group of patients who ingested rice oil contaminated with 

the structurally related PCBs and CDFs (Yu-Cheng incident) (Imamura and Tung 1984). The authors 

suggested that fasting may stimulate mobilization of the chemicals from adipose tissue to the liver where 

they are then metabolized, which would facilitate excretion and reduce body burden. However, the findings 

of that study should be interpreted with caution because a control group was not used, small number of 

subjects were evaluated, the patients volunteered for the study, and some of the end points that were 

evaluated were subjective. Promotion of fecal excretion of CDFs and PCBs by cholestyramine, a 

hypercholesterolemia therapeutic agent used in the treatment of poisoning by chlorinated organic 

agricultural chemicals, was inconclusive in a clinical trial with six Yusho patients (Iida et al. 1991; Murai et 

al. 1991). 

In experimental animals, administration of a diet containing 2.5 or 5% activated charcoal substantially 

reduced mortality due to a single lethal oral dose of 2,3,7,8-TCDD in rats, mice, and guinea pigs (Manara et 

al. 1984). Also, feed with 0.25 or 0.5% cholic acid had a similar protective action in mice (Manara et al. 

1984). The effect of activated charcoal was attributed to increased clearance of unabsorbed 2,3,7,8-TCDD 

from the body; the mechanism of protection by cholic acids was unclear. 

2.10.3 Interfering with the Mechanism of Action for Toxic Effects 

There are no established methods for interfering with the mechanism of action of CDDs. Many of the toxic 

effects of 2,3,7,8-TCDD and structurally related halogenated aromatic hydrocarbons have been shown to be 

mediated through the Ah receptor (see Section 2.4 for details). The sequence of events associated with the 

receptor-mediated mechanism involve entry of 2,3,7,8-TCDD into the cell, binding to the cytosolic Ah 

receptor, binding of the receptor-ligand complex to DNA recognition sites, and expression of specific genes 

and the translation of their protein products. Although speculative, it is possible that interference with this 

mechanism may lead to a more specific treatment for reducing some of the toxic effects of 2,3,7,8-TCDD 

and structurally related chemicals. Future research on Ah receptor antagonists may provide new insights 
CDDs 330 

2. HEALTH EFFECTS 

for clinical treatment of the Ah receptor-mediated toxicity of 2,3,7,8-TCDD and other Ah receptor agonists. 

2.11 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 CDDs 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 CDDs. 

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. 

2.11.1 Existing Information on Health Effects of CDDs 

The existing data on health effects of inhalation, oral, and dermal exposure of humans and animals to CDDs 

are summarized in Figures 2-5 and 2-6. The purpose of this figure is to illustrate the existing information 

concerning the health effects of CDDs. Each dot in the figure indicates that one or more 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 (ATSDR 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. 

As seen in Figures 2-5 and 2-6, information is available regarding death, systemic, immunological, 

neurological, developmental, reproductive, and genotoxic effects and cancer in humans. Most of the 

available information is for 2,3,7,8-TCDD. Most of this information is negative or inconclusive except for 


 

CDDs 333 

2. HEALTH EFFECTS 

dermal effects and cancer. As previously mentioned, exposure to humans probably occurs by a 

combination of the inhalation, oral, and dermal routes. No information is available regarding effects of a 

single route of exposure in humans. However, food is the major source of human exposure in the general 

population; therefore, the oral route is the most significant exposure route. 

Oral and dermal studies in animals provide data on death, systemic effects after acute-, intermediate-, and 

chronic-duration exposure, and immunological, reproductive, and cancer effects. Furthermore, data exist 

regarding neurological, developmental, and genotoxic effects after oral exposure. No data were located 

regarding effects in animals after inhalation exposure to CDDs. 

2.11.2 Identification of Data Needs 

As discussed in Section 2.5, the EPA and other agencies and scientists are using the TEF scheme as an 

alternative interim approach for hazard evaluation of CDDs and CDFs. Since toxicological data for other 

CDD congeners is more limited, additional congener-specific studies would provide valuable data for 

validating the TEF approach. In vitro and short-term parenteral injection studies using sensitive end points 

(i.e., enzyme induction, immune alterations) have been used for this purpose, but studies using other end 

points, the oral route, and/or longer durations of exposure would be more informative. Since the database 

for CDD effects not mediated through the Ah receptor is limited, additional studies may be relevant to 

understanding whether acute versus chronic responses to 2,3,7,8-TCDD occur by different mechanisms. 

CDDs and the structurally related CDFs and dioxin-like PCBs are of concern to ATSDR because of the 

potential of these chemicals to harm health at relatively low doses. As discussed in Section 2.5, many of 

the toxic effects of these compounds appear to be mediated by a common mechanism, and CDDs frequently 

occur with CDFs in the environment. Therefore, due to the common mechanism of toxicity, total toxicity 

of a CDD/CDF mixture probably results from the added contribution (not necessarily linear) of both classes 

of chemicals. Because of this, the complex issue of appropriate methodology for quantitatively assessing 

health risks of CDDs and CDFs is currently being evaluated by ATSDR. Additional information on toxic 

interactions between CDDs and CDFs, as well as PCBs, would facilitate health risk assessment of this class 

of chemicals. 

Acute-Duration Exposure Acute exposure of humans to 2,3,7,8-TCDD can cause chloracne and 

hepatic effects (Goldman 1973; Reggiani 1980). Specifying the route of exposure in these human cases is 
CDDs 334 

2. HEALTH EFFECTS 

difficult because the individuals were probably exposed by a combination of routes. Furthermore, human 

data did not provide any information regarding exposure levels and co-exposure to other chemicals 

confounds the results. Also, in most cases, the exposed subjects were examined long after exposure 

occurred. Acute oral exposure to 2,3,7,8-TCDD caused delayed type of death in all animal species tested, 

and LD50 values have been determined for rats (NTP 1982b; Schwetz et al. 1973; Walden and Schiller 

1985), minks (Hochstein et al. 1988), rabbits (Schwetz et al. 1973), guinea pigs (McConnell et al. 1984; 

Schwetz et al. 1973), and hamsters (Henck et al. 1981). Furthermore, an acute LD50 was calculated for rats 

and mice exposed to a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980). No deaths 

were observed with other congeners (2,7-DCDD, 1,2,3,4,6,7,8-HpCDD, 1,2,3,4,6,7,8,9-OCDD) (NTP 

1982b) tested, and 2,3,7,8-TCDD proved to be the most toxic CDD. However, interspecies and interstrain 

differences were found in the susceptibility to CDDs. Systemic effects observed in animals after acute oral 

exposure to 2,3,7,8-TCDD included cardiovascular (Hochstein et al. 1988; McConnell et al. 1978b), 

gastrointestinal (Theobald et al. 1991), hematological (Christian et al. 1986), hepatic (Christian et al. 1986; 

Kelling et al. 1985; Walden and Schiller 1985), renal (Christian et al. 1986; McConnell et al. 1978b), 

endocrine (Bastomsky 1977; Bestervelt et al. 1993; Fan and Rozman 1995; Potter et al. 1986; Weber et al. 

1995), dermal effects (Greig 1984; McConnell et al. 1978b), and body weight loss (Kelling et al. 1985; 

Moore et al. 1985; Seefeld and Peterson 1984; Weber et al. 1994, 1995). Hepatic and body weight effects 

were the main signs of 2,3,7,8-TCDD toxicity and occurred also after exposure to a mixture of 1,2,3,6,7,8­

HxCDD and 1,2,3,7,8,9-HxCDD (NCI/NTP 1980). Furthermore, immunological effects were observed 

following low oral doses of 2,3,7,8-TCDD (Burleson et al. 1996; White et al. 1986), and an acute oral MRL 

was based on a NOAEL for immunological effects (Burleson et al. 1996). In addition, the dermal LD50 for 

2,3,7,8-TCDD has been determined in rabbits (Schwetz et al. 1973). Since only dermal changes were 

investigated following acute dermal exposure (Puhvel et al. 1982), further studies could provide useful 

information regarding additional endpoints; dermal contact is a relevant route of exposure at waste sites 

where CDDs may be stored. Limited data were located regarding effects in animals after acute inhalation 

exposure to CDDs (Diliberto et al. 1996; Nessel et al. 1992). Further studies by the inhalation route of 

exposure would be useful since toxicokinetic data in rats suggest that this could be an important route for 

systemic absorption of CDDs (Diliberto et al. 1996). 

No information was located regarding health effects of other congeners in humans, and limited data exist 

about effects caused by an acute exposure to these congeners in animals. The information would be useful 

for populations living near hazardous waste sites who may be exposed to CDDs for acute durations. 

Should a case of high acute exposure to 2,3,7,8-TCDD occur in humans, prompt comprehensive 
 

 

CDDs 335 

2. HEALTH EFFECTS 

examination of those exposed would provide greatly needed information. Furthermore, follow-up medical 

surveillance of such a population should be conducted for as long as possible. 

Intermediate-Duration Exposure. Intermediate-duration exposure of humans to CDDs has occurred 

after industrial accidents or in population groups (e.g., Vietnam veterans, Vietnamese, and pesticide 

production workers and applicators) exposed to CDD-contaminated herbicides. As stated above, the route 

of exposure and exposure levels cannot be exactly determined. Hepatic and dermal changes were the main 

effects noted, and an association between incidence of diabetes and exposure to 2,3,7,8-TCDD has been 

reported (Jirasek et al. 1976; USAF 1991). More toxicokinetic data for various routes of exposure with 

relevant congeners would be useful. These data would help in extrapolation from one route of exposure to 

another, since no information is available in humans on exposure via the oral route, which is the major 

exposure route to CDDs. The main adverse effects in animals following intermediate-duration oral and 

dermal exposure to 2,3,7,8-TCDD included chloracne (Allen et al. 1977; Berry et al. 1978; McNulty 

1984), wasting syndrome (DeCaprio 1986; NTP 1982b; Vos et al. 1973), and liver effects (Hebert et al. 

1990; NTP 1982a). Similar effects were observed with a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9­

HxCDD (NCI/NTP 1980), and 1,2,3,7,8-PeCDD and 1,2,3,4,7,8-HxCDD (Viluksela et al. 1998a, 1998b). 

As with acute-duration exposure, the immune system proved to be a very sensitive end point for 

intermediate-duration exposure to 2,3,7,8-TCDD, and an intermediate-duration oral MRL was derived 

from a NOAEL value for immunological effects (DeCaprio et al. 1986). No data were located regarding 

toxicity or toxicokinetics in animals after intermediate-duration inhalation exposure to CDDs. Information 

obtained from a 90-day inhalation exposure study would be relevant to people living near hazardous waste 

sites who may be exposed to CDDs for similar durations or much longer time periods. 

Chronic-Duration Exposure and Cancer. A number of epidemiology studies have examined the 

toxicity of CDDs following chronic exposure to phenoxy herbicides and chlorophenols contaminated with 

2,3,7,8-TCDD (Calvert et al. 1991, 1992, 1996, 1998; Cook et al. 1987b; Egeland et al. 1994; Henriksen et 

al. 1997; Pesatori et al. 1998; Sweeney et al. 1993). Although a number of effects have been observed, 

interpretation of the results is confounded by a number of factors including lack of adequate exposure 

information, long postexposure periods, concomitant exposure to other chemicals, and small cohorts. 

Follow-up medical surveillance of subjects with known past high exposure to 2,3,7,8-TCDD would provide 

information on the possibility that adverse effects could manifest themselves later in adult life when 

compounded by normal age-related changes. In addition, further research is needed in areas for which the 

animal data have demonstrated exposure related effects, but the human data are inconclusive. Such 
CDDs 336 

2. HEALTH EFFECTS 

research in exposed humans should focus on diseases of the circulatory system, reproductive effects, 

immunological effects, effects on serum lipids, and effects on thyroid function. Hepatic effects were 

observed in animals after chronic exposure to CDDs (including 2,3,7,8-TCDD, mixed HxCDD isomers, and 

2,7-DCDD) by the oral route (NCI/NTP 1979a, 1980; NTP 1982b) and to 2,3,7,8-TCDD by the dermal 

route (Schwetz et al. 1973). The 2,3,7,8-TCDD congener was the most toxic. Studies in monkeys 

demonstrated their high susceptibility to 2,3,7,8-TCDD-induced toxicity. Developmental behavioral effects 

were seen in offspring of monkeys chronically exposed to low oral doses (Bowman et al. 1989a; Schantz 

and Bowman 1989; Schantz et al. 1992). The lowest dose tested in this series of studies was used to derive 

a chronic-duration oral MRL. 

No studies were located regarding chronic effects of CDD exposure by the inhalation route. Toxicokinetic 

inhalation data and chronic-duration studies would be useful for assessing the risk levels for people living 

near municipal, medical, and industrial waste incinerators who can be exposed for chronic durations to 

CDDs by this route. 

Several epidemiological studies of phenoxy herbicide and chlorophenol producers found increases in cancer 

mortality in populations exposed to 2,3,7,8-TCDD (Fingerhut et al. 1991; Kogevinas et al. 1993; Manz et 

al. 1991; Zober et al. 1990). 2,3,7,8-TCDD exposure has been especially associated with the development 

of soft-tissue sarcoma after a prolonged latency period (Eriksson et al. 1981, 1990; Fingerhut et al. 1991; 

Hardell and Eriksson 1988; Hardell and Sandrom 1979; Kogevinas et al. 1995; Smith et al. 1984a). The 

human data suggest that 2,3,7,8-TCDD may be a human carcinogen; however, the interpretation of many of 

these studies is limited by confounding factors (e.g., small cohorts, short latency periods, co-exposure to 

other chemicals, inadequate exposure data). Since these factors are inherent to epidemiological studies, it is 

unlikely that new human studies would clarify this issue. There are no reliable human studies on the 

carcinogenicity of other CDDs. Animal studies provided sufficient evidence that 2,3,7,8-TCDD is a 

carcinogen after oral (Kociba et al. 1978a; NTP 1982b; Toth et al. 1979) and dermal (Della Porta et al. 

1987; Rao et al. 1988) exposure. Furthermore, 2,3,7,8-TCDD has promoting ability on tumors initiated by 

diethylnitrosourea (Hebert et al. 1990; Poland et al. 1982). Similarly, chronic oral exposure of rodents to a 

mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD or to 2,7-DCDD resulted in carcinogenic effects 

(NCI/NTP 1979a, 1980). No studies were located regarding cancer effects in animals following inhalation 

exposure to CDDs. However, at this time, it is unlikely that such a study would add any new information 

regarding the potential carcinogenicity of CDDs in animals. 
 

 

CDDs 337 

2. HEALTH EFFECTS 

Genotoxicity. Inconclusive results were obtained regarding genotoxicity of CDDs in human as well as 

in animal studies. Structural chromosomal changes were found in some groups of exposed individuals 

(Kaye et al. 1985). However, the studies were confounded by small cohorts and unknown exposures. 

Positive and negative results at the chromosomal level (Green et al. 1977; Loprieno et al. 1982; Meyne et al. 

1985) as well as at the gene level (Randerath et al. 1989; Wahba et al. 1989) were reported in animal 

studies. Furthermore, negative results were obtained in dominant-lethal tests (Khera and Ruddick 1973) 

and sex-linked recessive-lethal tests in rats and Drosophila (Zimmering et al. 1985), respectively. In 

addition, mostly negative results were obtained in prokaryotic organisms (Geiger and Neal 1981; Gilbert et 

al. 1980; Toth et al. 1984). Some studies indicated that the covalent binding of 2,3,7,8-TCDD to DNA is 

low, and that this mechanism does not operate in CDD genotoxicity. Further studies on the mechanism of 

CDDs would be useful to evaluate the best possible method for detecting CDD genotoxicity. 

Reproductive Toxicity. Data from studies on reproductive effects in humans (Aschengrau and Monson 

1989; Egeland et al. 1994; Forsberg and Nordstrom 1985; Henriksen et al. 1996; Phuong et al. 1989a; 

Smith et al. 1982; USAF 1991; Wolfe et al. 1985, 1995) are inconclusive and are limited by confounding 

factors such as small cohorts, co-exposure to other chemicals, and inadequate exposure data. Better 

controlled epidemiological studies measuring 2,3,7,8-TCDD exposure levels or 2,3,7,8-TCDD body 

burdens would be useful to assess the human reproductive toxicity risk. Reproductive effects have been 

observed in oral animal studies. Increased incidences of pre- and post-implantation losses were observed in 

2,3,7,8-TCDD-exposed rodents (Giavini et al. 1983; Neubert and Dillman 1972; Smith et al. 1976; 

Sparschu et al. 1971a), rabbits (Giavini et al. 1982), and monkeys (McNulty 1985). Adverse effects have 

also been observed in the reproductive organs (decreased weight), hormone levels, and gametes of male rats 

(Khera and Ruddick 1973; Moore et al. 1985) and non-pregnant female rats (Li et al. 1995a, 1995b). None 

of the acute-duration exposure studies assessed the potential of CDDs to impair fertility; data on fertility 

would be useful in assessing potential effects in humans exposed to CDDs for a short period of time. 

Reduced fertility (Bowman et al. 1989b; Hong et al. 1989; Murray et al. 1979; Schantz et al. 1992), 

increased incidence of abortions (Bowman et al. 1989b; Hong et al. 1989; McNulty 1984; Schantz et al. 

1992), altered estrus cycle (Umbreit et al. 1987), and endometriosis (Rier et al. 1993) were observed in 

animals exposed for intermediate or chronic durations. Reproductive effects have also been observed in 

animals exposed to mixed HxCDD (Schwetz et al. 1973), but not following exposure to 2-MCDD, 

2,3-DCDD, 2,7-DCDD, 1,2,3,4-TCDD, or OCDD (Khera and Ruddick 1973). Data on the reproductive 

toxicity of CDD following dermal exposure is limited to a single animal study which found no adverse 

effects on reproductive organs of mice chronically exposed to 2,3,7,8-TCDD (NTP 1982a). No animal 
 

 

CDDs 338 

2. HEALTH EFFECTS 

inhalation reproductive toxicity studies were located. Additional animal inhalation and dermal reproductive 

studies, particularly studies which assessed reproductive performance, would be useful to assess the 

possible risk in humans exposed to CDDs by these routes. 

Developmental Toxicity. Studies in humans and animals indicated that 2,3,7,8-TCDD can cross the 

placenta and is excreted in milk (Fürst et al. 1989b; Schecter et al. 1989d, 1989g, 1990). Studies on the 

developmental toxicity of 2,3,7,8-TCDD in humans are inconclusive. Some studies have found significant 

increases in the risk of certain birth defects (Aschengrau and Monson 1990; Erickson et al. 1984; Hanify et 

al. 1981; Nelson et al. 1979; Phuong et al. 1989a; Wolfe et al. 1985, 1995), while other studies have found 

no significant alterations (Bisanti et al. 1980; Mastroiacovo et al. 1988; Townsend et al. 1982). However, a 

number of limitations (e.g., lack of exposure data, small sample sizes, and the lack of reliable data for birth 

defects prior to 2,3,7,8-TCDD exposure) limits the interpretation of the results of these studies. 

Epidemiology studies which measure exposure concentrations or body burdens would be useful to 

determine if 2,3,7,8-TCDD and other CDD congeners are human developmental toxicants. Developmental 

toxicity has been observed in animals orally exposed to 2,3,7,8-TCDD (Abbott and Birnbaum 1989a; 

Abbott et al. 1992; Bjerke and Peterson 1994; Bjerke et al. 1994a, 1994b; Bowman et al. 1989a, 1989b; 

Brown et al. 1998; Courtney 1976; Couture-Haws et al. 1991b; Giaviani et al. 1983; Gordon et al. 1995; 

Gray and Ostby 1995; Gray et al. 1995; Håkansson et al. 1987; Huuskonen et al. 1994; McNulty 1985; 

Moore et al. 1973; Neubert and Dillman 1972; Roman et al. 1998a, 1998b; Schantz et al. 1992; Silkworth 

et al. 1989b; Smith et al. 1976; Thomas and Hinsdill 1979; Weber et al. 1985), 2,7-DCDD (Khera and 

Ruddick 1973; Schwetz et al. 1973), mixed HxCDD (Schwetz et al. 1973), and OCDD (Schwetz et al. 

1973). The most common effects were cleft palate, hydronephrosis, impaired development of the 

reproductive system, immunotoxicity, and death. No studies were located regarding developmental effects 

in animals after inhalation and dermal exposure. Such studies would be useful for extrapolating the 

possible risk to human populations exposed environmentally by these routes. 

Immunotoxicity. Studies in humans did not provide conclusive evidence regarding immunotoxicity of 

CDDs (Ernst et al. 1998; Jansing and Korff 1994; Jennings et al. 1988; Jung et al. 1998; Mocarelli et al. 

1986; Neubert et al. 1993, 1995; Reggiani 1980; Stehr et al. 1986; Svensson et al. 1994; Tonn et al. 1996; 

USAF 1991; Webb et al. 1989; Wolfe et al. 1985). Studies in animals indicated that CDDs are immunosuppressive (Kerkvliet 1995). 2,3,7,8-TCDD induced thymic atrophy or thymic weight changes after oral 

(Hanberg et al. 1989; McConnell et al. 1978b), dermal (Hebert et al. 1990), and parenteral exposure (Gorski 

et al. 1988b; Olson et al. 1980a). Bone marrow degeneration was reported in orally exposed 
 

CDDs 339 

2. HEALTH EFFECTS 

monkeys (Hong et al. 1989). Suppressed cell-mediated and humoral immunity was found in rodents after 

intermediate-duration exposure (Vos et al. 1973). Similarly, immunotoxic effects were found after oral 

exposure of rodents to 2,7-DCDD or to a mixture of 1,2,3,6,7,8-HxCDD and 1,2,3,7,8,9-HxCDD 

(Holsapple et al. 1986b; NCI/NTP 1980). At least in mice, differences in responsiveness to CDDs’ 

immunotoxicity in vivo segregated with the Ah locus (Nagayama et al. 1989; Vecchi et al. 1983a). 

Studies in animals aimed at identifying 2,3,7,8-TCDD-sensitive immune end points that can also be 

measured in humans would be valuable to determine correlative changes in the biomarker and immune 

function. However, this can be done only after establishing a database of normal values for the clinical 

immunology end points that may be used as biomarkers of immune-function in immunotoxicity 

assessments. It also important to determine in animals how well changes in lymphoid organs correlate with 

changes in the expression of lymphocyte subset/activation markers in peripheral blood. The role of the Ah 

receptor in the immunotoxicity of 2,3,7,8-TCDD needs to be researched in species other than mice. In 

addition, the role of Ah receptor-independent processes in 2,3,7,8-TCDD-induced immunotoxicity needs to 

be examined further. Such actions may include changes in intracellular calcium or in the activity of 

kinase/phosphatase systems, or interactions with hormone systems. A battery of immune function tests in 

human cohorts exposed to CDDs would be useful for detecting the immunotoxic responses in exposed 

individuals. The ability of CDD-exposed individuals to mount an integrated functional response to a novel 

antigen, such as hepatitis B vaccine, would provide a broad measure of immune function in exposed human 

populations. 

Neurotoxicity. Studies in Vietnam veterans could not conclusively demonstrate cognitive or other 

central nervous system deficits (Goetz et al. 1994). Neurological examinations revealed neurological 

effects in humans exposed to a CDD-contaminated environment (Pocchiari et al. 1979) and in occupational 

settings (Goldman 1973; Jirasek et al. 1976; Klawans 1987; Pazderova-Vejlupkova et al. 1981) shortly 

following exposure, but reports with comparison groups do not offer clear evidence that exposure to 

2,3,7,8-TCDD is associated with chronic peripheral neuropathy (Suskind and Hertzberg 1984; Sweeney et 

al. 1993). No notable neurological effects were found in laboratory animals after oral and dermal exposure. 

The existing information suggests that in adults, no long-term neurologic affects were even caused by high 

exposure to 2,3,7,8-TCDD-contaminated materials. However, the possibility exists that subtle central 

nervous system changes acquired in early adulthood could manifest themselves later in adult life when 

compounded by normal age-related changes in the central nervous system (Goetz et al. 1994). Thus, it 

would be of interest to include tests of neurological function in ongoing prospective studies of 
 

 

CDDs 340 

2. HEALTH EFFECTS 

2,3,7,8-TCDD-exposed populations to determine if neurological effects occur as the exposed population 

ages. 

Epidemiological and Human Dosimetry Studies. Epidemiology studies have investigated the 

toxicity of 2,3,7,8-TCDD in populations exposed in the workplace or in the contaminated environment 

(after industrial accidents or herbicide spraying) (Bertazzi et al. 1993; Calvert et al. 1992, 1996, 1998; 

Egeland et al. 1994; Eriksson et al. 1981, 1990; Fingerhut et al. 1991; Flesch-Janys et al. 1995; Hardell and 

Eriksson 1988; Hardell and Sandrom 1979; Manz et al. 1991; Mastoiacovo et al. 1988; Mocarelli et al. 

1991; Pesatori et al. 1993, 1998; Saracci et al. 1991; Smith et al. 1984a; Sweeney et al. 1993; Vena et al. 

1998; Zober et al. 1990) and in Vietnam veterans exposed to Agent Orange (Burton et al. 1998; Henriksen 

et al. 1997; USAF 1991; Wolfe et al. 1985, 1995). The interpretation of the results of most of these studies 

is confounded by such factors as unknown levels of exposure, too short or too long postexposure periods, 

and small cohorts. Well conducted epidemiological and occupational studies that quantify exposure levels 

would be useful to assess the risk for the main end points of concern (i.e., reproductive, developmental, 

immunotoxic effects, and cancer). Some of the more recent studies have measured the levels of 

2,3,7,8-TCDD and related compounds in serum lipid; these levels can then be used to estimate body burden 

at the time of exposure. There are a number of drawbacks associated with extrapolating body burdens back 

to the time of the original exposure using current serum 2,3,7,8-TCDD levels; these include uncertainty 

associated with 2,3,7,8-TCDD half-life in humans and having to use average serum 2,3,7,8-TCDD levels 

and average exposure durations and reference body weights and percentage of body fat. There is a lack of 

consensus on the half-life of 2,3,7,8-TCDD in humans, half-lives ranging from 5 to 12 years have been 

estimated (Pirkle et al. 1979; Schecter et al. 1994b; Wolfe et al. 1994). Additional human studies 

measuring 2,3,7,8-TCDD half-life would be useful in establishing dose-response relationships for human 

effects. All of the above limitations for assessing the body burden of 2,3,7,8-TCDD also apply to other 

CDDs where far less human toxicokinetic data are available. Thus, it would be useful to have congenerspecific human toxicokinetic data on other CDDs and related compounds. Furthermore, human dosimetry 

studies would be useful in occupational settings to obtain results regarding levels of CDDs in the 

environment as opposed to levels in serum or adipose tissues. 

Biomarkers of Exposure and Effect. 

Exposure. Several studies reported results of measurements of CDD levels in the lipid fraction of adipose 

tissue, milk, and serum from members of the general population with unknown CDD exposure (Andrews et 
 

CDDs 341 

2. HEALTH EFFECTS 

al. 1989; Ryan et al. 1985a; Schecter et al. 1987b). The gas chromatography-mass spectrometry (GS/MS) 

tests used to detect CDD levels are sensitive and specific. Analytical testing for levels in biological fluids 

and tissues can be used for monitoring exposed populations. While chloracne is a known, readily 

identifiable effect of exposure to CDDs, it is not useful as a biomarker of exposure because of its variable 

expression in individuals with even very high levels of exposure to these agents. Further information on 

how aging and changes in body composition can influence the distribution of CDDs in tissues and body 

fluids would be valuable. A reverse transcriptase polymerase chain reaction method has been used to 

quantify CYP1A1 mRNA levels on total RNA extracts from human blood lymphocytes (Van den Heuvel et 

al. 1993). This method was found to be much more sensitive than, for example, measuring EROD activity, 

and could potentially be used as a human exposure marker for CDDs and structurally related compounds. 

However, EROD activity measurements can be useful as a marker of exposure to the agents. 

Effect. There are no specific biomarkers of effects for CDDs. Exposure to relatively high concentrations of 

CDDs can lead to the development of chloracne in humans. However, while the presence of chloracne 

indicates CDD or similar halogenated-chemical exposure, lack of chloracne does not indicate that exposure 

has not taken place, as evidenced in a cohort from the Seveso incident (Mocarelli et al. 1991). Additional 

studies could evaluate the feasibility of using body burden as a biomarker for predicting other effects of 

CDDs. Although the results of an earlier study suggested that 2,3,7,8-TCDD may form adducts with DNA, 

albeit at an extremely low rate (Poland and Glover 1979), more recent studies that have rigorously looked 

for 2,3,7,8-TCDD-DNA adducts have been negative (Randerath et al. 1988; Turteltaub 1990). Expression 

of CYP1A1 mRNA, protein, and/or activity are sensitive biological responses in human tissues which can 

be observed following exposure to 2,3,7,8-TCDD and related compounds, and may be useful biomarkers of 

effects. Further studies to identify biomarkers of effects of CDDs would facilitate medical surveillance 

leading to early detection of potentially adverse health effects and possible treatment. 

Absorption, Distribution, Metabolism, and Excretion. There are no quantitative data regarding 

absorption in humans by the inhalation and dermal routes, but data from accidentally exposed individuals 

suggest that exposure by these routes may lead to a significant increase in body burden of CDDs (Patterson 

et al. 1994; Schecter 1994b). Results from one human study indicated that more than 87% of an oral 

2,3,7,8-TCDD dose in an oil vehicle was absorbed (Poiger and Schlatter 1986). Also, results from studies 

of absorption of CDDs from maternal milk by nursing infants showed that 90–95% of the dose of CDDs 

can be absorbed; hepta-substituted congeners and OCDD exhibited lower absorption rates (Abraham et al. 

1994, 1996; Dahl et al. 1995; McLachlan 1993; Pluim et al. 1993b). The data indicate that 2,3,7,8-TCDD 
CDDs 342 

2. HEALTH EFFECTS 

is effectively absorbed and that absorption is vehicle-dependent (Fries and Marrow 1975; Lucier et al. 1986; 

Poiger and Schlatter 1980); oil vehicles were most effective (Olson et al. 1980b; Piper et al. 1973). 

Transpulmonary absorption of 2,3,7,8-TCDD also occurs in animals (Diliberto et al. 1996; Nessel et al. 

1992). Dermal absorption of 2,3,7,8-TCDD in rats was found to be age-dependent (Banks et al. 1993). In 

rats, following single equivalent intratracheal, oral, and dermal 2,3,7,8-TCDD doses, absorption was 

calculated as 95, 88, and 40% of the administered dose, respectively (Diliberto et al. 1996). The available 

information shows that absorption of 2,3,7,8-TCDD has been fairly well characterized in animals. 

Based on analysis of CDDs in adipose tissue, milk, and blood, it appears that humans store exclusively 

2,3,7,8-chlorine substituted congeners (Fürst et al. 1987; Rappe et al. 1987; Van den Berg et al. 1986b). 

Data are available on tissue distribution of 2,3,7,8-TCDD in rats after inhalation, oral, and dermal exposure 

(Diliberto et al. 1996). The liver and adipose tissue are the major storage sites in animals. In general, 

distribution of CDDs is congener specific, and depends on the dose and route of administration (Diliberto et 

al. 1996; Van den Berg et al. 1994). Age was also a factor in the distribution of 2,3,7,8-TCDD in mice 

(Pegram et al. 1995). The distribution of 2,3,7,8-TCDD-derived radioactivity in subcellular liver fractions 

has also been studied (Santostefano et al. 1996). 2,3,7,8-Chlorine substituted CDDs are the predominant 

congeners retained in tissue and body fluids from humans, rodents, and monkeys (Abraham et al. 1989c; 

Van den Berg et al. 1983). Further dosimetry studies of various durations in which levels of 2,3,7,8-TCDD 

and related compounds are monitored in tissues suspected of being targets for 2,3,7,8-TCDD toxicity would 

provide valuable information. These data can be used to establish correlations between target-tissue doses 

and adverse effects. 

Data regarding the biotransformation of CDDs in humans are limited to a self-dosing experiment that 

provided some evidence that 2,3,7,8-TCDD is partially excreted in the feces in the form of metabolites 

(Wendling et al. 1990). The use of human cell systems in culture might be considered a useful addition to 

whole-animal studies for examining the metabolic fate of CDDs. Biotransformation of CDDs has been 

examined in several species, but the structure of metabolites has been elucidated only in the rat and dog 

(Poiger and Buser 1984). Although information regarding metabolism following inhalation or dermal 

exposure is lacking, there is no reason to believe that different pathways would operate after exposure by 

these routes. 

Only one study was located that provide limited evidence of fecal excretion of 2,3,7,8-TCDD metabolites in 

adult humans (Wendling et al. 1990). Several studies provided information regarding fecal excretion of 
 

CDDs 343 

2. HEALTH EFFECTS 

CDDs in infants exposed through breast milk (Abraham et al. 1994; McLachlan 1993; Pluim et al. 1993b). 

Elimination of CDDs through maternal milk is well documented (Fürst et al. 1994; Rappe et al. 1985; 

Schecter and Gasiewicz 1987a; Schecter et al. 1989d, 1989e). Fecal excretion is the main route of excretion 

of CDDs in animals after all routes of exposure (Diliberto et al. 1996). Estimates of 2,3,7,8-TCDD half-life 

in humans are available (Pirkle et al. 1989; Poiger and Schlatter 1986; Wolfe et al. 1994), but further 

information regarding the relationships between aging, fat redistribution, and half-lives in humans would be 

valuable. 

Comparative Toxicokinetics. CDDs are efficiently absorbed from the gastrointestinal tract of 

mammals, but the vehicle plays an important role (Olson et al. 1980b; Piper et al. 1973; Poiger and 

Schlatter 1986; Van den Berg et al. 1987c). Distribution data in orally exposed rodents indicated that the 

highest postexposure levels were in the liver followed by the fat (Diliberto et al. 1996; Khera and Ruddick 

1973; Olson 1986), but distribution is highly dose- and species-dependent. The studies to date suggest that 

compared with rodents, primates, including humans, accumulate significantly less CDDs in the liver than in 

adipose tissue (Neubert et al. 1990a; Ryan et al. 1986; Van Miller et al. 1976). With the exception of the 

guinea pig, mammals retain only 2,3,7,8-substituted congeners. The high liver retention of 2,3,7,8­

substituted congeners by rodents has been attributed to the presence of inducible storage sites, presumably 

CYP1A2 (Leung et al. 1990b). In all mammalian species studied, exposure by breast-feeding has a much 

greater contribution to the offspring 2,3,7,8-TCDD body burden than placental transfer. Metabolic 

capacities are species-dependent. Rats, hamsters, and mice metabolize and eliminate CDDs much faster 

than the guinea pig. The metabolites were excreted predominantly via the bile and feces, with minor 

amounts excreted in the urine in all species (Diliberto et al. 1996; Fries and Marrow 1975; Weber and 

Birnbaum 1985). Whole-body half-lives ranged from 11 days in hamsters (Olson et al. 1980b) to more than 

1 year in monkeys (Bowman et al. 1989b; McNulty et al. 1982), and approximately 7–12 years in humans 

(Wolfe et al. 1994). The toxicity of CDDs has been associated with the parent compound and not the 

metabolites (Mason and Safe 1986a; Weber et al. 1982); therefore, metabolism and excretion represent a 

detoxification process. The data collected in recent years indicate differences in species susceptibility to 

CDDs cannot be explained by differences in toxicokinetics alone; it is likely that genetic factors have an 

important role. Based on this information, species-, congener-, and dose-specific toxicokinetic data need to 

be factored in human risk assessment for CDDs. Several models that describe the disposition of 

2,3,7,8-TCDD in animals and humans were identified from the literature (Andersen et al. 1993, 1997a, 

1997b; Carrier et al. 1995a, 1995b; Kissel and Robarge 1988; Kohn et al. 1993; Leung et al. 1988, 
 

 

CDDs 344 

2. HEALTH EFFECTS 

1990b). Although each new model that is published usually fills data gaps identified in earlier models, 

further research is necessary to increase their reliability for use in human risk assessment. 

Methods for Reducing Toxic Effects. The mechanism by which CDDs enter the blood stream in 

humans is not known; consequently, there are no established methods for reducing absorption. In 

experimental animals, however, administration of a diet containing activated charcoal reduced mortality in 

an acute-duration study presumably by preventing gastrointestinal absorption and the reabsorption of the 

chemical from biliary secretions (Manara et al. 1984). Identification of additional substances that could 

prevent or delay absorption and that do not represent a toxic risk per se would be valuable. Increasing the 

fact content of the diet by ingesting non-absorbable lipids has been suggested as a method for increasing the 

elimination rate (Rohde et al. 1997). These authors estimated that if the normal feces excretion of 5 g 

fat/day was quadrupled and the lipid based distribution of CDDs/CDFs between the body and the intestine 

stayed the same, the overall elimination rate would at least double. There are no established methods for 

reducing body burden in humans, but data from a study of Vietnam veterans suggested that persons with 

more fat tend to eliminate 2,3,7,8-TCDD more slowly (Wolfe et al. 1994). It was suggested that metabolic 

or other factors that change with age (i.e., redistribution of fat stores) affect 2,3,7,8-TCDD elimination. 

Studies examining the effect of fasting in animals exposed to CDDs would provide useful information that 

can be used to characterize the effectiveness of this approach better. Although, in recent years, great 

advances have been made related to the understanding of the mechanism of action of CDDs, no methods 

exist to block the toxic response due to exposure to CDDs. Further characterization of the Ah receptor 

protein and understanding of physiological effects of interfering with the chain of events that follow 

binding of CDDs to the Ah receptor would be useful for the possible identification of blockers of those 

events. Further studies aimed at elucidating the non Ah receptor-mediated mechanisms of action of CDDs 

would also be valuable. There are no established methods for mitigation of health effects resulting from 

exposure to CDDs. 

Children’s Susceptibility. A limited number of human studies have examined health effects of CDDs 

in children. Data from the Seveso accident suggest that children may be more susceptible to the dermal 

toxicity of 2,3,7,8-TCDD (chloracne), but it is not known if this would be the case for other effects. 

Follow-up medical surveillance of the Seveso children (including measurement of serum 2,3,7,8-TCDD 

levels) would provide information on whether childhood exposure would pose a risk when the individual 

matures and ages. The available human and animal data provide evidence that 2,3,7,8-TCDD can cross the 

placenta and be transferred to an infant via breast milk. Although information on the developmental 
CDDs 345 

2. HEALTH EFFECTS 

toxicity of CDDs in humans is limited, there are extensive animal data that the developing organism is very 

sensitive to the toxicity of 2,3,7,8-TCDD. Several human studies have found significant alterations in 

markers of liver, thyroid, immune, and neurological function in young breast-fed infants of mothers with 

higher current background or general population CDD levels. Recent data suggest that the neurological 

effects are reversible; prospective studies of the breast-fed individuals would provide useful information on 

whether these children are at risk of developing additional effects as they age. Further data needs relating to 

developmental effects are discussed above under Developmental Toxicity. 

In general, the available toxicokinetic data did not examine potential differences between adults and 

children; toxicokinetic studies examining how aging and changes in body composition can influence 

distribution and turnover rates would be useful in assessing children’s susceptibility to CDD toxicity. Most 

of the available mechanism of action data suggest that the toxicity of 2,3,7,8-TCDD is mediated through the 

Ah receptor. We do not know if there are any age-related differences in receptor binding or expression; 

studies in animals would be valuable to fill this information gap. No age-specific biomarkers of exposure 

or effect were identified for CDDs; the long half-life of 2,3,7,8-TCDD in humans, suggests that there may 

not be a way to assess whether adults were exposed as children to 2,3,7,8-TCDD. Additionally, there are 

no data to determine whether there are any interactions with other chemicals which would be specific for 

children. There is very little available information on methods for reducing 2,3,7,8-TCDD toxic effects or 

body burdens; it is likely that research in adults would also be applicable to children. 

Child health data needs relating to exposure are discussed in Section 5.8.1, Identification of Data Needs: 

Exposures of Children. 

2.11.3 Ongoing Studies 

Ongoing studies regarding the health effects of CDDs were reported in the Federal Research in Progress 

File (FEDRIP 1998). Table 2-18 presents a summary of this information. 











CDDs 357 

3. CHEMICAL AND PHYSICAL INFORMATION
 

CDDs are a class of related chlorinated hydrocarbons which are structurally similar. The basic structure is a 

dibenzo-p-dioxin (DD) molecule, which is comprised of 2 benzene rings joined at their para carbons by 

2 oxygen atoms. There are 8 homologues of CDDs, monochlorinated through octachlorinated. The class of 

CDDs contains 75 congeners, consisting of 2 monochlorodibenzo-p-dioxins (MCDDs), 10 dichlorodibenzop-dioxins (DCDDs), 14 trichlorodibenzo-p-dioxins (TrCDDs), 22 tetrachlorodibenzo-p-dioxins (TCDDs), 

14 pentachlorodibenzo-p-dioxins (PeCDD), 10 hexachlorodibenzo-p-dioxins (HxCDDs), 2 heptachlorodibenzo-p-dioxins (HpCDDs), and a single octachlorodibenzo-p-dioxin (OCDD) (Ryan et al. 1991). 

The general structure of the dibenzo-p-dioxins is shown below. The numbers indicate the positions for 

chlorine substitutions, excluding, of course, positions 5 and 10. 

O 

O 

10 1 

2 

3 

6 45

7 

8 

9 

Not all congeners have been studied for their chemical and physical properties, but basic properties are 

known for the CDDs as a chemical family and for the homologous groups. Chlorinated dioxins exist as 

colorless solids or crystals in the pure state. They have a low solubility in water and a low volatility. 

Chlorinated dioxins have an affinity for particulates and readily partition to particles in air, water, and soil. 

The more toxic compounds appear to be the 2,3,7,8-substituted tetra-, penta-, and hexachloro compounds 

(i.e., 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 1,2,3,4,7,8-HxCDD, 1,2,3,6,7,8-HxCDD, and 1,2,3,7,8,9-HxCDD). 

These are also the congeners, along with OCDD, that have the greatest tendency to bioaccumulate. One of 

the most toxic congeners in mammals is believed to be 2,3,7,8-TCDD; this compound has also been the 

most studied of the TCDD congeners. 

3.1 CHEMICAL IDENTITY 

Information regarding the chemical identities of CDDs is presented in Table 3-1. 

3.2 PHYSICAL AND CHEMICAL PROPERTIES 

Information regarding the physical and chemical properties of CDDs is presented in Table 3-2. 











CDDs 369 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

4.1 PRODUCTION 

CDDs are not manufactured commercially in the United States except on a small scale for use in chemical 

and toxicological research. CDDs are unique among the large number of organochlorine compounds of 

environmental interest in that they were never intentionally produced as desired commercial end products 

(Zook and Rappe 1994). Typically, CDDs are unintentionally produced during various uncontrolled 

chemical reactions involving the use of chlorine (EPA 1990c) and during various combustion and 

incineration processes (Zook and Rappe 1994). In the process of making white paper products, for 

example, chlorine or chlorine derivatives are often used as the primary bleaching agent. As a result, several 

chlorinated organic compounds are formed, including small amounts of CDDs (EPA 1990c). These 

chlorinated compounds not only leave the mills in the pulp and paper products, they are also released 

through waste waters (effluents from the mills) and sludge produced as a result of waste water treatment 

(EPA 1990c). CDDs are also produced as undesired by-products during the manufacture of chlorinated 

phenols such as pentachlorophenol, 2,4,5-trichlorophenol, and related chemicals, and during incineration of 

chlorinated wastes (IARC 1977; NTP 1989; Podoll et al. 1986). By far the greatest unintentional 

production of CDDs occurs via various combustion and incineration processes including all forms of waste 

incineration (municipal, industrial, and medical), many types of metal production (iron, steel, magnesium, 

nickel, lead, and aluminum), and fossil fuel and wood combustion (Czuczwa and Hites 1986a, 1986b; 

Oehme et al. 1987, 1989; Zook and Rappe 1994). More extensive information on sources of CDDs released 

to the environment can be found in Chapter 5. 

In general, there are two conventional methods for the preparation of CDDs for research purposes: 

condensation of a polychlorophenol and direct halogenation of the parent dibenzo-p-dioxin or a 

monochloro-derivative. For example, 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) is generally 

synthesized by the condensation of two molecules of 2,4,5-trichlorophenol in the presence of a base at high 

temperatures or by chlorination of dibenzo-p-dioxin in chloroform in the presence of iodine and ferric 

chloride (EPA 1987k; IARC 1977). Other methods of 2,3,7,8-TCDD synthesis include the following: 

pyrolysis of sodium α-(2,4,5-trichlorophenoxy) propionate at 500 EC for 5 hours; reaction of 

dichlorocatechol salts with o-chlorobenzene by refluxing in alkaline dimethyl sulfoxide; ultraviolet 

irradiation of CDDs of high chlorine content; Ullman reaction of chlorinated phenolates at 180–400 EC; 
CDDs 370 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

pyrolysis of chlorinated phenolates and chlorinated phenols; and heating 1,2,4-trichloro-5-nitrobenzene and 

4,5-dichlorocatechol in the presence of a base (EPA 1984a; IARC 1977). 

1,2,3,4-TCDD has been prepared by refluxing a mixture of catechol, potassium carbonate, pentachloronitrobenzene and acetone in nitrogen (IARC 1977). 

DCDD can be synthesized by two methods. In the first method, 2-bromo-4-chlorophenol and potassium 

hydroxide are dissolved in methanol and evaporated to dryness. The residue is then mixed with 

bis(2-ethoxyethyl) ether, ethylene diacetate, and a copper catalyst; and then heated, cooled, and eluted from 

a chromatographic column with chloroform. This residue is evaporated and then sublimed. DCDD can also 

be synthesized by heating the potassium salt of 2,4-dichlorophenol in the presence of copper powder in a 

vacuum sublimation apparatus (IARC 1977). 

1,2,4,6,7,9-HxCDD has been made by heating the potassium salt of 2,3,5,6-tetrachlorophenol with 

powdered copper and potassium carbonate in a vacuum sublimation apparatus (IARC 1977). 

1,2,3,4,7,8-HxCDD has been prepared by mixing 1,2,3,4-TCDD, ferric chloride, chloroform, and a crystal 

of iodine and then adding a solution of chlorine in carbon tetrachloride (IARC 1977). 

OCDD has been synthesized by the following methods: irradiation of aqueous solutions of CDD-free 

sodium pentachlorophenol with ultraviolet light; heating the potassium salt of pentachlorophenol; heating 

pentachlorophenol in the presence of an initiator, such as chlorine, bromine, iodine, or 2,3,4,4,5,6-hexachloro-2,5-cyclohexadienone; and heating hexachlorocyclohexadienone in an atmosphere of carbon dioxide 

for 30 minutes (Crosby and Wong 1976; EPA 1984a; IARC 1977). 

At present, the only reported producers of CDDs are Eagle Picher Industries, Inc., located in Lenexa, 

Kansas, and Cambridge Isotope Laboratories, located in Andover, Massachusetts. Eagle Picher Industries 

produces 2,3,7,8-TCDD and OCDD for research purposes (SRI 1991). Cambridge Isotope Laboratories 

produced unlabeled chlorodioxin standards (TCDD through HpCDD) and C13 labeled chlorodioxin 

standards (DCDD through OCDD) for use in chemical analyses and in toxicological research (Cambridge 

Isotope Laboratories 1995). 
CDDs 371 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Since CDD releases are not required to be reported under Superfund Amendments and Reauthorization Act 

(SARA) Section 313, there are no data on CDDs in the 1994 Toxics Release Inventory (TRI) (EPA 1995g). 

4.2 IMPORT/EXPORT 

2,3,7,8-TCDD is not imported into the United States (NTP 1989). There were no data located pertaining to 

the export of 2,3,7,8-TCDD or any other CDD for research purposes. 

4.3 USE 

The only reported use of CDDs/CDFs is as research chemicals (NTP 1989). A large diversified group of 

researchers use various CDDs in studies of toxicology, environmental fate, transformation, and transport, 

and in residue analysis of a variety of contaminated media. CDDs have been tested for use in flameproofing polymers such as polyesters and against insects and wood-destroying fungi; however, there are no 

data reporting its commercial production or use for these purposes (IARC 1977). 

4.4 DISPOSAL 

The 1994 estimates on the degree of TCDD contamination in the environment indicated that approximately 

500,000 tons of soil and sediment in the United States were contaminated with 2,3,7,8-TCDD (Hilarides et 

al. 1994). The development of treatment technologies for CDD-contaminated soils and wastes needed to 

address unique problems associated with CDDs: for example, they are insoluble in water, only slightly 

soluble in organic solvents, have a strong affinity for adsorption on organic matter, and are biologically and 

environmentally stable (U.S. Congress 1991). In order to meet the clean-up standards established for 

CDDs, the treatment system must be capable of removing the CDDs from the contaminated matrix (U.S. 

Congress 1991). Several treatment or disposal methods for CDDs and CDD-contaminated materials have 

been investigated, including land disposal, thermal destruction, and chemical and biological degradation. 

Each of these methods has limitations regarding economics, technical feasibility, and acceptability (HSDB 

1995). 

Land disposal of CDD-containing wastes is currently prohibited (EPA 1986f, 1988f). The Toxic 

Substances Control Act (TSCA) regulates the use, disposal, and distribution in commerce of process waste 

water treatment sludges intended for land application from pulp and paper mills employing chlorine or 
CDDs 372 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

chlorine derivative-based bleaching processes (EPA 1991b, 1991c). Also, under the Marine Protection 

Research and Sanctuaries Act, ocean dumping of CDD-containing wastes is prohibited except when only 

trace amounts are present (EPA 1977a, 1977b). 

Thermal destruction technologies offer the most straightforward approach to treating or disposing of CDDcontaminated materials because under the appropriate conditions the breakdown of the CDDs is assured 

(U.S. Congress 1991). The thermal treatment technologies that are currently used to treat waste containing 

hazardous or toxic constituents and that have demonstrated potential use toward the treatment of CDDcontaminated waste include rotary kiln incineration, liquid injection incineration, fluidized-bed incineration, 

advanced electric reactor (AER), infrared incineration, plasma arc pyrolysis incineration, supercritical water 

oxidation, and in situ vitrification (U.S. Congress 1991). In addition to kiln incinerators, the technologies 

that have been field-tested for treating CDD-contaminated media under EPA’s Superfund Innovative 

Technology Evaluation (SITE) program include dechlorination, stabilization, and in situ vitrification (U.S. 

Congress 1991). Although some alternatives look promising and have been shown effective in the 

laboratory or in application to other pollutants, more development and testing is needed to demonstrate 

viability for large-scale treatment of CDD contamination. 

Incineration, involving the high-temperature oxidation of CDD molecules, is the most extensively tested 

method for disposal of CDDs. CDDs such as TCDD, PeCDD, and HxCDD are classified by EPA as 

Principal Organic Hazardous Constituents (POHCs) and are required to be incinerated under conditions that 

achieve a destruction and removal efficiency of 99.99% (EPA 1990b; Sedman and Esparza 1991). 

Incinerator operating conditions currently considered adequate for destruction of 2,3,7,8-TCDD and most 

other chlorinated organics require a temperature of at least 1,500–2,600 EF, with residence times of at least 

30 minutes (although 1.5 hours is a more common residence time) to ensure complete destruction (EPA 

1990a). Thermal destruction of CDDs that are adsorbed on fly ash can be accomplished through the use of 

a rotary kiln furnace combined with a baghouse filter for the recycling of entrained fly ash and an activated 

carbon filter for adsorption of CDD traces transported in the gas phase. This method is capable of 

destroying 99.5% of CDDs in fly ash, which is considered a high level of efficiency (Kahr et al. 1990). 

EPA's Mobile Incineration System, a transportable rotary kiln system, was judged to be more than adequate 

for detoxifying CDD-contaminated solids and liquids after it was performance-tested with a variety of 

uncontaminated soils and other solid wastes, and thus could be expected to accomplish a successful CDD 

trial burn. The system, which has been extensively modified for field use, consists of a 
CDDs 373 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

rotary-kiln, a secondary combustion chamber, an air pollution control unit, and separate continuous stackgas analysis capabilities (HSDB 1995). In 1977, the U.S. Air Force disposed of Agent Orange 

contaminated with 2,3,7,8-TCDD by high temperature incineration at sea (Bumb et al. 1980). The high 

flame temperature reached 1,500 EC in the incinerator, and EPA determined a combustion efficiency of 

99.9% for 2,3,7,8-TCDD. 

Kiln incinerators have been used to treat a variety of containerized and noncontainerized solid and liquid 

wastes. Since the waste can be treated individually or simultaneously, the versatility of this technology has 

made it popular in the United States for disposing of hazardous waste. For the disposal of CDD-containing 

waste, however, kiln incineration is more commonly practiced in Europe than in the United States (U.S. 

Congress 1991). Although liquid injection incineration has been used for ocean-based incineration of 

Agent Orange, certain limitations must be considered before applying the technology to treating CDD 

contamination. These limitations include the applicability of the technology only to combustible lowviscosity liquids and slurries that can be pumped; atomizing the waste prior to injection into the combustor; 

and the importance of particle size because burners are susceptible to clogging (U.S. Congress 1991). 

Fluidized-bed combustion (FBC) systems have traditionally been used to treat the sludge produced by 

municipal waste treatment plants and waste generated from oil refineries, pulp and paper mills, and the 

pharmaceutical industry. The system consists of a vertical refractory-lined vessel which holds a perforated 

plate. A bed of granular material, usually sand, is placed on the perforated plate. The system uses forced 

hot air to fluidized the bed and cause a highly turbulent zone that ensures the mixing of the waste with bed 

particles and the combustion air. Combustion is facilitated by an overhead burner (U.S. Congress 1991). 

The type and size of materials to be treated are critical because variations in gravity and density could be 

deleterious to the process (U.S. Congress 1991). Modification of the traditional FBC system for treatment 

of chlorinated wastes continues to be investigated by researchers in the private sector. A modified system 

designed by Waste-Tech Services, Inc. uses a granular bed composed of a mixture of combustion catalyst 

and limestone. The results of the trial burn for the Waste-Tech Services system which used chlorinated 

waste containing carbon tetrachloride, tetrachloroethane, p-dichlorobenzene and some CDDs and CDFs, 

showed no measurable amount of any of the chlorinated pollutants treated and no 2,3,7,8-TCDD in any of 

the samples tested (U.S. Congress 1991). In situ vitrification (ISV), which treats waste in place, solidifies 

all materials not volatilized or destroyed. Bench-scale testing of ISV on soils containing 10 ppb CDDs 

showed destruction removal efficiency (DRE) values of 99.9999% (U.S. Congress 1991). 
CDDs 374 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Since the early 1970s, several chemical methods have been investigated for the degradation of CDDs. 

Treatment of CDD-contaminated materials with alkali polyethylene glycolate (APEG) reagents at 

hazardous waste sites has been demonstrated to successfully destroy CDDs in liquid wastes and to be viable 

even under difficult circumstances. This method involves the reaction of potassium hydroxide with 

polyethylene glycol to form an alkoxide that reacts with one of the chlorine atoms on the CDD to produce 

an ether and potassium chloride. Bioassays indicate that the by-products produced by treating 

2,3,7,8-TCDD with APEG reagents do not bioaccumulate or bioconcentrate, do not cause mutagenicity, and 

are far less toxic than 2,3,7,8-TCDD (Klee 1988). Cleavage of the ether linkages with the formation of 

halophenols may be achieved by treatment with strong acids or quaternary ammonium salts, but the 

dibenzodioxin nucleus is resistant to chemical attack. Oku et al. (1995) investigated the dechlorination of 

polychlorinated dibenzo-p-dioxins (CDDs) and polychlorinated dibenzofurans (CDFs) using a modified 

alkali-metal hydroxide method. The destruction reagent, prepared by dissolving either potassium hydroxide 

or sodium hydroxide in 1,3-dimethyl-2-imidazolidinone (DMI) destroyed all components regardless of the 

difference in the number of chlorine atoms or isomers of CDDs and CDFs (Oku et al. 1995). The efficiency 

of the methods was evaluated under varying conditions; in the presence and absence of water, at 90 and 

50 EC, for 0.5 and 5 hours. Although the degree of CDD destruction (99.95–99.80%) was less than that for 

CDFs (99.99–99.98%), overall, the investigators considered the DMI reagent to be more useful than the 

polyethylene glycols because of its stability under strongly basic conditions and its efficiency in the 

presence of water (Oku et al. 1995). 

Ruthenium tetroxide treatment can cause oxidative degradation of CDDs. This method can be used for 

detoxification of glassware and artifacts, or for the periodic purging of industrial reactors to counteract the 

accumulation of CDD residues (HSDB 1995). There is no available evidence on the nature of fragments 

formed during oxidation of the CDDs; however, the related chlorophenols undergo extensive 

decomposition to yield chlorine ions and no significant levels of organic products (HSDB 1995). Other 

chemical methods of detoxification include exposure to ultraviolet light or gamma radiation, the use of 

ozone or special chloroiodide compounds, and the use of solvents or adsorbents to concentrate CDDs into 

smaller volumes for final disposal by incineration (HSDB 1995). 

Dougherty et al. (1993) conducted a theoretical analysis of a proposed in situ method for decontaminating 

soil by photodegradation. Up to 86% of TCDD in the soil can be degraded by this process (Zhong et al. 

1993). Because of its extremely low water solubility and volatility, TCDD is a very persistent soil 

contaminant. With the method, based on the physical properties that facilitate photolysis of TCDD by 
CDDs 375 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

sunlight, an organic solvent mixture (2:1 w/w) of tetradecane and 1-butanol is applied to the contaminated 

soil (Dougherty et al. 1993). The controlling factors in TCDD photodegradation are desorption of the 

compound from the soil, the transport mechanism to the soil surface, and the availability of sunlight. As the 

solvents remove the tightly bound TCDD from the soil, convective upward movements of the compound 

are caused by the evaporation of the solvent (Dougherty et al. 1993; Zhong et al. 1993). The effectiveness 

of the process also depends on a balance between the convective movement and sunlight availability for 

degradation (Dougherty et al. 1993). Modeling conducted by Zhong et al. (1993) identified and quantified 

the controlling factors governing the TCDD photodegradation process. Following the concentration 

variation of TCDD in the top 2 mm of soil through sunlight/night cycles over an exposure period of 

15 days, the model showed that during the daytime of the first few days, there is little accumulation of 

TCDD as the losses due to photodegradation were almost equal to the convective flux in magnitude but 

with different signs. Although the losses due to photodegradation drop to zero at night, the convective flux 

effected a build-up of TCDD. The losses due to photodegradation held steady while the convective 

movements decreased as evaporation slowed down (Zhong et al. 1993). A balance between the build-up of 

TCDD concentration at night and the drop in concentration during the day did not occur until the eleventh 

day of exposure (Zhong et al. 1993). 

Hilarides et al. (1994) investigated degradation of TCDD in the presence of surfactants. Their results 

indicated that radiolytic destruction of TCDD using γ radiation can be achieved. Greater than 92% of the 

TCDD was destroyed in soils amended with 100 ppb TCDD, 25% water, and 2% nonionic surfactant using 

60Co at high radiation doses (800 kGy or 80 Mrad). The use of 60Co as a source avoids the temperature 

increases and power requirements of other sources of ionizing radiation such as an electron beam. It is also 

better suited for soil application because of its greater penetration depths (Hilarides et al. 1994). 

Biotreatment systems which use microorganisms for degradation of refractory organopollutants, like CDDs, 

are also being considered. Phanerochaete chrysosporium, a white rot fungus, has shown the ability to 

slowly degrade 2,3,7,8-TCDD in the laboratory (Bumpus et al. 1985; Des Rosiers 1986). The ability of this 

fungus to metabolize 2,3,7,8-TCDD is thought to be related to its extracellular lignin degrading enzyme 

system (Bumpus et al. 1985; Des Rosier 1986). 

Other proposed methods of disposal are burial in salt mines and inclusion of these chemicals with nuclear 

fission by-products in secured cavities (HSDB 1995). 
.
 
CDDs 377 

5. POTENTIAL FOR HUMAN EXPOSURE 

5.1 OVERVIEW 

Chlorinated dioxins (CDDs) are a family of compounds that includes some extremely toxic and potent 

congeners. The two most toxic of the CDDs in mammals are 2,3,7,8-TCDD and 1,2,3,7,8-PeCDD (Buser 

1987; Poland and Knutson 1982; Safe 1986; WHO 1997). In general, the more toxic congeners to mammals 

appear to be the 2,3,7,8-substituted tetra-, penta-, and hexachloro- compounds, (e.g., 2,3,7,8-TCDD, 

1,2,3,7,8-PeCDD, 1,2,3,4,7,8-HxCDD, 1,2,3,6,7,8-HxCDD, and 1,2,3,7,8,9-HxCDD) (Poland and Knutson 

1982; Safe 1986; WHO 1997). A more detailed discussion of the relative toxicities of the different CDD 

congeners is given in Section 2.5, Relevance to Public Health. 

CDDs usually occur in the environment concurrently with other chemicals such as chlorinated dibenzofurans (CDFs). CDDs and CDFs are highly persistent compounds and have been detected in air, water, 

soil, sediments, animals, and foods. CDFs include 135 congeners, which are structurally similar to CDDs 

and which elicit a number of similar toxicological and biochemical responses in animals (for more 

information on CDFs see ATSDR 1994). CDDs and CDFs are released to the environment during 

combustion processes (e.g., municipal solid waste, medical waste, and industrial hazardous waste 

incineration, and fossil fuel and wood combustion); during the production, use, and disposal of certain 

chemicals (e.g., PCBs, chlorinated benzenes, chlorinated pesticides); during the production of bleached 

pulp by pulp and paper mills; and during the production and recycling of several metals (Buser et al. 

1985; Czuczwa and Hites 1986a, 1986b; Oehme et al. 1987, 1989; Zook and Rappe 1994). The EPA 

has developed procedures for estimating risks associated with exposures to mixtures of CDDs and CDFs 

in environmental matrices (EPA 1989e). This approach is based on the assignment of 2,3,7,8-TCDD 

toxic equivalence factors (TEFs) to CDD/CDF congeners or homologues in complex mixtures. The 

rationale behind the use of TEFs is explained in Section 2.5, Relevance to Public Health. Although the 

focus of this profile is CDDs, it should be recognized that most exposure scenarios involve exposure 

to CDDs, CDFs, and the non-ortho polychlorinated biphenyls (PCBs) that have CDD-like toxicity; 

many of these exposure scenarios are discussed in this chapter. These exposures are usually reported 

in TEQs (for more information see Section 2.5, Relevance to Public Health, Toxic Equivalency Factors 

[TEFs] and Toxic Equivalents [TEQs]). Over the past several years sets of TEFs have been 
CDDs 378 

5. POTENTIAL FOR HUMAN EXPOSURE 

developed, varying slightly from one to another. The reader is encouraged to consult the original literature 

for specific details on TEQs computation. 

CDDs (TCDD, PeCDD, HxCDD, HpCDD, OCDD) are ubiquitous in the environment (Podoll et al. 1986). 

Although all of the sources or processes that contribute to CDDs in the environment have not been 

identified, CDDs are known to be formed in the manufacture of chlorinated intermediates and pesticides, 

during smelting of metals (EPA 1998j), in the incineration of municipal, medical, and industrial wastes 

(Podoll et al. 1986), and from the production of bleached wood pulp and paper (Fletcher and McKay 1993). 

CDDs are also found in emissions from the combustion of various other sources, including coal-fired or oilfired power plants, wood burning, and home heating systems (Chiu et al. 1983; Czuczwa and Hites 1984; 

EPA 1998j; Gizzi et al. 1982; Thoma 1988). Generally, the more highly chlorinated CDDs are the most 

abundant congeners present in the emissions from these combustion sources. CDDs also occur in other 

combustion products (e.g., cigarette smoke) (Bumb et al. 1980; Lofroth and Zebuhr 1992; Muto and 

Takizawa 1989), automobile exhaust from cars running on leaded gasoline with chlorine scavengers and to 

a lesser extent from cars running on unleaded gasoline (Bingham et al. 1989; Marklund et al. 1987, 1990), 

and diesel exhaust (Jones 1995; Cirnies-Ross et al. 1996). CDDs/CDFs can form during the synthesis and 

combustion of chlorine-containing materials, such as polyvinylchloride (PVC), in the presence of naturally 

occurring phenols, vegetation treated with phenoxy acetic acid herbicides, paper and wood treated with 

chlorophenols, and pesticide-treated wastes (Arthur and Frea 1989). 

CDDs occur as contaminants in the manufacture of various pesticides and, as a result, have been released 

to the environment during use of these pesticides. 2,3,7,8-TCDD is a by-product formed in the manufacture of 2,4,5-trichlorophenol (2,4,5-TCP) (Arthur and Frea 1989). 2,4,5-TCP was used to produce the 

bactericide, hexachlorophene, and the chlorophenoxyherbicide, 2,4,5-trichlorophenoxy acid (2,4,5-T). 

Trichlorophenol-based herbicides have been used extensively for weed control on crops, rangelands, 

roadways, right-of-ways, etc. Various formulations of 2,4-dichlorophenoxy acetic acid (2,4-D) 

contaminated mainly with higher chlorinated CDDs/CDFs and 2,4,5-T contaminated mainly with 

2,3,7,8-TCDD were used extensively for defoliation and crop destruction by the American military during 

the Vietnam War. Although six herbicides were used (Orange, Purple, Pink, Green, White, and Blue), 

herbicide Orange (Agent Orange) was the primary defoliant (Wolf et al. 1985). Hexachlorophene use has 

been restricted by the FDA and its disposal is regulated by EPA under the Resource Conservation and 

Recovery Act (RCRA). In 1983, EPA canceled registration for all chlorophenoxy herbicides used on 

foods, rice paddies, pastures, and rangelands (IARC 1986b). 2,4,5-T can no longer be used legally in the 
CDDs 379 

5. POTENTIAL FOR HUMAN EXPOSURE 

United States for any purpose (IARC 1986b). Other countries, including Canada, Sweden, the Netherlands, 

Australia, Italy, and the Federal Republic of Germany, have also canceled registrations for 2,4,5-T (IARC 

1986b), but many other countries have not. Currently, 2,4,5-T can be produced with lower 2,3,7,8-TCDD 

concentrations than were previously possible. 2,4,5-TCP production has been discontinued in many 

countries, including the United States, Canada, the United Kingdom, the Federal Republic of Germany, 

and Austria (IARC 1986a). HxCDD, HpCDD, and OCDD are known contaminants of pentachlorophenol 

(PCP), primarily a wood preservative and pesticide, which was used extensively in the 1970s and is still 

used today (to a lesser extent) in the lumber industry. PCP is currently registered as a restricted-use 

pesticide in the United States (Sine 1990). 

Although little definitive data exist to prove or disprove that CDDs form during natural processes, results 

from dated sediment cores have shown that there were significant increases in CDDs and CDFs after about 

1940 (Czuczwa and Hites 1984, 1986b, 1986b) and lower levels of CDDs are currently found in persons 

from less industrialized countries (Schecter et al.1991a). The congener/homologue profile of the 

sediments was similar to that of atmospheric samples, strongly suggesting that combustion processes were 

the source of CDDs in the sediments. The historical increase in CDDs/CDFs also was similar to the trends 

for the production, use, and disposal of chlorinated organics, suggesting that accumulation of these 

compounds in the environment is a recent phenomenon related to the production, use, and subsequent 

incineration of chlorinated organic chemicals (Schecter et al. 1988). 

CDDs are ubiquitous in the environment and are found at low background levels (parts per trillion [ppt] or 

parts per quadrillion [ppq]) in the air, water, and soil. Lower levels are found in biological and environmental samples from less industrialized rural regions than in those from more industrialized urban regions 

(Czuczwa and Hites 1986a; Des Rosiers 1987; Edgerton et al. 1989; Schecter et al. 1989e, 1989g, 1991a, 

1994d; Tiernan et al. 1989b). HpCDD and OCDD are the most common CDDs found in environmental 

samples (Christmann et al. 1989b; Clement et al. 1985, 1989; Pereira et al. 1985; Reed et al. 1990; Tashiro 

et al. 1989a; Tiernan et al. 1989b). 

The environmental fate and transport of CDDs involve volatilization, long-range transport, wet and dry 

deposition, photolysis, bioaccumulation, and biodegradation (Kieatiwong et al. 1990). CDDs strongly 

partition to soils and sediments. Due to their low vapor pressure and low aqueous solubility and their 

strong sorption to particulates, CDDs are generally immobile in soils and sediments. Although most 

biological and nonbiological transformation processes are slow, photolysis has been shown to be relatively 
CDDs 380 

5. POTENTIAL FOR HUMAN EXPOSURE 

rapid. Photolysis is probably the most important transformation process in environmental systems into 

which sunlight can penetrate (Kieatiwong et al. 1990). Estimates of the half-life of 2,3,7,8-TCDD on the 

soil surface range from 9 to 15 years, whereas the half-life in subsurface soil may range from 25 to 

100 years (Paustenbach et al. 1992). CDDs have been shown to bioaccumulate in both aquatic and 

terrestrial biota. CDDs have a high affinity for lipids and, thus, will bioaccumulate to a greater extent in 

organisms with a high fat content. 

Over the past decade, typical concentrations of CDDs in urban air in the United States have averaged 

2.3 pg/m3

, with OCDD and HpCDD homologues predominating and 2,3,7,8-TCDD being the least 

common congener (Smith et al. 1992). CDD concentrations range as follows: OCDD, 0.44–3.16 pg/m3

; 

HpCDD, 0.21–4.4 pg/m3

; HxCDD, 0.6–0.63 pg/m3; PeCDD, not detected to 0.1 pg/m3; and 2,3,7,8-TCDD, 

<0.04–0.18 pg/m3

 (Edgerton et al. 1989; Eitzer and Hites 1989a, 1989b; Hunt and Maisel 1992; Smith et 

al. 1992). Although 2,3,7,8-TCDD has been detected in some urban air, it is rarely detected in rural areas 

(Reed et al. 1990). Ambient air concentrations of 2,3,7,8-TCDD detected in the vicinity of a Superfund 

clean-up site were on the order of 1 pg/m3

 (Fairless et al. 1987). CDDs have been detected almost 

exclusively in raw surface waters, rather than in finished drinking water (Jobb et al. 1990). This is not 

unexpected because CDDs are hydrophobic, and the compounds tend to be adsorbed onto particulate 

matter in the water column. Conventional water treatment processes appear to be effective in removing the 

CDDs along with the particulates (Jobb et al. 1990; Meyer et al. 1989). OCDD is the congener most often 

detected in water supplies at concentrations ranging from 9 to 175 ppq (raw water) and from 19 to 46 ppq 

(finished water). 2,3,7,8-TCDD concentrations have not been detected in finished drinking water, but 

were detected in one raw sample at a concentration of 1.7 ppq (Meyer et al. 1989). Concentrations of 

2,3,7,8-TCDD in most soils are <12 ppt (Des Rosiers 1987; Nestrick et al. 1986); however, levels in 

contaminated soils can be several orders of magnitude higher (1,750 ppb) (Tiernan et al. 1985). 

2,3,7,8-TCDD and other CDDs, have also been detected in measurable amounts in the sediments of 

industrialized waterbodies throughout the United States (Bopp et al. 1991; Wenning and Erickson 1994; 

Wenning et al. 1992, 1993a, 1993b). 

In the National Study of Chemical Residues in Fish conducted by the EPA between 1986 and 1989, four 

CDDs (2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 1,2,3,6,7,8-HxCDD, and 1,2,3,4,6,7,8-HpCDD) were detected at 

over 50% (54 to 89%) of 388 sites surveyed nationwide (EPA 1992). The most frequently detected CDD 

compound (1,2,3,4,6,7,8,-HpCDD) was found in fish tissues at 89% of the sites. This compound was also 

detected at the highest concentrations of 249 ppt (mean 10.52 ppt) wet weight. 2,3,7,8-TCDD 
CDDs 381 

5. POTENTIAL FOR HUMAN EXPOSURE 

and 1,2,3,7,8-PeCDD, the CDDs currently believed to be most toxic to vertebrates (WHO 1997), were 

found in fish tissue at 70% and 54% of the sites, respectively. 2,3,7,8-TCDD was found at a mean 

concentration of 6.9 ppt and a maximum concentration of 204 ppt, and 1,2,3,7,8-PeCDD was found at a 

mean concentration of 2.38 ppt and a maximum concentration of 54 ppt. With respect to source 

categories, fish collected near pulp and paper mills using chlorine had the highest median 2,3,7,8-TCDD 

concentration (5.66 ppt), compared to the second highest median 2,3,7,8-TCDD concentrations of 1.82 ppt 

at refinery/other industrial sites, and the third highest median 2,3,7,8- TCDD concentration of 1.27 ppt at 

Superfund sites. Similarly, with respect to source categories, fish collected near pulp and paper mills using 

chlorine had the highest median 1,2,3,7,8-PeCDD concentration (1.52 ppt), compared to the second 

highest median concentrations of 1.35 ppt at refinery/other industrial sites, and the third highest median 

concentration of 1.09 ppt at industrial/urban sites. 

The detection of CDDs in blood, adipose tissue, breast milk, and other tissue samples from the general 

population indicates universal exposure to CDDs from environmental sources (Fürst et al. 1994; Orban et 

al. 1994; Patterson et al. 1986a; Ryan et al. 1986, 1993a; Schecter and Gasiewicz 1987a, 1987b; Schecter 

et al. 1986b, 1989e; Stanley 1986; Stanley et al. 1986). The general population is exposed to CDDs 

released from industrial and municipal incineration processes; exhausts from automobiles using leaded 

gasoline; cigarette smoke; and foods, including human milk (Pohl and Hibbs 1996; Schecter et al. 1994e). 

The major source (>90%) of exposure for the general population, however, is primarily associated with 

meat, dairy products, and fish (Beck et al. 1989a; Schaum et al. 1994; Schecter et al. 1994d, 1994e, 

1996a). CDDs are transferred through the placenta to the fetus, by breast milk to infants and young 

children, and by lifelong dietary ingestion. Workers involved with incineration operations or those who 

have been or may be involved in the production, use, or disposal of trichlorophenol, phenoxyherbicides, 

hexachlorophene, pentachlorophenol and other compounds that contain impurities of CDDs are at a greater 

risk from exposure to CDDs and TEQs (Päpke et al. 1992; Schecter and Ryan 1988; Schecter et al. 1991). 

Individuals in the general population who may be exposed to potentially higher levels of CDDs include 

recreational and subsistence fishers (including many native Americans) and their families living in CDDcontaminated areas who consume large quantities of fish from contaminated waters (CRITFC 1994; Ebert 

et al. 1996), subsistence hunters such as the Inuit of Alaska who consume large quantities of wild game 

(particularly marine mammals) (Dewailly et al. 1993; Hebert et al. 1996; Norstrom et al. 1990), 

subsistence farmers and their families living in areas contaminated with CDDs who consume their own 

farm-raised beef and dairy products (EPA 1996b; McLachlan et al. 1994), individuals who live in the 

vicinity of an industrial or municipal incinerator, or individuals who live in the vicinity of the 
CDDs 382 

5. POTENTIAL FOR HUMAN EXPOSURE 

126 hazardous waste sites where CDDs (and more especially where 2,3,7,8-substituted CDDs) have been 

detected (Gough 1991; Liem et al. 1991; Pohl et al. 1995; Riss et al. 1990; Wuthe et al. 1993). 

2,3,7,8-TCDD has been identified in at least 91 of 1,467 current or former EPA National Priorities List 

(NPL) hazardous waste sites (HazDat 1998). However, the number of sites evaluated for 2,3,7,8,-TCDD is 

not known. The frequency of these sites within the United States can be seen in Figure 5-1. Of these sites, 

90 are located in the United States and 1 is located in the Commonwealth of Puerto Rico (not shown). 

Total CDDs (including TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD) have been identified in 126, 

105, 34, 43, 49, and 53 sites, respectively, of the 1,467 hazardous waste sites on the NPL. The frequency 

of these sites within the United States for total CDDS, TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD, 

respectively, can be seen in Figures 5-2 through 5-7. Of the 126 sites with total CDD detections, 125 are 

located in the United States and 1 site is located in the Commonwealth of Puerto Rico (not shown). Of the 

105 sites with total TCDD detections, 104 are located in the United States and 1 site is located in the 

Commonwealth of Puerto Rico (not shown). Of the sites with PeCDD, HxCDD, HpCDD, and OCDD 

detections, all 34, 43, 49, and 53 sites, respectively, are located in the United States. 

5.2 RELEASES TO THE ENVIRONMENT 

CDDs have been measured in all environmental media including ambient air, surface water, groundwater, 

soil, and sediment. While the manufacture and use of chlorinated compounds, such as chlorophenols and 

chlorinated phenoxy herbicides, were important sources of CDDs to the environment in the past, the 

restricted manufacture of many of these compounds has substantially reduced their current contribution to 

environmental releases. It is now believed that incineration/combustion processes are the most important 

sources of CDDs to the environment (Zook and Rappe 1994). Important incineration/combustion sources 

include: medical waste, municipal solid waste, hazardous waste, and sewage sludge incineration; industrial 

coal, oil, and wood burning; secondary metal smelting, cement kilns, diesel fuel combustion, and 

residential oil and wood burning (Clement et al. 1985; Thoma 1988; Zook and Rappe 1994). 

Emissions to the atmosphere from incineration and combustion sources result in the wide-spread 

distribution of CDDs. Consequently, CDDs are found at low levels in rural soils as well as in sediments of 

otherwise pristine waterbodies. Much of the CDD deposits from wet and dry deposition ultimately 

become components of urban runoff which enter rivers, streams, and estuaries directly or through 

stormwater outfalls and combined sewer overflows (CSOs). In a recent study, Huntley et al. (1997) used 

statistical 







 

CDDs 390 

5. POTENTIAL FOR HUMAN EXPOSURE 

pattern matching techniques (principal components analysis) to evaluate CDD congener patterns in 

sediment samples collected adjacent to several CSOs. According to these authors, the presence of these 

unique CDD/CDF congener patterns in sediment adjacent to CSOs suggested that these CSOs were a likely 

source given the industrial, residential, and stormwater inputs to the combined sewer overflow system. 

Such statistical techniques have been applied elsewhere to CDD congener pattern matching in an effort to 

identify specific sources of CDDs. Wenning et al. (1993a, 1993b) also applied principal components 

analysis to Newark Bay Estuary sediments and found that most of the congener fingerprint patterns were 

related to combustion/incineration sources. More recently, Ehrlich et al. (1994) applied polytopic vector 

analysis, a fingerprinting technique that “unmixes” the CDD/CDF patterns, and concluded that the primary 

sources of CDD/CDFs in Newark Bay Estuary sediments were combustion/incineration, sewage-related 

sources, and PCB-related sources. Statistical techniques that have proven useful for identifying sources of 

CDDs have recently been reviewed (Wenning and Erickson 1994). Future efforts to reduce the release of 

CDDs to the environment will require additional analysis of the distributional patterns of CDDs in 

environmental media, which may also provide information on sources still to be identified. 

5.2.1 Air 

The key sources of CDD releases to air are from anthropogenic combustion processes and the production 

and use of chemicals contaminated with CDDs. Some evidence suggests that natural combustion 

processes (e.g., forest fires or volcanic activity) may also be sources of CDDs, but to a much smaller 

extent. Toxics Release Inventory (TRI) data are not available for CDDs since CDD releases are not 

required to be reported (EPA 1995g). 

Combustion Processes. Combustion processes generate CDDs, CDFs, and other halogenated 

aromatic compounds (Czuczwa and Hites 1984, 1986a, 1986b). Most of the direct releases of CDDs and 

CDFs from combustion processes are to the air (Czuczwa and Hites 1984, 1986a, 1986bc). CDDs and 

CDFs may be found in particulates released from the combustion of most types of organic material and 

limited evidence suggests that they may also result from trace chemical reactions in fire (Bumb et al. 1980; 

Crummett 1982; Safe 1990). The processes involved in the formation of CDDs and CDFs consist of 

numerous chemical reactions that occur during combustion of organic compounds in the presence of 

chlorinated material. The EPA has recently identified stationary source categories that release 

2,3,7,8-TCDD TEQ to the atmosphere (EPA 1998j). The percentage contribution of the five highest 

source categories are: 68% from municipal waste incineration, 12.3% from medical waste incineration, 
CDDs 391 

5. POTENTIAL FOR HUMAN EXPOSURE 

8.9% from Portland cement manufacture hazardous waste kilns, 3.5% from secondary aluminum smelting, 

and 3.0% from other biological incineration. These five source categories account for 95.9% of all 

stationary emissions of 2,3,7,8-TCDD TEQ to the air. 

The "Trace Chemistries of Fire Hypothesis" suggests that CDDs and CDFs can also form during a variety 

of combustion processes including natural ones, such as forest fires and volcanic eruptions (Crummett 

1982). However, there is very limited evidence suggesting that such natural processes could be minor 

sources of these compounds in the environment. Only data from one study were found that directly 

measured CDD/CDFs in actual emissions from forest fires. Tashiro et al. (1990) detected the concentration of total CDD/CDFs in air ranging from 15 to 400 pg/m3

. The samples were collected from fixed 

collectors 10 m above the ground and from aircraft flying through the smoke. Soil samples collected 

before the burn detected 43 ppt of OCDD in 1 of 4 samples tested. After the burn, OCDD was detected in 

3 of 4 soil samples at concentrations of 46, 100, and 270 ppt. Because the small sample size precluded 

statistical analysis, no further conclusions were drawn by the authors. Thomas and Spiro (1995), however, 

estimated that forest and agricultural burning accounted for the third largest emission of CDD/CDF in the 

United States (30 kg/year), behind municipal waste incineration (200 kg/year) and hospital incinerators 

(40 kg/year) although the inclusion of agricultural burning, which may include acreage treated with longlived organochlorine pesticides, may skew the values higher than would be expected from forest fires 

alone. Failure to find CDDs in ancient mummies or ancient frozen Eskimo tissues is another indication 

that the “Trace Chemistries of Fire Hypothesis” may have little bearing on human exposure (Ligon et al. 

1989; Schecter et al. 1988; Tong et al. 1990). The EPA recently found elevated levels of 2,3,7,8-TCDD in 

two chickens that were traced to clay (used as an anti-caking additive in soybean animal meal) derived 

from clay deposits mined at the Kentucky-Tennesse Ball Clay Company in Crenshaw, Mississippi. 

(Chemical Regulation Reporter 1997a, 1997b). However, no information on the origin of the 2,3,7,8­

TCDD, either natural or anthropogenic, was presented. 

The issue of natural sources of CDD/CDF is interesting, but historical deposition records strongly 

implicate anthropogenic activity as the major source of CDD/CDFs (Thomas and Spiro 1996). These 

authors further suggest that the historic record on CDD/CDF deposition provided by sediment cores 

strongly implies that anthropogenic sources have been overwhelmingly dominant. Sediment cores from 

Siskwit Lake on a remote island in northern Lake Superior, provide a historic record of atmospheric CDD 

fluxes (Czuczwa and Hites 1986a). An 8-fold increase in the CDD/CDF deposition rate (from approx 

imately 4–30 pg/cm2

/year) occurred between 1940 and 1970, corresponding to a great expansion in the 
CDDs 392 

5. POTENTIAL FOR HUMAN EXPOSURE 

industrial use of chlorine (Thomas and Spiro 1996). The decrease in deposition rate of about 30% (from 

30 to 24 pg/cm2 /year) from 1970 to the mid 1980s parallels decreased production and use of chlorophenols 

(pesticide registrations for 2,4,5-T and Silvex were discontinued in 1983 and 1984, respectively) (IARC 

1977; Sine 1990) and reductions in municipal incinerator emission resulting from improvements in design, 

pollution controls, and operation of these facilities (Thomas and Spiro 1996). It is difficult to reconcile 

these trends with predominantly natural sources, especially since the total area of U.S. forests consumed by 

forest fires diminished by more than a factor of 4 between 1940 and 1970 through more effective fire 

control (Thomas and Spiro 1996). 

Although the production of CDDs during combustion processes are highlighted here, most samples from 

combustion sources show a complex mixture of isomers and congeners of CDDs and CDFs which vary in 

their relative concentrations (Kolenda et al. 1994; Nestrick and Lamparski 1983; Vikelsoe et al. 1994). 

CDDs have been detected in emissions (flue gas and fly ash) from municipal, hazardous waste, and 

industrial incinerators (Buser 1987; Oppelt 1991; Sedman and Esparza 1991; Schecter 1983). Combustion 

of materials, such as vegetation treated with phenoxy acetic acid herbicides, paper and wood treated with 

chlorophenols, pesticide-treated wastes, and polyvinylchloride (PVC) in the presence of naturally 

occurring phenols, may lead to CDDs and CDD precursors (Arthur and Frea 1989). PVC is known to 

yield a small amount of chlorobenzene upon pyrolysis, which in turn thermally decomposes to CDDs and 

CDFs (Lustenhouwer et al. 1980). CDDs have also been detected in fly ash from an oil-fired power plant, 

in city dust, in commercial sludge fertilizer, in particulate deposits in car and truck mufflers, in exhaust 

from vehicles powered with leaded and unleaded gasoline and diesel fuel, in cigarette smoke, and in soot 

from home fireplaces and from PCB and chlorinated benzene contaminated transformer fires (Bumb et al. 

1980; Hutzinger et al. 1985; Lofroth and Zebuhr 1992; Marklund et al. 1987, 1990; Muto and Takizawa 

1989; Schecter 1983; Thoma 1988). Dichloroethane, the chlorinated additive in leaded gasoline, is also a 

source of CDDs (Marklund et al. 1987). The dichloroethane acts as a scavenger to prevent the deposition 

of lead compounds in engines (Safe 1990). Although the data indicate that CDDs result from diverse 

processes, the relative contributions of these sources and other unidentified sources to the presence of 

CDDs in the atmosphere are not known. 

A mixture of CDDs (TCDD, PeCDD, HxCDD, HpCDD, OCDD) has been found in emissions (both 

particles and flue gases) from various combustion sources, including municipal incinerators, power plants, 

wood burning, home heating systems, and petroleum refining (Chiu et al. 1983; Czuczwa and Hites 1984; 

Gizzi et al. 1982; Nessel et al. 1991; Thoma 1988; Thompson et al. 1990). In individual samples of 
CDDs 393 

5. POTENTIAL FOR HUMAN EXPOSURE 

emissions from an urban incinerator, HxCDDs and OCDD were often the most abundant CDDs found, 

although the homologue pattern can be quite variable (Gizzi et al. 1982). Emission of TCDD from 

municipal waste combustion ranged from 0.018 ng/m3

 to 62.5 ng/m3 depending on the type of combustion 

facility (Roffman and Roffman 1991). A municipal solid waste incinerator sampled in 1988 contained an 

average TCDD concentration of 0.0012 ng/m3

, where OCDD was present at 1.2 ng/m3, and HxCDD was 

present at >1 ng/m3

 (Nessel et al. 1991). In another study, no TCDDs were found in emissions from 

hazardous waste or municipal waste incinerators; the levels of PeCDD found in the emissions from 

municipal waste incinerators were three orders of magnitude higher than from hazardous waste 

incinerators (Oppelt 1991). Fly ash from a municipal incinerator and from coal-fired power plants was 

analyzed to study the CDD congener distributions typical of combustion samples (Czuczwa and Hites 

1984). OCDD was the most abundant CDD in all fly ash samples. Coal fly ash samples differed 

significantly from municipal incinerator fly ash samples. Although some CDDs were detected in coal fly 

ash, no TCDDs or PeCDDs were detected. CDDs were present in much lower concentrations in fly ash 

from coal-fired power plants than in fly ash from a municipal incinerator. The levels of OCDD in the coal 

fly ash samples (2.2 ppb and 3.8 ppb) were at least 100 times lower than those found in the municipal 

incinerator fly ash (400 ppb). No isomers of TCDD were detected in municipal incinerator fly ash samples 

with a detection limit of 100 ppt (Czuczwa and Hites 1984). 

CDDs have been detected in chimney soot samples from various home heating systems using unleaded 

heating oil, coal, and wood in Germany (Thoma 1988). A Canadian study of wood-burning stoves 

detected only OCDD in particulates from the stack emissions (Wang et al. 1983). Open-air burning of 

PCP-treated wood produced levels of CDDs ranging from 2 ppb (TCDD) to 187 ppb (OCDD) (Chiu et al. 

1983). Combustion of untreated wood also produces CDDs (TCDD, PeCDD, HxCDD, HpCDD, OCDD) 

(Clement et al. 1985). Samples of bottom ash and chimney ash from 2 wood-burning stoves, 1 open 

fireplace, and outdoor open-air burning had detectable levels of CDDs ranging from 0.3 to 33 ppb. For 

each homologous class, the total concentrations ranged from not detectable to 11 ppb. Detection limits 

were equal to 10 ppt for TCDD and PeCDD and 50 ppt for HxCDD, HpCDD, and OCDD. The open-air 

burning ash produced the highest total CDD concentration of 33 ppb, with HpCDD (11 ppb) and OCDD 

(10 ppb) being the most abundant (Clement et al. 1985). 

Fires involving capacitors or transformers containing chlorobenzene and PCBs are also sources of CDDs 

and CDFs. For example, in the transformer fire in the New York State Office Building in Binghamton, 

NY, TCDD, PeCDD, HxCDD, HpCDD, and OCDD were found in soot samples at levels ranging from 
 

CDDs 394 

5. POTENTIAL FOR HUMAN EXPOSURE 

<2 ppm (PeCDD) to 7 ppm (HpCDD) (Tiernan et al. 1985). CDFs were more abundant and were detected 

at higher concentrations ranging from 28 to 1,920 ppm in soot. 

Recently, Wikstrom et al. (1996) studied formation of CDDs, CDFs, and chlorobenzenes in the 

combustion process. These authors monitored combustion of an artificial fuel where the chlorine level and 

source were varied in the artificial waste. Different levels of organic chlorine (PVC) and inorganic 

chlorine (CaCl2@6H20) were added to the fuel. When the level of chlorine in the fuel was <1%, there was 

no correlation between the quantities of CDDs, CDFs, and chlorobenzenes present. However, when the 

chlorine level was >1%, increased formation rates were noted for CDDs, CDFs, and chlorobenzenes. 

Production and Use of Contaminated Chemicals and Certain Herbicides. CDDs are known 

trace contaminants of certain chlorinated industrial chemicals like chlorophenols (Buser 1987). CDDs can 

inadvertently form as by-products during the manufacture of chlorophenols. Since the 1930s, PCP and the 

tri- and tetrachlorophenols have gained recognition as fungicides, herbicides, insecticides, and precursors 

in the synthesis of other pesticides. 

PCP was developed primarily for use as a wood preservative but has also been used as an herbicide on 

pineapple and sugarcane plantations. It has also been employed as a molluscicide against schistosomiasis, 

a severe human parasitic disease prevalent in much of tropical Asia, Africa, and South America (Hutzinger 

et al. 1985). The major contaminant of commercial PCP is OCDD, which may be present at concentrations 

between 500 and 1,500 mg/kg (ppm) (Dobbs and Grant 1979; Miller et al. 1989a). PCP may also contain 

mixed isomers of HxCDD and HpCDD (Pereira et al. 1985). It is currently registered as a restricted-use 

pesticide for use as a wood preservative (Sine 1990). 

2,3,7,8-TCDD forms during the manufacture of 2,4,5-TCP. 2,4,5-TCP has been used in cooling towers 

and in paper, pulp, and leather processing (Hutzinger et al. 1985). 2,4,5-TCP was used to produce the 

bactericide hexachlorophene and phenoxy-herbicides like 2,4,5-trichlorophenoxy acids (2,4,5-T). 2,4,5-T, 

in turn, was used in the production of a wide variety of herbicides including Silvex (2-[2,4,5-trichlorophenoxy]propionic acid) and Agent Orange (Hutzinger et al. 1985). Hexachlorophene, which is currently 

under EPA suspension, is reported to contain <15 μg/kg (ppb) 2,3,7,8-TCDD (IARC 1977; Sine 1990). 

2,3,7,8-TCDD is an unwanted by-product formed during the production of hexachlorophene (Freeman et 

al. 1986). The 2,3,7,8-TCDD produced is primarily contained in still-bottom waste (waste oils) remaining 

after hexachlorophene is purified (Freeman et al. 1986). Still-bottom waste and other oils were used in the 
CDDs 395 

5. POTENTIAL FOR HUMAN EXPOSURE 

early 1970s for dust control on roads, parking lots, horse arenas, and other sites around Missouri (Freeman 

et al. 1986). The herbicide 2,4,5-T produced commercially prior to 1965 contained up to 30 mg/kg (ppm) 

or more 2,3,7,8-TCDD (IARC 1977). The level of 2,3,7,8-TCDD in commercial 2,4,5-T was reduced to 

<0.05 mg/kg (ppm), and most of the commercial 2,4,5-T available before its registration was discontinued 

in the United States in 1983 contained <0.02 mg/kg (ppm) 2,3,7,8-TCDD (IARC 1977; Sine 1990). 

Chlorophenoxy herbicides, such as 2,4-D, are typically formulated as esters or amine salt derivatives 

(IARC 1986b). Of 16 samples of 2,4-D formulations from Canada analyzed for CDDs in the early 1980s, 

8 of 9 ester formulations and 4 of 7 amine salt formulations were contaminated (IARC 1986b). The 2,4-D 

ester formulations contained 0.2–1.8 mg/kg (ppm) 1,3,6,8-TCDD (the only TCDD isomer detected), while 

the 2,4-D amine salt formulations contained 0.02–0.3 mg/kg (ppm) 1,3,6,8-TCDD (IARC 1986b). It 

should be noted that 1,3,6,8-TCDD is not one of the toxic CDDs with respect to mammals; however, 

2,3,7,8-substituted CDDs/CDFs have been reported in 2,4-D from Russia (Schecter et al. 1993). 

Agricultural and wartime uses of trichlorophenol-based herbicides such as 2,4,5-T and Silvex also have 

resulted in release of 2,3,7,8-TCDD at low concentrations in many countries (EPA 1987k). 2,4,5-T was 

used in aerial spraying operations for weed control on crops, along fence rows, ditch banks, farm roadways, pastures, and rangeland (Bovey 1980). Non-farm uses of 2,4,5-T included tree and bush control on 

rights-of-way, roadways, fire lanes, and railroads (Bovey 1980). Agent Orange, used as a defoliant in the 

Vietnam War from 1962 to 1970, was contaminated with an average of 2 ppm of 2,3,7,8-TCDD (Czuczwa 

and Hites 1986a, 1986b; Wolfe et al. 1985). An estimated 10–11 million gallons were applied in South 

Vietnam (EPA 1987k; Wolfe et al. 1985). This volume of Agent Orange contained an estimated 

368 pounds of 2,3,7,8-TCDD (Wolfe et al. 1985). Agent Orange is an equal parts mixture of the butyl 

esters of 2,4,5,-T and 2,4-dichlorophenoxyacetic acid (2,4-D) (Josephson 1983). These herbicides were 

used extensively in silviculture for control of deciduous trees in conifer forests before their use was 

discontinued (EPA 1987k). The use of Silvex, a herbicide closely related to 2,4,5-T, was discontinued in 

the United States in 1984 (Sine 1990). 

Industrial accidents have also released high levels of CDDs into the air. In 1976, at least 1.3 kg 

(2.87 pounds) of 2,3,7,8-TCDD was released into the air as a result of an industrial accident at the 

ICMESA chemical plant near Seveso, Italy, that was involved in 2,4,5-TCP synthesis (Cerlisi et al. 1989; 

Mocarelli et al. 1991). The 2,3,7,8-TCDD release contaminated a populated area of about 2.8 km2 

(1.08 mi2) (Mocarelli et al. 1991). 
CDDs 396 

5. POTENTIAL FOR HUMAN EXPOSURE 

2,3,7,8-TCDD has been detected in air samples collected at 9 of the 91 NPL hazardous waste sites where it 

has been detected in some environmental media (HazDat 1998). Total CDDs have been detected in air 

samples collected at 10 of the 126 NPL sites where they have been detected in some environmental media. 

Total TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD have been detected in air samples at 10, 3, 3, 3, 

and 1 sites of the 105, 34, 43, 49, and 53 sites, respectively, where they have been detected in some 

environmental media (see Table 5-1). 

5.2.2 Water 

CDDs can enter water by a number of different mechanisms including urban runoff, combined sewer 

overflows (CSOs), and direct discharge by industrial facilities and publicly-owned treatment works 

(POTWs); deposition of particulates from combustion sources, runoff and drift from the use of 

chlorophenol-based pesticides; and leaching from chlorophenol-containing waste sites (Huntley et al. 

1997; Muir et al. 1986a; Periera et al. 1985; Shear et al. 1996). Direct application or drift of 2,4,5-T or 

Silvex into water has also resulted in release of TCDD to surface water (Norris 1981); however, the 

contribution of CDDs from pesticide drift is now negligible since most CDD-containing pesticides have 

been banned. The migration of chemical wastes containing CDDs from disposal sites has resulted in 

contamination of surface water and groundwater (HazDat 1998). 

CDDs/CDFs, specifically 2,3,7,8-TCDD and 2,3,7,8-TCDF, are also present in effluent and sludges from 

pulp and paper mills that employ the bleached kraft process (Clement et al. 1989; EPA 1991b; Swanson et 

al. 1988). 2,3,7,8-TCDD was detected in 7 of 9 bleached pulps at concentrations ranging from not 

detected (<1 ppt) to 51 ppt (median 4.9 ppt; mean 13 ppt) (Amendola et al. 1989). It was also detected in 

waste waters from 4 of 5 paper mills at levels ranging from not detected (<0.006 ppt) to 3.6 ppt (Amendola 

et al. 1989). 

During 1988, the EPA and the U.S. pulp and paper industry jointly conducted a survey of 104 pulp and 

paper mills in the United States to measure concentrations of CDDs in effluent, sludge, and paper (EPA 

1990d). This study is commonly called the 104-Mill Study and includes all U.S. mills where wood pulps 

are bleached with chlorine or chlorine derivatives. Higher chlorinated CDDs/CDFs are typically found in 

effluent when chlorine dioxide is used, but not when elemental chlorine is used. In 1992, the pulp and 

paper industry conducted its own survey (NCASI 1993). As part of an effort to develop revised effluent 

guidelines and standards for the pulp and paper industry, the EPA recently published the development 

CDDs 398 

5. POTENTIAL FOR HUMAN EXPOSURE 

document for the guidelines and standards being proposed for this industry (EPA 1993a). This 

development document presents estimates of annual discharges of two congeners, 2,3,7,8-TCDD and 

2,3,7,8-TCDF in effluents (from wastewater treatment systems) from this industry as of January 1993. 

The joint EPA/paper industry study of 104 pulp and paper mills provides an estimate of the release of 

2,3,7,8-TCDD and 2,3,7,8-TCDF in bleached pulp, waste water sludge, and waste water effluent from the 

U.S. pulp and paper industry as of mid-to-late 1988 (EPA 1990d). This was a time in the industry’s 

history when only limited use of pulping and bleaching technologies and operating practices that 

demonstrated potential to reduce the formation of TCDDs and TCDFs had been implemented. In this 

study 2,3,7,8-TCDD was detected at 90 and 56% of the kraft and sulfite mills, respectively, that were 

surveyed, and no mill was found to be free of 2,3,7,8-TCDD/TCDF. For bleached pulp, the mean 

2,3,7,8-TCDD concentration was 7.5 ppt (maximum 56 ppt) for kraft hardwoods, 12 ppt (maximum 

116 ppt) for kraft softwoods, 7.1 ppt (maximum 15 ppt) for sulfite hardwoods, and 3.5 ppt (maximum 

3.5 ppt) for sulfite softwoods. Mean waste water effluent concentrations of 2,3,7,8-TCDD were 0.076 ppt 

for kraft mills (maximum 0.64 ppt) and 0.013 ppt (maximum 0.023 ppt) for sulfite mills. Waste water 

sludges contained mean 2,3,7,8-TCDD concentrations of 101 ppt for kraft mills (maximum 1,390 ppt) and 

13 ppt (maximum 58 ppt) for sulfite mills. Furthermore, for all kraft mills, about 38% of the 

2,3,7,8-TCDD was partitioned to pulps, 33% to waste water sludges, and 29% to waste water effluents. 

The NCASI (1993) report found that <10% of pulp and paper mills had 2,3,7,8-TCDD and 2,3,7,8-TCDF 

concentrations in effluent above the detection limits of 10 ppq and 100 ppq, respectively; however, none of 

the more highly congener groups were measured. Similar results were obtained in the short- and long-term 

sampling reported for 18 mills (EPA 1993a). 2,3,7,8-TCDD and 2,3,7,8-TCDF were detected at four and 

nine mills, respectively. Waste water sludges at 75–90% of all mills contained <10 ppt of 2,3,7,8-TCDD 

and <100 ppt of 2,3,7,8-TCDF (NCASI 1993). Similar results were reported in the EPA (1993a) report 

except that 2,3,7,8-TCDD and 2,3,7,8-TCDF were found in sludges in 64 and 85%, respectively, of the 

mills sampled. NCASI (1993) reported that almost 90% of bleached pulps contained <2 ppt of 

2,3,7,8-TCDD and <20 ppt of 2,3,7,8-TCDF. For bleached pulps, the mean 2,3,7,8-TCDD concentration 

was 0.9 ppt (maximum 10 ppt) and the mean 2,3,7,8-TCDF concentration was 6 ppt (maximum 323 ppt). 

The mean waste water effluent concentration of 2,3,7,8-TCDD was 0.006 ppt (maximum 0.08 ppt) and 

0.031 ppt (maximum 0.510 ppt) for 2,3,7,8-TCDF. Waste water sludges contained a mean 2,3,7,8-TCDD 

concentration of 11 ppt (maximum 133 ppt) and 11 ppt (maximum 735 ppt) for 2,3,7,8-TCDF. In this 

study, mean pulp, waste water effluents, and waste water sludge concentrations of 2,3,7,8-TCDD all 
CDDs 399 

5. POTENTIAL FOR HUMAN EXPOSURE 

declined by a factor of about 10 from those cited in the 104 Mill Study (EPA 1990d). Overall, NCASI 

(1993) reports a 90% reduction in TEQs generated by pulp and paper mills from 1988 to 1992 for all 

2,3,7,8-TCDDs and 2,3,7,8-TCDFs. 

2,3,7,8-TCDD has been detected in surface water and groundwater samples collected at 9 and 15 sites of 

the 91 NPL hazardous waste sites where it has been detected in some environmental media (HazDat 1998). 

Total CDDs have been detected in surface and groundwater samples collected at 14 and 32 of the 126 NPL 

sites where they have been detected in some environmental media. Total TCDDs, PeCDDs, HxCDDs, 

HpCDDs, and OCDD have been detected in surface water samples collected at 10, 1, 4, 4, and 6 sites and 

in groundwater samples collected at 21, 3, 10, 14, and 16 of the 105, 34, 43, 49, and 53 NPL sites, 

respectively, where these homologues have been detected in some environmental media (see Table 5-1). 

5.2.3 Soil 

Historically, CDDs have been deposited onto soil through pesticide applications and disposal of CDDcontaminated industrial wastes, and via land application of paper mill sludges (EPA 1991b). Currently, 

however, atmospheric fall-out of CDD-laden particulates and gases appears to be the predominant source 

of CDDs to soil (Hutzinger et al. 1985). 

The commercial production of trichlorophenol, as well as various derivative products such as 2,4,5-T and 

other biocides, has yielded large quantities of waste products containing substantial concentrations of 

CDDs. Extensive contamination of the environment with 2,3,7,8-TCDD occurred in Missouri in the early 

1970s as a result of the spraying of horse arenas, roads, and parking lots with mixtures of used oil and 

chemical waste (Tiernan et al. 1985). The chemical waste, formed during the manufacture of 2,4,5-TCP 

and then used to make hexachlorophene, contained several hundred ppm of 2,3,7,8-TCDD (Tiernan et al. 

1985). Several thousand gallons of this waste were dispersed over a sizable area of southwestern and 

eastern Missouri during the 1970s. Concentrations of 2,3,7,8-TCDD in soil samples from selected 

contaminated sites throughout Missouri ranged from 30 to 1,750 ppb (Tiernan et al. 1985). Concentrations 

of 2,3,7,8-TCDD in soil samples from Times Beach, Missouri, which had been heavily contaminated, 

ranged from 4.4 to 317 ppb (Tiernan et al. 1985). 

In Seveso, Italy, an explosion occurred during the production of 2,4,5-T and a cloud of toxic material 

including 2,3,7,8-TCDD was released (Cerlisi et al. 1989; MMWR 1988; Mocarelli et al. 1991). Debris 
CDDs 400 

5. POTENTIAL FOR HUMAN EXPOSURE 

from the cloud covered an area of approximately 700 acres (2.8 km2). The total amount of 2,3,7,8-TCDD 

released during the accident was estimated to be 1.3 kg. Soil samples from this industrial accident were 

measured in three areas: Zone A, the most contaminated zone where residents were evacuated; Zone B, the 

moderately contaminated area where residents were advised not to eat locally raised produce; and Zone R, 

where 2,3,7,8-TCDD contamination in soil was lowest of the three areas. Mean soil concentrations in 

these 3 areas were: 230 μg/m2

 (maximum 5,477 μg/m2) in Zone A, 3 μg/m2 (maximum 43.9 μg/m2) in 

Zone B, and 0.9 μg/m2

 (maximum 9.7 μg/m2) in Zone R (MMWR 1988). 

The migration of chemical waste containing CDDs from disposal sites has also resulted in environmental 

contamination of sediment. For example, at Love Canal in Niagara Falls, New York, where an estimated 

200 tons of 2,4,5-TCP production waste were disposed of during the 1940s and early 1950s, 

2,3,7,8-TCDD was detected at high concentrations (up to several hundred ppb) in storm sewer sediments 

(Smith et al. 1983; Tiernan et al. 1985). 

2,3,7,8-TCDD has been detected in soil and sediment samples collected at 61 and 17 sites of the 91 NPL 

hazardous waste sites where it has been detected in some environmental media (HazDat 1998). Total 

CDDs have been detected in soil and sediment samples collected at 94 and 31 of the 126 NPL sites where 

they have been detected in some environmental media. Total TCDDs, PeCDDs, HxCDDs, HpCDDs, and 

OCDD have been detected in soil samples at 71, 21, 29, 34, and 38 sites and in sediment samples at 22, 7, 

10, 9, and 13 sites of the 105, 34, 43, 49, and 53 NPL sites, respectively, where these homologues have 

been detected in some environmental media (see Table 5-1). 

5.3 ENVIRONMENTAL FATE 

Combustion generated CDDs may be transported long distances (as vapors or associated with particulates) 

in the atmosphere (Czuczwa and Hites 1986a, 1986b; Tysklind et al. 1993). They may eventually be 

deposited on soils, surface waters, or plant vegetation as a result of dry or wet deposition. CDDs 

(primarily MCDD, DCDD, TrCDD) will slowly volatilize from the water column, while the more highly 

chlorinated CDDs will adsorb to suspended particulate material in the water column and be transported to 

the sediment (Fletcher and McKay 1993; Muir et al. 1992). CDDs deposited on soils will strongly adsorb 

to organic matter. CDDs are unlikely to leach to underlying groundwater but may enter the atmosphere on 

soil dust particles or enter surface waters on soil particles in surface runoff. Low water solubilities and 

high lipophilicity indicate that CDDs will bioconcentrate in aquatic organisms, although as a result of their 
CDDs 401 

5. POTENTIAL FOR HUMAN EXPOSURE 

binding to suspended organic matter the actual uptake by such organisms may be less than predicted. This 

is also true of uptake and bioconcentration by plants, although foliar deposition and adherence may be 

significant. 

5.3.1 Transport and Partitioning 

Combustion processes appear to have contributed to the ubiquity of CDDs in the environment (Hites and 

Harless 1991; Tysklind et al. 1993). CDDs have relatively long residence times in the atmosphere, and 

combustion-generated CDDs associated with particulates can become distributed over large areas 

(Tysklind et al. 1993). During transport in the atmosphere, CDDs are partitioned between the vapor phase 

and particle-bound phase (Hites and Harless 1991). However, because of the very low vapor pressure of 

CDDs, the amount present in the vapor phase generally is negligible as compared to the amount adsorbed 

to particulates (Paustenbach et al. 1991). The two environmental factors controlling the phase in which the 

congener is found are the vapor pressure and the atmospheric temperature (Hites and Harless 1991). 

Congeners with vapor pressure <10-8 mm Hg will be primarily associated with particulate matter while 

congeners with a vapor pressure >10-4 mm Hg will exist primarily in the vapor phase. Those chemicals 

with vapor pressures between these values can be found in both the vapor phase and associated with 

particulates (Eisenreich et al. 1981). With a reported vapor pressure ranging from 7.4x10-10 to 

3.4x10-5 mm Hg, 2,3,7,8-TCDD falls into the intermediate category. 

The detection of CDDs in sediments from Siskiwit Lake, Isle Royale, suggests that CDDs can be 

transported great distances in air (Czuczwa and Hites 1986a, 1986b). Because this lake is landlocked on a 

wilderness island in Lake Superior, the only way that CDDs could reach these sediments is by atmospheric 

fall-out (i.e., by wet and dry deposition). Similar amounts of CDDs were also found in Lake Huron and 

Lake Michigan sediments, which indicates that atmospheric transport is a source of CDDs found on these 

Great Lake sites (Czuczwa and Hites 1986a, 1986b; Hutzinger et al. 1985). Atmospheric deposition of 

TCDD to Lake Erie may contribute up to 2% of the annual input of TCDD to the lake (Kelly et al. 1991). 

Through pattern analysis of herring gull monitoring data, Hebert et al. (1994) provided evidence that the 

sources of CDDs in Great Lakes food chains were mainly atmospheric, with the exception of 

2,3,7,8-TCDD in Lake Ontario, and several CDDs in Saginaw Bay in Lake Huron where point sources 

were implicated. 
CDDs 402 

5. POTENTIAL FOR HUMAN EXPOSURE 

CDDs are physically removed from the atmosphere via wet deposition (scavenging by precipitation), 

particle dry deposition (gravitational settling of particles), and gas-phase dry deposition (sorption of CDDs 

in the vapor phase onto plant surfaces) (Rippen and Wesp 1993; Welschpausch et al. 1995). Precipitation 

(rain, sleet, snow) is very effective in removing particle-bound CDDs from the atmosphere (Hites and 

Harless 1991; Koester and Hites 1992). Table 5-2 summarizes the average ppt scavenging ratios and 

percentage of washout due to particulates for congener groups of both CDDs and CDFs collected at two 

sites in Indiana. The scavenging ratio is the ratio of the concentration of the congener group in rain to the 

atmospheric concentration of the congener group and is a measure of the effectiveness of rain in removing 

the congener groups from the atmosphere. Table 5-2 also summarizes the percentages of the congener 

groups scavenged as particles in rain rather than as dissolved solutes in rain. Total rain scavenging ratios 

ranged from 10,000 to 150,000; HpCDDs and OCDD (the congeners most strongly associated with 

particulates) were the congeners scavenged most efficiently (Hites and Harless 1991; Koester and Hites 

1992). 

Environmental fate modeling of CDDs requires knowledge of a number of fundamental physical and 

chemical parameters, such as water solubility, vapor pressure, Henry's law constant, octanol-water 

partition coefficient (Kow), and organic carbon partition coefficient (Koc). CDDs are a class of high 

molecular weight, highly hydrophobic compounds. Although the class contains 8 homologues (congener 

groups) and 75 congeners, solubility values are available for only a handful of these congeners (Doucette 

and Andren 1988). CDDs have very low water solubilities, with solubility decreasing with increasing 

chlorine substitutions (Doucette and Andren 1988). The water solubility of 2,3,7,8-TCDD ranges from 

7.9x10-6 to 33.2x10-4 mg/L (Shiu et al. 1988). See Table 3-2 for the water solubilities for specific 

congeners. Water solubilities at 25 EC for the congener groups have been estimated as follows: MCDD, 

0.278–0.417 mg/L; DCDD, 3.75x10-3–1.67x10-2 mg/L; TrCDD, 4.75x10-3–8.41x10-3; TCDD, 7.9x10-6 to 

6.3x10-4 mg/L; PeCDD, 1.18x10-4 mg/L; HxCDD, 4.42x10-6 mg/L; HpCDD, 2.4x10-6–1.9x10-3 mg/L; and 

OCDD, 0.1x10-9–7.4x10-8 mg/L (ASTER 1995; Doucette and Andren 1988; HSDB 1997; McCrady and 

Maggard 1993; Shiu et al. 1988). 

CDDs generally exhibit very low vapor pressures, with the tendency of decreasing vapor pressure with 

increasing chlorine substitution (Friesen et al. 1985; Rordorf 1986, 1989). At 25 EC, the vapor pressure of 

2,3,7,8-TCDD ranges from 7.4x10-10 to 3.4x10-5 mm Hg (HSDB 1997; Rordorf 1989). See Table 3-2 for 

the vapor pressures of specific congener groups. Vapor pressures at 25 EC for the other congener groups 

have been estimated as follows: MCDD, 9.0x10-5–1.3x10-4 mm Hg; DCDD, 9.0x10-7–2.9x10-6 mm Hg; 

CDDs 404 

5. POTENTIAL FOR HUMAN EXPOSURE 

6.46x10-8–7.5x10-7; TCDD, 7.4x10-10–4.0x10-3 mm Hg; PeCDD, 6.6x10-10 mm Hg; HxCDD, 3.8x10-11 mm 

Hg; HpCDD, 5.6x10-12–7.4x10-8 mm Hg; and OCDD, 8.25x10-13–1.68x10-12 mm Hg (HSDB 1997; 

McCrady and Maggard 1993; Rordorf 1989; Shiu et al. 1988). CDDs can be found in both the vapor and 

particle-bound phases (Eitzer and Hites 1989a; Hites and Harless 1991), with the low vapor pressure of 

OCDD resulting in its enrichment in the particulate phase in the atmosphere. When this particulate matter 

is deposited on water, OCDD-enriched sediments will result (Eitzer 1993). The less chlorinated CDD 

congeners (TCDD and PeCDD) occur in greater proportion in the vapor and dissolved phases of air and 

rain, whereas the more chlorinated congeners (HpCDD and OCDD) are associated with the particulatebound phases (EPA 1991d). Data from one study of CDDs in the ambient atmosphere of Bloomington, 

IN, found that vapor-to-particle ratios for individual CDDs ranged from 0.01 to 30 and were dependent on 

the ambient temperature and the compound's vapor pressure (Eitzer and Hites 1989b). Since the lesschlorinated CDDs have higher vapor pressures, they are found to a greater extent in the vapor phase (Eitzer 

and Hites 1989a). As air moves, photodegradation of the vapor-phase CDDs occurs and they are lost more 

readily than the particulate-bound CDDs. Vapor-phase CDDs are not likely to be removed from the 

atmosphere by wet or dry deposition (Atkinson 1991), although this is a primary removal process for 

particulate-bound CDDs. Wet or dry deposition could result in greater concentrations of the more 

chlorinated CDDs reaching soil or water surfaces and eventually sediment (EPA 1991d). All CDDs are 

found to some extent in both the vapor phase and bound to particulates. At warmer temperatures (28 EC), 

CDDs, particularly the MCDDs, DCDDs, TrCDDs, and TCDDs will have a greater tendency to exist in the 

vapor phase. At cooler temperatures (16–20 EC and <3 EC), all CDDs will have less propensity to exist in 

the vapor phase and greater propensity to adsorb to particulates (Shroy et al. 1985). At a constant 

temperature, there is a positive relationship between increasing numbers of chlorine atoms on the molecule 

and decreased propensity to exist in the vapor phase relative to particulate adsorption (Eitzer and Hites 

1989b; Paustenbach et al. 1991; Shroy et al. 1985). 

CDDs are removed from the water column to a minor extent by volatilization to the atmosphere, with 

binding to particulates and sediment, or bioaccumulation by aquatic biota being more significant processes 

(Fletcher and McKay 1993; Muir et al. 1992; Paustenbach et al. 1992). CDDs have Henry's law constants 

ranging from 1.31×10-6 to 146×10-6 atm-m3

/mol (Shiu et al. 1988). These values indicate that volatilization from water is likely to be a slow, with the transfer rate controlled by the gas-phase resistance (i.e., 

the rate is controlled by slow diffusion through the air) (Lyman et al. 1982; Shiu et al. 1988). The more 

chlorinated homologous classes (TCDD, PeCDD, HxCDD, HpCDD, OCDD) have lower Henry's law 

constant values than the less chlorinated homologous classes (MCDD, DCDD, TrCDD). Thus, 
CDDs 405 

5. POTENTIAL FOR HUMAN EXPOSURE 

volatilization from the water column is not expected to be a very significant loss process for the TCDD 

through OCDD congeners as compared to adsorption to particulates. In general, the Henry's law constants 

decrease with increasing chlorine number as a result of the decrease in vapor pressure and water solubility 

(Shiu et al. 1988). Volatilization half-lives for 2,3,7,8-TCDD were calculated for ponds and lakes 

(32 days) and for rivers (16 days) (Podoll et al. 1986). The primary removal mechanism for CDDs from 

the water column is sedimentation, with 70–80% of the CDDs being associated with the particulate phase 

(Muir et al. 1992). The remainder was associated with dissolved organic substances. CDDs bound to 

sediment particles may be resuspended in the water column if the sediments are disturbed. This could 

increase both the transport and availability of the CDDs for uptake by aquatic biota (Fletcher and McKay 

1993). 

Generally, CDDs are characterized by low vapor pressure, low aqueous solubility, and high hydrophobicity, suggesting that these compounds strongly adsorb to soil and that their vertical mobility in the 

terrestrial environment is low (Eduljee 1987b). In general, higher chlorinated CDDs also volatilize more 

slowly from soil and water surfaces than do lower chlorinated ones (Hutzinger et al. 1985). Nash and 

Beall (1980) reported that only 12% of 2,3,7,8-TCDD applied to bluegrass turf as a component of 

emulsifiable Silvex volatilized over a 9-month period. Because CDDs (particularly the more highly 

chlorinated PCDD, HxCDD, HpCDD, and OCDD) strongly adhere to soil and exhibit low solubility in 

water, leaching of CDDs would be unlikely if water were the only transporting medium. Instead, wind and 

erosion can cause the mixing and transport of CDD-contaminated soil. As a result of erosion, surface soil 

contaminated with CDDs is either blown away by wind or washed via surface water runoff into rivers, 

lakes, and streams, with burial in the sediments being the predominant fate of CDDs sorbed to soil 

(Hutzinger et al. 1985). 

Adsorption is an important process affecting transport of hydrophobic compounds such as CDDs. The 

organic carbon fraction of the soil is believed to be the most important factor governing the degree of 

adsorption of hydrophobic organic contaminants. CDDs adsorb more strongly to soils with a higher organic 

carbon content than to soils with low organic carbon content (Yousefi and Walters 1987). Because of their 

very low water solubilities and vapor pressures, CDDs found below the surface soil (top few mm) are 

strongly adsorbed and show little vertical migration, particularly in soil with high organic carbon content 

(Yanders et al. 1989). Vertical movement of CDDs in soil may result from the saturation of sorption sites 

of the soil matrix, migration of organic solvents, or human or animal activity (Hutzinger et al. 1985). 

Adsorption/desorption of 2,3,7,8-TCDD in contaminated soils was studied by Des Rosiers (1986). Soil 

samples were taken from an abandoned 2,4,5-T manufacturing facility and a scrap metal yard in New 
CDDs 406 

5. POTENTIAL FOR HUMAN EXPOSURE 

Jersey and from horse arenas, roadways, and residential property in Missouri. Historically, these samples 

were contaminated with either chemical residues or waste oils containing 2,3,7,8-TCDD. Mean log organic 

carbon partition coefficient (Koc) values ranged from 7.39 to 7.58 (Des Rosiers 1986). This Koc range 

indicates that 2,3,7,8-TCDD is immobile in soil (Swann et al. 1983). However, the mobility of 

2,3,7,8-TCDD in soil will increase if organic co-solvents that can solubilize 2,3,7,8-TCDD are present in 

the soil (Podoll et al. 1986). This situation might occur at a hazardous waste site. In one study, only 1.5% 

of the CDDs applied to soil surfaces had leached to a depth of 2.5 cm below the soil surface after 

15 months. Leaching of the CDDs through the soil was primarily associated with carriers such as petroleum 

oil (Orazio et al. 1992). 

Most CDDs entering surface waters are associated with particulate matter (dry deposition of atmospheric 

particles) and eroded soil particulates contaminated with CDDs (Hallett and Brooksbank 1986). In the 

aquatic environment, significant partitioning of CDDs from the water column to sediment and suspended 

particulate organic matter may occur. Dissolved CDDs will partition to suspended solids and dissolved 

organic matter (detritus, humic substances) and are likely to remain sorbed once in the aquatic environment. 

From suspended sediment and water data collected from the Niagara River on the New York-Canada 

border, it was found that CDDs were strongly associated with suspended sediment (Hallett and Brooksbank 

1986). Concentrations of total TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD in raw water ranged 

from below detection limits to 3.6 pg/L (3.6 ppq), while the concentration of these same homologue groups 

in suspended sediments ranged from below detected limits to 228 pg/g (ppt) (Hallett and Brooksbank 1986). 

The more highly chlorinated congeners (HxCDD, HpCDD, and OCDD) predominated in both water and 

suspended sediment samples. 

A model has been developed to describe the vertical transport of low-volatility organic chemicals in soil 

(Freeman and Schroy 1986). The model was used to make predictions on the transport of 2,3,7,8-TCDD at 

the Eglin Air Force Base Agent Orange biodegradation test plots (Freeman and Schroy 1986). Trenches 

10 cm deep were dug in the soil, and Agent Orange containing 40 ppb of 2,3,7,8-TCDD was applied to the 

trench bottom. The model predicted a vertical movement of 2,3,7,8-TCDD, buried in 1972, through the soil 

column. Soil-column-profile data confirm the vertical movement of 2,3,7,8-TCDD from core samples taken 

in 1984 (Freeman and Schroy 1986). The 2,3,7,8-TCDD in the Eglin Air Force Base biodegradation plots 

moved through the entire 10 cm of the soil column in 12 years (Freeman and Schroy 1986). The rates of 

migration and loss of 2,3,7,8-TCDD in contaminated soil were studied under natural conditions in 

experimental plots at the Dioxin Research Facility, Times Beach, Missouri (Yanders et al. 1989). The 
CDDs 407 

5. POTENTIAL FOR HUMAN EXPOSURE 

TCDD concentration profiles of sample cores taken at Times Beach in 1988 (mean range 78–160 ppb) were 

virtually the same as those in cores taken in 1984 (mean range 76–162 ppb). The results show that little 

movement and essentially no loss due to volatilization of 2,3,7,8-TCDD had occurred in the experimental 

plots in the four years since the Dioxin Research Facility was established (Yanders et al. 1989). 

CDDs are characterized by low water solubilities and high lipophilicities. Kow values range from 104

 to 1012 

for MCDD through OCDD, with Kow values increasing relative to increasing chlorination (Table 3-2). 

Because of these physicochemical properties, CDDs are expected to adsorb to bedded and suspended 

sediments and to bioaccumulate in aquatic organisms. 

The bioconcentration factor (BCF) is the ratio of the concentration of CDDs in an organism over the 

concentration of CDDs in water. The BCF values for CDDs can be estimated from their Kow values, and a 

number of regression equations are available for this purpose (Bysshe 1990). Experimentally measured 

BCFs for selected CDD congeners in various aquatic species are summarized in Table 5-3. Measurements 

of the bioconcentration of CDDs tend to increase with the degree of chlorination up to TCDDs, and then 

decrease as chlorination continues to increase up to the OCDD congener (Loonen et al. 1993). The more 

highly chlorinated congeners, such as OCDD, appear to have the lowest bioconcentration potential either 

because they are less bioavailable because of their rapid adsorption to sediment particles (Servos et al. 

1989a, 1989b) or because their large molecule size may interfere with transport across biological 

membranes (Bruggeman et al. 1984; Muir et al. 1986a, 1986b). 

The hydrophobic nature of CDDs, combined with their great affinity for organic carbon, suggests that a 

major proportion of CDDs in the aquatic environment is sorbed to organic matter and sediment. Because 

only a minute fraction of CDDs are dissolved in the natural environment, bioconcentration is not the 

primary route of exposure for most aquatic organisms. Whereas the term bioconcentration is defined as the 

uptake of a chemical from water only, the term bioaccumulation refers to the combined uptake of a 

chemical from both dietary sources (e.g., food) and water. A bioaccumulation factor (BAF) that includes 

the ingestion route of uptake can be calculated based on fish uptake from water, food, and sediment 

(Sherman et al. 1992). 

The primary route of exposure to CDD congeners for lower trophic organisms (e.g., phytoplankton and 

various aquatic invertebrates) is uptake from the water column or from interstitial water (between sediment 


CDDs 410 

5. POTENTIAL FOR HUMAN EXPOSURE 

particles). Certain benthic organisms accumulate highly lipophilic compounds (e.g., PCBs and 

CDDs/CDFs) from water at the water/sediment interface (the concentration of a lipophilic compound is 

generally higher at this interface than in the water column) and via intake of phytoplankton, zooplankton, 

and suspended particulate materials that contain higher concentrations of these chemicals than the 

surrounding water (Porte and Albaiges 1993; Pruell et al. 1993; Secor et al. 1993). For the higher trophic 

level organisms, such as foraging fish, predaceous fish, and piscivorous wildlife, the predominant route of 

exposure is via food chain transfer, with negligible contributions from CDDs in water and sediment (Muir 

and Yarechewski 1988). Exposure through direct consumption of CDD-contaminated sediment and detritus 

may occur in some bottom-feeding species such as carp and white suckers (Kuehl et al. 1987a, 1987b; 

Servos et al. 1989a, 1989b). Under natural conditions, in which a high proportion of these hydrophobic 

CDD compounds are sorbed to suspended and dissolved organic matter, direct uptake of these CDDs from 

water is not expected to be substantial (Muir et al. 1986a, 1986b). The estimated BCFs in such cases may 

not be a good indicator of the experimental bioaccumulation measured in the field. Another reason for the 

difference between estimated BCFs and experimentally measured bioaccumulation values is the ability of 

some aquatic organisms to metabolize and eliminate specific CDD congeners from their bodies and thereby 

change the congener profile pattern in their tissues. 

Preferential bioconcentration and bioaccumulation of 2,3,7,8-TCDD and other 2,3,7,8-substituted CDDs 

by aquatic organisms have been reported (Branson et al. 1985; Kuehl et al. 1985, 1987a, 1987b, 1987c; 

Opperhuizen 1986; Paustenbach et al. 1992). In water-only exposure studies, BCF values for fish exposed 

to 2,3,7,8-TCDD ranged from 37,900 to 128,000 (Cook et al. 1991; Mehrle et al. 1988). Much lower BCF 

values ranging from 1,400 to 5,840 and 34 to 2,226 have been reported for fish exposed to 1,3,6,8,-TCDD 

and OCDD, respectively (Muir et al. 1986a, 1986b). These BCF values are approximately two orders of 

magnitude less than would be predicted using the Kow values. Similarly, the lower BCFs for HpCDD in 

fathead minnows and OCDD in rainbow trout fry relative to the other CDDs tested resulted from lower 

uptake efficiencies from water. Elimination half-lives for TCDDs and PeCDDs were similar and rapid, 

averaging about 2.6 days in trout fry and 3 days in minnows. Elimination half-lives for HxCDD and 

HpCDD were longer, averaging about 16 days in rainbow trout and 20 days in fathead minnows (Muir et 

al. 1986b). The results of these studies also indicate that BCFs of the higher chlorinated CDDs (HxCDD, 

HpCDD, OCDD) from water are much lower than would be predicted based on their Kow values. Servos et 

al. (1989a, 1989b) also noted that the BCF values were less than predicted based on the Kow values, and 

these authors suggest that BCFs reported in the literature may underestimate the true BCF, unless the BCFs 

were calculated using truly dissolved CDD concentrations in the water column rather than 
CDDs 411 

5. POTENTIAL FOR HUMAN EXPOSURE 

total dissolved concentrations, which would include complexes with large molecules of dissolved organic 

carbon. 

BCF values measured in fish exposed to both water and sediment were much lower than equivalent 

exposures to water only and ranged from 2,500 to 5,800 (Adams et al. 1986; Cook et al. 1991; Tsushimoto 

et al. 1982) (Table 5-3). Loonen et al. (1993) also reported that bioaccumulation of CDDs was reduced in 

the presence of sediment and that the effects of sediment increased with increasing hydrophobicity (degree 

of chlorination) of the congeners. BCFs were reduced by 15–82% for various CDD/CDF congeners, with 

the greatest reduction associated with OCDD. 

The bioavailability of CDDs/CDFs from municipal incinerator fly ash and sediment to freshwater fish has 

been studied in experimental situations. Like the BCF and BAF values, the biota-sediment-accumulation 

factor (BASF) (ratio of contaminant concentration in the organism normalized to lipid content to the 

concentration in fly ash or sediment, normalized to organic carbon content) generally decreased with an 

increasing degree of chlorination (Kuehl et al. 1985, 1987b, 1987c). The BASF values for benthic 

(bottom-dwelling) fish (e.g., carp, catfish) are generally higher than for those pelagic (water column) 

species (e.g., bass, trout, sunfish) because of the higher lipid content and increased exposure to 

contaminated sediments for the benthic species (Paustenbach et al. 1992). 

Several authors have studied the disposition and metabolism of CDDs in fish. Studies on the disposition of 

2,3,7,8-TCDD in rainbow trout and yellow perch indicate that fatty tissues (visceral fat, carcass, skin, and 

pyloric caeca) typically contain the bulk of 2,3,7,8-TCDD (78–90%) with only a small percentage (2–5%) 

associated with the skeletal muscle (Kleeman et al. 1986a, 1986b). For other congeners, such as 

1,3,6,8-TCDD and OCDD, the greatest proportion of the total body burden is concentrated in the bile, with 

lesser concentrations in liver > caeca > kidney > spleen > skin > muscle (Muir et al. 1986a, 1986b). 

Differences in the distribution among various species may be a function of the exposure pathway (i.e., 

dietary versus water uptake) and differences in metabolic breakdown rates. For example, both the parent 

compound and metabolites of 2,3,7,8-TCDD and 1,3,6,8-TCDD were present in the bile of fish exposed 

under laboratory conditions (Branson et al. 1985; Muir et al. 1986a, 1986b). Kleeman et al. (1986b) 

reported the presence of several polar metabolites in the gall bladder of yellow perch exposed to a single 

dose of 14 C- 2,3,7,8-TCDD. One week later, the gall bladder, skin, skeletal muscle, and kidneys were 

removed. In contrast to liver, muscle, and kidney where the parent compound accounted for 96–99% of the 

extractable 14 C, the gall bladder contained almost entirely 2,3,7,8-TCDD metabolites, at least one of which 
CDDs 412 

5. POTENTIAL FOR HUMAN EXPOSURE 

was a glucuronide conjugate. Although the metabolic breakdown was slow, it is clear that CDDs can be 

transformed by fish to polar metabolites that are subsequently excreted in the bile. 

Freshwater aquatic invertebrates have been shown to bioaccumulate CDDs/CDFs through water, sediment, 

and food pathways (Isensee 1978; Muir et al. 1983; Yockim et al. 1978). The range in experimentally 

determined BCF values for freshwater invertebrates is presented in Table 5-3. As discussed previously, 

exposure to CDDs from sediment and water containing dissolved organic material markedly decreases the 

BCF values, especially for the more highly chlorinated CDDs. Sediment-dwelling organisms (e.g., 

Chironomous sp. larvae and Hexagenia sp. nymphs), stoneflies, and other predaceous nymphs showed poor 

accumulation of OCDD in comparison to 1,3,6,8-TCDD (Muir et al. 1983). The lower bioaccumulation of 

OCDD was attributed to greater adsorption of the OCDD onto sediment particles and organic matter, and 

the reduced uptake across biological membranes due to large molecular size. The potential ingestion of 

sediments during burrowing activities by sediment-dwelling insects was believed to result in greater tissue 

concentrations of CDDs than those observed for predaceous insects. It is also possible that predaceous 

insects may metabolize 1,3,6,8-TCDD more effectively, leading to a greater rate of elimination. Sedimentdwelling organisms are important food sources for fish and other predaceous insects; consequently, if rapid 

elimination of 1,3,6,8-TCDD and low accumulation of OCDD occur in the natural environment, 

bioaccumulation of these congeners in trophically higher-level organisms may not be significant (Muir et al. 

1983). 

Marine invertebrates have also shown an ability to bioaccumulate CDDs/CDFs to varying degrees in their 

tissues (Brown et al. 1994; Cai et al. 1994; Conacher et al. 1993; Hauge et al. 1994; Rappe et al. 1991), 

although no information on BCF values was found in the literature. Interestingly, several investigators have 

reported that shellfish species (crustaceans and molluscs) are better indicators of CDD/CDF contaminant 

levels than fish because their tissues contain larger numbers and higher residues of CDD/CDF congeners in 

addition to the 2,3,7,8-TCDD congeners and other 2,3,7,8-substituted congeners that are selectively 

accumulated in fish species (Brown et al. 1994; Conacher et al. 1993; Rappe et al. 1991). This is in contrast 

to what is observed in fish and fish-eating birds, in which there is selective retention of congeners with the 

2,3,7,8-substitution positions occupied, which may be due to an increased ability to metabolize and 

eliminate non-2,3,7,8-substituted CDD/CDF congeners (Brown et al. 1994; Rappe et al. 1991). The use of 

shellfish species as target organisms in CDD/CDF-monitoring studies is recommended as these species 

provide a better overall representation of both the magnitude and congener-specific nature of the 

environmental contamination (Petreas et al. 1992). Conacher et al. (1993) present an example where 
CDDs 413 

5. POTENTIAL FOR HUMAN EXPOSURE 

use of a shellfish species provides a much higher estimate of exposure to CDDs/CDFs as well as to total 

CDD equivalent toxicity (TEQs) than use of a fish species. This difference in congener bioaccumulation 

profiles between fish and shellfish species is a result of the ability of fish to metabolize CDDs/CDFs. Both 

the parent congeners and metabolites of 2,3,7,8-TCDD and 1,3,6,8-TCDD were present in the bile of fish 

exposed under laboratory conditions (Branson et al. 1985; Muir et al. 1986a). Kleeman et al. (1986a, 

1986b) reported the presence of several polar metabolites, including glucuronide conjugates, in various fish 

exposed to 2,3,7,8-TCDD. Despite the slowness of the metabolic breakdown processes, it is clear that 

CDDs can be transformed within fish to polar metabolites that are subsequently excreted with the bile. It 

does not appear from the results obtained in studies conducted to date that shellfish species have the same 

ability to metabolize and eliminate non-2,3,7,8-substituted CDDs/CDFs (Brown et al. 1994; Cai et al. 

1994). 

It is apparent from the available data regarding the substantial bioaccumulation potential of CDDs/CDFs in 

aquatic organisms (particularly the 2,3,7,8-substituted congeners) as well as data on the extent of 

contamination of fish and shellfish in various freshwater and marine waterways, that ingestion of 

contaminated fish and shellfish is an important exposure pathway for CDDs/CDFs in humans. 

CDDs have been found to accumulate in both surface and rooted aquatic vegetation, with BCF values 

ranging from 208 to 2,083 (Table 5-3) (Isensee 1978; Tsushimoto et al. 1982; Yockim et al. 1978). Corbet 

et al. (1983) reported that a rooted plant species (Potemageton pectimatus) and a surface-dwelling 

duckweed (Lemna sp.) accumulated concentrations of 1,3,6,8-TCDD of 280 and 105 ng/g (dry weight), 

respectively, following exposure to water containing 1,000 ng/L (ppt). The maximum concentrations were 

observed 8 days post-application and represented 6% of the total TCDD applied. These results are similar 

to those reported by Tsushimoto et al. (1982) in an outdoor pond study, in which a maximum bioaccumulation of 2,3,7,8-TCDD in the pond weeds Elodea nuttali and Ceratophyllon demersum equivalent to a BCF 

of 130 occurred after 5 days of exposure. In both studies, the tissue concentrations reached equilibrium in 

approximately 20 days and remained constant until the end of the experiment (approximately 58 and 

170 days, respectively). These experimental data indicate that CDDs can accumulation in aquatic plant 

species through waterborne exposure. 

Like many fish, several species of fish-eating birds have shown the ability for preferential bioaccumulation 

of 2,3,7,8-TCDD and other 2,3,7,8-substituted CDDs and TCDFs. Jones et al. (1994) monitored TEQ 

values for 2,3,7,8-TCDD in double-crested cormorants from three of the Great Lakes: Superior, Michigan, 
CDDs 414 

5. POTENTIAL FOR HUMAN EXPOSURE 

and Huron. Biomagnification factors (BMF, the ratio of the concentration of TCDD-equivalents in bird 

eggs to concentrations in forage fish) were found to range from 11.7 to 56.8 (mean, 31.3). In another study, 

all of the CDDs and CDFs detected in double-crested cormorant and Caspian tern eggs were 2,3,7,8­

substituted (Yamashita et al. 1992). Concentrations of 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 1,2,3,4,7,8­

HXCDD, 1,2,3,6,7,8-HXCDD, 1,2,3,7,8,9-HXCDD, 1,2,3,4,6,7,8-HpCDD, and OCDD ranged from 5.3 to 

20, 3.2 to 9.4, 10 to 20, 3.6 to 11, and 7.8 to 16 pg TEQ/g, respectively, for double-crested cormorant eggs, 

and 8.2 to 22, 3.3 to 6.4, 8.7 to 17, 2.4 to 6.0, and 9.7 to 21 pg TEQ/g, respectively, for Caspian tern eggs. 

This same pattern was also reported to occur in California peregrine falcons and their eggs (Jarman et al. 

1993). For this species, mean concentrations of 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 1,2,3,4,7,8-HxCDD, 

1,2,3,6,7,8-HxCDD, 1,2,3,7,8,9-HxCDD, 1,2,3,4,6,7,8-HpCDD and OCDD in eggs were 5.7, 11, 2, 11, 1.3, 

3.8, and 5.3, respectively. Fish-eating birds are exposed to CDDs primarily through their diet. A rapid 

decline in contaminant levels in eggs of fish-eating birds, therefore, reflects a rapid decrease in contaminant 

levels of their prey. This has been shown to occur in Great blue heron chicks in British Columbia 

(Sanderson et al. 1994) in areas where CDD/CDF levels in pulp and paper mill effluents decreased 

substantially within a few years. The Great blue heron chicks also showed an increased hepatic microsomal 

ethoxyresorufin O-deethylase (EROD) activity in the areas of highest contamination. This indicates that the 

induction of cytochrome P-450 1A1 has occurred, and that the Ah-receptor-mediated process, by which 

2,3,7,8-TCDD and related chemicals exert their toxicities, has been activated. 

Ankley et al. (1993) studied the uptake of persistent polychlorinated hydrocarbons by four avian species at 

upper trophic levels of two aquatic food chains. Concentration of 2,3,7,8-TCDD toxic equivalents (TEQs) 

were evaluated in Forster’s tern and common tern chicks and in tree-swallow and red-winged-blackbird 

nestlings from several areas in the watershed. Young birds accumulated small concentrations of 

2,3,7,8-TCDD and several other 2,3,7,8-substituted CDDs and CDFs, including 1,2,3,6,7,8-HxCDD, 

2,3,7,8-TCDF, 1,2,3,6,7,8-HxCDF, 1,2,3,4,6,7,8-HpCDF, 1,2,3,7,8-PeCDD, 1,2,3,4,6,7,8-HpCDD, and 

OCDD. The general trend in concentrations of CDDs from the greatest to least was Forster’s tern ­

common tern > tree swallow > red-winged blackbird. The similarity in concentrations between the two tern 

species is expected given that they are both piscivores and their similar life histories and the close proximity 

of the two colonies. The greater concentrations in the tree swallows than in the red-winged blackbirds were 

somewhat unexpected given the presumed similarity of the diets (both species are insectivores). The 

authors suspect that the red-winged blackbirds foraged more on relatively uncontaminated upland food 

sources than the tree swallows, which fed primarily on chironomids emerging from the bay. 
CDDs 415 

5. POTENTIAL FOR HUMAN EXPOSURE 

2,3,7,8-TCDD is generally considered to be bioavailable to terrestrial birds primarily through ingestion of 

TCDD-laden food items and soil particles (Nosek et al. 1992). These authors, using H3

TCDD-administered 

suspensions in various environmental matrices, found that 30% of the dose absorbed from suspensions of 

earthworms, 33% absorbed from soil suspensions, 41% absorbed from suspensions of paper mill sludge 

solids, and 58% absorbed from a suspension of crickets. These authors also reported that the percentage of 

the cumulative TCDD dose translocated to an individual egg was 1.1% for the first 15 eggs laid and that the 

percentage was not affected by the order in which the eggs were laid. Assuming an adult female could lay 

30 eggs, 35% of the hen TCDD body burden could be translocated to all eggs laid. Results of these studies 

suggest that TCDD can be orally bioavailable from earthworms and crickets, important dietary sources for 

this species and other terrestrial species, as well as from nonfood items such as orally ingested soil and 

paper mill sludge solids. 

For terrestrial mammals, the BCF value is the quotient of the concentration of CDD in the tissues divided 

by the concentration in food (Geyer et al. 1986a, 1986b). BCF values for 2,3,7,8-TCDD were calculated in 

the liver and/or fat of rats, cows, and monkeys (Geyer et al. 1986a; Kociba et al. 1978a). BCF values 

ranged from 10.9 to 24.5 in liver tissue and from 3.7 to 24.5 in fat tissue of rats fed 2,200, 210, or 22 ng/kg 

of 2,3,7,8-TCDD in their diet for 2 years (Geyer et al. 1986a; Kociba et al. 1978a). The BCF value 

calculated for this rat study, increased as the concentration in the animals’ food decreased. In a cattlefeeding study, 24 ng 2,3,7,8-TCDD in the diet was fed to cows for 28 days after which time the BCF of 

2,3,7,8-TCDD in the liver was 0.7 and in the fat was 3.5. Using a linear one compartment model, Geyer et 

al. (1986a) calculated that a steady state would be reached in 499 days and that the cattle fatty tissue would 

contain 594 ng/kg. The calculated BCF value for 2,3,7,8-TCDD would then be 24.8 (Geyer et al. 1986a; 

Jensen et al. 1981). This value is in good agreement with the BCF of 24.5 calculated for rats that received 

22 ng TCDD/kg in their diet for years. This is a much higher BCF than has been reported by Fries and 

Paustenbach (1990). After 4 years of chronic exposure to 25 ng/kg 2,3,7,8-TCDD in their diet, the 

calculated BCF in fatty tissue of monkeys ranged from 24 to 40 (Geyer et al. 1986a). Using the 

2,3,7,8-TCDD concentration in human adipose tissue (10.7 ppt whole weight) and in food 

(0.052–0.103 ng/kg), the calculated BCF is between 104 and 206 on a whole-weight basis, or between 

115 and 229 on a lipid basis (90% lipid) (Geyer et al. 1986a). Using a pharmacokinetics model, the 

calculated BCF value is 153 (Geyer et al. 1986a). The authors further point out that the calculated BCFs for 

2,3,7,8-TCDD in human adipose tissue are of the same order of magnitude as those calculated for PCBs, 

DDT, and hexachlorobenzene which are also persistent compounds with comparable lipophilicity 

(n-octanol/water partition coefficients). Based on this BCF range, 2,3,7,8-TCDD was ranked as having a 
CDDs 416 

5. POTENTIAL FOR HUMAN EXPOSURE 

high bioconcentration potential in human adipose tissue (Geyer et al. 1986b). The half-life in humans was 

estimated to be approximately 7 years (Pirkle et al. 1989). 

The primary mechanisms by which CDDs enter terrestrial food chains are by atmospheric wet and dry 

deposition of vapor-phase and particulate-bound chemicals (McCrady and Maggard 1993). Uptake of 

CDDs from soils by vegetables and other plants may occur (Schroll and Scheunert 1993). Accumulation of 

CDDs on vegetation may involve both of these mechanisms. Since 2,3,7,8-TCDD is lipophilic, adsorbs 

strongly to soil, and is not very soluble in water, root uptake and translocation to upper plant parts is only a 

minor source of vegetative contamination (Travis and Hattemer-Frey 1987) except perhaps for plant species 

belonging to the Cucurbitaceas (e.g., zucchini and pumpkin). For zucchini and pumpkin plants, root uptake 

of CDD/CDFs and subsequent translocation to the shoots and into the fruits is a main contamination 

pathway (Hulster et al. 1994). Hulster and Marschner (1993) reported that CDD levels in foliage were not 

related to CDD levels in soil. The contamination of plant foliage via atmospheric deposition is a more 

important contamination mechanism than root uptake and translocation to plant foliage (McCrady et al. 

1990). Welschpausch et al. (1995) determined that dry deposition was the main pathway of uptake in grass 

of CDDs/CDFs from the atmosphere. Particles <2.9 μm in diameter were not important in atmospheric 

deposition, but large particles may contribute to HpCDD and OCDD accumulation. McCrady et al. (1990) 

conducted experiments with plants growing in nutrient solutions containing TCDD in a closed-laboratory 

system. These authors demonstrated that translocation from roots to shoots did not occur, but shoot 

contamination was associated with foliar uptake from the air. In general, there is little bioaccumulation of 

CDDs in plants (Hutzinger et al. 1985). BCFs for TCDD in plants have been estimated to be 0.0002, 

although most absorption occurs in the plant root with little or no translocation through the plant to the 

foliage (Wild and Jones 1992). A concentration of 0.06 ppm 2,3,7,8-TCDD was applied to the soil and root 

uptake from soil was then measured in oats and soybeans (Kearney et al. 1971). Oat and soybean plants (at 

all growth stages) accumulated very small quantities of 2,3,7,8-TCDD. A maximum of 0.15% (0.12 ppm) 

of 2,3,7,8-TCDD present in soils was translocated to the aerial portion of the oat and soybean plants. No 

detectable amounts of the compound were found in the oat or soybean plants harvested at maturity. The 

amount of 2,3,7,8-TCDD applied to these soils was many thousands of times greater than that which would 

occur in soils from herbicide applications containing a few ppm of 2,3,7,8-TCDD as an impurity. Even 

upon exposure to these high concentrations in the soil, significant amounts of 2,3,7,8-TCDD could not be 

measured in the plants (detection limit not reported) (Kearney et al. 1971). 
CDDs 417 

5. POTENTIAL FOR HUMAN EXPOSURE 

Maize (corn) and bean cultivations grown in soils spiked with 22–1,066 ppt 2,3,7,8-TCDD showed 

2,3,7,8-TCDD concentrations in roots ranging from 16 to 1,278 ppt for maize and from 37 to 1,807 for 

beans (Fachetti et al. 1986). The soil-grown crops did not show a significant increase of 2,3,7,8-TCDD in 

above-ground parts, either as a function of time or with increasing concentration of the pollutant in the soil 

(Fachetti et al. 1986). 

Uptake of 14C-labeled OCDD was studied in a closed, aerated-soil plant system for 7 days after application 

of the OCDD to soil (Schroll et al. 1994). The BCF (concentration of 14C equivalent to the OCDD in plant 

dry matter divided by 14C-labeled OCDD in dry soil) was 0.742 in carrot root and 0.085 in carrot shoots 

grown on OCDD-contaminated soil as compared to a BCF of not determinable and 0.084 in the control 

carrot root and shoots, respectively. There was no transport of 14C-labeled OCDD between the roots and 

shoots or vice versa. The residues in roots were due only to root uptake from the soil; those in shoots were 

due only to foliar uptake from the air. 

Muller et al. (1993) studied transfer pathways of CDD/CDFs to fruit. These authors found that homologue 

patterns of CDDs/CDFs in soil were different from those in both apples and pears grown in the contaminated soil. Concentrations of CDDs/CDFs ranged from 1 to 4 ng/kg (fresh weight) and were 4–8 times 

higher in the peel than in the pulp. These authors suggest that airborne CDDs/CDFs are a major source of 

contamination of fruits grown in contaminated soil. Muller et al. (1994) conducted field studies of CDD 

transfer pathways from soil to several edible plant varieties (carrots, lettuce, and peas). Plants were grown 

in soil with 5 ng TEQ/kg or total CDD/CDF concentrations of 363 ng/kg dry weight (control plots) and 

56 ng TEQ/kg or total CDD/CDF concentrations of 3,223 ng/kg dry weight on the contaminated plots. 

CDD/CDF concentrations in carrot peels were three times higher on the contaminated plots than on the 

control plots. This was the result of a 10-fold increase in the CDD/CDF levels in the carrot peel. 

CDD/CDF concentrations in lettuce (17.7 and 21.1 ng/kg dry weight) and in peas (7.1 ng/kg dry weight) 

were not any higher when grown on the contaminated plot as compared to the control plots and were much 

lower than concentrations in the carrots (47.3 and 47.5 ng/kg, dry weight). This indicates that the 

CDD/CDFs in the lettuce and peas from both plots were of atmospheric origin. The CDD/CDF homologue 

pattern in the contaminated soil showed OCDFs and HpCDFs were the two most prevalent congeners, while 

the CDD/CDF homologue pattern from the peel of carrots grown on the contaminated plots contained 

TCDF, PeCDF, and HxCDF. Levels of TCDD were the lowest of all CDD/CDF homologues in both 

contaminated soils and carrot peels. The homologue profile in lettuce samples was largely dominated by 

lower chlorinated CDFs (TCDF and PeCDF) and higher chlorinated CDDs (HpCDD and OCDD), a 
CDDs 418 

5. POTENTIAL FOR HUMAN EXPOSURE 

profile often found in samples of atmospheric deposition (Eitzer and Hites 1989a, 1989b). The lowest 

CDD/CDF levels of this study were found in peas with pea pods showing higher levels than seeds. The 

homologue profiles was dominated by lower chlorinated CDFs and higher chlorinated CDDs similar to the 

profile found in lettuce. 

Since most of the CDDs released into the atmosphere settle onto water and soil surfaces, foliar deposition is 

the major route of vegetative contamination (Travis and Hattemer-Frey 1987). The translocation of foliarapplied 2,3,7,8-TCDD has been studied (Kearney et al. 1971). Labeled 2,3,7,8-TCDD was applied to the 

center leaflet of the first trifoliate leaf of 3-week-old soybean plants and the first leaf blade of 12-day-old 

oat plants. The compound was applied in an aqueous surfactant solution to enhance leaf adsorption and to 

keep the water-insoluble TCDD in solution. Plants were harvested 2, 7, 14, and 21 days after treatment, 

dissected into treated and untreated parts, and analyzed. 2,3,7,8-TCDD was not translocated from the 

treated leaf to other plant parts. Very little 2,3,7,8-TCDD was lost from soybean leaves, while a gradual 

loss (38% in 21 days) did occur from oat leaves (Kearney et al. 1971). The authors considered 

volatilization to be a possible mechanism for removal of 2,3,7,8-TCDD, but photolysis may also have 

contributed to the loss. 

McCrady and Maggard (1993) measured the uptake and elimination mechanisms for 2,3,7,8-TCDD applied 

to grass foliage in a closed-laboratory system using [3

H]TCDD. The [3H]2,3,7,8-TCDD was injected into 

the chamber as a vapor originating from a [3

H]2,3,7,8-TCDD generator. The total recovered radioactivity 

was 74%. Plant foliage accounted for 59% and the air and other chamber components accounted for 6 and 

9%, respectively. This indicated that plant foliage was a major sink for [3

H]2,3,7,8-TCDD vapor. Less than 

0.2% was recovered from the soil and associated with root tissues, further verifying an airborne mechanism 

of [3

H]2,3,7,8-TCDD uptake and negligible translocation. The authors also demonstrated that both 

photodegradation and volatilization were primary loss mechanisms for [3

H]2,3,7,8-TCDD. The 

photodegradation half-life (first-order kinetics) of 2,3,7,8-TCDD sorbed to grass and exposed to natural 

sunlight was 44 hours, while the half-life for volatilization of 2,3,7,8-TCDD from grass foliage was 

128 hours. 

In conclusion, CDDs may be transported long distances in the atmosphere. They eventually may be 

deposited on soils or surface water as a result of wet or dry deposition. CDDs will slowly volatilize from 

the water column or, more likely, will adsorb to suspended particulate materials in the water column and be 

transported to the sediment. CDDs deposited on soils will strongly adsorb to organic matter. They are 
CDDs 419 

5. POTENTIAL FOR HUMAN EXPOSURE 

unlikely to leach to underlying groundwater, but may enter the atmosphere on soil or dust particles or enter 

surface water in runoff. Low water solubilities and high lipophilicity indicate that CDDs will bioconcentrate in aquatic organisms, although as a result of their binding to suspended organic matter, actual 

uptake by these organisms may be less than predicted. This is also true of uptake and bioconcentration by 

plants, although foliar deposition and adherence may be significant. 

5.3.2 Transformation and Degradation 

CDDs belong to a class of highly lipophilic compounds with low water solubility and low chemical 

reactivity that are resistant to microbial degradation. The dominant transformation processes affecting their 

fate have been shown to be surface photolysis and gas-phase diffusion/volatilization with subsequent 

photolysis (Yanders et al. 1989). 

5.3.2.1 Air 

The primary transformation reaction for CDDs in the atmosphere depends on whether the CDD is in the 

vapor or particulate phase. Vapor-phase CDDs are not likely to undergo reactions with atmospheric ozone, 

nitrate, or hydroperoxy radicals; however, reactions with hydroxyl radicals may be significant, particularly 

for the less-chlorinated congeners (MCDD through TCDD) (Atkinson 1991). Based on the photolysis 

lifetimes of CDDs in solution, it is expected that vapor-phase CDDs will also undergo photolysis in the 

atmosphere, although reactions with hydroxyl radicals will predominate. For TCDD, the photolytic lifetime 

ranges from 1.3 to 7.1 days, depending on the season (faster in summer), whereas the hydroxyl radical 

reaction lifetime is estimated to be 2 days (Atkinson 1991). A half-life of 8.3 days was estimated for the 

gas-phase reaction of 2,3,7,8-TCDD with photochemically produced hydroxyl radicals in the atmosphere 

(Podoll et al. 1986). Using the gas-phase hydroxyl radical reaction rate constant of 1×10-11 cm3

-molecule-1 

sec-1 and an average 12-hour daytime hydroxyl radical concentration of 1.5×106 molecules cm-3, the 

atmospheric lifetimes of CDDs are estimated to range from 0.5 days for MCDD to 9.6 days for OCDD, with 

TCDD having a lifetime of 0.8–2 days (Atkinson 1991). 

Particulate-bound CDDs are removed by wet or dry deposition with an atmospheric lifetime $10 days 

(Atkinson 1991) and, to a lesser extent, by photolysis. Miller et al. (1987) measured photolysis of 

2,3,7,8-TCDD sorbed onto small-diameter fly ash particulates suspended in air. The results indicated that 

fly ash confers photostability to the adsorbed 2,3,7,8-TCDD. The authors reported little (8%) to no loss of 
CDDs 420 

5. POTENTIAL FOR HUMAN EXPOSURE 

2,3,7,8-TCDD on the fly ash samples after 40 hours of illumination in simulated sunlight. Koester and 

Hites (1992) studied the photodegradation of CDDs naturally adsorbed to five fly ash samples (two from 

coal-fired plants, two from municipal incinerators, and one from a hospital incinerator). Although the 

authors reported that CDDs underwent photolysis in solution and on silica gel, no significant degradation 

was observed in 11 photodegradation experiments conducted for periods ranging from 2 to 6 days. 

The selected transformation of the more and less chlorinated CDDs has been demonstrated by the analysis 

of CDDs found in soil samples compared with atmospheric concentrations of CDDs at the emission source 

(Marklund et al. 1991; Yamamoto and Fukushima 1993). Soil samples contained progressively greater 

concentrations of HpCDD and OCDD with increasing distance from the emission source, indicating that 

photolysis of the less chlorinated congeners was occurring (Eitzer 1993). In the air, the low vapor pressure 

of OCDD results in its partitioning primarily to the particulate phase rather than the vapor phase; therefore, 

atmospheric photodegradation is less likely to occur for this tightly bound congener (Eitzer 1993). 

5.3.2.2 Water 

Photolysis is the major route of CDD disappearance in aqueous solutions (Hutzinger et al. 1985). While 

photolysis is a relatively slow process in water, CDDs are rapidly photolyzed under certain conditions, (i.e., 

when exposed to ultraviolet light of the appropriate wavelength and in the presence of an organic hydrogen 

donor). These hydrogen donors can be expected to be present in chlorophenol pesticides either as 

formulation solvents (e.g., xylene or petroleum hydrocarbons), as active constituents of the formulation 

(e.g., the alkyl esters of 2,4-D and 2,4,5-T), or as natural organic films on soils (Crosby et al. 1973). The 

photolytic behavior of CDDs in an organic solvent or in a water-organic solvent, however, may not 

accurately reflect the photolytic behavior of these compounds in natural waters (Hutzinger et al. 1985). For 

example, Choudry and Webster (1989) reported that photolysis of 1,3,6,8-TCDD was slower in natural 

pond-water solutions than was predicted from studies with laboratory solutions. Conversely, Friesen et al. 

(1990) reported that photolysis of PeCDD and HpCDD proceeds faster in a pond or lake-water solutions 

than was predicted or measured in a laboratory solution. In general, however, lower chlorinated CDDs are 

degraded faster than higher chlorinated congeners. Chlorine atoms in the lateral positions (e.g., 2, 3, 7, 8) 

are also more susceptible to photolysis than are chlorine atoms in the para positions (e.g., 1, 4, 6, 9) 

(Choudhry and Hutzinger 1982; Crosby et al. 1973; Hutzinger et al. 1985). 
CDDs 421 

5. POTENTIAL FOR HUMAN EXPOSURE 

Podoll et al. (1986) used the quantum yield data of Dulin et al. (1986) for a water:acetonitrile solution to 

calculate seasonal half-life values for dissolved 2,3,7,8-TCDD at 40 degrees north latitude in clear nearsurface waters. Photolysis half-lives for dissolved 2,3,7,8-TCDD in sunlight range from 118 hours in 

winter, to 51 hours in fall, to 27 hours in spring, to 21 hours in summer (Podoll et al. 1986). Choudhry and 

Webster (1989) studied photolysis of a series of CDDs in a water:acetonitrile solution (2:1 v/v). These 

authors estimated the midday midsummer sunlight photolysis half-lives values at 40 degrees north latitude 

in clear near-surface waters as follows: 1,3,6,8-TCDD (0.3 days), 1,2,3,7-TCDD (1.8 days), 

1,2,3,4,7-PeCDD (15 days), 1,2,3,4,7,8-HxCDD (6.3 days), 1,2,3,4,6,7,8-HpCDD (47 days), and OCDD 

(18 days) near the surface of water bodies (Choudhry and Webster 1989). Sunlight photolysis half-lives 

were also reported for the spring, fall, and winter for 1,2,3,4,6,7,8-HpCDD (57, 88, and 156 days, 

respectively) and for OCDD (21, 31, and 50 days, respectively) (Choudhry and Webster 1989). Photolysis 

half-lives for 1,2,3,4,6,7,8-HpCDD and OCDD in water-acetonitrile solutions irradiated at 313 nm were 

reported to be 8 and 7.7 days, respectively (Choudhry and Webster 1987, 1989). The half-lives of 

1,3,6,8-TCDD and OCDD in lake water are 2.6 and 4 days, respectively, with removal by partitioning to the 

lake sediments (Servos et al. 1992). 

The photodegradation profiles of 2,3,7,8-TCDD, 1,3,6,8-TCDD, and 1,2,3,4-TCDD in 1,4-dioxane 

solutions at various wavelengths under xenon lamp irradiation were studied (Koshioka et al. 1989a, 1989b, 

1989c). Reductive dechlorination reactions were observed in the photolysis of TCDD isomers. After 

200 minutes of irradiation with a xenon lamp, 2,3,7,8-TCDD formed 2,3,7-TrCDD, 2,7-DCDD, 2,8-DCDD, 

2-MCDD, and DD. Photodegradation half-lives of 2,3,7,8-TCDD at the maximal photodegradation 

wavelengths of 252.6 nm and 318.6 nm were 72.6 minutes and 29.7 minutes, respectively (Koshioka et al. 

1989b, 1989c). After 267 minutes of irradiation with a xenon lamp, 1,3,6,8-TCDD formed 1,3,6-TrCDD, 

1,3-DCDD, 1,6-DCDD, 1-MCDD, 2-MCDD, and DD, while 1,2,3,4-TCDD formed 1,2,3-TrCDD, 

1,2,4-TrCDD, 1,2-DCDD, 1,3-DCDD, 1,4-DCDD, 2,3-DCDD, 1-MCDD, 2-MCDD, and DD (Koshioka et 

al. 1989a). 

The photolytic half-life of 2,3,7,8-TCDD in isooctane was estimated to be 40 minutes with a light source at 

0.5 meters and 3 hours with a light source at 1 meter (Stehl et al. 1973). Very little change was observed in 

OCDD on exposure to artificial sunlight. Approximately 20% photolysis of OCDD was observed in 

isooctane at the end of 18 hours and about 6% photolysis of OCDD after 20 hours of exposure in 1-octanol 

(Stehl et al. 1973). Irradiation of pentachlorophenol (PCP) dissolved in sodium hydroxide at a wavelength 
CDDs 422 

5. POTENTIAL FOR HUMAN EXPOSURE 

of 300 nm (equivalent to sunlight) for 16 hours produced OCDD (Crosby and Wong 1976). OCDD then 

underwent photoreduction to HpCDD as a PCP photolysis product (Crosby and Wong 1976). 

Under equivalent light exposure conditions, photolytic half-lives were determined for each of the individual 

TCDD isomers in dilute hydrocarbon solution and as a diffuse molecular dispersion on a clean soft-glass 

surface (Nestrick et al. 1980). The photolytic behavior of 2,3,7,8-TCDD was atypical compared to other 

TCDD isomers. In a hydrocarbon solution, 2,3,7,8-TCDD had the fastest decomposition rate (half-life 

56.8 minutes) and 1,4,6,9-TCDD had the slowest decomposition rate (half-life 8,400 minutes [5.8 days]). 

The half-lives of the remaining TCDD isomers ranged from 153 to 1,388 minutes (2.55–23.1 hours). 

However, as a diffuse molecular dispersion on a glass surface, the 2,3,7,8-TCDD had the slowest 

decomposition rate (half-life 8,400 minutes [5.8 days]), and 1,4,6,9-TCDD had the second slowest 

decomposition rate (half-life 830 minutes [13.8 hours]). The half-lives of the remaining TCDDs ranged 

from 121 to 560 minutes (2–9.3 hours). The majority of TCDD isomers photolytically decomposed faster 

on a glass surface than in a hydrocarbon solution under conditions of equivalent light intensity. 

2,3,7,8-TCDD and 1,4,6,9-TCDD possess the highest degree of symmetry within the group, and these 

isomers demonstrated the largest change in the photodecomposition rate for surface and solution reactions, 

with the changes being in opposite directions. Additional photolysis tests were conducted using more 

highly chlorinated CDD congeners. In a hydrocarbon solution, the half-lives of 1,2,3,4,6,7,8-HpCDD, 

1,2,3,4,6,7,9-HpCDD, and OCDD were 1,800 minutes (1.3 days), 3,300 minutes (2.3 days), and 

1,460 minutes (1.01 days), respectively, and 3,140 minutes (2.18 days), 2,400 minutes (1.67 days), and 

48,900 minutes (33.96 days), respectively, on a glass surface (Nestrick et al. 1980). 

2,3,7,8-TCDD decomposed rapidly when dissolved in methanol and exposed to ultraviolet (UV) light 

(Plimmer et al. 1973). Rate measurements showed that 2,3,7,8-TCDD is more rapidly photolyzed in 

methanol than OCDD (Plimmer et al. 1973). The photolysis half-lives for 2,3,7,8-TCDD, 

1,2,3,4,6,7,8-HpCDD, 1,2,3,4,6,7,9-HpCDD, and OCDD in n-hexadecane solution were 56.8 minutes, 

1,800 minutes (1.25 days), 3,300 minutes (2.29 days), and 1,460 minutes (1.01 days), respectively 

(Mamantov 1984). 

Solution-phase photolysis of HpCDD and OCDD has been reported (Dobbs and Grant 1979). Solutions of 

these CDDs in hexane (approximately 1 μg/mL) were exposed to natural sunlight as well as to fluorescent 

blacklight. The photolytic half-life for OCDD exposed to both types of radiation was 16 hours. HpCDD 

was generated by photolysis of OCDD (Dobbs and Grant 1979). The photolytic half-lives of 
CDDs 423 

5. POTENTIAL FOR HUMAN EXPOSURE 

1,2,3,4,6,7,9-HpCDD and 1,2,3,4,6,7,8-HpCDD were 28 hours and 11 hours, respectively (Dobbs and 

Grant 1979). 

It has been suggested that the potential for biological degradation of 2,3,7,8-TCDD in a wide variety of 

environmental samples is low (Arthur and Frea 1989). The fate of 2,3,7,8-TCDD in sediment and water 

from two lakes in Wisconsin was examined (Ward and Matsumura 1978). After incubation periods of up to 

589 days, little metabolism of 2,3,7,8-TCDD was detected. The slight metabolism that was detected was 

stimulated by the presence of sediment and the addition of nutrients (Ward and Matsumura 1978). Also, 

2,3,7,8-TCDD does not hydrolyze in water (Mabey et al. 1982; Miller et al. 1987). 

5.3.2.3 Sediment and Soil 

Photolysis of 2,3,7,8-TCDD on soils is a relatively slow process compared to photolysis in an aqueous 

media (Kieatiwong et al. 1990). 2,3,7,8-TCDD applied to soil or a solid surface seems to be extremely 

resistant to the action of sunlight and decomposes very slowly (Plimmer et al. 1973). A methanol solution 

of 2,3,7,8-TCDD (2.4 ppm) applied to glass plates coated with soil and illuminated 96 hours with a 

fluorescent UV lamp remained unchanged at the end of the period (Plimmer et al. 1973). Organic solvents 

added to the soil, however, can enhance the extent of photolysis. Use of a solvent mixture of tetradecane 

and 1-butanol to TCDD-treated soil, combined with exposure to sunlight, resulted in 61–85% 

photodegradation of TCDD after 60 days. The solvent was effective in transporting TCDD from deeper in 

the soil column (60 cm) to the soil surface via evaporation. At the soil surface, photodegradation could 

occur. TCDD concentrations at 60 cm decreased from 23.8 ng/g (ppb) to 7.1 ng/g (ppb) after 60 days 

(McPeters and Overcash 1993). 

Photolysis of OCDD (10 mg/kg) on soils resulted in production of the lower chlorinated CDDs, notably 

2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, three HxCDD isomers substituted at the 2,3,7,8-positions, and 

1,2,3,4,6,7,8-HpCDD. Photolysis of OCDD occurred in mean soil depths between 0.06 and 0.13 mm 

(Miller et al. 1989). Approximately 30–45% of OCDD was lost by day 5 of irradiation; no further 

significant loss of OCDD was observed following 10 additional days of irradiation. Although photolysis 

only occurred at shallow soil depths and the conversion of OCDD to the more toxic TCDD, PeCDD, and 

HxCDD homologues was small (0.5–1%) compared with the photodechlorination to HpCDD (67%), 

photolysis of OCDD may represent a significant source of these toxic isomers (Miller et al. 1989). 
CDDs 424 

5. POTENTIAL FOR HUMAN EXPOSURE 

The loss of 2,3,7,8-TCDD in contaminated soil has been studied under natural conditions in experimental 

plots at the Dioxin Research Facility, Times Beach, Missouri (Yanders et al. 1989). The 2,3,7,8-TCDD 

concentration profiles of sample cores taken at Times Beach in 1988 were virtually the same as those in 

cores taken in 1984. The authors concluded that the loss of 2,3,7,8-TCDD due to photolysis at Times 

Beach was minimal in the 4 years covered by the study (Yanders et al. 1989). Estimates of the half-life of 

TCDD on the soil surface range from 9 to 15 years, whereas the half-life in subsurface soil may range from 

25 to 100 years (Paustenbach et al. 1992). 

A white rot fungus (Phanerochaete chrysosporium) has demonstrated the ability to degrade 2,3,7,8-TCDD 

in laboratory experiments (Bumpus et al. 1985; Des Rosiers 1986). In cultures containing 1.25 nmol of the 

2,3,7,8-TCDD substrate, 27.9 pmol were mineralized to CO2 in 30 days (2.23% metabolism) increasing to 

49.5 pmol in 60 days (3.96% metabolism) (Des Rosiers 1986). It was suggested that the ability of this 

fungus to metabolize 2,3,7,8-TCDD is dependent on its extracellular lignin-degrading enzyme system 

(Bumpus et al. 1985; Des Rosiers 1986). More recently, Valli et al. (1992) reported that 2,7-DCDD also 

was degraded by P. chrysosporium via the removal of both aromatic chlorines before aromatic ring 

cleavage took place. 

Cultures of Pseudomonas testosteroni, of an unidentified bacterium isolated from soil from Seveso, Italy, 

and of a mixture of 6 unidentified bacterial strains isolated from Seveso soil were incubated aerobically 

with 14C-2,3,7,8-TCDD for 35, 54, and 12 weeks, respectively (Philippi et al. 1982). Results showed the 

occurrence of a metabolite of 14C-2,3,7,8-TCDD in all three cultures. The polar metabolite amounted to 

approximately 1% of the input material and was found to be a hydroxylated derivative of 14C-2,3,7,8-TCDD 

(Philippi et al. 1982). 

Approximately 100 strains of pesticide-degrading microorganisms were tested for their ability to degrade 

2,3,7,8-TCDD (Matsumura and Benezet 1973). The organisms were maintained in liquid axenic culture, 

and the production of metabolites from ring-labeled 14C-2,3,7,8-TCDD was measured. Five strains were 

identified that showed some ability to degrade 14C-2,3,7,8-TCDD. The degradative organisms included a 

fungus (Trichoderma viride), a bacterium (Pseudomonas putida), and three organisms referred to by coded 

numbers (Matsumura and Benezet 1973). 

To determine the persistence of 2,3,7,8-TCDD, concentrations of 1, 10, and 100 ppm of unlabeled 

2,3,7,8-TCDD were added to 300-g samples of silty loam and sandy soils and then assayed periodically for 
CDDs 425 

5. POTENTIAL FOR HUMAN EXPOSURE 

residues (Kearney et al. 1971). Measurements of 2,3,7,8-TCDD residues after 20, 40, 80, 160, and 

350 days of incubation at 28 EC in foil-sealed beakers indicated a relatively slow degradation process in 

both soils. After 350 days, 56% of the initially applied 2,3,7,8-TCDD was recovered from the sandy soil, 

while 63% was recovered from the silty clay loam for all concentrations (Kearney et al. 1971). 

Parsons (1992) studied the influence of suspended sediment on the biodegradation of several CDDs. In this 

study, aqueous solutions of a mixture of 2-chloro-, 1,3-dichloro, 2,8-dichloro-, and 1,2,4-trichloro CDDs 

were incubated for 24 days with 100 mg/L suspended sediment. Subsequently, the degradation of the 

CDDs in the sediment suspensions by Alcaligenes sp. strain JB1 was compared to that in solutions without 

sediment. The amounts of all four CDD compounds degraded in the sediment suspensions after 7 days 

were greater than those initially present in the dissolved phase, based on their calculated sediment-water 

partition coefficients. The sorbed fractions were, therefore, sufficiently desorbed to be partly degraded. 

However, the biodegradation rates were slower in the sediment suspensions than in the solutions. The 

results indicate that sorbed fractions of CDDs formed after relatively short incubation periods are 

sufficiently labile to be available for biodegradation after desorption. Evidence that the presence of 

sediment lowers biodegradation rates in sediment suspension, however, implies that longer residence times, 

such as those observed under field conditions, may also lead to a significant lowering of the biodegradation 

rates in soil. This will apply even more to the more highly chlorinated CDD congeners. In another study, 

the degradation of highly chlorinated CDD congeners (5–7 chlorine/molecule) was studied for a period of 

6 months in anaerobic microcosm incubations using PCB-contaminated Hudson River sediments and 

creosote-contaminated aquifer samples from Pensacola, Florida (Adriaens and Grbic-Galic 1994). The 

authors reported (pseudo-first-order) half-life values for 1,2,3,4,6,7,8-HpCDD of 4.1 and 2.9 years for the 

Hudson River and Pensacola aquifer-incubated microcosm samples, respectively. The half-life values for 

1,2,3,4,7,8-HxCDD were 2 and 2.9 years for the Hudson River and Pensacola aquifer-incubated microcosm 

samples, respectively. The 1,2,4,6,8,9/ 1,2,4,6,7,9-HxCDD congeners were found not to be degraded, 

which was presumably due to the low concentration spiked. The authors reported that tentative 

identification of the degradation products indicate that para-dechlorination was the preferential route of 

reduction, as has been observed with 1,2,3,4,5,6,7,8-HpCDD in aquifer microcosms. This observation is 

contrary to photolytic dechlorination patterns of soil-sorbed CDDs. 

Beurskens et al. (1995) reported that an anaerobic microbial consortium enriched from Rhine River 

sediments was able to remove chlorine substituents from CDDs. A model CDD, 1,2,3,4-TCDD, was 

reductively dechlorinated to both 1,2,3- and 1,2,4-TrCDD. These TrCDD compounds were further 
CDDs 426 

5. POTENTIAL FOR HUMAN EXPOSURE 

dechlorinated to 1,3- and 2,3- DCDD and trace amounts of 2-MCDD. The TrCDD compounds were 

detected at low concentrations, but the 1,3- and 2,3- DCDD were detected at higher concentrations. The 

anaerobic culture dechlorinates 1,2,3,4-TCDD at a relatively rapid rate with a half-life value estimated at 

15.5 days (first-order kinetics). The formation of metabolites with a conserved 2,3-substitution pattern from 

1,2,3,4-TCDD indicates that dechlorination of highly chlorinated CDDs may result in metabolites that are 

potentially more toxic than the parent compounds. 

5.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT 

Reliable evaluation of the potential for human exposure to CDDs depends in part on the reliability of 

supporting analytical data from environmental samples and biological specimens. Historically, CDD 

analysis has been both complicated and expensive, and the analytical capabilities to conduct such analysis 

have been available through only a relatively few analytical laboratories. Limits of detection have 

improved greatly over the past decade with the use of high-resolution mass spectrometry, improvements in 

materials used in sample clean-up procedures, and with the use of known labeled and unlabeled chemical 

standards. Problems associated with chemical analysis procedures of CDDs in various media are discussed 

in greater detail in Chapter 6. In reviewing data on CDD levels monitored or estimated in the environment, 

it should be noted that the amount of the chemical identified analytically is not necessarily equivalent to the 

amount that is bioavailable (see Section 2.3) and that every measurement is accompanied with a certain 

analytical error. 

5.4.1 Air 

Indoor household dust samples gathered by a vacuum cleaner from rooms with furniture treated with a 

wood-preserving formulation were analyzed for CDDs (Christmann et al. 1989b). The wood-preserving 

formulation contained PCP, which is known to be contaminated with CDDs, particularly HxCDD, HpCDD, 

and OCDD. OCDD was the most abundant congener found in the dust samples at an average concentration 

of 191 μg/kg (ppb), followed by HpCDD (20 μg/kg), HxCDD (2.5 μg/kg), PeCDD (0.9 μg/kg), and TCDD 

(0.2 μg/kg) (Christmann et al. 1989b). 

Indoor air concentrations of CDD/CDFs were measured in kindergarten classrooms in West Germany to 

evaluate releases from wood preservatives (e.g., PCP) that may have been used in building materials (Päpke 

et al. 1989a). Measured indoor air concentrations of total CDD/CDF ranged from 1.46 to 
CDDs 427 

5. POTENTIAL FOR HUMAN EXPOSURE 

4.27 pg/m3, while measured outdoor air concentrations ranged from 0.61 to 78.97 pg/m3. The 2,3,7,8­

substituted congeners predominated with mean concentrations as follows: OCDD (131.5 pg/m3

), 

1,2,3,4,6,7,8-HpCDD (77 pg/m3

), 1,2,3,4,6,7,8-HpCDF (51 pg/m3), and OCDF (25.3 pg/m3). 

Measured indoor air samples collected in an office building in Binghamton, New York, 2 years after a fire 

in an electrical transformer that contained PCBs and tri- and tetra-chlorobenzenes had concentrations of 

2,3,7,8-TCDD ranging from 0.23 to 0.47 pg/m3

 (0.017–0.036 ppq) (Smith et al. 1986a). The 2,3,7,8-TCDD 

isomer constituted 23–30% of the 1.0–1.3 pg/m3

 (0.076–0.099 ppq) total TCDDs. The limit of detection for 

these samples was approximately 0.003 pg/m3

 (Smith et al. 1986a). 

Background levels of CDD in air were measured in a semi-rural location in Elk River, Minnesota, located 

about 25 miles northwest of Minneapolis-St. Paul (Reed et al. 1990). No major industrial or commercial 

activity occurred in the area at the time of the study. Ambient air samples were collected in the winter and 

summer of 1988. 2,3,7,8-TCDD was not detected in any of the ambient air samples taken in the summer 

(detection limits for 2,3,7,8-TCDD ranged from 0.005 to 0.065 pg/m3

 [0.0004–0.0046 ppq]). 

2,3,7,8-TCDD was noted in a wintertime sample at concentrations of 0.015 pg/m3

 (0.0011 ppq) and 

0.019 pg/m3

 (0.0014 ppq). Detection limits in the remaining wintertime samples for 2,3,7,8-TCDD ranged 

from 0.005 to 0.01 pg/m3

 (0.0004–0.0007 ppq). Wintertime CDD concentrations were greater than those 

observed for summertime. The authors noted that this may be a result of increased numbers of combustion 

sources operating during the winter months. The wintertime CDD congener profile showed increasing 

concentrations with increasing chlorine substitutions. Average wintertime ambient air concentrations of 

HpCDD and OCDD ranged from approximately 0.5 to 4.1 pg/m3

 (0.029–0.236 ppq) and 0.74 to 8.2 pg/m3 

(0.039–0.436 ppq), respectively (Reed et al. 1990). Average summertime ambient air concentrations of 

HpCDD and OCDD ranged from approximately 0.204 to 0.246 pg/m3

 (0.011–0.014 ppq) and 0.018 to 

0.024 pg/m3

 (0.001–0.0013 ppq), respectively (Reed et al. 1990). The authors found that, in general, the 

more highly chlorinated congeners were present at higher concentrations that the less chlorinated 

congeners. 

A long-term study (1985–1988) of CDDs in the ambient atmosphere of Bloomington, IN (a suburban area), 

was carried out in order to provide base-line data against which the impact of a future incinerator on local 

CDD concentrations could be judged (Eitzer and Hites 1989b). Ambient air samples were analyzed for the 

presence of CDDs in both the particulate-bound phase and the vapor-phase forms. At the four sites 

sampled, the concentrations of CDDs (TCDD, PeCDD, HxCDD, HpCDD, and OCDD) increased with an 

increasing level of chlorination. All sites showed that the less chlorinated CDDs have a higher vapor-phase 
 

CDDs 428 

5. POTENTIAL FOR HUMAN EXPOSURE 

fraction than the more chlorinated CDDs. In addition, all sites show OCDD to be the most abundant CDD, 

averaging from 0.44 to 0.69 pg/m3

 (0.023–0.032 ppq) (detection limit 0.001 pg/m3 [5.3×10-5 ppq]) (Eitzer 

and Hites 1989b). A seasonal effect was seen on the proportion of the total atmospheric burden present in 

the vapor phase. During the warm summer months, the total vapor-to-particle bound ratio (V/P) was as 

great as 2, whereas in the winter it was <0.5. At warm temperatures, most of the less chlorinated CDDs are 

found in the vapor phase, whereas at cooler temperatures more of the CDDs were associated with the 

particle phase (Eitzer and Hites 1989b). 

CDDs have been found in urban air particulates from Washington, D.C., and St. Louis, Missouri; OCDD 

was the predominant congener at concentrations of 200 ppb and 170 ppb for Washington and St. Louis, 

respectively (Czuczwa and Hites 1986a). Combustion of municipal and chemical wastes was the most 

likely source of these compounds. CDDs were detected in air samples from Albany, Binghamton, Utica, 

and Niagara Falls, NY (Smith et al. 1990b). Concentrations of CDD congener groups for all 4 cities were 

as follows: total TCDD, not detected (<0.21 pg/m3

 [0.016 ppq]); total PeCDD, <0.04–0.62 pg/m3 

(<0.003–0.043 ppq); total HxCDD, 0.10–2.4 pg/m3 (0.007–0.15 ppq); total HpCDD, <0.21–4.4 pg/m3 

(0.012–0.25 ppq); and OCDD <0.54–4.6 pg/m3 (0.029–0.244 ppq) (Smith et al. 1990b). In 1988–89, total 

CDDs measured downwind from an industrial source in Niagara Falls, NY, ranged from 0.3 pg/m3

 to 

133 pg/m3

 and were approximately 2.5 times higher than upwind concentrations (Smith et al. 1990b). 

Between 1986 and 1990, total CDD concentrations averaged 2.3 pg/m3

, of which 65% was OCDD (Smith et 

al. 1992). 

An extensive multi-year monitoring program for CDDs/CDFs was conducted at eight sampling locations in 

the Los Angeles South Coast Air Basin from 1987 to 1989 (Hunt and Maisel 1992). The monitoring 

network, which monitored for both vapor and particulates, included several sites situated in residential areas 

as well as sites in the vicinity of suspected CDD/CDF sources. Monitoring results indicated that 

2,3,7,8-TCDD was virtually undetected. The most commonly detected 2,3,7,8-substituted congener was 

OCDD followed by 1,2,3,4,6,7,8-HpCDD. The predominance of 1,2,3,4,6,7,8-HpCDD as the most 

persistent congener is associated with stationary or mobile combustion source emissions. 

1,2,3,4,6,7,8-HpCDD was found at all 7 sampling sessions at a mean concentration of 1.140 pg/m3

. OCDD 

also was found at all 7 sampling sessions at a mean concentration of 2.883 pg/m3

. The mean total TCDD 

concentration was 0.114 pg/m3

 and was measured during only 3 sampling sessions (Hunt and Maisel 1992). 
CDDs 429 

5. POTENTIAL FOR HUMAN EXPOSURE 

The concentrations of CDDs in the ambient air at several sites in metropolitan Dayton, Ohio, have been 

determined (Tiernan et al. 1989b). No CDDs (TCDD, PeCDD, HxCDD, HpCDD, and OCDD) were found 

in rural regions, with average detection limits ranging from 0.03 pg/m3

 (TCDD) to 1.44 pg/m3 (OCDD). 

The rural area was outside the impact zone of air pollutants from any regional industrial sources. CDDs in 

the industrialized regions appear to originate from a combination of sources, including municipal waste 

incinerators, motorized vehicles, and a PVC-coated metal incinerator, the latter being a major source of 

these pollutants. Suburban/roadside area samples were taken at ground level at a distance of about 3 meters 

from a street intersection through which approximately 60,000 cars passed each day. Other sampling 

sources were on the roofs of buildings in the downtown Dayton area, which lay in the emissions path from 

municipal solid-waste incinerators. TCDDs and PeCDDs (detection limits 0.01 and 0.03 pg/m3

, 

respectively) were not detected in the suburban/roadside area but were detected in the municipal wasteincinerator areas at 0.24 and 0.38 pg/m3

, respectively. HpCDD was detected in both the suburban/roadside 

areas and the municipal waste-incinerator areas at concentrations of 0.41 pg/m3

 (0.024 ppq) and 3.34 pg/m3 

(0.19 ppq), respectively. OCDD was also detected in the suburban/roadside areas (1.09 pg/m3 [0.058 ppq]) 

and the municipal waste incinerator areas (4.69 pg/m3

 [0.25 ppq]). Concentrations of HxCDD were lower 

than HpCDD and OCDD, 0.05 pg/m3

 (0.003 ppq) in the suburban/ roadside areas and 2.56 pg/m3 

(0.160 ppq) in the vicinity of the municipal waste incinerators (Tiernan et al. 1989b). 

Air samples were collected in Ohio in 1987 at an industrial area, an urban area downwind of a municipal 

incinerator, a high-traffic density area, and a rural area (Edgerton et al. 1989). No 2,3,7,8-TCDD was 

detected in any of the air samples with detection limits of <0.24 pg/m3

 (0.02 ppq) in any of the areas. The 

ambient concentrations of CDDs collected in the urban area were as follows: total HpCDD, 1.0–1.1 pg/m3 

(0.058–0.063 ppq); OCDD, 1.0–1.2 pg/m3 (0.053–0.064 ppq); PeCDD, 0.1 pg/m3 (0.03 pg/m3); and total 

HxCDD, 0.6–0.63 pg/m3

 (0.038–0.039 ppq) (detection limit not specified). Concentrations of CDDs in the 

industrial area were: total HpCDD, 0.41–1.0 pg/m3

 (0.024–0.058 ppq), OCDD, 0.51–1.1 pg/m3 

(0.027–0.058 ppq), and total HxCDD, 0.43–0.78 pg/m3 (0.027–0.049 ppq). Concentrations of total 

HpCDD, OCDD, total HxCDD in the high-traffic density area were 0.56 pg/m3

 (0.032 ppq), 0.96 pg/m3 

(0.051 ppq), and 0.15 pg/m3 (0.008 ppq), respectively. Ambient air concentrations of total HpCDD, 

OCDD, and total HxCDD in the rural area were 0.48 pg/m3

 (0.028 ppq), 0.5 pg/m3 (0.027 ppq), and 

0.33 pg/m3

 (0.021 ppq), respectively. PeCDD was not detected in the industrial, high-traffic, or rural areas 

(Edgerton et al. 1989). 
CDDs 430 

5. POTENTIAL FOR HUMAN EXPOSURE 

Air monitoring at Windsor, Ontario, downwind of a proposed municipal solid-waste incinerator in Detroit, 

Michigan, between 1987 and 1988 found a mean total CDD concentration of 2.12 pg/m3

. A sampling 

station located in a rural area 30 miles away provided background total CDD concentrations of 0.51 mg/m3

. 

At both stations, the primary congeners were HpCDD and OCDD in the particulate phase, whereas TCDD 

and PeCDD were not detected in the vapor or particulate phases above the detection limit (Bobet et al. 

1990). 

A mixture of CDDs (TCDD, PeCDD, HxCDD, HpCDD, OCDD) has been found in emissions from the 

combustion of various sources, including municipal incinerators, power plants, wood burning, househeating systems, and petroleum-refining operations (Chiu et al. 1983; Clement et al. 1985; Thoma 1988; 

Thompson et al. 1990). CDDs were found in stack and fly ash samples from the following combustion 

sources (ranges given): municipal incinerator, 8 ppb (OCDD) to 390 ppb (HxCDD) (TCDD was found at 

10 ppb); open-air burning of PCP-treated wood, 2 ppb (TCDD) to 187 ppb (OCDD); coal-fired power plant, 

1 ppb (TCDD) to 6 ppb (PeCDD and HxCDD); hydroelectric power plant, 0.5 ppb (OCDD) to 5.2 ppb 

(TCDD) (Chiu et al. 1983); and petroleum refining, 0.8 (OCDD) to 3.4 ng/m3

 (PeCDD) (Thompson et al. 

1990). Samples of ash from wood-burning stoves, a fireplace, and open-air wood burning contained 

detectable levels of CDDs ranging from 0.3 to 33 ppb (Clement et al. 1985). The open-air burning ash 

contained the highest total CDD concentration (33 ppb), with HpCDD being the most abundant homologue 

(11 ppb). The total CDD concentrations in 4 samples from wood-burning stoves ranged from 0.3 to 15 ppb, 

with the relative amounts of each homologue varying for each sample. Ash samples from the fireplace 

contained total CDD concentrations ranging from 3.1 to 5.4 ppb, with HxCDD (0.3–1.7 ppb) and OCDD 

(0.4–3.1 ppb) being the most abundant homologues present (Clement et al. 1985). TCDD was present in 

ash samples from open-air burning (0.8 ppb) and was detected in ash from the fireplace. 

Ambient air monitoring in the vicinity of a Superfund clean-up site detected 2,3,7,8-TCDD levels on the 

order of 1 pg/m3

 (0.08 ppq) (Fairless et al. 1987). The surface and subsurface soils at the site were tested 

and found to contain 2,3,7,8-TCDD at concentrations above 1 ppb at most locations within the site. 

2,3,7,8-TCDD has been detected in air samples (concentrations unspecified) collected at 9 of the 91 NPL 

hazardous waste sites where it has been detected in some environmental media (HazDat 1998). CDDs have 

been detected in air samples (concentrations unspecified) collected at 10 of the 126 NPL sites where they 

have been detected in some environmental media. TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD 
CDDs 431 

5. POTENTIAL FOR HUMAN EXPOSURE 

have been detected in air samples (concentrations unspecified) at 10, 3, 3, 3, and 1 sites of the 105, 34, 43, 

49, and 53 sites, respectively, where they have been detected in some environmental media (see Table 5-1). 

In conclusion, most of the measurements of CDDs in air tend to be very close to current detection limits. 

CDDs are found at the greatest concentrations in urban air with OCDD being the most prevalent congener 

(up to 0.100 ppq), HpCDDs being the next most common congener, and 2,3,7,8-TCDD being the least 

common congener (0.014 ppq). Concentrations of all CDDs are highest in the air near industrial areas. 

Rural areas usually have very low or unquantifiable levels of all CDDs. In urban and suburban areas, 

concentrations of CDDs may be greater during colder months of the year when furnaces and wood stoves 

are used for home heating. 

5.4.2 Water 

Precipitation samples collected in a rural location (Dorset, Ontario) over an 8-month period between 1986 

and 1987 were analyzed for CDDs (Tashiro et al. 1989a, 1989b). No TCDDs were found in any samples at 

detection limits of 4–30 ppq. OCDD concentrations were found in 3 samples in the 60–1,200 ppq range. 

Lower concentrations of HpCDD (70 ppq) were also found (Tashiro et al. 1989a). Precipitation samples 

were also collected in 1987–88 in urban and rural locations in Canada (Tashiro et al. 1989b). Varying 

levels of OCDD were detected throughout the sampling period, mainly at the rural location. OCDD was the 

only CDD detected at the rural site. OCDD concentrations ranged from 35 to 230 ppq, with the median 

value being slightly below 100 ppq. No seasonal pattern of OCDD concentrations was observed. OCDD 

was detected in only 2 of the urban precipitation samples at concentrations of 33 and 15 ppq (Tashiro et al. 

1989b). Rain collected at Bloomington, IN, between June 1987 and July 1988 showed low concentrations 

of total CDDs, although OCDD was the most prominent congener in all samples at concentrations ranging 

from below the detection limit of 0.1 pg/L to 220 pg/L. Total TCDD was detected in only 3 of 28 samples 

at concentrations <9 pg/L (EPA 1991d). 

An analysis of EPA's STORET (STOrage and RETrieval) database for 1980–82 showed that based on the 

statistical criteria used, 2,3,7,8-TCDD was detected but at concentrations too low to be quantified in 

surface-water samples collected at sampling sites (Staples et al. 1985). The sampling sites in the STORET 

database included both ambient and pipe sites. Ambient sites included streams, lakes, ponds, wells, 

reservoirs, canals, estuaries, and oceans and were intended to be indicative of general U.S. waterway 

conditions. Pipe sites referred to municipal or industrial influents or effluents (Staples et al. 1985). 
CDDs 432 

5. POTENTIAL FOR HUMAN EXPOSURE 

Treated effluents from various Ontario pulp and paper plants using either the bleached kraft (8 mills) or 

sulfite bleaching process (2 mills) were analyzed for CDDs (Clement et al. 1989). 2,3,7,8-TCDD was not 

detected in any of the effluent samples with detection limits ranging from 0.07 to 0.7 ppt. A few samples 

contained a TCDD isomer (not 2,3,7,8-TCDD) at concentrations ranging from 0.06 to 0.12 ppt. PeCDD 

(0.07 ppt) was detected in one effluent sample, and OCDD (0.05–0.79 ppt) was detected in 4 effluent 

samples. Suspended particulates were collected from the final effluent from two plants. 2,3,7,8-TCDD and 

OCDD were detected in the particulates at a concentration range of 200–660 ppt and 180–210 ppt, 

respectively. The concentration of 2,3,7,8-TCDD determined in the particulates represents levels in the 

final effluent of 5–10 ppq, suggesting that 2,3,7,8-TCDD is associated with suspended particulate materials 

in the effluents (Clement et al. 1989). 

In 1983, Jobb et al. (1990) conducted a survey of 49 drinking water supplies in Ontario, Canada, including 

supplies in the vicinity of chemical plants and pulp and paper mills. OCDD was detected in 36 of 

37 positive samples ranging from concentrations of 9 to 175 ppq in raw samples (33 positive samples) and 

from 19 to 46 ppq in treated (filtered) water samples (4 positive samples). These low concentrations were 

found primarily in water obtained downstream of industrial areas in the St. Clair/Detroit River system. 

Concentrations of 2,3,7,8-TCDD were not detected in any sample. Because CDDs are hydrophobic, 

concentrations of these compounds in water tend to be adsorbed onto particulate matter in water. 

Conventional water treatment processes are expected to be effective in removing the CDDs along with the 

particulates. This is substantiated by the fact that only 4 of the 37 positive detections were found in treated 

drinking water, while 33 detections were found in raw water samples. 

During 1986, a survey of 20 community water systems throughout the state of New York was conducted to 

evaluate CDD/CDF concentrations (Meyer et al. 1989). The sampling sites selected were representative of 

major surface water sources in the state used to obtain drinking water. The sites included surface water 

sources receiving industrial discharges and those known to contain CDD-contaminated fish, as well as 

water sources from more remote areas. Raw water sampled at the Lockport, NY, facility contained 

concentrations of TCDDs (1.7 ppq) as well as concentrations of TCDFs to OCDFs (18, 27, 85, 210, and 

230 ppq, respectively). These data show that the CDF congener group concentrations increased with 

increasing chlorine numbers. TCDFs were also detected in finished water sampled at the Lockport facility 

(duplicate samples contained 2.1 and 2.6 ppq). Except for a trace of OCDF detected at one other location, 

no other CDDs/CDFs were detected in finished water at any of the other 19 community water systems 

surveyed. 
CDDs 433 

5. POTENTIAL FOR HUMAN EXPOSURE 

Utility telecommunication and railway right-of-ways may be contaminated by leaching of CDDs associated 

with chlorophenol-treated railway ties and utility poles. A study in British Columbia showed that CDDs 

and CDFs were not detected in parkland ditch water (control area), but were detected in farmland, utility, 

and railway right-of-way ditch water (Wan and van Oostdam 1995). Total mean CDD concentrations 

(mainly OCDD and HpCDD) measured in farm ditch water, and railway ditch water, without and with 

utility poles were 2.22 μg/L, 45 μg/L, and 9,627 μg/L respectively. Mean total concentrations of CDDs 

were much higher in ditch water adjacent to utility poles (13,142 ng/L) than in ditch water 4 meters 

downstream (4,880 ng/L) or 4 meters upstream of the utility poles (2.72 ng/L). The authors concluded 

that utility poles and railway ties are a potential constant source of CDD/CDF contamination to both water 

and sediment in aquatic environment through ditch runoff. 

2,3,7,8-TCDD has been detected in surface water samples collected at 9 sites of the 91 NPL hazardous 

waste sites where it has been detected in some environmental media (HazDat 1998). CDDs have been 

detected in surface water samples collected at 14 of the 126 NPL sites where they have been detected in 

some environmental media. TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD have been detected in 

surface water samples at 10, 1, 4, 4, and 6 sites of the 105, 34, 43, 49, and 53 NPL sites, respectively, where 

they have been detected in some environmental media (see Table 5-1). 

Groundwater in the vicinity of an abandoned wood treatment facility was sampled from monitoring wells 

constructed at depths ranging from 6.1 to 30.5 meters and was analyzed for CDDs in January 1984 (Pereira 

et al. 1985). Concentrations of HxCDD, HpCDD, and OCDD in groundwater samples taken from wells at a 

depth of 6.1 meters were 61 ppt, 1,500 ppt, and 3,900 ppt, respectively. The authors noted that the high 

concentrations of CDDs in the sample from a depth of 6.1 meters probably resulted from the presence of 

microemulsions of oil that were difficult to separate from the sample. Groundwater samples collected from 

deeper wells (12.2–30.5 meters) contained HxCDD, HpCDD, and OCDD at concentration ranges of not 

detected to 21 ppt, not detected to 34 ppt, and not detected to 539 ppt, respectively (Pereira et al. 1985). 

2,3,7,8-TCDD has been detected in groundwater samples collected at 15 sites of the 91 NPL hazardous 

waste sites where it has been detected in some environmental media (HazDat 1998). CDDs have been 

detected in groundwater samples collected at 32 of the 126 NPL sites where they have been detected in 

some environmental media. TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD have been detected in 
 

CDDs 434 

5. POTENTIAL FOR HUMAN EXPOSURE 

groundwater samples at 21, 3, 10, 14, and 16 sites of the 105, 34, 43, 49, and 53 NPL sites, respectively, 

where they have been detected in some environmental media (see Table 5-1). 

In conclusion, CDDs are rarely detected in drinking water at ppq levels or higher. Raw water samples 

generally have higher concentrations of CDDs (9–175 ppq) than finished drinking water samples 

(19–46 ppq) because conventional water treatment processes remove the CDDs along with the particulates 

from raw water. CDDs have been detected in treated effluent samples collected at pulp and paper mills 

using the bleach kraft or sulfite bleaching process. In groundwater samples collected near industrial sites, 

CDDs have been detected at concentrations up to 3,900 ppt. 

5.4.3 Sediment and Soil 

Soil. As part of this National Dioxin Study, EPA conducted a 2-year nationwide study to assess the 

extent of 2,3,7,8-TCDD contamination (EPA 1987n). Environmental samples (including soil, sediment, 

water, and fish) were analyzed for 2,3,7,8-TCDD concentrations at seven different tiers of sites (including 

NPL, various industrial, urban, and pristine rural sites). Soil concentrations at most of the Tier 1 and 2 sites 

(i.e., sites classified as or expected to be classified as NPL sites) were in the ppb range, although at a few of 

the sites where 2,4,5-TCP production waste storage or disposal occurred, concentrations were as high as 

2,000 ppm. Offsite soil contamination of concern (in the ppb range) was confirmed at 7 of these 100 Tier 

1 and 2 sites. At 11 of 64 Tier 3 sites (facilities and associated disposal sites where 2,4,5-TCP and its 

derivatives were formulated into pesticide products), soil concentrations exceeded 1 ppb, but in 7 of the 

11 sites where contamination was found, only 1 or 2 samples exceeded 1 ppb. At 15 of 26 Tier 5 sites 

(areas where 2,4,5-TCP and other pesticide derivatives had been or were currently being used), soil 

concentrations were generally above 1 ppt with one detection at 6 ppb. Two-thirds of all detections at the 

Tier 5 sites were below 5 ppt. At 3 of 18 Tier 6 sites (organic chemical and pesticide manufacturing 

facilities where production processes could have resulted in 2,3,7,8-TCDD being introduced into the waste 

streams), soil concentrations exceeded the 1 ppt detection limit, although these concentrations were limited 

to one or two samples per site. In general, 2,3,7,8-TCDD was detected infrequently and at very low 

concentrations in background soil samples taken at sites (urban and rural areas) that did not have previously 

known sources of 2,3,7,8-TCDD contamination (1 ppt detection limit). Only 17 of 221 urban sites and 1 of 

138 rural sites in Tier 7 (background sites not expected to have contamination) had detectable levels of 

2,3,7,8-TCDD, with 11.2 ppt being the highest concentration reported (Des Rosiers 1987; EPA 1987n). 
CDDs 435 

5. POTENTIAL FOR HUMAN EXPOSURE 

Background levels of CDDs in soil were measured at Elk River, Minnesota, a semi-rural area located about 

25 miles northwest of Minneapolis-St. Paul (Reed et al. 1990). No major industrial or commercial activity 

occurred in the area at the time of the study. The soil data reflected generally low background 

concentrations of CDDs. 2,3,7,8-TCDD, total TCDD, and PeCDD were not detected (detection limit range 

0.75–2.9 ppt). OCDD represented the highest baseline levels, ranging from 340 to 3,300 ppt. Levels of 

total HpCDD ranged from 62 to 640 ppt, while levels of total HxCDD ranged from 12 to 99 ppt (Reed et al. 

1990). 

Birmingham (1990) analyzed soil samples from industrial, urban, and rural sites in Ontario, Canada, and 

some Midwestern U.S. states for CDDs and CDFs. The concentrations of CDD/CDF in rural soils were 

generally not detectable, although HpCDDs and OCDD were found in a few samples. In urban soils, the 

tetra- through octa-congener groups were measured for both CDDs and CDFs. The HpCDDs and OCDD 

dominated the homologue profile and were two orders of magnitude greater than concentrations in rural 

soils. These urban soils also contained measurable quantities of TCDDs, PeCDDs, and HxCDDs. 

Industrial soils did not contain any TCDDs or PeCDDs, but they did contain the highest concentrations of 

the HpCDDs, OCDD, TCDFs, HpCDFs, and OCDFs. In an earlier study, soil concentrations of 

2,3,7,8-TCDD were measured in industrialized areas of a group of mid-western and mid-Atlantic states 

(Illinois, Michigan, New York, Ohio, Pennsylvania, Tennessee, Virginia, and West Virginia) (see 

Table 5-4) (Nestrick et al. 1986). Many of the samples were taken within one mile of major steel, 

automotive, or chemical manufacturing facilities or of municipal solid-waste incinerators. The data show 

that in these typical industrialized areas, 2,3,7,8-TCDD soil concentrations are below 0.01 ppb (range, 

ND–9.4 ppt). The widespread occurrence of 2,3,7,8-TCDD in U.S. urban soils at levels of 0.001–0.01 ppb 

suggests that local combustion sources, including industrial and municipal waste incinerators, are the 

probable sources of the trace 2,3,7,8-TCDD soil concentrations found in those locations (Nestrick et al. 

1986). Soil samples collected in the vicinity of a sewage sludge incinerator were compared with soil 

samples from rural and urban sites in Ontario, Canada (Pearson et al. 1990). Soil in the vicinity of the 

incinerator showed a general increase in CDD concentration with increasing degrees of chlorination. Of the 

CDFs measured, only OCDFs was detected (mean concentration, 43 ppt). Rural woodlot soil samples 

contained only OCDD (mean concentration, 30 ppt). Soil samples from undisturbed urban parkland 

revealed only concentrations of HpCDDs and OCDD, but all CDF congener groups from TCDF to OCDF 

were present. The parkland samples showed an increase in concentrations from the HpCDDs to OCDD and 

PeCDFs to OCDF. The TCDFs were found at the highest concentration (mean, 29 ppt) of all the CDF 

congener groups. 

CDDs 437 

5. POTENTIAL FOR HUMAN EXPOSURE 

A large-scale environmental survey was conducted by the Dow Chemical Company to determine soil levels 

of 2,3,7,8-TCDD on the Dow Midland Plant site and in the city of Midland, Michigan (Nestrick et al. 

1986). The Dow Midland Plant site manufactures a variety of chlorophenolic compounds. Soil samples 

were taken from three different types of areas: locations known to be directly associated with current or 

historic chlorophenolic production and handling, locations known to be associated with incineration of 

chemical and conventional wastes and with ash storage, and locations away from established 2,3,7,8-TCDD 

sources that provide a measure of general background levels of 2,3,7,8-TCDD surface soil within the Dow 

property. Soil samples taken from chlorophenolic production areas showed a range of 2,3,7,8-TCDD 

concentrations from 0.041 to 52 ppb. Two localized areas of elevated concentrations (above 5 ppb) were 

identified with peaks at 34 ppb and 52 ppb. All other samples taken around this area had 2,3,7,8-TCDD 

soil concentrations below 1 ppb. Two of 10 surface soil samples with 2,3,7,8-TCDD concentrations above 

1 ppb (2.0 and 4.3 ppb) were found near the waste incinerator. The concentrations observed there 

(0.018–4.3 ppb) closely matched the 2,3,7,8-TCDD content of the ash produced by the incinerator, which 

ranged up to 10 ppb. The background levels of 2,3,7,8-TCDD (0.0065–0.59 ppb) within the Dow Midland 

Plant site were well below 1 ppb. Soil samples taken within the city of Midland showed 2,3,7,8-TCDD soil 

concentrations below the 1 ppb concern level established by the U.S. Public Health Service, Centers for 

Disease Control and Prevention (CDC), for residential areas (Kimbrough et al. 1984). 2,3,7,8-TCDD soil 

concentrations in the city of Midland (0.6–450 ppt) were higher in areas nearer the Dow Chemical 

Company Midland Plant site (22–450 ppt) (Nestrick et al. 1986). This gradient suggests that operations on 

the Midland Plant site are associated with the appearance of the trace levels of 2,3,7,8-TCDD in the nearby 

environment. 

Several studies have analyzed soil samples in the State of Missouri for 2,3,7,8-TCDD contamination and all 

reported values are comparable. Concentrations of 2,3,7,8-TCDD in soil samples from contaminated sites 

throughout Missouri ranged from 30 to 1,750 ppb and concentrations in Times Beach, MO, a heavily 

contaminated site ranged from 4.4 to 317 ppb (Tiernan et al. 1985). In another study, soil core samples 

taken from a roadside in Times Beach, MO, contained levels of 2,3,7,8-TCDD ranging from 0.8 to 274 ppb. 

Many roadways in Times Beach had been sprayed with waste oil containing CDDs for dust control 

(Freeman et al. 1986). In a third study conducted by Hoffman et al. (1986), 2,3,7,8-TCDD was measured in 

soil samples from the Quail Run Mobile Home Park in Gray Summit, MO. A maximum soil concentration 

of 2,200 ppb (single non-composited sample) was detected at one site; however, concentrations typically 

ranged from 39 to 1,100 ppb in composite soil samples. 
 

CDDs 438 

5. POTENTIAL FOR HUMAN EXPOSURE 

2,3,7,8-TCDD has been detected in soil samples collected at 61 sites of the 91 NPL hazardous waste sites 

where it has been detected in some environmental media (HazDat 1998). CDDs have been detected in soil 

samples collected at 94 of the 126 NPL sites where they have been detected in some environmental media. 

TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD have been detected in soil samples at 71, 21, 29, 34, and 

38 sites of the 105, 34, 43, 49, and 53 NPL sites, respectively, where they have been detected in some 

environmental media (see Table 5-1). 

In conclusion, soil concentrations of CDDs are typically higher in urban areas than in rural areas. Soil 

concentrations associated with industrial sites are clearly the highest, with CDD levels ranging from the 

hundreds to thousands of ppt. In general, as the degree of chlorination increases, the concentrations 

increase. HpCDD and OCDD congeners are generally found at higher concentrations in soil and sediments 

than the TCDD, PeCDD, and HxCDD congeners. 

Sediment. Highly stratified sediments from Green Lake in upstate New York had CDD concentrations 

that could be correlated with atmospheric deposition. CDDs could be detected as far back as 1860–1865 at 

a total CDD concentration of 7 ppt; 98% of all CDDs detected were OCDD. The CDD sediment profile 

showed a strong increase after 1923 and continued to increase until 1984 (the last year analyzed), with a 

maximum concentration of >900 ppt, of which 75% was OCDD (Smith et al. 1992). 

In another study, surficial (surface) sediment samples taken from the Saginaw River and Bay and from 

southern Lake Huron showed that CDDs are ubiquitous in the samples studied, including the most remote 

locations (Czuczwa and Hites 1984). The concentrations were highest in those sediments collected closest 

to urban areas and lowest in open-lake cores. This indicates that the most of the CDDs found in these 

samples are anthropogenic in origin (Czuczwa and Hites 1984). The CDDs found closest to urban areas 

may be related to point source industrial inputs as well as atmospheric deposition, while CDDs found at the 

remote sites are likely to be only atmospheric in origin. In dated sediment cores, CDDs were absent before 

1940. Thus, the authors suggest that accumulation of CDDs in the environment is a recent phenomenon and 

is related to industrial activities (Czuczwa and Hites 1986a, 1986b). Surface sediments taken from the 

Great Lakes showed that CDDs were ubiquitous in the sediments. OCDD was predominant at concentrations ranging from 560 to 4,800 ppt (dry weight) (Czuczwa and Hites 1986a, 1986b). The sediments also 

contained relatively high concentrations of HpCDD. The less chlorinated CDDs were not found in the 

sediments (Czuczwa and Hites 1986a). Sediment samples were collected from five sampling stations in the 

western basin of Lake Ontario near the mouth of the Niagara River and were analyzed for 2,3,7,8-TCDD 
CDDs 439 

5. POTENTIAL FOR HUMAN EXPOSURE 

(Onuska et al. 1983). Measurable quantities of 2,3,7,8-TCDD were present in sediment at two of the 

stations. The highest concentration of 2,3,7,8-TCDD (13 ppt) was found at a depth of 3–5 cm, followed by 

a concentration of 4 ppt at a depth of 3 cm, and 3 ppt at a depth of 13–14 cm. Concentrations of 

2,3,7,8-TCDD in the rest of the sediment samples were below the detection limit (0.1 ppt) (Onuska et al. 

1983). 

Surficial sediments collected from Jackfish Bay on the north shore of Lake Superior, near a pulp and paper 

manufacturer, contained moderate concentrations of the TCDFs (range of geometric mean, 2.4–6,223 pg/g) 

and OCDD congeners (range of geometric mean, 12–250 pg/g), with trace (< 60 ppt) concentrations of 

other congeners (Sherman et al. 1990). The OCDF and OCDD profile for a sediment core collected from 

Moberly Bay was similar to the surficial sediment pattern. These congener groups predominated at all 

sediment depths where detectable concentrations occurred. Low concentrations of the HpCDD, PeCDF, 

and HpCDF congeners also were detected. The concentration profile of the HpCDF congener group 

showed a relatively high value that dropped abruptly to nondetectable (<60 ppt) below a sediment depth of 

10 cm. This abrupt change corresponded to a date of 1973 that reflected an operational change at the pulp 

mill. 

Surficial harbor sediments collected near a PCP wood preserving plant in Thunder Bay, Ontario, Canada on 

the north shore of Lake Superior were found to contain CDDs and CDFs (McKee et al. 1990). The highest 

concentrations were detected at stations closest to the plant docking area and lower concentrations occurred 

at stations further from the source (McKee et al. 1990). No CDDs or CDFs were detected below the 

surficial layer. Concentrations of TCDD and PeCDD congeners were below detection limits (<1 ppt) in all 

samples. The concentrations of the HxCDD and OCDD congeners increased with the degree of 

chlorination. The maximum concentrations of the HxCDDs to OCDD ranged from 5,600 ppt (HxCDDs) to 

980,000 ppt (OCDD). As with the CDD distribution profile, the concentrations of HxCDFs and OCDFs 

increased with the degree of chlorination. 

Sediment samples taken from Love Canal storm sewers in Niagara Falls, NY, contained from 0.9 to 

312 ng/g of 2,3,7,8-TCDD (0.9–312 ppb) (Smith et al. 1983). The highest concentration of 2,3,7,8-TCDD 

(312 ppb) was found immediately adjacent to the canal at its southern end; the next highest concentration 

(120 ppb) was found just upstream. A sample taken one street away from the canal, near the high altitude 

division of the storm sewer system where only a small amount of canal runoff occurs, contained only 

0.9 ppb 2,3,7,8-TCDD (Smith et al. 1983). 
CDDs 440 

5. POTENTIAL FOR HUMAN EXPOSURE 

Surface sediment samples were collected from various estuaries in the United States: Black Rock Harbor 

(an urban industrialized estuary in Connecticut), Long Island Sound (a relatively clean reference site in 

New York), Narragansett Bay (an estuary affected by input from chemical industries in Rhode Island), New 

Bedford Harbor (an estuary within a Superfund site boundary in Massachusetts), and Eagle Harbor (a 

creosote wood-treatment facility in Washington) (Norwood et al. 1989). 2,3,7,8-TCDD was detected in 

sediment from Black Rock Harbor (56–57 ppt), Narragansett Bay (15–19 ppt), and in some of the sediment 

samples found in New Bedford Harbor (4.2–4.6 ppt). 1,2,3,7,8-PeCDD was detected in estuarine sediments 

from Black Rock Harbor (79–95 ppt), New Bedford Harbor (21–29 ppt), and Eagle Harbor (5 ppt). 

HxCDD, HpCDD, and OCDD were also detected in sediments from all estuaries at concentrations ranging 

from approximately 10–100 ppt, 500–3,000 ppt and 2,000–37,000 ppt, respectively. The highest 

concentrations of HpCDD (>1,000 ppt) were detected in Narragansett Bay sediments, while the highest 

concentration of OCDD (37,000 ppt) was detected in Eagle Harbor sediments. The levels of CDDs 

reported for all samples were for dry weight (air dried) concentrations (Norwood et al. 1989). 

Sediment samples collected in 1985–86 from estuarine areas (Passaic River and Newark Bay), near a 

Newark, NJ, facility that manufactured 2,4,5-T between 1948 and 1969, contained high concentrations of 

2,3,7,8-TCDD and OCDD (Bopp et al. 1991). Concentrations of OCDD in the sediment were many times 

higher than concentrations of 2,3,7,8-TCDD. The study indicated that there probably was a significant 

regional source (i.e., combustion and/or use of the wood preservative PCP) for OCDD, a source that is 

lacking in significant concentrations of 2,3,7,8-TCDD relative to the local industrial source. A high 

correlation was found between 2,3,7,8-TCDD and 2,3,7,8-TCDF concentrations, suggesting that the 

industrial site was a major source of 2,3,7,8-TCDF to the natural waters of the area. Sediment core samples 

from a depth of 108–111 cm contained 2,3,7,8-TCDD at a concentration of 21,000 ppt, the highest 

concentration measured in the study. This residue value corresponds to deposition of sediments that 

occurred during the late 1950s to early 1960s during active 2,4,5-T production at the industrial site. 

Maximum concentrations of TCDD in the sediment cores corresponded to the period of maximum 2,4,5-T 

production, with more recently deposited sediments containing lower concentrations of TCDD. This study 

established the persistence of 2,3,7,8-TCDD and 2,3,7,8-TCDF in anaerobic sediments on a time scale of 

several decades (Bopp et al. 1991). 

There has been considerable discussion about historic releases of CDDs/CDFs in Newark Bay, New Jersey. 

Bopp et al. (1991) suggested that a single source (pesticide production facility) is responsible for the 

presence of 2,3,7,8-TCDD/TCDF in the watershed. Recently, Wenning et al. (1992, 1993a, 1993b), 
CDDs 441 

5. POTENTIAL FOR HUMAN EXPOSURE 

using chemometric comparisons of CDD/CDF residues in surficial sediments, found that congener patterns 

in Newark Bay were closely related to those found in sediments from other industrialized waterways with 

several different pollutant sources (e.g., New Bedford Harbor, Massachusetts; Black Rock Harbor, 

Connecticut; Providence River, Rhode Island; Eagle Harbor, Washington, as well as several European 

waterways). The similarities and differences among the various waterways examined in the analysis 

suggests that the presence of 2,3,7,8-substituted CDDs/CDFs in surficial sediments from Newark Bay are 

more likely due to multiple sources of contamination. Most recently, Ehrlich et al. (1994) identified the 

relative contributions of various sources of CDDs/CDFs to recently deposited sediments of Newark Bay 

using polytopic vector analysis, a multivariate statistical technique. These authors also concluded that the 

2,3,7,8-substituted CDD/CDF patterns in the sediments of Newark Bay are consistent with discharges from 

multiple sources. In a recent study, Huntley et al. (1997) reported that combined sewer overflows may 

contribute substantially to surface sediment contamination of the nearby Passaic River. Several such 

sources that have existed over the past century in the vicinity include scrap metal refineries, pulp and paper 

mills, copper smelters, chemical manufacturing plants, municipal sewage treatment plants, and industrial/ 

municipal incinerators (EPA 1987n). 2,3,7,8-TCDD sediment concentrations ranged from below the 

detection limit (22 ppt) to 21,000 ppt (21 ppb), whereas OCDD concentrations ranged from 3.1 ppb to 

42,000 ppt (42 ppb), although other sources of OCDD were thought to contribute to the elevated levels of 

OCDD (Bopp et al. 1991; Wenning et al. 1992). 

Sludges from various Ontario pulp and paper plants using either the bleached kraft (8 mills) or sulfite 

bleaching process (2 mills) were analyzed for CDDs (Clement et al. 1989). 2,3,7,8-TCDD was detected in 

sludge samples at a concentration range of 170–370 ppt. Only one other TCDD isomer (180 ppt) was 

detected in a sludge sample, but it was not identified. PeCDDs and HxCDDs were not detected in any 

sludge samples, whereas HpCDD (400 ppt) was found in 1 sludge sample and OCDD (120–1,800 ppt) was 

found in 6 sludge samples (Clement et al. 1989). 

Utility telecommunication and railway right-of-ways may be contaminated by leaching of CDDs associated 

with chlorophenol-treated railway ties and utility poles. A study in British Columbia showed that CDDs 

and CDFs were not detected in parkland ditch sediments (control area), but were detected in farmland, 

utility, and railway right-of-way ditch sediments (Wan and van Oostdam 1995). Total mean CDD 

concentrations (mainly OCDD and HpCDD) ranged from 18.8 to 277 ng/kg (ppt) (dry weight) in ditch 

sediments and ballasts respectively. Concentrations of CDDs were much higher in ditch sediment adjacent 

to utility poles (mean 2,576 ng/kg (ppt) [dry weight]) than in sediment 4 meters downstream (14 ng/kg 
 

CDDs 442 

5. POTENTIAL FOR HUMAN EXPOSURE 

[dry weight]) or 4 meters upstream of the utility poles (not detected). CDD concentrations in ditch water 

were also higher close to the poles (mean 13,142 ng/L [ppt] than 4 meters downstream of the poles (mean 

4,880 ng/L [ppt]). The authors concluded that utility poles and railway ties are a potential constant source of 

CDD/CDF contamination to both water and sediment in aquatic environment through ditch runoff. 

2,3,7,8-TCDD has been detected in sediment samples collected at 17 sites of the 91 NPL hazardous waste 

sites where it has been detected in some environmental media (HazDat 1998). CDDs have been detected in 

sediment samples collected at 31 of the 126 NPL sites where they have been detected in some environmental media. TCDDs, PeCDDs, HxCDDs, HpCDDs, and OCDD have been detected in sediment samples 

at 22, 7, 10, 9, and 13 sites of the 105, 34, 43, 49, and 53 NPL sites, respectively, where they have been 

detected in some environmental media (see Table 5-1). 

In conclusion, CDD congener profiles in sediment generally reflect those exhibited by the contamination 

source or sources. High concentrations of HxCDDs, HpCDDs, and OCDDs in sediment are usually the 

result of anthropogenic inputs via industrial processes and releases or urban runoff, and concentrations 

generally increase with the degree of chlorination, but decrease with distance from the source (McKee et al. 

1990). 

5.4.4 Other Environmental Media 

Foods. The FDA has conducted limited analyses for the higher chlorinated CDDs (HxCDD, HpCDD, 

and OCDD) in market-basket samples collected from 1979 to 1984 under the FDA's Total Diet Program 

(Firestone et al. 1986). Food samples found to contain PCP residues >0.05 μg/g (ppm) were analyzed for 

1,2,3,4,6,7,8-HpCDD and OCDD. In addition, selected samples of ground beef, chicken, pork, and eggs 

from the market-basket survey were analyzed for these CDD congeners (wet weight basis), regardless of the 

results of the PCP analysis. HxCDD was not found in any of the foods sampled; however, the detection 

limit (10–40 pg/g [ppt]) was very high. Generally low concentrations (<300 pg/g [ppt]) of HpCDD and 

OCDD were found in bacon, chicken, pork chops, and beef liver. Several beef livers had higher concentrations of OCDD residues (614–3,830 pg/g), and one beef liver contained 428 pg/g (ppt) of HpCDD. 

HxCDD, HpCDD, and OCDD were not detected in milk, ground beef, or seafood samples, but the detection 

limits (10–40 ppt) were very high. No CDDs were found in 17 egg samples collected in various parts of the 

United States. OCDD was detected in 2 of 18 pork samples (27 ppt and 53 ppt) and in 2 of the 16 chicken 

samples (29 ppt and 76 ppt). One chicken sample with PCP residues (>0.05 μg/g) contained 
CDDs 443 

5. POTENTIAL FOR HUMAN EXPOSURE 

concentrations of 1,2,3,4,6,7,8-HpCDD (28 ppt) and OCDD (252 ppt). The CDD residues (21–1,610 pg/g) 

in eggs from Houston, Texas, and Mena, Arkansas, with PCP residues >0.05 μg/g collected in 1982 and 

1983–84, respectively, contained 1,2,3,4,6,7,8-HpCDD concentrations ranging from 21 to 588 ppt, and 

OCDD concentrations ranging from 80 to 1,610 ppt. These residues were attributed to local PCP 

contamination problems in these areas (Firestone et al. 1986). Milk samples contaminated with PCP at 

levels ranging from 0.01 μg/g to 0.05 μg/g PCP contained no detectable CDDs. It should be noted that the 

reported limits of detection (10–40 ppt) for the FDA analyses from these older samples, are higher than 

concentrations of CDDs observed in foods from more recent studies. Samples of beef liver, pork chops, 

chicken, ground beef, and eggs collected in the United States and analyzed for HpCDD and OCDD 

contained average concentrations of HpCDD (2.2–9.6 ppt) and OCDD (6.3–47.6 ppt) (Jasinski 1989). Eggs 

contained the lowest levels of HpCDD and OCDD, and beef liver contained the highest levels. 

LaFleur et al. (1990) analyzed the concentration of 2,3,7,8-TCDD and 2,3,7,8-TCDF (wet weight basis) in a 

variety of food products collected randomly from grocery stores located in the southern, Midwestern, and 

northwestern regions of the United States. Concentrations of 2,3,7,8-TCDD ranged from 17 to 62 pg/kg for 

ground beef, were not detectable in ground pork, ranged from 12 to 37 pg/kg for beef hot dogs, and ranged 

from 7.2 to 9.4 pg/kg for canned corned beef hash on a whole-weight basis. Concentrations of 

2,3,7,8-TCDF were generally much less than concentrations of 2,3,7,8-TCDD, with the exception of ground 

pork and corned beef hash. For ground pork, TCDD concentrations were not detectable and 2,3,7,8-TCDF 

concentrations ranged from 13 to 20 pg/kg and for corned beef hash, concentrations of TCDD ranged from 

7.2 to 9.4 pg/kg, while concentrations of TCDF ranged from 9.8 to 10 pg/kg. 

A study conducted by the province of Ontario, Canada analyzed concentrations of CDDs/CDFs in a variety 

of foods locally grown in Canada or imported from the United States and New Zealand (Birmingham et al. 

1989). Concentrations of OCDD in various foods ranged from 3 to 210 ppt (wet weight basis). OCDD 

concentrations were detected in U.S. eggs (8 ppt), U.S. beef (24 ppt), Canadian hamburger (3 ppt), and 

Canadian chicken samples (210 ppt). Chicken also contained 15 ppt HpCDD, but no other CDDs were 

detected in these samples. Fruits and vegetables were generally free of CDD and CDF residues (detection 

limit = 1 ppt). OCDD concentrations ranging from 0.6 to 8 ppt (wet weight) were also detected in samples 

of U.S. potatoes (3 ppt), apples (8 ppt), peaches (0.6 ppt), and wheat (0.7 ppt). For these food items, OCDD 

was the only homologue detected. No 2,3,7,8-TCDD was detected in any of the food samples tested 

(detection limits of 1–4 ppt). 
CDDs 444 

5. POTENTIAL FOR HUMAN EXPOSURE 

Beck et al. (1989a) analyzed the concentrations of CDDs/CDFs in 22 samples of foodstuffs collected in the 

Federal Republic of Germany. Twelve randomly collected food samples (chicken, eggs, butter, pork, 

redfish (ocean perch), cod, herring, vegetable oil, cauliflower, lettuce, cherries, and apples) were purchased 

in various stores in West Berlin. The highest 2,3,7,8-TCDD levels were observed in fish samples with 

concentrations of 4.7, 23, and 2.8 ppt (lipid basis) for herring, cod, and redfish, respectively. Concentrations 

of 2,3,7,8-TCDD (on a lipid basis) in meat, poultry, and dairy products were lower; with concentrations 

ranging from 0.01 ppt in sheep, 0.03 in pork, 0.2 in eggs, 0.3 in chicken, 0.6 in cattle, and 0.02 in cow's 

milk and 0.08 in butter (see Table 5-5). The EPA TEQ values for CDDs/CDFs calculated for these 

products ranged from 20.0–39.7 ppt in fish, 0.14–1.31 ppt in meat and poultry, and from 0.43 to 0.86 ppt in 

dairy products. In all samples tested, the 2,3,7,8-substituted congeners predominated in the samples and 

non-2,3,7,8-substituted congeners were not detected in fish, chicken and eggs. For meat, poultry, and dairy 

samples, the congener profile showed high concentrations of 1,2,3,4,6,7,8-HpCDD and OCDD with 

concentrations of most other congeners at or below 1 ppt (lipid basis). In fish samples, high concentrations 

of the 2,3,7,8-substituted congeners, TCDDs (range 2.8–23 ppt), PeCDD (range 1.3–12 ppt), HxCDD 

(range 0.01–17 ppt), and OCDD (range 11–83 ppt) resulted in TEQ values for CDDs/CDFs ranging from 

20–40 ppt (lipid basis). In the five food samples of plant origin, no CDDs/CDFs were detected on a whole 

weight basis (detection limit . 0.01 ppt). 

Congener-specific analyses for CDDs and CDFs were performed on 18 dairy, meat, and fish products 

obtained from a supermarket in upstate New York (Schecter et al. 1994d). Total CDD concentrations (on a 

wet weight basis) ranged from 0.35 to 2.91 ppt in fish, 0.6–59.3 ppt for meats, and 0.6–14 ppt in dairy 

products. A summary of the CDD/CDF concentrations and TEQ concentrations calculated for the 18 foods 

is presented in Table 5-6. The TEQ for both the CDDs and CDFs on a wet weight basis for these food 

samples ranged from 0.02 to 1.5 ppt, 0.02–0.13 ppt for fish products, 0.03–1.5 ppt for meat products, and 

0.04–0.7 ppt for dairy products, with the highest TEQ found in ground beef. 

Recently, the EPA and U.S. Department of Agriculture (USDA) completed the first statistically designed 

surveys of the occurrence and concentrations of CDDs/CDFs in beef fat (Ferrario et al. 1996; Winters et al. 

1996), pork fat (Lorber et al. 1997), poultry fat (Ferrario et al. 1997), and the U.S. milk supply (Lorber et al. 

1998). The congener specific results for various foods are shown in Table 5-7. It is clear from the results, 

that two congeners (1,2,3,4,6,7,8-HpCDD and 1,2,3,4,6,7,8-OCDD) were typically found at the highest 

concentrations in all food samples. Concentrations of 2,3,7,8-TCDD were highest in heavy fowl (0.43 ppt) 

and young turkeys (0.24 ppt); much lower concentrations were found in beef (0.05 ppt), pork 



CDDs 448 

5. POTENTIAL FOR HUMAN EXPOSURE 

(0.10 ppt), young chickens (0.16 ppt), light fowl (0.03 ppt) and milk (0.07 ppt). The total concentrations of 

CDDs/CDFs were highest in pork fat (75.67 ppt) and milk (15.43 ppt), and ranged from 5.68 to 14.09 ppt 

for all other types of foods tested. The TEQ value for CDDs/CDFs combined was highest for pork fat (1.30 

ppt), heavy fowl (0.98 ppt), young turkeys (0.93 ppt), and beef fat (0.89 ppt), with lower TEQ values of 

0.40–0.82 ppt for young chickens, light fowl, and milk. 

CDDs have been found in infant formulas purchased in the United States (Schecter et al. 1989c). The infant 

formulas were derived from cow's milk or soybeans. In general, both types of infant formula had very low 

concentrations of CDDs. 2,3,7,8-TCDD and PeCDD were not detected in cow's milk or soybean formula at 

detection limits ranging from 0.5 to 1.0 ppt. HxCDD was not detected in soybean formula at the same 

detection limits. Whole and lowfat (2% fat) cow's milk contained total HxCDD at lipid-adjusted 

concentrations of 3.6 and 3.3 ppt, respectively. Lipid-adjusted levels of HpCDD were found in whole cow's 

milk formula (6.5 ppt), lowfat (2%) cow's milk formula (8 ppt), and soybean formula (2.3–3.0 ppt). OCDD 

was the most abundant congener in both cow's milk and soybean formula. Concentrations of OCDD (lipidadjusted) were as follows: cow's milk formula (15 ppt), low fat (2%) cow's milk formula (21 ppt), and 

soybean formula (21–36 ppt) (Schecter et al. 1989c). 

In comparison, a study by LaFleur et al. (1990) reported the concentration of 2,3,7,8-TCDD and 

2,3,7,8-TCDF in whole milk and half and half. These authors also measured the additional exposure that 

resulted from migration of these compounds from bleached paperboard containers into the milk over 

various storage periods. The concentrations of 2,3,7,8-TCDD in whole milk ranged from 24 to 25 pg/kg 

and in half-and-half ranged from 13 to 14 pg/kg. The corresponding concentrations of 2,3,7,8-TCDF 

ranged from 260 to 280 pg/kg for whole milk and 146 to 195 pg/kg for half and half. These authors also 

determined the concentration of 2,3,7,8-TCDD and TCDF for cow’s milk obtained directly from a dairy and 

for milk stored for various time periods in bleached paperboard cartons. On a lipid basis, the concentration 

of 2,3,7,8-TCDD of control milk obtained directly from the dairy was 0.48 pg/g, and milk stored in 

paperboard cartons for 24, 48, 120, and 288 hours was 0.95, 1.4, 1.9, and 2.7 pg/g, respectively. The 

2,3,7,8-TCDD and 2,3,7,8-TCDF concentrations in the paperboard carton were 4.3 and 25 ppt, respectively. 

Concentrations of 2,3,7,8-TCDF in the control milk was not detectable, but increased in milk stored in 

cartons for 24, 48, 120, and 288 hours to 6.8, 10.2, 20.1, and 35.1 pg/g, respectively. The percent migration 

of the 2,3,7,8-TCDD ranged from 2 to 6%, while the percentage of migration of the 2,3,7,8-TCDF ranged 

from 4 to 18% over the same period (LaFleur et al. 1990). 
 

CDDs 449 

5. POTENTIAL FOR HUMAN EXPOSURE 

Similar levels of CDD contamination were reported in two European studies. CDDs were detected in 

8 samples of cow's milk in Germany at concentrations ranging from 0.2 ppt for 2,3,7,8-TCDD (detection 

limit 0.2 ppt) to <10 ppt of OCDD (detection limit not significantly higher than blanks) (Beck et al. 1987). 

In a Swedish study, only 1 of 10 samples of milk held in either glass bottles or paper cartons contained a 

detectable level of 2,3,7,8-TCDD (0.46 pg/g milk fat; paper carton; detection limit 0.4 pg/g). Other CDDs 

were also detected (maximum 7.8 pg/g for OCDD) with the highest concentrations associated with milk 

packaged in paper cartons, indicating that leaching of CDDs from the paper carton into the milk can occur 

(Rappe et al. 1990). 

Fish and Wildlife. A survey of 2,3,7,8-TCDD contamination in benthic (bottom feeding) and predator 

fish from major U.S. watersheds was conducted for the EPA National Dioxin Study (Kuehl et al. 1989). It 

was observed that 17 of 90 (19%) samples collected at sites statistically selected by the EPA had detectable 

levels of 2,3,7,8-TCDD, whereas 95 of 305 (31%) samples from sites chosen by EPA regional laboratories 

had detectable levels (detection limits 0.5–2 ppt on a wet weight basis). Of the 112 sites where 

2,3,7,8-TCDD was detected, 74 samples (67%) were below 5 ppt, 34 samples (32%) were between 5 and 

25 ppt, and 4 samples (1%) were above 25 ppt. A subset of samples collected at sites near the discharges 

from pulp/paper manufacturing facilities (n=28) had a higher frequency of 2,3,7,8-TCDD contamination 

above 5 ppt (38%). This subset of samples also contained the sample with the highest level of 

2,3,7,8-TCDD contamination (85 ppt). Of the 29 samples collected in the Great Lakes region, 23 (79%) of 

the sites were found to have detectable levels of 2,3,7,8-TCDD. The most highly contaminated sample, 

with a concentration of 41 ppt, was collected from Lake Ontario near Oswego, NY. Four of 57 (7%) 

estuarine or coastal sites had detectable 2,3,7,8-TCDD levels in either fish or shellfish. The level of 

contamination in these 4 samples ranged from 1.08 to 3.5 ppt (Kuehl et al. 1989). In another study, fish 

sampled downstream from a bleached kraft paper mill were found to contain higher concentrations of CDDs 

compared with fish sampled upstream of the paper mill (Hodson et al. 1992). TCDD concentrations in the 

fish ranged from 1.47 pg/g (wet weight basis) in upstream areas to 15.6 pg/g in fish sampled 2 km 

downstream. Fish sampled 95 km downstream contained only about half the residues (8.87 pg/g TCDD) of 

those collected immediately downstream of the facility (Hodson et al. 1992). 

Travis and Hattemer-Frey (1991) analyzed data collected as part of the National Dioxin Study regarding 

2,3,7,8-TCDD concentrations in fish. The TCDD levels measured in fish from lakes and rivers in the 

United States confirm that 2,3,7,8-TCDD is bioaccumulating in fish and that low-level contamination of 

fish is widespread (EPA 1987n). The fish survey included 304 urban areas in the vicinity of population 
CDDs 450 

5. POTENTIAL FOR HUMAN EXPOSURE 

centers or areas with known commercial fishing activity, including sites in the Great Lakes region. The 

results of this study indicate that only 29% of fish fillets collected at urban sites had detectable 

concentrations of 2,3,7,8-TCDD (detection limit =1 ppt). The geometric mean for these fillet samples was 

0.3 ppt (wet weight basis). Fish samples from the Great Lakes area contained higher concentrations of 

2,3,7,8-TCDD than fish from urban areas (e.g., 67 versus 29% contained detectable levels, respectively). In 

the Great Lakes area, the geometric mean concentrations of 2,3,7,8-TCDD in fish fillets (2.3 ppt) was 

almost 7 times higher than the concentrations in the fillets from fish collected from urban areas (0.3 ppt). 

Comparable concentrations of 2,3,7,8-TCDD were detected in bottom-feeding and predator species from 

the Great Lakes region. Approximately 74% of the fish fillet samples collected from sites near pulp and 

paper mills contained detectable concentrations of 2,3,7,8-TCDD. The geometric mean concentration for 

these fillet samples was 0.9 ppt. This geometric mean is 3 times higher than for urban fillet concentrations 

(0.3 ppt) but is approximately 2 times lower than for TCDD concentrations in fillets from the Great Lakes 

Region (2.3 ppt). 

From 1986 to 1989, the National Study of Chemical Residues in Fish (NSCRF) was conducted by the EPA 

as a follow-on study to the National Dioxin Study (EPA 1992). The purpose of the NSCRF was to assess 

the concentrations of 60 toxic pollutants (including CDDs and CDFs) in the tissues of benthic and game 

fish nationwide. Benthic species were analyzed as whole-body samples, while game species were analyzed 

as fillet samples and all concentrations were on a wet weight basis. A summary of the prevalence and 

concentrations of 6 CDDs and 9 CDFs detected at 388 sites surveyed nationwide in the NSCRF is presented 

in Table 5-8. Four of the CDDs and three of the CDFs analyzed were detected at over 50% (58–89%) of 

the sites surveyed. The most frequently detected CDD/CDF compounds (1,2,3,4,6,7,8-HpCDD and 

2,3,7,8-TCDF) were both found at 89% of the sites. These compounds were also detected at the highest 

concentrations: 1,2,3,4,6,7,8-HpCDD at 249 ppt and 2,3,7,8-TCDF at 404 ppt (wet weight). The mean 

concentrations of these 2 compounds were substantially lower at 10.5 and 13.6 ppt, respectively. The CDD 

(2,3,7,8-TCDD) believed to be the most toxic congener to mammals, was found at 70% of the sites at a 

maximum concentration of 204 ppt and a mean of 6.8 ppt (wet weight basis). The NSCRF report further 

shows that pulp and paper mills using chlorine bleach pulp appeared to be the dominant source of the 

2,3,7,8-TCDD and 2,3,7,8-TCDF. Fish collected at sites downstream of pulp and paper mills had 

significantly higher concentrations of 2,3,7,8-TCDD than fish collected near all other source categories. 

The statistical tests also showed the same result for 2,3,7,8-TCDF, with the exception that fish residue 

concentrations downstream of Superfund sites also marginally met the statistical criteria. With respect to 

source categories, the NSCRF data showed that fish collected downstream of pulp and paper mills (using 

CDDs 452 

5. POTENTIAL FOR HUMAN EXPOSURE 

chlorine bleaching processes) had the highest median 2,3,7,8-TCDD concentrations (5.66 ppt), compared to 

the next highest source category, refinery/other industrial sites (1.82 ppt), industrial/urban sites (1.40 ppt), 

Superfund sites (1.27 ppt), and background sites (0.5 ppt). Source categories with the highest 

2,3,7,8-TCDD concentrations in fish also had the highest TEQ values. OCDD and OCDFs were not 

analyzed in tissue because at the time the NSCRF study was initiated (1986), the TEFs were zero for these 

compounds. In 1989, TEFs for OCDD and OCDFs were increased to 0.001. Consequently, TEQ values 

presented in the NSCRF report may be underreported for samples collected at sites with sources of 

OCDD/OCDFs (e.g., wood preservers) (EPA 1992). 

De Vault et al. (1989) collected samples of lake trout and walleye for CDD and CDF analysis from each of 

the Great Lakes and Lake St. Clair. One of the conclusions of the National Dioxin Study was that fish from 

the Great Lakes region were among the most severely contaminated in the United States. Fish were 

analyzed for 8 congeners of CDDs and 10 congeners of CDFs. Total CDD concentrations ranged from 

7.2 ng/kg in lake trout from Lake Superior to 64.5 ng/kg in Lake Ontario (wet weight basis). Concentrations of 2,3,7,8-TCDD ranged from 1 ng/kg in lake trout from Lake Superior to 48.9 ng/kg in lake trout 

from Lake Ontario. The dominant congener in all but Lake Ontario was 1,2,3,7,8-PeCDD at concentrations 

ranging from 2.3 ng/kg in Lake Superior to 16.7 ng/kg in Lake Michigan. The only other congener that 

significantly contributed to the total CDD concentration was 1,2,3,6,7,8-HxCDD, which ranged from 

1.3 ng/kg in Lake Superior to 10.9 ng/kg in Lake Michigan. Substantial interlake differences exist in the 

percentage of total CDD contributed by the various congeners. The 2,3,7,8-TCDD congener contributes a 

relatively small percentage of the total CDD in fish from Lakes Superior, Michigan, and Erie. It is 

comparatively more important in Lake Huron (32%) and Lake St. Clair (36%) and contributes 76% of the 

total CDD in Lake Ontario. The results of this study support the widespread contamination of the Great 

Lakes ecosystem and clearly show that both the concentration of individual congeners and the congener 

composition of total CDDs in Great Lakes fish vary significantly between lakes and in Lake Michigan 

between sites. The authors suggest that these differences may be associated with different sources and 

loadings of these compounds to each of the Great Lakes (De Vault et al. 1989). This is confirmed by the 

analysis of sources of CDDs in the Great Lakes which appear to be both from atmospheric deposition and 

industrial point sources (Hebert et al. 1994). 

In another study, CDDs and CDFs were measured in four species of salmonids (coho salmon, lake trout, 

rainbow trout, and brown trout) collected from Lake Ontario (Niimi and Oliver 1989a). Total CDD 

concentrations ranged from 46 to 290 ng/kg (ppt) in whole fish and 60–366 ng/kg in muscle composite 
CDDs 453 

5. POTENTIAL FOR HUMAN EXPOSURE 

samples. Levels of 2,3,7,8-TCDD in whole fish ranged from 60–20 ppt (wet weight basis). This 

represented 60% of total TCDDs and 10% of total CDDs. The HxCDD congener group was most dominant 

in all fish species and represented approximately 39% of the total CDD concentrations. High 

concentrations of OCDD were also detected in coho salmon (160 ng/kg whole fish and 280 ng/kg muscle) 

and lake trout (89 ng/kg whole fish and 28 ng/kg muscle) but not in brown or rainbow trout. The authors 

could not explain this difference; however, OCDD is typically the CDD present at the highest concentrations in Lake Ontario water, suspended sediments, and sediments. Although the total CDF concentrations were 75% lower than the total CDD concentrations, the levels of 2,3,7,8-TCDF (11–20 ppt) were 

comparable to levels of 2,3,7,8-TCDD (6–20 ppt). Results of another study by the same authors found that 

TEQ values for PCB concentrations in Lake Ontario salmonids were several fold higher than TEQ values 

for CDDs and CDFs in the same fish species (Niimi and Oliver 1989b). 

Background concentrations of CDDs in fish were measured in the Mississippi River and Lake Orono in Elk 

River, Minnesota, a semi-rural location (Reed et al. 1990). No major industrial or commercial activity 

occurred in the area at the time of the study, and the survey was conducted as a baseline study prior to the 

operation of the Elk River Electric Generating Station (powered by refuse-derived fuel). None of the fish 

collected contained measurable amounts of 2,3,7,8-TCDD; however, one of the composites from the 

Mississippi River contained 3.9 ppt of total TCDD (wet weight basis). Detection limits ranged from 

0.28 ppt to 6.6 ppt on a congener-specific and sample-specific basis and were not individually reported for 

each result. OCDD was the most abundant congener (average 59 ppt, range 56–62 ppt), followed in 

decreasing order by total HpCDD (average 19.3, range 15–22 ppt), total HxCDD (average 6.87 ppt, range 

2.3–11 ppt), and total PeCDD (average 3.9 ppt, range 3.5–4.5 ppt) (Reed et al. 1990). Lake Orono showed 

the same pattern, with OCDD being the most abundant congener (average 39 ppt, range 35–43 ppt), 

followed by total HpCDD (average 10.5, range 10–11 ppt), and total HxCDD (3.0 ppt). PeCDDs were not 

detected in the Lake Orono samples (Reed et al. 1990). 

Contamination of the Spring River in southwest Missouri by 2,3,7,8-TCDD is believed to have resulted 

from several well defined point-source waste disposal sites (Crunkilton et al. 1987). Analysis of 31 fish 

samples (11 different fish species) collected from 1981 to 1983 demonstrated a rapid decline in 

2,3,7,8-TCDD concentrations in fish at increasing distances both upstream and downstream from the area of 

contamination. Mean concentrations of 2,3,7,8-TCDD 0.5 km downstream from the area of contamination 

were 38 ppt in whole fish and 20 ppt in fish fillets (wet weight basis). Mean concentrations in fish 
CDDs 454 

5. POTENTIAL FOR HUMAN EXPOSURE 

caught more than 14 km downstream were below 4 ppt in both whole fish and fillet samples (Crunkilton et 

al. 1987). 

Fish samples (butterfish, flounder, hake, and herring) collected in 1984 from the Atlantic Ocean off Long 

Branch, NJ, contained no detectable levels of 2,3,7,8-TCDD (detection limit <10 pg/g) (wet weight basis) 

(Firestone et al. 1986). Cod caught in the northwest Atlantic in November 1990 did not have detectable 

levels of any CDDs in their muscles or ovaries, although 5 of 10 liver samples had OCDD at a mean 

concentration of 0.8 ppt and TCDD was found in 3 of 10 samples at 0.1 ppt (Hellou and Payne 1993). A 

4-year study of marine and freshwater fish and other edible aquatic organisms taken from Canadian waters 

that received effluents from pulp and paper mills indicated that 2,3,7,8-TCDD was the most prominent 

CDD found in the fish regardless of the tissue sampled or sampling location. The maximum 2,3,7,8-TCDD 

concentration detected in the edible organisms sampled was for crab hepatopancreas tissue (>500 pg/g) (wet 

weight basis). Whole fish samples also contained greater CDD concentrations than fillet samples (Whittle 

et al. 1993). 

Several studies have been conducted to monitor 2,3,7,8-TCDD concentrations in fish and shellfish in 

northern New Jersey in the vicinity of a pesticide manufacturing site that allegedly released an estimated 

4–8 kg of 2,3,7,8-TCDD over a 20-year period (Bopp et al. 1991). Samples of striped bass, blue crabs, 

and lobsters collected from Newark Bay and the New York Bight (marine waters directly offshore from 

New York Harbor) all contained high concentrations (up to 6,200 ppt) (wet weight basis) of 2,3,7,8­

substituted TCDD, PeCDD, and CDFs (Rappe et al. 1991). Concentrations of HxCDD and HpCDD 

ranged from <0.1 to 220.7 ppt and <0.7 to 244.9 ppt, respectively. The concentrations of 2,3,7,8-TCDD in 

these marine organisms were higher than any other New Jersey samples and represented the highest 

concentrations of 2,3,7,8-TCDD reported for aquatic species. The two crustaceans sampled in the study 

had similar congener patterns; they all contained both a large number and large amounts of CDD and 

CDF congeners in addition to the 2,3,7,8-substituted chlorinated compounds. In contrast, the striped bass 

samples contained primarily the 2,3,7,8-chlorine-substituted congeners. The concentrations of 

2,3,7,8-TCDD in crab hepatopancreas tissue ranged from 3,700 to 6,200 ppt and from 100 to 120 ppt in 

crab meat. Concentrations of 2,3,7,8-TCDD were lower in the lobster, ranging from 250 to 610 ppt in the 

hepatopancreas and 5 to 6 ppt in the meat. Concentrations of 2,3,7,8-TCDD in striped bass muscle tissue 

ranged from 84 to 730 ppt. In this study, the crustacean samples all contained very complex ion curves 

for the TCDDs showing 10 major and 5 minor peaks, while the striped bass samples primarily contained 

the 2,3,7,8-TCDD isomer and a few other isomers. With respect to the PeCDDs, the crustacean samples 
CDDs 455 

5. POTENTIAL FOR HUMAN EXPOSURE 

contained 5–6 peaks including the 1,2,3,7,8-PeCDD (100 ppt in hepatopancreas and 1–2 ppt in meat), 

while the major isomer found in the striped bass was 1,2,3,7,8-PeCDD (5–10 ppt). Regarding the 

HxCDDs, the crustacean samples contained 3 major peaks one of which was 1,2,3,6,7,8-HxCDD 

(100–300 ppt in the hepatopancreas); while the striped bass samples contained concentrations <1 ppt. 

The HpCDD congeners (1,2,3,4,6,7,9- and 1,2,3,4,6,7,8-) were detected in crustacean hepatopancreas 

tissue ranging from 31.7 to 411.9 ppt, while meat samples contained 0.00-8.5 ppt. Striped bass tissue 

samples contained 4–11.4 ppt. Concentrations of OCDD ranged from 50.5 to 94.6 ppt in crustacean 

hepatopancreas tissues, 6.3–78.8 ppt in meat samples, while concentrations in striped bass ranged from 

5.1–49.5 ppt (Rappe et al. 1991). 

Cai et al. (1994) analyzed blue crab tissue (hepatopancreas and muscle) from Newark Bay and some adjacent 

areas of the New York Bight for 2,3,7,8-TCDDs and 2,3,7,8-TCDF and several other 2,3,7,8-substituted 

CDDs/CDFs. These authors found 2,3,7,8-TCDD concentrations in hepatopancreas tissue to be 10–20 times 

higher than muscle concentrations. Lipid content of the hepatopancreas is 6–9% as compared to 1% for muscle 

tissue. The highest concentration of CDDs/CDFs was detected in crabs collected at the station closest to the 

pesticide production facility. Crabs also had higher CDDs/CDF concentrations in September after feeding all 

summer than crabs collected in June. Concentrations of 2,3,7,8-TCDD up to 1 ppb were detected in some crabs 

collected from Newark Bay (detection limit 0.5–1 ppt). Hauge et al. (1994) conducted further studies of blue 

crabs and lobsters from three distinct fisheries in the Hudson-Raritan estuary. The Ambrose fishery includes 

Raritan and Sandy Hook Bays and extends to a 7-nautical-mile radius from Ambrose Light near the entrance to 

New York Harbor. The Alongshore fishery is a box-shaped area extending from Long Branch, NJ, south to 

Point Pleasant, NJ, and then extending off-shore approximately 25 nautical miles. The Offshore fishery extends 

eastward from the 50-fathom line to the 100-fathom line approximately 100 miles seaward to the edge of the 

continental shelf. Combined muscle/hepatopancreas samples of blue crabs from Raritan Bay and the lower 

Hudson River had a mean 2,3,7,8-TCDD concentration of 71.5 pg/g (range not detected to 260 pg/g) and a 

mean 2,3,7,8,-TCDF concentration of 67.1 pg/g (range not detected to 110 pg/g). The mean total TEQ 

concentration was 78.2 pg/g (ppt) (wet weight basis). Both the mean 2,3,7,8-TCDD concentration and the mean 

TEQ values exceeded the FDA guidelines for "no consumption" (>50 ppt) (see Chapter 7). FDA “no 

consumption” guidelines advise consumers that fish and shellfish should not be consumed when CDD 

concentrations exceed 50 ppt. 2,3,7,8-TCDD and TCDF were detected in 53 and 67% of crabs, respectively. 

Levels of 2,3,7,8-TCDD in muscle and hepatopancreas of lobsters were similar in animals from the Ambrose 

and Alongshore fisheries, with the mean concentration in both areas ranging from 34 to 
 

CDDs 456 

5. POTENTIAL FOR HUMAN EXPOSURE 

41 pg/g of 2,3,7,8-TCDD/TCDF. The mean total TEQ values were 38.5 pg/g (Ambrose fishery) and 44.4 pg/g 

(Alongshore fishery). Mean 2,3,7,8-TCDD/TCDF levels thus exceeded the FDA "safe consumption" level 

(#25 ppt), but did not exceed the FDA "no consumption" level (>50 ppt). None of the lobsters from the 

Offshore fishery contained detectable TCDD/TCDF. Analysis of separate muscle and hepatopancreas tissues 

from individual lobsters yielded detectable concentrations of both contaminants only in the hepatopancreas 

(<26–410 pg/g for 2,3,7,8-TCDD and <33–380 pg/g for 2,3,7,8-TCDF). Concentrations in lobster muscle 

tissues were all below detection limits of 6–20 pg/g for 2,3,7,8-TCDD and 10–25 pg/g for 2,3,7,8-TCDF. 

Concentrations of CDDs/CDFs were also evaluated in a bivalve mollusc, the soft-shelled clam (Mya arenaria) 

in Newark Bay, Arthur Kill, and Raritan Bay (Brown et al. 1994). Clams from Newark Bay contained 

11–20 ppt TCDD, 3.5–5 ppt TCDF, and 13–25 ppt TEQ; those from Arthur Kill contained 4.8–7.7 ppt TCDD, 

3.1–5.1 ppt TCDF, and 6.8–11 ppt TEQ; and those from Raritan Bay contained 0.5–1.1 ppt TCDD, 2–4.6 ppt 

TCDF, and 1.2–2.1 ppt TEQ (wet weight basis). The maximum TEQ concentration of the Newark Bay clams 

(25 ppt) approached the upper limit of the FDA "safe consumption" level of #25 ppt. The FDA believes that 

consumption of fish or shellfish with CDD concentrations #25 ppt should not result in any serious health 

effects. Concentrations decreased with increasing distance from the suspected pesticide plant site near Newark 

Bay. The authors also showed that the clams could eliminate TCDD and TCDF when they were removed to 

clean water sites. The half-lives of the TCDD, TCDF, and TEQ were calculated to be 45, 111, and 66 days, 

respectively. 

Concentrations of CDDs and CDFs were also reported in wood ducks (Aix sponsa), species of migratory 

waterfowl collected near Bayou Meto, Arkansas (White and Hoffman 1995). The EPA identified a former 

2,4,5-T chemical manufacturing plant as the source of the contamination and subsequently listed the areas on 

the NPL of hazardous waste sites in 1982. Residues in wood duck eggs based on 2,3,7,8-TCDD (TEQs) ranged 

up to 611 ppt, and the egg arithmetic means were 90-fold higher at the site nearest the point source discharge 

compared to the reference site. The State of Arkansas has issued a wildlife consumption advisory for wood 

ducks in the Bayou Meto area (EPA 1998). 

CDDs were determined in pooled samples of ringed seal (Phoca hispida) blubber, beluga whale (Delphinapterus leucas) blubber, and polar bear (Ursus maritimus) liver and fat collected from several areas throughout 

the Canadian north (Norstrom et al. 1990). All seal samples and all but one polar bear sample had detectable 

levels of 2,3,7,8-TCDD (wet weight) ranging from 2 to 37 ppt, but 2,3,7,8-TCDD was not 
 

CDDs 457 

5. POTENTIAL FOR HUMAN EXPOSURE 

found in beluga blubber (<2 ppt). All seal samples and one of the three beluga whale samples contained 2,3,7,8­

TCDF (2-7 ppt), but 2,3,7,8-TCDF was not found in polar bear samples. OCDD concentrations in seal blubber 

and polar bear samples ranged from not detected (<8 ppt) to 43 ppt. No biomagnification of TCDD and OCDD 

occurred from seal to bear fat. The highest concentrations of 2,3,7,8-TCDD and OCDD in seals and bears were 

found in the central Canadian Arctic Archipelago, and the lowest concentrations were found in the Hudson Bay 

area. The reason for higher concentrations of 2,3,7,8-TCDD and OCDD in the Arctic than in sub-Arctic areas is 

thought to be transpolar movement of aerosols from combustion-related sources originating in Eurasia 

(Norstrom et al. 1990). CDDs and CDFs were determined in caribou tissue samples from 7 herds across the 

Canadian Arctic (Hebert et al. 1996). In contrast to marine mammals, concentrations for caribou were 

extremely low, sub-ng/kg (lipid basis), for all congeners except OCDD and 1,2,3,7,8-PeCDD in one herd. 

OCDD was found in most of the samples at concentrations ranging from < 0.2 ng/kg in fat to 4.7 ng/kg in 

adipose tissue. The one pooled liver sample analyzed from the Yukon had an OCDD concentration of 11 ng/kg 

(lipid basis). 2,3,7,8-TCDD was detected in adipose tissue samples of two herds in the eastern Canadian Arctic 

at levels of 0.73 and 0.14 ng/kg, but was not detected in tissue samples from other herds at detections limits as 

low as 0.03 ng/kg (lipid basis). CDF levels were sub-ng/kg in all cases. TEQs were dominated by non-ortho 

substituted PCBs in all cases, and ranged from 0.33 ng/kg to 3.29 ng/kg in adipose tissue. The authors 

concluded that caribou tissues are therefore less contaminated than tissues from marine mammals. 

Consumer products. 

Cigarettes and Cigarette Smoke. CDDs have been detected in cigarettes and cigarette smoke. In a recent 

study, Lofroth and Zebuhr (1992) detected CDD/CDF concentrations in both mainstream (collected directly on 

a glass fiber filter) and sidestream smoke (emitted into an acrylic box and then collected on a glass fiber filter) 

from a single brand of commercially available Swedish cigarettes. These authors reported that the mainstream 

smoke from 20 cigarettes contained about 18 pg TEQ (1 pg TEQ per cigarette), while sidestream smoke 

contained 39 pg TEQ (2 pg TEQ per cigarette). No particular isomer contributed more than 20% to the total 

TEQ value. Most isomers were not present at concentrations above the detection limits (0.3–1.3 pg), with the 

exception of 1,2,3,4,6,7,8-HpCDD (6.8 pg), 1,2,3,4,6,7,8-HpCDF (4 pg), and OCDD (7.3 pg). An earlier study 

that used low-resolution mass spectrometry for analysis of CDDs in cigarette smoke obtained by a continuous 

smoking process (all cigarette tobacco gave rise to mainstream smoke) found that HpCDD was the most 

abundant homologue detected, accounting for >90% of the total CDDs (Muto and Takizawa 1989). 
CDDs 458 

5. POTENTIAL FOR HUMAN EXPOSURE 

Paper Products. CDDs are formed during pulp bleaching, and as a result they have been found in many 

different types of paper products. 2,3,7,8-Substituted CDDs were determined in different samples of coffeefilter paper (Beck et al. 1988b, 1989d). 2,3,7,8-TCDD was the most abundant congener detected at a mean 

concentration of 3.85 ppt (range 1.6–7.3 ppt). OCDD was detected at a mean concentration of 2.05 ppt (range 

0.7–3.5 ppt). PeCDDs, HxCDDs, and HpCDDs were identified at concentrations ranging from 0.03 to 0.7 ppt. 

In an earlier study, HxCDD was the most abundant homologue detected in coffee filters (2.1 ppt) and 

2,3,7,8-TCDD was found at concentrations of 1 ppt (Beck et al. 1988b). Coffee brewed without filters did not 

contain any detectable CDDs; however, coffee brewed with one filter showed leaching of TCDDs from the 

paper into the coffee. Carryover (leaching) rates were 25% for 2,3,7,8-TCDD, indicating that people who drink 

coffee brewed with paper filters containing 5 ppt 2,3,7,8-TCDD may ingest small quantities (<5 pg) of 

2,3,7,8-TCDD per day (Beck et al. 1989d). Hashimoto et al. (1992) also analyzed CDD/CDF concentrations in 

coffee-filter paper available in Japan. These authors reported mean TEQ values of 0.89 pg/g (ppt) (range 

0.042–3.6 pg/g) for chlorine-bleached filters, 0.13 pg/g (range 0.079–0.18 pg/g) for oxygen-bleached filters, 

and 0.009 pg/g (range 0.00038–0.017 pg/g) for unbleached filters. Chlorine-bleached filters also gave the 

highest mean TCDD and TCDF elutions (range <0.043–2.1 and <0.081–6 pg/g, respectively). Oxygenchlorinated filters gave considerably lower elutions (range <0.085–0.14 and 0.23–0.26 pg/g, respectively), 

while unbleached filters produced elutions near the detection limits (<0.094 pg/g). Approximately one-third of 

the total CDD/CDF contamination was eluted from the filter paper into the coffee during brewing; however, 

almost the same amount was eluted from the filters with hot water. This leaching rate of approximately 30% 

agrees with that obtained by Beck et al. (1989d). The elution ratio was almost constant for all CDD/CDF 

congeners and isomers. Using the maximum TEQ value of 3.6 pg/g paper and the minimum TEQ value of 

0.00038 pg/g paper, the TEQ values for one cup of coffee were calculated to range from 0.000015 pg to 1.4 pg. 

The authors suggest that any potential health risk from CDD/CDF exposure from coffee-filter paper is small and 

can be further reduced by rinsing the filter prior to brewing the coffee. 

CDDs (TCDD, PeCDD, HxCDD, HpCDD, and OCDD) are also present in newsprint, facial (cosmetic) tissue, 

and recycled paper at levels ranging from <0.4 to 335 ppt (Beck et al. 1988b). OCDD was the most abundant 

congener detected in newsprint (37 ppt). HxCDD was the most abundant homologue detected in cosmetic 

tissue (79 ppt) and recycled paper (335 ppt). 2,3,7,8-TCDD was found at lower concentrations in cosmetic 

tissue (1.1 ppt), and recycled paper (0.6 ppt) (Beck et al. 1988b). Another study of CDDs in bleached and 

unbleached consumer paper products showed that the highest levels of 2,3,7,8-TCDD were found in bleached 

coffee filters (5 ppt), unbleached coffee filters (2.0 ppt), bleached shopping bags 
CDDs 459 

5. POTENTIAL FOR HUMAN EXPOSURE 

(1.3 ppt), and cigarette paper (1.4 ppt) (Wiberg et al. 1989). Concentrations were found to be lower in the 

unbleached products than in the corresponding bleached products (Wiberg et al. 1989). 2,3,7,8-TCDD has been 

measured in tea bags at concentrations ranging from not detected (0.36 ppt) to 4.79 ppt (Sullivan and Stanford 

1990). 

The use of newsprint for cow bedding was examined to determine if CDDs in the newsprint would find their 

way into the cow's milk as a result of the cow’s ingesting its bedding. Although HxCDD and OCDD were 

detected in the milk, TCDD was not detected (detection limit 0.5 ppt) (Shane et al. 1993). 

Ryan et al. (1992) analyzed the concentrations of CDDs/CDFs in Canadian bleached-paper milk containers 

from 1988 to 1989 and examined the resulting concentrations transferred to the milk. Milk-carton paper 

manufactured prior to 1989 tested positive for 2,3,7,8-TCDF and 2,3,7,8-TCDD, with levels on a TEQ basis 

varying between 1.4 and 55 ng/kg of paper. Bleached milk-carton paper produced after mid-1989 tested 

negative for these compounds at a limit of detection of 1 ng/kg paper. Storage of 3 types of milk in the 

pre-1989 low- and high-level cartons resulted in the transfer of the TCDD/TCDF into the milk, most of which 

occurred within the first 7 days. The TCDD/TCDF transfer varied between 3 and 25%, with whole and 2% fat 

milk accumulating about twice the concentrations of skim milk. On the basis of these results, milk stored for up 

to 14 days at 5 EC in currently produced bleached-paper containers with less than 1 ng TEQ/kg of paper would 

not contain any detectable CDDs/CDFs (<0.005 ng TEQ/kg milk). 

An FDA study of the migration of TCDD from paper products that come in contact with food found that TCDD 

was present in all paper products at concentrations ranging from 0.5 ppt for coated paper trays to 13 ppt for 

coated paper cups (average 2–8.5 ppt). Migration of TCDD from the paper into the food ranged from below 

detectable limits for coated juice cartons to 24% for coffee filters. Most CDDs migrated in the range of 4–8%. 

The TEQ estimated concentration values ranged from 1.5 ppt for coffee filters to 140 ppt for paper plates 

(Cramer et al. 1991). 

LeBel et al. (1992) analyzed a wide variety of paper products purchased from retail stores in Canada in 1988 

and 1991 for TCDD/TCDF through OCDD/OCDF. The congeners exhibiting the highest concentrations in 

most paper products were 2,3,7,8-TCDF, OCDD, and 2,3,7,8-TCDD. With respect to the TEQ values, the mean 

TEQ for disposable diapers increased from 1.4 to 2.0 pg/g from 1988 to 1991. The TEQ values decreased 

during the same period for facial tissues (5.2–4.0 pg/g), paper plates (6.4–2.2 pg/g), paper cups 

(22.7–10.5 pg/g), and coffee filters (3.7–0.1 pg/g). The CDD/CDF concentrations found in 
CDDs 460 

5. POTENTIAL FOR HUMAN EXPOSURE 

coffee filters in 1988 were similar to concentrations reported in the United States (Cramer et al. 1991) and in 

Germany (Beck et al. 1988b). The authors cautioned that, given the small number of both paper products and 

samples analyzed, these data do not permit them to draw any conclusions regarding trends in CDDs/CDF levels 

in paper products. 

Dyes and Pigments. Malisch (1994) reported the presence of CDDs/CDFs in colored candle wax produced 

with the dye pigment Violet 23, which is derived from chloranil. The three candle samples with the highest 

contamination contained 1.8, 1.4, and 0.8 ng TEQ/kg (ppt). The author also noted that candles of the same 

color could have highly different CDD/CDF concentrations based on the composition of dye pigments used in 

the manufacturing process. 

Three pigments used in fabric dyeing that are derived from chloranil include the dioxazine pigments Violet 23 

and Direct Blue 106 and 108 (Williams et al. 1992). Concentrations of the congeners OCDD and OCDF 

predominated in the pigment Blue 106 and ranged from 18,066 to 41,953 ng/g (ppb) for OCDD and 

1,006–12,463 ng/g (ppb) for OCDF. Pigment Blue 108 contained much lower concentrations of CDDs/CDFs, 

although OCDD and OCDF were also the predominant congeners detected at 23 and 11 ng/g, respectively. 

Violet 23 contained higher CDD/CDF concentrations than Direct Blue 108 but lower concentrations than Direct 

Blue 106. OCDD concentrations ranged from 806 to 11,022 ng/g (ppb), while OCDF concentrations ranged 

from 125 to 3,749 ng/g (ppb). The TEQ values for Direct Blue 106, Direct Blue 108, and Violet 23 were 35.4, 

0.1, and 9.1 ng/g (ppb), respectively. 

Textile Products. A recent study has identified sources of CDDs/CDFs found in textiles. Horstmann and 

McLachlan (1994a) detected CDD/CDF concentrations in new textile products ranging from less that 50 pg/g to 

as high as 290,000 pg/g. The authors believe that textile finishing processes are not the source of the high 

CDD/CDF concentrations because of the randomness of the textiles with high concentrations. Since PCP is still 

being used in developing countries, especially for purposes of preserving cotton during sea transport, the 

authors hypothesize that this is a likely source. 

Dry Cleaning Fluid Residues. Chemical analysis of dry cleaning solvent residues collected in Germany prior 

to 1993 indicated that residues from machines using perchloroethylene contained an average concentration of 

256 ppb CDD/CDF, with 2,3,7,8-TCDD being detected in 21 of 28 samples; however, the HpCDD and 

OCDD congeners comprised between 90 and 95% of the CDDs/CDFs found (Towara et al. 1992). 

Horstmann and McLachlan (1994b) detected CDD/CDF residues in used dry cleaning fluid and 
CDDs 461 

5. POTENTIAL FOR HUMAN EXPOSURE 

concluded that the source of the CDDs/CDFs residues in the dry cleaning fluid were introduced by dry 

cleaning new, unwashed textiles that had been treated with pentachlorophenol (PCP). 

Motor Vehicle Exhaust. CDDs have also been identified in automobile exhaust emissions (Marklund et al. 

1987, 1990). 2,3,7,8-TCDD was found in car exhaust from 4 Swedish cars running on leaded gasoline at 

levels ranging from <0.05 to 0.3 ng/24.8 km (0.002–0.01 ng/km) running cycle. PeCDD was also found in 

the exhaust of cars running on leaded gasoline at levels ranging from 6 to 98 ng/24.8 km (0.24–3.95 ng/km). 

No CDDs were found in samples where unleaded gasoline was used at detection limits of 0.05 ng 

(2,3,7,8-TCDD) and 0.3 ng (PeCDD) (Marklund et al. 1987). Another study of exhaust emissions from cars 

running on leaded and unleaded gasoline found total HpCDD concentrations ranging from not detected to 

0.482 ng/km and OCDD concentrations ranging from not detected to <0.510 ng/km for the cars running on 

leaded gasoline (Bingham et al. 1989). HpCDD was not detected in car exhaust emissions from a single car 

running on unleaded gasoline. OCDD concentrations ranged from not detected to <0.110 ng/km (Bingham et 

al. 1989). Most recently, however, Cirnies-Ross et al. (1996) reported that using copper in diesel fuels to 

reduce soot generation in engines was accomplished at the expense of increasing CDD/CDF during 

combustion. These authors reported that using the copper doped fuel significantly increased CDD/CDF 

particulate formation from < 20 ng/L TEQ in normal fuel to almost 60 ng/L TEQ in doped fuel at an engine 

output of 1 kW. The increased CDD/CDF particulate formation was most striking at low engine output (1 

kW). 

From the research conducted on CDD emissions from vehicles running on leaded and unleaded gasoline, it is 

clear that CDD emissions are typically less in cars running on unleaded gasoline. It should be noted however, 

that because the use of leaded gasoline is no longer permitted in the vast majority of domestic automobiles in 

the United States, this source of CDD emissions to the air should have been significantly reduced in recent 

years (EPA 1996a). 

5.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE 

5.5.1 General Population 

Currently, consumption of food (including human milk) is by far the most important pathway for exposure to 

CDDs for the general population representing over 90% of the total daily intake. Other pathways of exposure 

include inhalation of CDDs from municipal, medical, and industrial waste incinerators and other 
 

CDDs 462 

5. POTENTIAL FOR HUMAN EXPOSURE 

incineration and combustion processes (-2% of the daily intake), and ingestion of drinking water(<0.1% of 

the daily intake) (Travis and Hattermer-Frey 1987; Schaum et al.1994). 

Foods. Food is the major source (>90%) of human exposure to CDDs (Beck et al. 1989a; Hattemer-Frey 

and Travis 1989; Liem and van Zorge 1995; Rappe 1992; Schaum et al. 1994; Schecter et al. 1994d, 1994e, 

1996a). An estimate of the daily intake of 2,3,7,8-TCDD by adults in the general U.S. population from 

ingestion of contaminated food items and drinking water and inhalation of ambient air is given in Table 5-9. 

The average daily adult intake of 2,3,7,8-TCDD estimated by the model was 47 pg/day (Hattemer-Frey and 

Travis 1989) with a lower bound daily intake of 8 pg/day and an upper bound daily intake of 300 pg/day. 

Food, especially meat, and dairy products, accounted for 98% of the total daily intake of 2,3,7,8-TCDD. 

Hattemer-Frey and Travis (1989) estimated that the average daily intake of 2,3,7,8-TCDD for an adult in the 

United States from meat alone was 23 pg/day, accounting for 50% of the total daily intake of 2,3,7,8-TCDD 

from food sources. The average daily intakes of 2,3,7,8-TCDD from milk, produce, and fish were 13 pg/day 

(27%), 5 pg/day (11%), and 5 pg/day (10%), respectively of the total daily intake in the United States 

(Hattemer-Frey and Travis 1989). However, for certain subpopulations (recreational and subsistence fishers), 

fish consumption may be a more important source of CDDs. The maximum daily intake of 2,3,7,8-TCDD for 

residents of the Great Lakes region who regularly consume fish from the Great Lakes was estimated to range 

from 390 to 8,400 pg/day (EPA 1985a). Inhalation of ambient air and ingestion of water are not major 

pathways of human exposure, accounting for only 2% (1 pg/day) and <0.01% (6.5 x10-3 pg/day), 

respectively, of the total daily intake of 2,3,7,8-TCDD (Hattemer-Frey and Travis 1989). The percentage of 

daily intake of 2,3,7,8-TCDD estimated by Hattemer-Frey and Travis (1989) from each exposure pathway 

agrees closely with more recent estimates made by Schaum et al. (1994) for intakes of total CDDs/CDFs 

(Table 5-10). However, quantitatively, the estimates differ by a factor of 2–3 because Hattemer-Frey and 

Travis (1989) considered only 2,3,7,8-TCDD, while Schaum et al. (1994) based their estimates on all CDDs 

and CDFs. 

Based on their congener-specific analysis of 18 food samples collected in Binghamton, New York, 

Schecter et al. (1994d), estimated the U.S. mean daily exposure to CDD equivalents for an adult (65 kg 

body weight) to range from 18 to 192 pg TEQs depending on how not-detected values were treated. This 

is equal to a daily adult intake of CDDs/CDFs ranging from 0.3 to 3.0 pg TEQs/kg body weight. These 

authors reported that total CDDs ranged from 0.35 to 2.91 ppt (wet weight) in fish, from 0.6 to 59.3 ppt in 

meat products, and from 0.6 to 14 ppt in dairy products. The total CDD/CDF TEQ value ranged from 

0.023 to 0.13 ppt for fish, 0.03 to 1.5 for meat products, and 0.04 to 0.7 for dairy products. The authors 


CDDs 465 

5. POTENTIAL FOR HUMAN EXPOSURE 

reported that a vegetarian diet (vegan diet with no consumption of dairy products) might have health 

advantages by lowering daily intakes to only 2% of the level estimated for persons consuming fish, meat, 

and dairy products (Schecter et al. 1994d, 1984e). An ovo-lacto vegetarian diet that contains eggs and 

dairy products would not achieve this same reduction level. More recently, these same authors estimated 

the U.S. mean daily exposure to CDD equivalents based on an expanded analysis of 100 food samples 

collected in supermarkets in Binghamton, New York; Chicago, IL; Louisville, Kentucky; Atlanta, 

Georgia; and San Diego, California (Schecter et al. 1996a). For 1995, these authors report that the 

estimated U.S. mean daily exposure to CDDs/CDFs TEQs for an adult (65 kg body weight) ranges from 

34 to 167 pg TEQs. This is equivalent to a daily adult intake of CDDs/CDFs ranging from 0.52 to 2.57 

pg TEQs/kg body weight. If PCB TEQs are also considered (where TEF values are available), the daily 

adult intake ranges from 1.16 to 3.57 pg TEQ/kg body weight/day. A more recent survey of CDDs/CDFs 

in total diet food samples in Canada was conducted by Ryan et al. (1997). These authors found, through 

analysis of more than 100 food samples collected from commercial outlets in 1992 and 1993, that the total 

TEQ intake for CDDs/ CDFs was about 0.8 pg TEQs/kg/day. If all dioxin-like PCBs were also included, 

this TEQ value rose to approximately 1.2 pg TEQs/kg/day. 

In 1995, Schecter and Li (1997) conducted a congener-specific analysis of CDDs, CDFs, and dioxin-like 

PCBs in US fast foods. These authors reported TEQ values from 0.03–0.28 pg/g wet weight for 

McDonald’s Big Mac, 0.03–0.29 pg/g for Pizza Huts personal pan supreme pizza with all toppings, 

0.01–0.49 pg/g for Kentucky Fried Chicken 3 piece original recipe meal, and 0.3–0.31 pg/g for HaagenDaz chocolate-chocolate chip ice cream. Daily TEQ consumption per kilogram body weight assuming a 

65-kg adult, from one serving of each of the fast foods tested ranged between 0.046–1.556 pg/kg. This 

same value in a 20-kg child (6-year-old) ranged from 0.15 to 5.05 pg/kg. A child on average consumes 

three times more TEQs on a per kg/body weight basis as compared to adults eating any one of the fast 

foods tested. 

Studies conducted in other industrialized countries have reported similar values to those obtained for the 

United States. Estimated daily intakes of CDDs and CDFs from various foods were calculated in a 

Canadian study of foods domestically produced in Canada or imported from the United States (Birmingham 

et al. 1989). Based on contamination levels (CDDs and CDFs) in samples of meats, eggs, fruits, and 

vegetables from the United States and Canada, a total daily intake of 1.52 pg TEQs/kg body weight was 

calculated (for a 60-kg adult). The foods that contributed the most exposure to CDDs/CDFs TEQs were 

milk, eggs, and beef. Approximately one-half (0.81 pg TEQs/kg) was contributed by milk products. Eggs 
 

CDDs 466 

5. POTENTIAL FOR HUMAN EXPOSURE 

and beef were also estimated to make substantial contributions (0.28 and 0.27 pg TEQs/kg, respectively). 

The total contribution from these animal products to the daily dietary intake is 1.5 pg TEQ/kg body weight 

(99% of the total). Plant products (fruits, vegetables, and wheat products) contribute only 0.068 pg TEQ/kg 

body weight/day (1% of the total). The authors also estimated that consumption of freshwater fish was 

0.28 pg TEQ/day, thus the total daily intake of CDDs/CDFs amounted to 1.8 pg TEQ/kg body weight 

(Birmingham et al. 1989). 

In a German study where 22 samples of different foodstuffs (cow's milk, butter, pork, beef, lamb, chicken, 

eggs, and fish) were analyzed for CDDs/CDFs, the total average daily intake of 2,3,7,8-TCDD via food was 

0.35 pg/kg/day and for total CDDs/CDFs was 1.3 pg TEQs/kg body weight/day for a 70-kg adult (Beck et 

al. 1989a). Meat, milk and other dairy products, and fish were the most important food groups contributing 

17.9, 26.6, and 38.6 pg TEQ/day, respectively, to the body burden. Eggs, vegetable oil, vegetables, and 

fruits contributed 3.1, 0.3, 2.4, and 1.3 pg TEQ/day, respectively, to the body burden (Beck et al. 1989a). 

In general, vegetation contamination from airborne sources of CDDs results in more substantial exposures 

to grazing animals due to the proportionally higher accumulation from foliar (leaf) deposition as compared 

to root uptake (Travis and Hattemer-Frey 1987). OCDD is the CDD contaminant most concentrated by 

plants. This supports the contention that atmospheric deposition is the primary mechanism by which plants 

become contaminated, as OCDD is not readily available for root uptake or translocation in plants (Hulster 

and Marschner 1993; Muller et al. 1993). The concentration of 2,3,7,8-TCDD (due to root uptake and foliar 

deposition) on vegetation consumed by cows was estimated to be 0.11 ng/kg (97% due to foliar deposition). 

The estimated total concentration on exposed produce and vegetation consumed by humans was 0.02 ng/kg 

(67% resulting from foliar deposition) (Travis and Hattemer-Frey 1987). A model has been developed that 

estimated the CDD content in cow's milk based on emissions from a nearby municipal solid-waste 

incinerator. The model includes three components that predict atmospheric transport and deposition, soil 

and grass concentrations, and uptake and bioavailability from fodder to cows. Results indicate that models 

can be used to estimate CDD contamination in foods (Lorber et al. 1994; Slob and Jaarsveld 1993). 

Municipal and industrial incinerators and other combustion sources. Combustion processes 

are widely recognized as a source of CDDs/CDFs. Using a model, Hattemer-Frey and Travis (1989) 

estimated a total daily intake of CDD/CDF of 3×10-4 ng TEQs/day associated with exposure to a typical, 
 

CDDs 467 

5. POTENTIAL FOR HUMAN EXPOSURE 

state-of-the-art municipal solid-waste (MSW) incinerator, assuming a CDD/CDF emission rate based on the 

geometric mean from 11 proposed MSW facilities. Daily intakes of CDD/CDF in TEQs associated with 

exposure to a typical state-of-the-art municipal waste incinerator were estimated to be 1.3x10-4 ng/day from 

inhalation, 1.1x10-4 ng/day from total ingestion, 5.7x10-5 ng/day for mother’s milk and 2.2x10-6 ng/day from 

dermal absorption. This total daily intake value (3×10-4 ng TEQs/day) was 160 times lower than the 

estimated total daily background intake from all sources of CDDs (0.047 ng/day) to which the general U.S. 

population is exposed. Thus, the authors concluded that MSW incinerators will not substantially increase 

human exposure to CDDs/CDFs above normal background levels (Hattemer-Frey and Travis 1989). 

Table 5-11 shows estimated average daily intakes of CDD/CDF TEQs from various exposure pathways. 

Fries and Paustenbach (1990) evaluated the effects of 2,3,7,8-TCDD from incinerator emissions to humans. 

These authors also concluded that airborne emissions of CDDs/CDFs from modern waste incinerators that 

are equipped with appropriate air pollution devices should not pose a significant health hazard via 

inhalation of CDD contaminated particles or via contamination of foods regardless of the incinerator 

location. Hattemer-Frey and Travis (1989) focused on ideal state-of-the-art incinerators. In a more recent 

analysis, Travis and Hattemer-Frey (1991) estimated that the total daily intake of CDDs/CDFs (TEQs) by a 

maximally exposed individual living near a modern municipal solid waste incinerator was 0.7 pg/day (0.9% 

of total daily intake) and 92.8 pg/day (99.1%of total daily intake) was from all other background exposures. 

 These estimates are supported by recent data of Schecter et al. (1995) who found that workers who operate 

municipal waste incinerators have blood levels of TEQs which do not differ significantly from background 

levels. 

The presence of CDDs in cigarette smoke is also of importance with respect to inhalation exposure since 

cigarette smoke is inhaled directly into the lungs. Daily exposure to CDDs by smoking 20 cigarettes was 

estimated to be 18 TEQ pg/day equivalent to a daily intake of 0.26 pg/kg body weight/day (for a 70-kg 

adult) (Lofroth and Zebuhr 1992). 

Consumer products. The presence of CDDs in a variety of consumer products ranging from plastic 

packaging to colored candle wax, and from textiles to air filters for home-heating systems suggests that 

CDDs are virtually ubiquitous in the environment (Beck et al. 1989d; Berry et al. 1993; Horstmann and 

McLachlan 1994; Malisch 1994; Ryan et al. 1992). 2,3,7,8-TCDD and 2,3,7,8-TCDF have been found in 

many paper products, including coffee-filter paper, although present day paper products now contain less 

than 1 ng/kg TEQ. Under the preconditions of using 4 small coffee-filter papers (4×1 g) per day containing 

5 ppt 2,3,7,8-TCDD and 23 ppt 2,3,7,8-TCDF, which leaches into coffee, a daily exposure of 5 pg 

 

CDDs 469 

5. POTENTIAL FOR HUMAN EXPOSURE 

2,3,7,8-TCDD and 32 pg 2,3,7,8-TCDF or nearly 10 pg TEQs (for these two compounds combined) were 

calculated for a coffee drinker consuming all the coffee (this is a worst-case assumption that an individual 

would consume all the coffee brewed) (Beck et al. 1989d). TCDD, PeCDDs, HxCDD, HpCDD, and OCDD 

were found in all samples derived from consumer products (including plastic packaging, clothes dryer lint, 

vacuum cleaner dust, room and car air filters, and furnace filter dust), and bleached and unbleached paper 

products tested. In general, the more highly chlorinated congener groups (HxCDD, HpCDD and OCDD) 

exhibited the highest concentrations. The highest levels of most congeners were found in home-furnace 

filter dust, which contained HxCDD, HpCDD and OCDD at concentrations up to 135 ppt, 9,990 ppt, and 

24,600 ppt, respectively (Berry et al. 1993). Car air filters displayed a different CDD profile than the other 

products with the highest concentration detected for TCDD (2,080 ppt), PeCDD (1,320 ppt) and HxCDD 

(1,320 ppt). TEQ values for CDD/CDF were highest for the home furnace filter (170 ppt), car air filter (84 

ppt), and room air filter (29 ppt). 

Adipose tissue residues. The general population of the United States is continuously exposed to 

small amounts of CDDs, as exemplified by the fact that all human adipose tissue samples contain CDDs 

(Orban et al. 1994; Patterson et al. 1986a; Ryan et al. 1986; Schecter et al. 1986b; Stanley 1986; Stanley et 

al. 1986). Results of the National Human Adipose Tissue Survey (NHATS) conducted in 1982, which 

estimated the general population exposure to toxic organic chemicals, showed that 2,3,7,8-TCDD was 

detected in 35 of 46 (76%) composite samples with an average lipid-adjusted concentration of 6.2±3.3 ppt 

(Stanley 1986; Stanley et al. 1986). The average concentration of the other CDD compounds ranged from 

43.5 ppt for PeCDD (detected in 91% of the composites) to 694 ppt for OCDD (detected in 100% of the 

composites). The congener distributions found in adipose tissue are similar to those found in human milk 

(i.e., OCDD was the most abundant and 2,3,7,8-TCDD the least abundant congener). The analysis of 

46 composite adipose samples verified the prevalence of the 2,3,7,8-substituted tetra- through octa CDDs 

in the U.S. population (Stanley 1986; Stanley et al. 1986). The number of adipose samples in each 

composite was defined based on differences in age, gender, race, and regional affiliation of the individuals 

from whom the specimens were collected. The results also suggested that adipose tissue concentrations 

tended to increase with age for the congeners tested, with the exception of PeCDD. The NHATS study also 

showed regional differences in CDD concentrations in adipose tissue, with the greatest exposure occurring 

in the East North Central region of the United States (i.e., Ohio, Michigan, Indiana, Illinois, and 

Wisconsin). Exposure was also relatively high in the mid-Atlantic and East South Central regions (Phillips 

and Birchard 1991). 
 

CDDs 470 

5. POTENTIAL FOR HUMAN EXPOSURE 

Results of the more recent 1987 NHATS Study were summarized by Orban et al. (1994). Human adipose 

samples from autopsy cases were obtained through a network of pathologists to provide a representative 

sample of the general U.S. population. NHATS samples collected during 1987 were analyzed for 7 CDDs 

and 10 CDFs and the results are summarized in Table 5-12. Data were evaluated by census region, age 

group, sex, and racial group. The average concentration of 2,3,7,8-TCDD in adipose tissue in the U.S. 

population was estimated to be 5.38 pg/g (±6%). The 1987 survey data clearly show that nearly all of the 

CDD/CDF congeners increased with the age of the donor (i.e., the highest concentrations occur in the 

45+ age group and the lowest concentrations occur in the 0–14 age group). On a regional basis, only the 

average concentration of 2,3,4,7,8-PeCDF was statistically different in the Northeast (13.7 pg/g) compared 

to the national average (9.7 pg/g). Orban et al. (1994) also compared NHATS 1987 data to the NHATS 

1982 data. Because of slight differences in study design, the congeners that were most comparable between 

the two surveys were 2,3,7,8-TCDD and 2,3,7,8,-OCDD. Statistical analysis of the two survey data sets 

revealed no significant differences between the national average concentration of 2,3,7,8, TCDD 

determined in 1982 and 1987. There were also no significant differences in the profiles with respect to 

census region, sex, and race. With respect to age, however, there was a significant difference; the 1987 

NHATS data demonstrated that the concentration of 2,3,7,8-TCDD consistently increased with the age of 

the donor. The average concentration of 2,3,7,8-TCDD in the 1987 survey increased from 1.98 pg/g in the 

0–14-year-old group, to 4.37 pg/g in the 15–44-year-old group, to 9.4 pg/g in the 45+-year-old group. The 

average concentration of OCDD in the 1982 survey was 768 pg/g (±79.7 standard error) as compared to 

724 pg/g (± standard error 28.6) in the 1987 study. 

Analysis of human adipose tissue from 35 autopsy cases from Georgia and Utah found 2,3,7,8-TCDD in all 

of the samples at a concentration range (whole-weight) of 2.7–19 ppt (Patterson et al. 1986b). The 

geometric mean value for 2,3,7,8-TCDD in these samples on a whole-weight basis was 7.1 ppt. The 

geometric mean value for 2,3,7,8-TCDD in 31 of these samples on a lipid basis was 9.6 ppt. The histories 

of exposure to 2,3,7,8-TCDD were not known for any of the autopsy cases (Patterson et al. 1986b). 

Blood residues. CDDs/CDFs were measured in the blood (lipid basis) of 10 individuals in Germany 

with no prior CDD exposure (Päpke et al. 1989b). OCDD was the most abundant congener present (mean 

610.8 ppt; range 439–889 ppt), followed by 1,2,3,4,6,7,8-HpCDD (mean 88.2 ppt; range 30-142 ppt), 

HxCDD (mean 75.7 ppt; range 52–99.8 ppt), 1,2,3,7,8-PeCDD (mean 16.5 ppt; range 5.6–39 ppt), and 

2,3,7,8-TCDD (mean 4 ppt; range <1.5–9.1 ppt). Mean blood levels of CDFs ranged from 24 to 46.3 ppt. 

Detection limits for 2,3,7,8-TCDD were 1–4 ppt (extractable lipids) which corresponds to 0.005–0.02 ppt 

CDDs 472 

5. POTENTIAL FOR HUMAN EXPOSURE 

for whole blood with a lipid content of 0.5% (Päpke et al. 1989b). These authors reported that blood levels 

of CDDs/CDFs from persons with no known exposure to CDDs corresponded to those levels measured in 

adipose tissue of unexposed individuals. Needham et al. (1996) reported reference range data for CDDs 

and CDFs in blood of individuals who presumably had not been exposed occupationally to these 

compounds. The range of means in ppt (lipid basis) in blood for 7 CDDs, 11 CDFs, and 4 PCBs are 

summarized in Table 5-13. OCDD was the most abundant congener present (range of means 

560–1,000 ppt), followed by 1,2,3,4,6,7,8-HpCDD (range of means, 80.3–230 ppt), 1,2,3,6,7,8-HxCDD 

(range of means, 45–85 ppt), 1,2,3,7,8-PeCDD (range of means, 6.6–32 ppt), and 2,3,7,8-TCDD (range of 

means, 3.2–10.1 ppt). Mean blood levels of CDFs ranged from not detected to 27 ppt. The TEQ range in 

mean blood levels was 13.7–41.39 ppt for all CDDs and 15.1–58.0 ppt for CDDs/CDFs. 

Tepper et al. (1997) compared serum levels of CDD and CDF concentrations in 16 community residents 

who had no occupational exposure to these compounds. OCDD was the most abundant congener present 

(range, 285–1,489 ppt), followed by 1,2,3,4,6,7,8-HpCDD (range, 64.1–115 ppt), 1,2,3,6,7,8-HxCDD 

(range, 48.3–101 ppt), 1,2,3,7,8-PeCDD (range, 2–7.8 ppt), and 2,3,7,8-TCDD (range, 1.5–3.5 ppt). Mean 

blood levels of CDFs ranged from 0.8 to 31.3 ppt. The mean in TEQs was 13.5 ppt (range, 9.5–19.1) for all 

CDDs, 5.0 ppt (range, 3.4–8.8 ppt) for all CDFs, and 19.1 ppt (range, 12.9–25.9 ppt) for CDD/CDFs. 

Most recently, Michalek et al. (1998) measured levels of 2,3,7,8-TCDD in 1,302 unexposed Air Force 

Vietnam-era veterans. These veterans served as controls in the 20-year epidemiologic study of Air Force 

veterans of Operation Ranch Hand, the unit responsible for aerial spraying of Agent Orange in Vietnam. 

These authors reported mean 2,3,7,8-TCDD concentrations in blood of 4.32±2.53 ppt for the control group. 

The 99th percentile of the distribution was less than or equal to 10.4 ppt. 

Needham et al. (1991) also showed that human adipose tissue concentrations of CDDs may be correlated 

with blood serum levels after adjusting for total lipid content. On a lipid basis, total CDD/CDFs are higher 

in blood than adipose tissue. Partitioning is not identical in these tissues; 2,3,7,8-TCDD levels are almost 

identical in blood and adipose tissues, but OCDD levels are higher in blood. However, the presence of 

OCDD at levels of 5,000–10,000 pg/person when concentrations in food are generally in the low pg/g level 

suggests that the contribution of food to the OCDD body burden in humans requires further study (Rappe 

1993). 

CDDs 474 

5. POTENTIAL FOR HUMAN EXPOSURE 

5.5.2 Occupational Exposure 

Occupational exposure to CDDs occurs primarily through inhalation and dermal contact of fire fighters and 

cleanup workers involved with transformers containing PCBs and polychlorobenzenes; in workers involved 

in incineration operations; in workers in metal reclamation facilities, and in workers producing and handling 

pesticides, hexachlorophene, trichlorophenol, or other chlorinated compounds (e.g., pentachlorophenol) that 

may contain small impurities of 2,3,7,8-TCDD or other CDDs (Päpke et al. 1992). In addition, these 

authors reported that the CDD/CDF homologue profiles in whole blood of workers engaged in a variety of 

different chemical processes or in occupational accidents exhibited distinct CDD/CDF patterns (Päpke et al. 

1992). 

A dust sample collected from the ambient atmosphere of a municipal incineration plant in Europe over a 

6-day period was analyzed for CDDs. The concentrations of 2,3,7,8-substituted congeners ranged from 

0.9 ppb (2,3,7,8-TCDD) to 310 ppb (1,2,3,4,6,7,8-HpCDD) (Tong et al. 1989a). These results indicate that 

the ambient atmosphere of a municipal incinerator can be contaminated by CDDs by means of fly ash and 

possibly other incineration products; thus, municipal incinerator workers are at risk of exposure to CDDs 

(Tong et al. 1989a). Blood analysis of 10 workers at a municipal solid-waste incinerator in Germany 

showed elevated CDD levels of 4 workers whose work was associated with exposure to fly ash and slag 

(Päpke et al. 1993). In several individuals studied, the higher chlorinated CDDs/CDFs especially the 

2,3,7,8-substituted HxCDD, HpCDD, and OCDD congeners showed slightly elevated levels. 

Compared with background 2,3,7,8-TCDD levels (3.6 ppt), workers involved in trichlorophenol 

production had elevated 2,3,7,8-TCDD blood levels, with a mean concentration of 332 ppt (Päpke et al. 

1992). PCP manufacturing resulted in the greatest increases for workers with respect to all congeners, 

with OCDD blood levels of approximately 300,000 ppt. PeCDF, HxCDF, and HpCDF levels in the blood 

were elevated in workers at a metals reclamation plant (Päpke et al. 1992). Workers exposed to CDD as a 

result of an industrial accident had mean 2,3,7,8-TCDD blood levels of 53 ppt almost 36 years after the 

incident (Päpke et al. 1992). In one documented case, a U.S. domestic agricultural worker was exposed to 

2,3,7,8-TCDD during spraying of 2,4,5-T herbicide on pasture land and hay ground. A sample of the 

herbicide that he used contained 7.7 ppb 2,3,7,8-TCDD. 2,3,7,8-TCDD levels measured in the worker's 

adipose tissue 5 years post-exposure were 72 ppt (whole weight) or 77 ppt (lipid basis) (Tong et al. 

1989b). Thirty-two years after an industrial accident in a chemical plant manufacturing trichlorophenol, 

the average lipid-adjusted concentration of 2,3,7,8-TCDD in the adipose tissue of exposed workers who 
CDDs 475 

5. POTENTIAL FOR HUMAN EXPOSURE 

developed symptoms (chloracne and other illnesses) was 49 ppt (range of 11-141 ppt) (Schecter and Ryan 

1988). 

In a recent study by Tepper et al. (1997), serum levels of CDDs and CDFs were measured in pulp and 

paper mill workers in the United States. These authors reported that serum levels of CDDs and CDFs 

among 46 long-term workers at a pulp and paper mill were not appreciably different among three exposure 

groups studied (community residents, low-exposure-potential worker group, and high-exposure-potential 

worker group). Serum CDD TEQs were 13.5 ppt (range, 9.5–19.1 ppt), 15.9 ppt (range, 6.5–31.8 ppt), and 

13.3 ppt (range, 7.5–24.9 ppt) respectively. Total TEQ for both CDDs and CDFs were similar for the three 

groups at 19.1 ppt, 21.2 ppt and 18.1 ppt, respectively. Serum levels of CDDs and CDFs in this study 

were within the range previously reported for persons with no known occupational exposure. 

A series of adipose tissue samples collected from one exposed individual, as well as surgical and autopsy 

specimens from four control individuals, was analyzed for CDDs (TCDD, PeCDD, HxCDD, HpCDD, and 

OCDD) (Schecter et al. 1985a). All specimens were obtained from persons residing in urban or rural areas 

of upstate New York during 1983 or 1984. The worker who had been exposed to soot containing PCBs, 

CDFs, and small amounts of CDDS from the CDD/ CDF-contaminated Binghamton State Office Building 

in New York, had a total CDD concentration (whole-weight basis) of 1,015 ppt, whereas the average total 

CDD concentration for the controls was 765 ppt. Mean concentrations were highest for OCDD among all 

the CDD congener groups in both the controls (585 ppt) and the exposed person (690 ppt). 2,3,7,8-TCDD 

concentrations were lowest in both groups, an average of 6.3 ppt for the controls and 11.6 ppt for the 

exposed person. Intermediate levels were found for PeCDD (7.5–13.8 ppt), HxCDD (6.8–64.2 ppt), and 

HpCDD (2.6–119 ppt) in the control groups. Intermediate levels were also found in the exposed 

individual for PeCDD (15 ppt), HxCDD (7.3–72.6), and HpCDD (9.6–209 ppt) (Schecter et al. 1985a). 

An occupational study of workers exposed to CDDs at a Missouri chemical plant from 1968 to 1972 found 

a mean 2,3,7,8-TCDD concentration of 390 ppt in the adipose tissue of 4 exposed workers measured 

13–17 years post-exposure. The chemical plant made 2,4,5-trichlorophenol (2,4,5-TCP), which was used 

as a feedstock to produce butyl esters of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T esters) and hexachlorophene between 1968 and 1972. The 2,3,7,8-TCDD was generated as an unintended contaminant during 

the production of 2,4,5-TCP. Consequently, workers involved in these processes were potentially exposed 

to 2,3,7,8-TCDD. The mean concentration of 2,3,7,8-TCDD found in these workers was 45 times higher 

than the mean of 8.7 ppt reported for 7 unexposed Missouri residents (Patterson et al. 1989a). 
CDDs 476 

5. POTENTIAL FOR HUMAN EXPOSURE 

Another group that is occupationally exposed to CDDs includes analytical chemists involved in the 

synthesis of CDDs for research purposes or those involved in analysis of environmental samples 

contaminated with CDDs. Schecter et al. (1994b) reported that a chemist who had been involved in 

synthesizing 2,3,7,8-TCDD in 1956 at a university research laboratory had subsequently developed 

chloracne, headaches, backaches, and severe leg pains when he walked. During 1990–91, approximately 

35 years after the chemist's initial exposure, blood CDDs analysis was performed to determine whether 

residual CDD levels still remained. Schecter et al. (1994b) reported that blood TCDD levels (on a bloodlipid basis) were 20 ppt as compared to 3–5 ppt measured in a control population of 100 individuals. 

Analytical laboratory personnel who are involved in analyzing CDD-contaminated samples also may be 

exposed to higher levels of CDD contamination than the general population (Hesso et al. 1992; Oliver 

1975). 

In a study conducted by NIOSH, serum levels of 2,3,7,8-TCDD were measured in 27 U.S. chemical 

workers previously exposed to CDDs-contaminated products (Fingerhut et al. 1989). The workers were 

employed at two U.S. facilities that produced 2,4,5-TCP and 2,4,5-T between 1951 and 1972. Serum 

levels of 2,3,7,8-TCDD were also measured in 19 unexposed controls. A mean serum 2,3,7,8-TCDD level 

of 208.2 ppt found in the exposed workers exceeded (by more than 25 times) the mean background level of 

8.2 ppt found in the controls who did not produce these chemicals (Fingerhut et al. 1989). 

Workers who are involved with incineration operations may be exposed to levels of CDDs that are higher 

than background levels to which the general population is exposed. Schecter et al. (1991b) measured CDD 

and CDF blood levels on a lipid basis in pooled blood samples from a group of 56 New York City 

incinerator workers and 14 controls. The levels of 11 of the 18 CDD/CDF congeners measured were 

increased in the incinerator workers as compared to the controls. CDD levels in incinerator workers were 

48, 17, 27, 30, and 31% higher for 1,2,3,7,8-PeCDD, 1,2,3,6,7,8-HxCDD, 1,2,3,7,8,9-HxCDD, 

1,2,3,4,6,7,8-HpCDD, and OCDD, respectively. Only 2,3,7,8-TCDD and 1,2,3,4,7,8-HxCDD were lower 

in incinerator workers’ blood than in controls (5 and 15% lower, respectively). Overall, the total 

CDD/CDF level in workers’ blood was, 1,007.2 ppt (lipid basis) as compared to 747.3 ppt for the controls 

(Schecter et al. 1991b). In the past, workers involved in the production or use of hexachlorophene, 

trichlorophenol, 2,4,5-T, and other compounds that are no longer used were also exposed to 

2,3,7,8-TCDD. Workers in pulp and paper mills also have the potential for 2,3,7,8-TCDD exposure 

because of the occurrence of 2,3,7,8-TCDD in bleached kraft paper-making processes (Clement et al. 

1989; Kuehl et al. 1987a; Tepper et al. 1997); although exposure to this source has probably declined since 
 

CDDs 477 

5. POTENTIAL FOR HUMAN EXPOSURE 

1990 with the implementation of bleach plant modifications at many pulp and paper mills (NCASI 1993). 

Workers in the sawmill industry who handle treated lumber may be exposed to chlorophenols, particularly 

PCP; consequently, they may be exposed to higher levels of the more highly chlorinated CDDs 

(Kalliokoski and Kauppinen 1990). Workers employed at sites of improper chemical waste disposal 

(trichlorophenol, hexachlorophene, 2,4,5-T) have a greater potential for exposure to 2,3,7,8-TCDD via 

inhalation or via oral or dermal contact than the general population. 

A current estimate of the number of workers in the United States that are potentially exposed to CDDs is 

not available. At-risk worker populations include incinerator personnel, those involved in production or 

use of chlorinated compounds containing CDD contamination (e.g., hexachlorophene, PCP, 

2,4,5-trichlorophenol, and 2,4-D), analytical research chemists, and workers at chemical waste disposal 

sites, electrical utility workers, and firefighters. 

5.6 EXPOSURES OF CHILDREN 

This section focuses on exposures from conception to maturity at 18 years in humans and briefly considers 

potential pre-conception exposure to germ cells. Differences from adults in susceptibility to hazardous 

substances are discussed in Section 2.6, 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, and 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; they put things in their mouths; they 

may ingest inappropriate things such as dirt or paint chips; they spend more time outdoors. Children also 

are closer to the ground, and they do not have the judgement of adults in avoiding hazards (NRC 1993). 

Children are primarily exposed to CDDs in the same manner as adults in the general population (i.e., via 

consumption of foods contaminated with small amounts of CDDs, particularly meat, milk and other dairy 

products, and fish). Children that are at additional risk of exposure primarily through dietary habits, 

include: infants and young children who are breast-fed; children of recreational and subsistence fishers, 

who typically consume larger amounts of locally caught fish and shellfish than the general population; 
CDDs 478 

5. POTENTIAL FOR HUMAN EXPOSURE 

children of subsistence hunters, particularly those in the high latitudes, who typically consume large 

amounts of locally caught game especially marine mammals; and children of subsistence farmers living in 

areas contaminated with CDDs (either by waste incinerators or the use of CDD- contaminated sewage on 

their land) who exclusively consume their own farm-raised beef and dairy products (see Section 5.7). 

The human fetus is exposed to CDDs/CDFs through transplacental transfer from the mother. Schecter et 

al. (1990a) reported 2,3,7,8-TCDD concentrations in liver tissue of three still-born infants ranging from 

0.03 to 0.18 ppt (whole weight basis) and 1.3 to 4.3 ppt (lipid weight basis). Schecter et al. (1990a) also 

reported CDD/CDF concentrations in liver tissue of three stillborn infants ranging from 2.1 to 4.92 ppt 

(whole weight basis) and 98 to 104 ppt (lipid weight basis). The TEQ for CDD/CDFs combined ranged 

from 0.14 to 0.49 ppt (whole weight basis) and 6.4 to 12 ppt (lipid weight basis). In a more recent study, 

Schecter et al. (1996e) reported TEQs for CDDs/CDFs in placental material ranging from 8.4 to 17.6 ppt 

(lipid basis). In a pooled sample of fetal tissue (8–14 weeks), the TEQ was 5.3 ppt (lipid basis). 

Concentrations of 2,3,7,8-TCDD in adipose tissue and liver were also reported by Kreutzer et al. (1997) 

for stillborns at levels between 0.2 and 0.8 ppt and 0.3 to 0.7 ppt, respectively. Kreutzer et al. (1997) 

developed a pharmacokinetic model for 2,3,7,8-TCDD that predicted a decrease in body burdens during 

the first year for non-breast-fed infants and this was supported by empirical data (see Section 2.3.4.4, 

Transfer of CDDs Through the Placenta and Breast Milk). 

In addition to transplacental transfer, CDDs and CDFs have been found in human milk (Fürst et al. 1992; 

Ryan et al. 1993a; Schecter and Gasiewicz 1987b; Schecter et al. 1986a, 1989d, 1989e, 1989g, 1991a); 

human milk is thus a potential source of CDDs for nursing infants and children (see Section 5.5). In 

Binghamton, New York, and Los Angeles, California, human breast milk was found to contain almost 

identical levels of detectable CDDs on a lipid basis probably because food consumption and sources are 

similar across the United States (Schecter et al. 1989e). Mean values of two pooled samples (n=42) from 

both cities showed that OCDD was the most abundant congener present (233 ppt), followed in decreasing 

order by total HxCDD (42.65 ppt), 1,2,3,4,6,7,8-HpCDD (42 ppt), 1,2,3,6,7,8-HxCDD (30.5 ppt), 

1,2,3,7,8-PeCDD (6.7 ppt), 1,2,3,7,8,9-HxCDD (6.2 ppt), 1,2,3,4,7,8-HxCDD (4.95 ppt), and 

2,3,7,8-TCDD (3.3 ppt). The total CDDs value was reported as 327 ppt. The TEQ for CDDs/CDFs, but 

not PCBs in breast milk in the United States was 17 ppt (Schecter et al. 1989e). Between 1986 and 1987, 

concentrations of CDDs found in breast milk sampled from Canadian women ranged from 2.2 ng/kg (ppt) 

(lipid basis) for TCDDs to 173 ppt for OCDD. In addition, the combined CDD/CDF mean TEQ of 

15.6 ppt (lipid basis) declined from a TEQ of 24.7 ppt measured in 1981–1982 (Ryan et al. 1993a). 
CDDs 479 

5. POTENTIAL FOR HUMAN EXPOSURE 

CDD/CDF concentrations also have been measured in breast milk in several foreign studies. Concentrations of CDDs/CDFs were also measured in preserved breast milk from 505 persons (both primiparas and 

multiparas) in Japan from 1978 to 1984 (Ogaki et al. 1987). OCDD was a major component in primiparas' 

milk (789 ppt) and multiparas' milk (518 ppt) (Ogaki et al. 1987). For the primiparas, average 

concentrations found for HpCDD, HxCDD, PeCDD, total TCDD, and 2,3,7,8-TCDD were 150, 76, 15, 37, 

and 13 ppt, respectively. The average CDD concentrations found in the milk of multiparas were generally 

lower than the average concentrations found in the milk of primiparas. For the multiparas, average 

concentrations for HpCDD, HxCDD, PeCDD, and TCDD (not 2,3,7,8) were 75, 56, 11, and 19 ppt, 

respectively (Ogaki et al. 1987). A similar study in Germany between 1984 and 1991 found mothers 

nursing their second child had 20% less CDD TEQs in their milk than did primigravidae (Fürst et al. 

1992). CDD concentrations in human milk can be directly correlated with the age of the mother and the 

amount of animal (but not vegetable) fat and protein consumed, suggesting that meat, milk and other dairy 

products, and fish are the major sources of CDD intake (Pluim et al. 1993a). The fact the CDD concentrations in milk fat were significantly related to age is in agreement with the results of Stanley et al. (1986) 

and Orban et al. (1994) who reported a strong correlation between age group and CDD levels in adipose 

tissue in the general U.S. population. The positive correlation can be expected because of the long half-life 

of CDDs in humans (7–11.3 years) (Pirke et al. 1989; Wolfe et al. 1994). 

Estimated daily intakes of CDD/CDF TEQs by nursing infants in the United States have been reported by 

Schecter and Gasiewicz (1987a). The daily intake by nursing infants in the United States was estimated to 

be 83.1 pg TEQs/kg body weight/day. To determine this daily intake, various assumptions were made 

regarding infant body weight (10 kg), duration of nursing, average amount of milk consumed, and 

gastrointestinal absorption. It was also assumed that breast milk was the only source of CDDs while the 

infant was nursing during the first year of life. From results of earlier studies that determined the 

concentrations of CDDs/CDFs in human breast milk in the United States (Schecter et al. 1989e) and in 

cow’s milk and soybean-derived infant formula sold in the United States (Schecter el al. 1989c) (see 

Section 5.4.4), Schecter et al. (1994e) estimated slightly lower intakes of 35–53 pg TEQ/kg of body 

weight/day for infants (7.3 kg) that were breast-fed within the first year of life as compared to 0.07–0.16 

pg TEQ/kg of body weight for infants who were fed soy formula. 

Exposure of infants and young children to CDDs may be very high because of their relatively high 

consumption of milk, including breast milk (ECETOC 1992). Schecter et al. (1994e) evaluated the intake 

of CDDs/CDFs from human breast milk and estimated that high levels reported for breast milk in the 
CDDs 480 

5. POTENTIAL FOR HUMAN EXPOSURE 

United States (.17 ppt TEQ on a lipid basis) contribute 35–53 pg TEQ/kg of body weight per day to the 

nursing infant in its first year of life (Schecter et al. 1989e). The CDD concentrations in cow’s milk and 

soy-based formula were much lower than in breast milk (327 ppt) (Schecter et al. 1991a). The following 

concentrations for CDDs (on a lipid basis) were reported: cow’s milk (25.1 ppt), 2% cow’s milk (32.3 ppt), 

SimilactTM infant formula (39 ppt), IsomilTM infant formula (23.3 ppt), and ProsobeeTM infant formula (42.7 

ppt) (Schecter et al. 1989c). The TEQ values for cow's milk and soy-based infant formula were also much 

lower than for human breast milk (.17 ppt). The corresponding TEQ values for CDDs/CDFs (on a lipid 

basis) were reported: cow's milk (2.1 ppt), 2% lowfat cow's milk (0.79 ppt), SimilacTM infant formula 

(0.08 ppt), IsomilTM infant formula (0.05 ppt), and ProsobeeTM infant formula (0.127 ppt) (Schecter et al. 

1989c). Schecter and Gasiewicz (1987a, 1987b) calculated TEQ values for CDDs/CDFs in human milk in 

two populations in Vietnam and in the general population in the United States. These authors reported 

mean values during the 1980s of 1.04 pg TEQ/g (whole milk basis) for the United States (maximum 

4.72 pg TEQ/g), a mean of 1.11 pg TEQ/g for South Vietnamese (maximum value 4.38 pg TEQ/g) 

exposed to Agent Orange sprayed between 1962 and 1970, and a mean of 0.065 pg TEQ/g (maximum 

value 0.18 pg TEQ/g) for a North Vietnamese population that was not exposed to Agent Orange. These 

authors concluded that some infants in the United States (whose mothers had CDD milk concentrations in 

the upper range of measured values) were being exposed to mean concentrations comparable to levels 

observed in the South Vietnamese population exposed to Agent Orange (Schecter and Gasiewicz 1987a, 

1987b). 

The highest exposure to CDD-contaminated breast milk reported was associated with the widespread use 

of Agent Orange as a defoliant during the Vietnam War. Human milk specimens from Ho Chi Minh City 

and Song Be Province in South Vietnam had lower 2,3,7,8-TCDD values in the late 1980s (7.1 and 17 ppt 

lipid basis) (TEQ values of 18.5 and 31.7 ppt), respectively, than they did in the 1970s when Agent Orange 

spraying occurred (Schecter et al. 1989e). A 1970 mean value for 2,3,7,8-TCDD in human milk in South 

Vietnam was reported to be 484.9 ppt (range, not detectable to 1,450 ppt) (Baughman and Meselson 1973; 

Schecter et al. 1986a). These values serve as reference values for the highest levels of 2,3,7,8-TCDD 

documented in human milk (Schecter et al. 1989e). Estimated daily intakes of TEQs by nursing infants 

from Vietnam have been reported (Schecter and Gasiewicz 1987a). The estimated daily intake by nursing 

infants in southern Vietnam in 1970 was 908 pg TEQs/kg body weight/day, whereas the daily intakes in 

southern and northern Vietnam in 1984 were 88.7 and 5.1 pg TEQs/kg body weight/day, respectively. 

Analysis of 9 milk samples from individuals living in northern Vietnam showed no detectable 

concentrations of 2,3,7,8-TCDD (detection limit 2 ppt) (Schecter and Gasiewicz 1987a). To determine 
CDDs 481 

5. POTENTIAL FOR HUMAN EXPOSURE 

these daily intakes, various assumptions were made regarding infant weight, duration of nursing, average 

amount of milk consumed, and gastrointestinal absorption. It was also assumed that breast milk was the 

only lifetime source of exposure to CDDs during the first year of life. In another study, Tarkowski and 

Yrjanheikki (1989) evaluated the health risks associated with human milk. These authors concluded that 

levels of CDD/CDFs in breast milk did not present a health risk to infants and children and that there was 

no justification for limiting breast-feeding. However, these authors believed there was a need for primary 

prevention of CDD/CDF exposure in humans. Because of the relatively short period of intake and the 

accepted benefits of breast-feeding, the World Health Organization did not recommend limitations on 

breast-feeding at the levels of background exposures to CDDs and CDFs (WHO 1991). More recently, 

Pohl and Hibbs (1996) reviewed recent studies indicative of a possible link between development of subtle 

health effects in children and their exposure to CDDs and CDFs from maternal milk. It is the ATSDR 

position that for background exposures, the benefits of breast feeding outweigh any potential risk 

associated with exposure. For higher CDDs levels in breast milk, the safety of breast-feeding may be of 

concern in some cases. 

Two recent studies have looked at ways to reduce CDD exposure in breast-fed infants. Koppe (1995) 

reported that exposure before and after birth to CDDs and PCBs has given rise to subtle abnormalities 

(disturbed cognitive development and delayed motor development) in approximately 10% of newborns 

in the Netherlands. This author examined possibilities of reducing this exposure by influencing the 

diet of the lactating mother. Mobilization of fatty acids from adipose tissue will cause release of stored 

CDDs which will then be secreted in breast milk. Two maternal diets were tested for their ability to 

reduce concentrations of CDDs in human milk. One diet was a low fat/high carbohydrate/low CDD 

diet, while the second was a high fat/low carbohydrate/low CDD diet. Despite significant changes in 

fatty acid profiles of the milk, no significant changes in CDD concentrations in breast milk were 

observed. The author concluded that short-term dietary measures will not reduce CDDs in breast milk. 

A lowering of CDD intake must occur years before the woman becomes pregnant. An important food 

source for the women is cow’s milk and other dairy products and these are responsible for about half 

of the daily exposure CDDs and PCBs in women in the Netherlands, so levels of the compounds in 

dairy foods must be lowered. In addition, the author believes that a lowering of CDD concentrations 

in fish is also necessary. Based on the results of his dietary study, Koppe (1995) reported that daily 

dietary intake of CDDs during lactation represents only 14% of the daily secretion of CDD in breast 

milk, while 86% was derived from CDDs stored in adipose tissue. Thus, reducing dietary intake of 

CDDs during lactation would only reduce CDDs in milk by 14%. Schlaud et al. (1995) also reported 

that to reduce organochlorine residue levels including 
CDDs 482 

5. POTENTIAL FOR HUMAN EXPOSURE 

CDDs in human breast milk in the short-term, nursing mothers should be advised not to try to reduce 

their body weight until after lactation. These authors reported statistically significant positive 

associations between breast milk contamination and average dietary fat intake per week (P=0.001) and 

proximity of residence to hazardous waste sites (P<0.05) for CDDs. These authors believe that public 

promotion of a lower dietary fat intake may reduce the lifetime accumulation of CDDs in human fatty 

tissues and in the long-term, resulting in lower concentrations in breast milk as well. 

In addition to exposure to CDDs through consumption of breast milk, cow’s milk, and soy-based 

infant formula, older children can be exposed through dietary practices similar to those of adults in the 

general population (see Section 5.4.4). One study has looked at the exposure that might occur in a 

6-year-old child who consumes “fast foods.” In 1995, Schecter and Li (1997) conducted a congenerspecific analysis of CDDs, CDFs, and dioxin-like PCBs in U.S. fast foods. These authors reported a 

CDD/CDF TEQ value, depending on the treatment of not detected congeners, from 0.03–0.28 pg/g wet 

weight for one McDonald’s Big Mac, 0.03–0.29 pg/g for one Pizza Hut personal pan pizza supreme 

with all toppings, 0.01–0.49 pg/g for one Kentucky Fried Chicken 3-piece original recipe meal, and 

0.3–0.31 pg/g for one Haagen-Daz chocolate-chocolate chip ice cream. The daily intake from one 

serving of each of the fast foods tested, assuming a 20-kg child (6 years old), ranged between 0.15 and 

5.05 pg TEQ/kg body weight. These authors calculated that, on average, a child (6 years old) 

consumes 3 times more TEQs on a per kg/body weight basis than an adult eating any one of the fast 

foods tested. 

As a result of the transfer of CDDs through the placenta to the fetus, by breast milk to infants and 

young children, and by lifelong dietary intakes from the consumption of meat, milk and dairy 

products, and fish, CDDs are found to be widespread in the adipose tissue of members of the general 

population (Orban et al. 1994). Human adipose samples from the recent 1987 NHATS Study provide 

a representative sample of CDD body burden in the general U.S. population (see Section 5.5.1). The 

average concentration of 2,3,7,8-TCDD in the U.S. population was estimated to be 5.38 pg/g (±6%). 

The 1987 survey data clearly show, however, that nearly all of the CDD/CDF congeners in adipose 

tissue increased with the age of the donor (i.e., the highest concentrations occur in the 45+ age group 

and the lowest concentrations occur in children in the 0–14 age group). The average concentration of 

2,3,7,8-TCDD in the 1987 survey increased from 1.98 pg/g in the 0–14-year-old group, to 4.37 pg/g in 

the 15–44-year-old group, to 9.4 pg/g in the 45+-year-old group. 
CDDs 483 

5. POTENTIAL FOR HUMAN EXPOSURE 

Children may be exposed to CDDs through a variety of lifestyle practices of their parents or of their 

own. For example, CDD/CDF concentrations have been reported in cigarette smoke (Lofroth and 

Zebuhr 1992; Muto and Takizawa 1989) (see Section 5.4.4). Young children and infants may be 

exposure to CDDs indirectly by inhalation of room air contaminated from cigarette smoking of their 

parents. In addition, older children and teenagers, may be directly exposed if they become smokers 

themselves. Malisch (1994) reported that some colored candle wax produced with certain dye 

pigments contained CDDs/CDFs. By burning these candles, CDDs could be released into room air 

and be an additional source of inhalation exposure for children. 

Children may also be exposed to CDDs by dermal contact with some new, unwashed clothing, 

particularly those manufactured in some developing countries or from fabric shipped from developing 

countries where pentachlorophenol (PCP) is used for preserving cotton fabrics during sea transport 

(Horstmann and McLachlan 1994). Exposures can be reduced by washing new clothes prior to 

wearing. 

Children could potentially be exposed to CDDs at home from a variety of incineration sources. For 

example, if their parents routinely burn domestic garbage containing scrap wood treated with PCP 

(Chiu et al. 1983) or untreated wood (Clement et al. 1985), old pesticide containers that may have 

contained 2,4,5 T or 2,4-D or Silvex (Arthur and Frea 1989), or polyvinylchloride (PVC) pipes or 

other plastics items (Lustenhouwer et al. 1980), or extensively use a wood stove (Clement et al. 1985), 

children may be exposed to higher levels of CDDs in outdoor and/or indoor air. Time spent in a garage 

where cars or trucks are being repaired and the engines are running, exposes children and teenagers to 

exhaust products and engine soot that may also contain CDDs (Bingham et al. 1989; Cirnies-Ross et 

al. 1996). 

Although there are many studies on the effects of CDDs on adults that receive occupational exposures 

(Fingerhut et al. 1989; Hesso et al. 1992; Patterson et al. 1989a; Schecter et al. 1985a, 1994b; Tepper 

et al. 1997), no information was located on the potential for workers in the United States to bring 

CDDs home on their clothing or shoes, thus contaminating other family members, including children. 

It is conceivable, however, that because CDDs are present in a variety of diverse occupational settings 

(see Section 5.5.2 Occupational Exposures), that poor occupational hygiene could result in CDDs 

being brought home and contaminating domestic dwellings. 

Children in populations with potentially high exposure living in the vicinity of former or current production 

sites where CDDs are released as by-products, (e.g., incinerators, other waste disposal facilities, and 
CDDs 484 

5. POTENTIAL FOR HUMAN EXPOSURE 

hazardous waste sites) may be exposed to CDDs by several pathways (see Section 5.7). Children may be 

exposed to CDDs in CDD-contaminated soils. Dermal absorption from contaminated soil, however, is 

likely to be inefficient (Poiger and Schlatter 1980; Shu et al. 1988; Weber et al. 1991c). Young children are 

potentially exposed to CDDs because of their tendency, through hand-to-mouth activity, to ingest soils 

(pica) that may be contaminated with CDDs (see Section 5.7 for further details) (Fries and Paustenbach 

1990; Kimbrough et al. 1984; Paustenbach et al. 1992; Pohl et al. 1995). La Goy (1987) estimated the 

following average soil ingestion rates for children: age 0–1 years old, 50 mg/day (maximum 250 mg/day); 

1–6 years old, 100 mg/day (maximum 500 mg/day); 6–11 years old, 50 mg/day (maximum 250 mg/day); 

and over 11 years old, 50 mg/day (maximum 100 mg/day). If children ingest between 50 and 100 mg of 

soil per day (LaGoy 1987) and the soil they ingest contains 1 pg/g (1 ppt) of CDDs, a child may be exposed 

to 0.05-0.1 pg of CDDs per day by this pathway alone (see Section 5.7). 

Children in high risk populations include children of recreational or subsistence fishers, children of 

subsistence hunters particularly those that consume tissues of marine mammals, and children of subsistence 

farmers that consume meat, milk and/or dairy products from their own farm raised animals (see Section 5.7 

for further details). For example, Native American and other subsistence fishing communities may be at 

greater health risks from CDDs in fish and children in these population often consume larger amounts of 

fish than adult members of the general population (CRITFC 1994; Mott 1995). Children of recreational and 

subsistence fishers who routinely consume locally caught fish from CDD-contaminated waterbodies can be 

exposed to higher CDD concentrations than children who consume similar or larger amounts of 

commercially marketed fish from a variety of sources (Ebert et al. 1996; EPA 1995c; Mott 1995). The 

exposure to CDDs will also be highest among children who regularly eat fish as compared to those who 

only occasionally or never eat fish. Several recent studies have documented the higher fish consumption 

rates among subsistence fishers some of which are Native American populations (CRITFC 1994; Nobmann 

et al. 1992; Wolfe and Walker 1987). A study of fish consumption patterns among the Umatilla, Nez Perce, 

Yakama, and Warm Springs tribes of the Columbia River Basin in Washington and Oregon (CRITFC 1994) 

found that the consumption rate for these Native American children (5 years and younger) from these 

four tribes was 19.6 g/day (a consumption rate over 3 times higher than that for adults in the general 

population (6.5 g/day). 

This increased exposure has been demonstrated by serum CDD levels, which are found to be several times 

higher in people who regularly eat fish as compared to those who occasionally or never eat fish (Anderson 

et al. 1998; Svensson et al. 1991) (see Sections 5.5 and 5.9). In addition, this same situation also applied 
CDDs 485 

5. POTENTIAL FOR HUMAN EXPOSURE 

for consumption of wildlife, specifically marine mammals (Ayotte et al. 1997; Dewailly et al. 1992). 

Similar dietary situations exist for children of subsistence hunters that tend to consume tissues of marine 

mammals and children of subsistence farmers that consume beef, milk and other dairy products from their 

own farm raised animals. In the case of subsistence fishers, subsistence hunters, and subsistence farmers, 

all three populations share one problem, that the source of their fish, meat, and/or milk and other dairy 

products, is typically restricted to a localized area, and if these food sources are contaminated with CDDs, 

adults and children in these populations will be exposed to higher levels of CDDs than members of the 

general population (see Section 5.7 for additional details on these populations at risk). 

In order to reduce exposure from consumption of CDD-contaminated fish and wildlife, consumption 

advisories are issued by states recommending that individuals restrict their consumption of specific fish, 

shellfish, and wildlife species from certain waterbodies where CDD concentrations in tissues of these 

species exceed the human health level of concern (EPA 1995c) (see Section 5.7 for additional information). 

Recreational and subsistence fishers typically consume larger quantities of fish and shellfish than the 

general population and frequently fish the same waterbodies routinely. Because of this, children living in 

these populations are at greater risk of exposure to CDDs and other chemical contaminants if the waters 

they fish are contaminated. Currently, 66 advisories have been issued by 21 states restricting the 

consumption of CDD-contaminated fish and shellfish (EPA 1998b) and one state Arkansas also has issued a 

consumption advisory for wood ducks, a species of migratory waterfowl. Three states (New Jersey, New 

York and Maine) also have statewide advisories for CDDs in their marine waters (EPA 1998a). The 

number of waterbodies under advisory for CDD in each state is shown in Figure 5-8. 

5.7 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES 

In addition to individuals who are occupationally exposed to CDDs (Section 5.5), there are several groups 

within the general population with potentially high exposures (higher-than-background-normal levels) to 

CDDs. Historically, populations that have been exposed to higher-than-normal background levels of CDDs 

in the air, water, soil, and/or food have included those who were exposed to 2,3,7,8-TCDD as a result of 

industrial accidents (e.g., Nitro, West Virginia; and Seveso, Italy) and those exposed through environmental 

contamination (e.g., Times Beach, MO; Binghamton, NY; Love Canal, NY; Newark, NJ; and Vietnam) 

(Kahn et al. 1988; Schecter 1985; Schecter and Tiernan 1985; Schecter et al. 1987a, 1989a; Umbriet et al. 

1986a, 1986b; Zook and Rappe 1994). 

 

CDDs 487 

5. POTENTIAL FOR HUMAN EXPOSURE 

Currently, individuals living in proximity to sites where CDDs are produced as chemical by-products or 

sites where CDD-contaminated chemicals are disposed, individuals living near municipal and industrial 

incinerators, and individuals living near one of the 110 NPL hazardous waste sites where CDDs have been 

detected in some environmental media (HazDat 1998) are at risk of receiving potentially higher-thannormal background levels of exposure. Other populations at risk of exposure primarily through dietary 

habits, include recreational and subsistence fishers who typically consume larger amounts of locally caught 

fish and shellfish than the general population, subsistence hunters, particularly those in the high latitudes, 

who typically consume large amounts of locally caught game including marine mammals; and subsistence 

farmers and their families living in areas contaminated with CDDs who exclusively consume their own 

farm-raised beef and dairy products. 

Individuals exposed through industrial accidents or environmental contamination. Very 

extensive residential contamination by 2,3,7,8-TCDD occurred in Seveso, Italy, when a 2,4,5-TCP reactor 

exploded in 1976 (Mocarelli et al. 1991). The contaminated area was divided into three zones based on the 

concentration of 2,3,7,8-TCDD in the soil. Families in zone A, the most heavily contaminated area based 

on soil 2,3,7,8-TCDD levels, were evacuated within 20 days of the explosion and measures were taken to 

minimize exposure of residents in nearby zones. A recent analysis of 19 blood samples from residents of 

zone A, which were collected and stored shortly after the accident, showed serum lipid levels of 

2,3,7,8-TCDD that ranged from 828 to 56,000 ppt. These serum lipid levels are among the highest ever 

reported for humans (Mocarelli et al. 1991). 

In a study conducted in Missouri, 2,3,7,8-TCDD was measured in the adipose tissue of 39 volunteers with a 

history of residential, recreational, or occupational exposure (14 years post-exposure) and in 57 persons in a 

control group (Patterson et al. 1986a). Based on questionnaire responses, the eligible exposed group for 

this study consisted of people who were exposed either to areas with 2,3,7,8-TCDD concentrations in soil 

between 20 and 100 ppb for 2 or more years or to 2,3,7,8-TCDD concentrations >100 ppb for at least 

6 months. Persons who met these criteria were classified as having one of three types of exposure: 

residential (either living in close proximity to areas with 2,3,7,8-TCDD-contaminated soil or having 

evidence of contamination inside the home), recreational (riding or caring for horses in 2,3,7,8-TCDDcontaminated stable arenas at least one time per week), or occupational (working either in a 

hexachlorophene production facility or at truck terminals where the grounds had been sprayed with 

2,3,7,8-TCDD-contaminated waste oil). All study participants had detectable levels of 2,3,7,8-TCDD in 

their adipose tissue, but the group with known previous exposures had significantly higher levels than 
CDDs 488 

5. POTENTIAL FOR HUMAN EXPOSURE 

controls. Nineteen (49%) of the 39 exposed persons had levels higher than the highest 2,3,7,8-TCDD 

concentration (20.2 ppt) detected in the 57 controls. Six (15%) of the 39 exposed persons had 

2,3,7,8-TCDD concentrations >100 ppt. Five of the 6 values >100 ppt were from persons exposed to 

2,3,7,8-TCDD during the production of hexachlorophene. The other high value (577 ppt) was found in a 

man exposed to 2,3,7,8-TCDD while horseback riding in a contaminated arena. 2,3,7,8-TCDD 

concentrations measured in the occupational group (average 136.2 ppt; range 3.5–750 ppt) were, in general, 

higher than those in the residential group (average 21.1 ppt; range 2.8–59.1 ppt), the recreational group 

(average 90.8 ppt; range 5.0–577 ppt), and the control group (average 7.4 ppt; range 1.4–20.2 ppt) 

(Patterson et al. 1986a). 

2,3,7,8-TCDD has been detected at concentrations ranging from 20 to 173 ppt in adipose tissue from 

3 Vietnam veterans reported to have been heavily exposed to Agent Orange (Gross et al. 1984). Except for 

these few men, however, 2,3,7,8-TCDD concentrations in American Vietnam and non-Vietnam veterans 

were nearly identical with mean serum levels of approximately 4 ppt (CDC 1988). Concentrations of 

2,3,7,8-TCDD in the controls (those who never served in Vietnam) ranged from not detected (4 ppt) to 

20 ppt. The veterans had served in Vietnam in 1967 and 1968 in areas where Agent Orange had been 

heavily used (CDC 1988). In another study, 2,3,7,8-TCDD was detected in adipose tissue of 14 Vietnam 

veterans and 3 control patients at levels ranging from not detected (2–13 ppt) to 15 ppt. No significant 

differences in the tissue levels of Vietnam veterans and the controls were found in this study (Weerasinghe 

et al. 1986). Air Force personnel associated with Operation Ranch Hand (spraying of Agent Orange) in 

Vietnam from 1962 to 1971 had serum CDD levels up to 10 ppt (521 persons). A correlation was found 

between CDD concentrations and increased body fat (USAF 1991). The median half-life of 2,3,7,8-TCDD 

in 36 veterans was estimated to be 7.1 years (Pirke et al. 1989). In 1987, many of the exposed Air Force 

personnel had serum CDD concentrations >50 ppt and several had concentrations exceeding 300 ppt (CDC 

1987). Wolfe et al. (1994) reported a half-life value of 11.3 years for Air Force personnel involved in 

Operation Ranch Hand. Using individuals from two rice oil poisoning episodes (Yusho [Japan] and YuCheng [Taiwan]), Ryan et al. (1993a) have shown that the elimination of related CDFs is not constant, but 

variable, with faster clearance at higher doses followed by a slowing down in the rate of loss as body 

burden decreases. By analogy, the same may be true for CDDs. It is also likely that individual congeners 

or those with the same degree of chlorination are excreted from the body at rates that differ from those 

estimated for 2,3,7,8-TCDD. Because the rate of clearance is not constant, uncertainty in determining the 

half-life measurement may result, especially for estimates of the changing body burden of total CDDs 

measured as TEQs. 
 

CDDs 489 

5. POTENTIAL FOR HUMAN EXPOSURE 

Individuals living in proximity to production or disposal sites. Individuals living in the vicinity 

of former or current production sites where CDDs are released as by-products, (such as incinerators, coalfired electric generating facilities, other waste disposal facilities, and hazardous waste sites) may be exposed 

to CDDs from several exposure pathways. CDDs have been detected in soil at 94 of the 126 sites where 

they have been detected in some environmental media (HazDat 1998). 

Children and adults may receive higher CDD exposures from dermal contact if they play or work with 

CDD-contaminated soils. Several studies have examined the bioavailability of 2,3,4,7-TCDD for uptake by 

dermal exposure. In a study using in vitro human skin tissue, 2,3,7,8-TCDD did not readily penetrate into 

the human skin and the vehicle of exposure played an important role in the dermal penetration (Weber et al. 

1991c). These authors reported that when the exposure vehicle was acetone, the maximum 2,3,7,8-TCDD 

penetration into the in vitro human stratum corneum (30–45% of the dose) was reached within 100 minutes, 

with a tendency to decrease after 1,000 minutes. Using mineral oil as the exposure vehicle, absorption of 

2,3,7,8-TCDD leveled off at 10% of the dose, and it took more than 300 minutes to reach the maximum. 

The data suggest that the rates of absorption of 2,3,7,8-TCDD into in vitro human skin are moderate (worstcase scenario) to low when the 2,3,7,8-TCDD is applied in acetone; when applied in mineral oil, the 

adsorption rate was further reduced. Shu et al. (1988) reported that in rats, dermal absorption of 

2,3,7,8-TCDD in a soil vehicle was only 1% of the administered dose. Similarly, Poiger and Schlatter 

(1980) reported that in rats, dermal absorption of 2,3,7,8-TCDD was almost eliminated when soil or 

activated carbon was used as vehicles. These data support the original Kimbrough et al. (1984) risk 

assessment of a contaminated site in which the authors estimated that the additional lifetime uptake of 

TCDD from soil above background uptake will consist of 95% from soil ingestion, 3% from dermal 

exposure to soil (assuming 1% dermal absorption), and 2% from inhalation of soil particles. 

Children and adults also may receive potentially higher oral exposures from ingestion of CDDcontaminated soils from their unwashed hands while playing or working in CDD-contaminated areas (Fries 

and Paustenbach 1990; Kimbrough et al. 1984; Paustenbach et al. 1992; Pohl et al. 1995). Bioavailability is 

an integral factor in the estimation of the internal dose (or dose at the target tissue) of the chemical. Like 

dermal absorption, gastrointestinal absorption of 2,3,7,8-TCDD and related compounds is variable, 

incomplete, and congener- and vehicle-specific (see Section 2.3.1). More lipid soluble congeners, such as 

2,3,7,8-TCDF, are almost completely absorbed, while the extremely insoluble OCDD is poorly absorbed. 

However, laboratory data suggest that there are no major interspecific differences in the gastrointestinal 

absorption of CDDs and CDFs. Results from animal studies indicate that bioavailability of 2,3,7,8-TCDD 
 

CDDs 490 

5. POTENTIAL FOR HUMAN EXPOSURE 

from soil varies between sites because CDDs bind tightly to soil, and increasingly so with the passage of 

time and clay content of the soil (Gough 1991; Umbreit et al. 1986a;1986b). Therefore, 2,3,7,8-TCDD soil 

concentrations alone may not be indicative of the potential for human health hazard from contaminated 

soils, and site-specific evaluation may be essential. In their risk assessments, Kimbrough et al. (1984) 

assumed 30% bioavailability from ingestion of soil, but they point out that animal studies with 

contaminated Missouri soil indicated absorption as high as 30 to 50% (McConnell et al. 1984). Pohl et al. 

(1995) assumed 40% bioavailability of 2,3,7,8-TCDD from soil. In contrast, Paustenbach et al. (1986) 

assumed only 10–30% bioavailability. However, unless toxicokinetic studies that use soil samples from the 

specific site are available, it is difficult to speculate on how much 2,3,7,8-TCDD as well as other CDDs will 

be bioavailable. ATSDR’s policy on CDDs contaminated soils is in Appendix B of this profile. 

Individuals may also receive higher doses from routine consumption of CDD-contaminated fish from local 

waters receiving runoff or leachates from the waste site (Paustenbach et al. 1992). CDDs have been 

detected in fish collected at 12 of the 126 NPL sites where they have been detected in some environmental 

media (HazDat 1998). 

Lastly, individuals living near incinerator or hazardous waste sites may inhale vapors or particulates 

contaminated with CDDs from ambient outdoor air. This, however, would be a relatively minor exposure 

pathway as only about 50% of all particles are of inhalable size (<10μm) (Fries and Paustenbach 1990; 

Paustenbach et al. 1992). These authors concluded that since uptake of 2,3,7,8-TCDD from foods will be 

approximately 500–1,000-fold greater than that due to inhalation, that inhalation exposure was a relatively 

insignificant exposure pathway even for individuals living in proximity to an incinerator. 

Recreational and subsistence fishers. In general, concentrations of CDDs in sport fish and 

shellfish from CDD-contaminated waters can be at least an order of magnitude higher than in commercial 

fish and shellfish purchased in a supermarket (see Section 5.4.4). Since CDDs have been found in fish from 

contaminated lakes and streams (Crunkilton et al. 1987; De Vault et al. 1989; EPA 1987n, 1992; Kuehl et 

al. 1989, 1994; Niimi and Oliver 1989a, 1989b; Reed et al. 1990) and fish and shellfish from estuarine 

waters (Bopp et al. 1991; Brown et al. 1994; Cai et al. 1994; Hauge et al. 1994; Rappe et al. 1991), 

populations that consume large quantities of fish or shellfish from these contaminated waters are also likely 

to have higher exposures to CDDs, although the method of preparation (edible fillets versus skin-on fillets 

or whole fish) and cooked versus raw consumption may substantially reduce the amount of CDDs ingested 

(Paustenbach et al. 1992; Schecter et al. 1996c). 
CDDs 491 

5. POTENTIAL FOR HUMAN EXPOSURE 

Recreational (sport) and subsistence fishers (including some native American peoples) who consume locally 

caught fish from CDD-contaminated waterbodies can be exposed to higher CDD concentrations than 

individuals who consume similar or larger amounts of commercially marketed fish from a variety of sources 

(Ebert et al. 1996; EPA 1995c). The exposure to CDDs will also be highest among people who regularly 

eat large amounts of fish as compared to those who only occasionally or never eat fish. This increased 

exposure has been demonstrated by serum CDD levels to be several times higher in people who regularly 

eat fish as compared to those who occasionally or never eat fish (Anderson et al. 1998; Svensson et al. 

1991). 

Several studies suggests there is a correlation between consumption of CDD-contaminated fish and/or 

marine mammal tissues and elevated levels of CDDs in blood (Anderson 1998; Ayotte et al. 1998; 

Svensson et al. 1991). Svensson et al. (1991) reported elevated blood levels of CDDs in Swedish fish 

consumers living near the Baltic Sea. Three distinct groups of consumers were studied: individuals who did 

not consume any fish, moderate fish consumers (consumption rate of 220–500 g of fish/week 

[31–71 g/day]), and high fish consumers (consumption rate of 700–1,750 g of fish/week [100–250 g/day]). 

The highest fish consuming group was composed of fishermen or workers in the fishing industry who 

consumed primarily salmon (30–90 pg TEQ/g) and herring (8–18 pg TEQ/g) from the Baltic Sea. The TEQ 

blood level was found to average 63.5 pg TEQ/g lipid for the high consumption group, 25.8 pg TEQ/g lipid 

for the moderate consumption group, and 17.5 pg TEQ/g lipid for the nonfish-eating group. With respect to 

2,3,7,8-TCDD blood levels, the mean blood level detected was 1.8 pg/g (lipid basis) for individuals that 

consumed no fish, 2.5 pg/g for the moderate consumers, and 8.0 pg/g for the high consumer group. It 

should be noted that even in those individuals who consumed no fish, detectable levels of CDDs were 

present in their blood. This indicates that other food sources (e.g., meat, milk, and other dairy products) are 

likely to be important contributors to the total body burden of CDDs (Rappe 1992). 

Recently, Anderson et al. (1998) completed a preliminary study of the levels of 8 CDDs, 10 CDFs, 36 

PCBs, and 11other persistent organochlorine pesticides in human serum samples from Great Lakes sport 

fish consumers. Overall, the 31 fishers on average consumed 49 Great Lakes sport fish meals per year, for a 

mean of 33 years. This is in contrast to the general population in the Great Lakes basin that typically 

consumes 6 meals of Great Lakes sport fish per year. A summary of the distribution of CDDs is provided 

in Table 5-14. CDD congeners detected most often were 1,2,3,4,6,7,8-HpCDD (31 detects), OCDD (31 

detects), 1,2,3,6,7,8-HxCDD (30 detects), 2,3,7,8-TCDD (25 detects) and 1,2,3,7,8-PeCDD (20 detects). 

The overall mean concentration for 2,3,7,8-TCDD was 6.6 ppt. Total CDD concentrations were highest 

CDDs 493 

5. POTENTIAL FOR HUMAN EXPOSURE 

for Lake Huron fish consumers (1,259 ppt), intermediate for Lake Michigan consumers (1,087 ppt) and 

lowest for Lake Erie consumers (844 ppt). The comparison group serving as a control included individuals 

residing in Arkansas and had a total CDD serum concentration of 1,198 ppt. With respect to the TEQ 

values for CDDs, the pattern among Great Lakes fish consumers was similar to that for total CDD 

consumers with TEQs for Lake Huron fish consumers of 36 ppt, Lake Michigan consumers of 25.9 ppt, and 

Lake Erie consumers of 20.7 ppt. The TEQ values for the three Great Lakes sport fish consumer groups 

were statistically different (p<0.03). Although the comparison population had CDD concentrations within 

the range of the Lake Michigan and Lake Huron fish consumers, the TEQ value for CDDs for this 

population was the lowest of the four groups at 15.5 ppt. The authors concluded that Great Lakes anglers 

who are life-long frequent consumers of sport fish represent a subpopulation with the potential for 

significant exposure to CDDs as well as CDFs and PCBs. The levels of CDDs, CDFs, and PCBs found in 

sportfish and human tissue residues were above those in the general population. 

Ayotte et al. (1998) measured concentrations of CDDs/CDFs and PCBs in plasma of adult Inuit living in 

Arctic Quebec, Canada. The Inuit consume large amounts of fish and marine mammal tissue. The mean 

concentration of 2,3,7,8-TCDD was 8.4 ppt (range 2.5 to 36.0 ppt) in the Inuit population and < 2 ppt 

(range <2) for the control population in Southern Quebec. The TEQ values for all CDDs/CDFs was 39.6 ppt 

(range 17.1 to 81.8 ppt) in the Inuit population and 14.6 ppt (range 11.5 to 18.9 ppt) for the control 

population. When PCBs and CDDs/CDFs are considered together, the mean TEQ values for all dioxin-like 

compounds was 184.2 ppt in the Inuit population (range 55.8 to 446.7 ppt) and 26.1 ppt (range 20.1 to 31.7 

ppt) for the control population. 

Several recent studies have documented the higher fish consumption rates among subsistence fishers some 

of which are Native American populations. In a study of Alaskan subsistence economies, Wolf and Walker 

(1987) reported daily fish consumption rates ranging from 6 to 1,536 g/day, with an average consumption 

rate of 304 g/day. This average consumption rate for subsistence fishers is more than 46 times higher than 

the mean fish consumption rate of 6.5 g/day estimated for the general population (EPA 1995c). In a study 

of 11 Alaskan communities, Nobmann et al. (1992) reported an average daily fish consumption rate of 

109 g/day. This is more than 16.8 times higher than the mean fish consumption rate of 6.5 g/day estimated 

for the general population (EPA 1995c). A recent study of fish consumption patterns among the Umatilla, 

Nez Perce, Yakama, and Warm Springs tribes of the Columbia River Basin in Washington and Oregon 

(CRITFC 1994) found that adults in these 4 tribes consume an average of 58.7 g/day of fish and the 

95th percentile of fishers consume 170 g/day of fish. This mean consumption rate is more than nine times 
 

CDDs 494 

5. POTENTIAL FOR HUMAN EXPOSURE 

higher than the mean fish consumption rate estimated for the general population (EPA 1995c). Fitzgerald et 

al. (1995) conducted a study to establish patterns of fish consumption of nursing Mohawk Indians residing 

in the vicinity of three hazardous waste sites located near Akwesasne, NY, compared to a control 

population of Caucasian women in New York state. The dietary data showed that there was a significant 

past prevalence of local fish consumption among Mohawk mothers (23.5 meals per year) more than a year 

before pregnancy as compared to the controls (14.1 meals per year). 

In order to reduce CDD exposure from consumption of CDD-contaminated fish and shellfish, consumption 

advisories are issued by states recommending that individuals restrict their consumption of specific fish and 

shellfish species from certain waterbodies where CDD concentrations in fish and shellfish tissues exceed 

the human health level of concern. This level of concern is set by individual state agencies, but many states 

use the FDA tolerance levels of >25 ppt, but <50 ppt to advise consumers to restrict consumption and levels 

>50 ppt to issue advisories recommending no consumption of contaminated fish and shellfish (see 

Table 7-1). These values are used despite the fact that they were designed to protect consumers from the 

health risks associated with consumption of fish and shellfish that are shipped in interstate commerce and 

are purchased in commercial markets. In 1995, the EPA Office of Water issued guidance to states on 

sampling and analysis procedures to use in assessing the health risks from locally caught fish and shellfish. 

The risk assessment method proposed by EPA was specifically designed to assist states in developing fish 

consumption advisories for recreational and subsistence fishers (EPA 1995c). These two groups within the 

general population consume larger quantities of fish and shellfish than the general population and 

frequently fish the same waterbodies routinely. Because of this, these populations are at greater risk of 

exposure to CDDs and other chemical contaminants if the waters they fish are contaminated. The EPA 

advises states to use a screening value of 7x10-7 ppm (0.7 ppt) of total TEQ value (wet weight) in fillets for 

the general population as a criteria to evaluate their fishable waterbodies (EPA 1995c). Currently, 

66 advisories have been issued by 21 states restricting the consumption of CDD-contaminated fish and 

shellfish (EPA 1998b). Three states (New Jersey, New York, and Maine) also have statewide fish 

advisories in effect for their marine coastal waters. In addition, the State of Arkansas also has issued a 

wildlife advisory for wood ducks, a species of waterfowl in Bayou Meto, a site contaminated by a point 

source discharge (White and Hoffman 1995). The number of waterbodies under advisory for CDD in each 

state is shown in Figure 5-8. 

Subsistence hunters. Native American populations such as the Inuit of Alaska and other subsistence 

hunters (particularly those living in high latitude areas of the United States) may have higher exposure to 
CDDs 495 

5. POTENTIAL FOR HUMAN EXPOSURE 

CDDs in marine mammals, but not terrestrial mammals (i.e., caribou). CDDs have been detected in seal 

and polar bear fat and liver tissues (Norstrom et al. 1990). Recently, Ayotte et al. (1997) reported that in 

20 pooled samples of Inuit blood, the TEQ concentration supplied by CDDs/CDFs (combined) was 39.6 ppt 

as compared to 14.6 ppt in 3 pooled samples of blood from individuals in the general population in 

Southern Quebec. The mean 2,3,7,8-TCDD concentrations were 8.4 ppt for Inuit and <2 ppt for members of 

the general population. When planar PCBs, mono- and di- ortho PCBs were added, the mean TEQs were 

184.2 ppt and 26.1 ppt, respectively, for Inuit and members of the general population in Southern Quebec. 

In a related study, Dewailly et al. (1992) found that breast milk of these Inuit was more contaminated by 

CDDs, CDFs, and PCBs than milk of women in Southern Quebec. For CDDs and CDFs, differences were 

less impressive than for PCBs. However, mean OCDD concentrations were 292 ppt versus 132 ppt and 

2,3,7,8-TCDD concentrations were 6.2 ppt versus 2.3 ppt for Inuit women and women in Southern Quebec, 

respectively. Concentrations of OCDD and 2,3,7,8-TCDD were 2 to 3 times greater among the native Inuit 

population that consumed large amounts of fish and marine mammal tissue. No data were located 

specifically for CDD concentrations in the adipose tissue, blood, or breast milk of native American 

populations in the United States. However, by analogy to CDDs in Canadian Inuit populations, it is 

anticipated that CDD concentrations in these tissues are likely to be higher among individuals who 

routinely consume large quantities of wild game species, especially marine mammal than among members 

of the general population. 

In a study of subsistence economies in the State of Alaska, Wolfe and Walker (1987) reported that total 

annual per capita consumption of wild game species (including land mammals, marine mammals, and fish) 

ranged from 10 to 1,498 pounds (median harvest of 252 pounds) as compared to 222 pounds of meat, fish, 

and poultry consumed each year by individuals in the western United States. In the 1980s, the 98 Alaskan 

subsistence communities surveyed harvested wild game at levels from one-half to 4 times the U.S mean. 

The average daily per capita consumption was 0.67 pounds for fish and 0.23 pounds for land mammals 

based on all 98 communities, and 0.2 pounds for marine mammals based on the 41 coastal communities 

surveyed. Land mammals consumed in these communities included moose, caribou, deer, black bear, 

snowshoe and tundra hare, beaver, and porcupine; while marine mammals included seal, walrus and whales. 

Subsistence hunters and their families are a population at potentially higher risk of CDD exposure when the 

wild game species particularly marine mammals they consume are contaminated with CDDs. It should be 

noted that concentrations of CDDs/CDFs were recently determined in caribou tissue samples from 7 herds 

across the Canadian Arctic (Hebert et al. 1996). In contrast to levels of 2,3,7,8-TCDD found in marine 

mammals which ranged from 2 to 37 ng/kg (ppt wet weight) (Norstrom et al. 1990), 
 

CDDs 496 

5. POTENTIAL FOR HUMAN EXPOSURE 

concentrations of 2,3,7,8-TCDD in caribou were extremely low, sub ng/kg (lipid basis). 2,3,7,8-TCDD was 

detected in adipose tissue samples of two herds in the eastern Canadian Arctic at levels of 0.73 and 

0.14 ng/kg, but was not detected in tissue samples from other herds at detections limits as low as 0.03 ng/kg 

(lipid basis). CDFs were detected at sub-ng/kg levels in all cases. TEQs were dominated by non-ortho 

substituted PCBs in all cases, and ranged from 0.33 ng/kg to 3.3 ng/kg in adipose tissue. The authors 

concluded that caribou tissues are therefore less contaminated than tissues from marine mammals. 

Subsistence farmers that consume their own farm-reared meat and dairy products. 

Subsistence farmers and their families living on farms where CDD concentrations may be high who 

exclusively eat meat and dairy products produced on their own farms may be exposed to higher levels of 

CDD in these foods than the general population. Grazing cattle in farming areas downwind of municipal or 

industrial incinerators where CDDs may be deposited as particulates on soil or forage crops or grazing 

cattle in areas with CDD-contaminated soil or soil amended with municipal sewage sludges or paper mill 

sludges contaminated with CDDs may result in higher CDD tissue residues in the animals (EPA 1991b; 

Fries and Paustenbach 1990). In an evaluation of potential transmission of 2,3,7,8-TCDD from incinerator 

emissions to humans via foods, Fries and Paustenbach (1990) determined that the amount of 2,3,7,8-TCDD 

accumulated in soil from airborne emissions was less important that the amount deposited in forage. These 

authors further concluded that the airborne emissions of CDDs/CDFs from modern waste incinerators that 

have appropriate air pollution devices should not pose a significant health hazard regardless of the 

incinerator location. The authors, however, acknowledged that it would be desirable to measure 

2,3,7,8-TCDD in soil and crops around existing facilities to better evaluate their assessment results, 

although they felt it was likely that concentrations would be too low to reliably quantify. More recently, 

Fürst et al. (1993) reported that soil levels up to 30 pg TEQ/g dry matter did not result in elevated CDD 

concentrations in cow’s milk. However, these authors did show increasing concentrations of CDDs in grass 

resulted in slightly higher CDD concentrations in cow’s milk. These authors, like Fries and Paustenbach 

(1990), believe that the pathway of air to grass to cow is more important than the pathway of soil to grass to 

cow. 

In a European survey of cow’s milk samples collected on dairy farms, elevated CDD/CDF milk 

concentrations were used to pinpoint the existence of certain “hotspots” of CDD contamination. These 

“hotspots” were generally found near CDD/CDF emitting sources, such as cable waste incinerators or 

metal refining industries (Beck et al. 1990; Liem et al. 1991; Rappe et al. 1987). Riss et al. (1990) 

analyzed blood from one farmer in a CDD/CDF-contaminated location in proximity to a metal reclamation 
CDDs 497 

5. POTENTIAL FOR HUMAN EXPOSURE 

plant who consumed milk produced from his own farm, and found 2,3,7,8-TCDD blood levels (55 pg/g on 

a lipid basis) above expected background levels. 

Exposure to 2,3,7,8-TCDD from land application of municipal sewage sludge or paper mill sludge also can 

occur through the dietary pathway if people consume food grown or animals grazed on sludge-amended 

lands (EPA 1991b). Wild et al. (1993), using a human exposure assessment model, predicted that if all the 

produce consumed by a human is derived from agricultural land to which sewage sludge is applied at a rate 

of 10 tons/hectare and 0.5 ng/kg dry weight concentration, this would increase exposure to 2,3,7,8-TCDD 

by 0.0332 ng/day or 39% over background conditions. This scenario assumes that the poultry, eggs, and 

fish CDD concentrations are unaffected by the sludge application. Most recently, McLachlan et al. (1994) 

reported that the prolonged use of sewage sludge as a soil amendment on English farms under some 

conditions can lead to an increase in the concentrations of CDDs/CDFs in both the soil and in cow’s milk. 

Subsistence farmers and their families are a population at potentially higher risk of CDD exposure because 

meat and dairy products are substantial sources of CDDs in the U.S. diet (Schecter et al. 1994e, 1996a). 

5.8 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 CDDs is available. Where adequate information is not available, 

ATSDR, in conjunction with the 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 

CDDs. 

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. 
 

 

 

CDDs 498 

5. POTENTIAL FOR HUMAN EXPOSURE 

5.8.1 Identification of Data Needs 

Physical and Chemical Properties. The physical and chemical properties of 2,3,7,8-TCDD are 

sufficiently characterized to predict the environmental fate of 2,3,7,8-TCDD (IARC 1977; Sax and Lewis 

1987; Schroy et al. 1985; Shiu et al. 1988). Of all the CDDs, 2,3,7,8-TCDD has been the compound most 

studied. Not all isomers within each homologous class have been equally well studied for many of the 

physical and chemical properties. Information on physical and chemical properties of certain congeners 

(particularly 1,2,3,7,8,-PeCDD and 1,2,3,6,7,8-HxCDDs) would be helpful in better understanding the 

different fate and transport pathways of the homologous groups. 

Production, Import/Export, Use, Release, and Disposal. CDDs are not manufactured 

commercially in the United States except on a laboratory scale for use in chemical and toxicological 

research (Cambridge Isotope Laboratories 1995). They are produced as undesired by-products during the 

manufacture of chlorophenols (e.g., PCP and 2,4,5-trichlorophenol) and during combustion processes 

(IARC 1977; NTP 1989; Podoll et al. 1986). CDDs are ubiquitous in the environment and have been 

found at low levels (ppt or lower) in air, water, soil, sediment, and foods. Current disposal methods are 

efficient and are subject to EPA and state regulations. 

According to the Emergency Planning and Community Right-To-Know Act of 1986, 42 U.S.C. Section 

11023, industries are required to submit chemical release and off-site transfer information to the EPA. The 

Toxics Release Inventory (TRI), which contains this information for 1994, became available in May of 

1996. This database will be updated yearly and should provide a list of industrial production facilities and 

emissions. However, there are no TRI data for CDDs since CDD releases are not required to be reported 

(EPA 1995g). 

Environmental Fate. CDDs are subject to atmospheric transport and both wet and dry deposition 

(Kieatiwong et al. 1990). They are partitioned to air, water, sediment, and soil, and they accumulate in 

both aquatic and terrestrial biota. CDDs can volatilize to the atmosphere from water and soil surfaces. 

They adsorb strongly to soils and are not likely to leach into groundwater (Eduljee 1987b). In the aquatic 

environment, CDDs partition to sediment or suspended particulates. TCDD, HpCDD, and OCDD are 

subject to photolysis in air, water, and soil (Plimmer et al. 1973). 2,3,7,8-TCDD is biodegraded very 

slowly in soil and thus is likely to persist in the soil. A better understanding of environmental behavior of 

CDDs is needed with respect to the importance of vapor-phase versus particulate transport, the 
 

 

 

CDDs 499 

5. POTENTIAL FOR HUMAN EXPOSURE 

environmental behavior of different congeners, and the significance of processes that reintroduce CDDs 

into the atmosphere after deposition. Information regarding the degradation of other congeners, 

specifically OCDD, and their degradation products in water, sediment, and soil would be useful in 

evaluating the various pathways of human exposure. 

Bioavailability from Environmental Media. Toxicokinetic data in humans regarding adsorption of 

CDDs following oral and dermal exposure are very limited (Poiger and Schlatter 1986). CDDs can be 

absorbed following oral exposure in both humans and animals (Birnbaum and Couture 1988; Fries and 

Marrow 1975; Koshakji et al. 1984; Norback et al. 1975; Olson et al. 1980b; Piper et al. 1973; Poiger and 

Schlatter 1980). The more highly chlorinated CDD congeners are absorbed to a lesser extent than 

2,3,7,8-TCDD (Koshakji et al. 1984). Also, limited information is available on the bioavailability from fly 

ash (Van den Berg et al. 1983, 1985). 2,3,7,8-TCDD can be adsorbed following dermal contact (Banks 

and Birnbaum 1991; Poiger and Schlatter 1980; Shu et al. 1988); however, dermal absorption of 

2,3,7,8-TCDD from soil is very low (Shu et al. 1988). More information is needed regarding oral and 

dermal exposure to determine the bioavailability of CDDs from food, water, and soil. Additional 

information is needed to examine the discrepancy noted in the mass balance from CDDs ingested from 

foods and eliminated in feces. For inhalation exposure, information on the bioavailability from fly ash and 

sediments would be useful. Information is also needed on the selective uptake of the 2,3,7,8-substituted 

CDD congeners. 

Food Chain Bioaccumulation. CDDs are bioconcentrated in aquatic organisms, plants, and 

terrestrial animals. Shellfish (including crustaceans and bivalve mollusks) appear to accumulate CDDs 

nonselectively to relatively high concentrations in their tissues (Bopp et al. 1991; Brown et al. 1994; Cai et 

al. 1994; Conacher et al. 1993; Hauge et al. 1994; Rappe et al. 1991). In contrast, finfish appear to 

selectively accumulate primarily 2,3,7,8-TCDD and other 2,3,7,8-substituted isomers in their tissues 

(Rappe et al. 1991). Information from a larger number of species on the retention of 2,3,7,8-substituted 

CDD congeners and general information on retention and distribution of other CDDs would be useful in 

better understanding both aquatic and terrestrial food chains. 

Exposure Levels in Environmental Media. CDDs have been detected in air, water, soil, sediment, 

plant material, and foods. Environmental monitoring studies show that the higher chlorinated CDDs are 

usually the ones most commonly found in environmental samples (Christmann et al. 1989b; Clement et al. 

1985, 1989; Pereira et al. 1985; Reed et al. 1990; Tashiro et al. 1989a; Tiernan et al. 1989b). Current 
 

CDDs 500 

5. POTENTIAL FOR HUMAN EXPOSURE 

monitoring studies are needed to determine CDD levels in media surrounding hazardous waste sites. 

Using a model, the total average daily intake of 2,3,7,8-TCDD (by air, water, and food) for the general 

population was estimated to be 0.05 ng/day (range 0.008–0.3 ng/day) (Travis and Hattemer-Frey 1987). 

Schecter et al. (1994d, 1994e, 1996a) and Schecter and Li (1997) have provided current information on 

CDD exposures from food. Food consumption accounts for over 90% of background human exposure to 

2,3,7,8-TCDD and other CDDs/CDFs in the general U.S. population (Hattemer-Frey and Travis 1989; 

Schaum et al. 1994). The average daily intake by nursing infants in the United States has been estimated 

to be 83 pg TEQs/kg (Schecter and Gasiewicz 1987a, 1987b). 

Reliable monitoring data for the levels of CDDs in contaminated media at hazardous waste sites are needed 

so that the information obtained on levels of CDDs in the environment can be used in combination with the 

known body burdens of CDDs to assess the potential risk of adverse health effects in populations living in 

the vicinity of hazardous waste sites. 

Exposure Levels in Humans. CDDs/CDFs have been found in blood (Fingerhut et al. 1989; 

Needham et al. 1991; Päpke et al. 1989b, 1992, 1993), adipose tissue (Orban et al. 1994; Patterson et al. 

1986a; Ryan et al. 1986; Schecter et al. 1986b; Stanley 1986; Stanley et al. 1986), and breast milk of both 

the general population and workers exposed through industrial accidents or environmental contamination 

(Fürst et al. 1992; Pluim et al. 1993a; Ryan et al. 1993b; Schecter and Gasiewicz 1987b; Schecter and 

Tiernan 1985a; Schecter et al. 1986a, 1986b; 1989e). Levels of 2,3,7,8-TCDD as well as other CDDs are 

generally higher in occupationally exposed individuals or those individuals exposed through industrial 

accidents or environmental contamination (Kahn et al. 1988; Schecter et al. 1986b; Schecter and Tiernan 

1985; Schecter et al. 1987a; Umbriet et al. 1986a, 1986b). CDDs have also been detected in breast milk 

and blood of Canadian populations of native Inuit that consume large amounts of fish and marine 

mammals (Ayotte et al. 1997; Dewailly et al. 1992). Additional biological monitoring data are needed, 

however, for those U.S. populations surrounding hazardous waste sites or municipal, medical, or industrial 

incinerators, for urban versus rural exposures, and for other potentially exposed populations including 

subsistence fishers and hunters (Liem et al. 1991; Startin et al. 1989; Wuthe et al. 1993). Information on 

tissue levels in the general population worldwide are for the most part lacking (Schecter et al. 1991a). As 

they are identified, exposed populations should be evaluated to characterize exposure levels and health 

effects. This information is necessary for assessing the need to conduct health studies on these 

populations. 
 

 

CDDs 501 

5. POTENTIAL FOR HUMAN EXPOSURE 

Exposures of Children. Children in the general population are exposed to CDDs primarily through 

dietary exposures in utero via placental blood and in newborn infants via breast-feeding. Despite the fact 

that studies on the concentrations of CDDs in human breast milk have been conducted in various other 

countries, there is a need to determine the levels of CDDs in human milk in the United States. Additional 

exposure studies also are needed to determine whether dietary modifications in mothers can reduce total 

CDD exposures in newborns and whether dietary modifications of the infant can also reduce lifetime 

exposure. For children in populations with potentially high exposure to CDDs, the primary exposure 

pathway is through their diet; however, additional exposure to CDDs via consumption of contaminated 

groundwater, contaminated soil, and dermal exposure to contaminated soil may increase their exposure 

levels. Studies of workers in various industrial settings that are exposed to CDDs (i.e., elevated CDD 

levels in adipose or blood serum) should be conducted to determine whether CDDs are routinely brought 

home by these workers on their clothing and shoes to assess in order to determine whether this is an 

important exposure route for children. 

Schecter and Li (1997) have calculated weight-adjusted intakes of CDDs derived from consumption of 

four types of fast foods for 6-year-old children. Additional information on dietary intake of CDDs from 

other types of foods should be conducted for various age groups of children to help identify the magnitude 

and sources of dietary exposure during childhood. Studies to verify these calculations would be helpful in 

assessing health risks to children. 

The primary childhood specific means to decrease exposure to CDDs involves placing the infant on a 

cow’s milk or soy-based baby formula and on maintenance of children on a long-term diet that is lower in 

animal fats (meat, dairy products, and fish) and higher in grains, fruits, and vegetables. It should be noted 

however, that because of the relatively short period of intake and the accepted benefits of breast-feeding 

that maintenance of children on long-term diet low in animal fat would likely be more beneficial in 

decreasing total lifetime CDD body burdens than cessation of breast-feeding. Additional means of 

reducing CDD exposures also should be investigated. 

Exposure Registries. Approximately 250 members were enrolled in the 2,3,7,8-TCDD Subregistry 

of the National Exposure Registry in 1991 (ATSDR 1996). These individuals were chosen because they 

participated in one or more of the Missouri Dioxin Health Studies and were reportedly exposed to CDDs at 

one of the four Times Beach, MO area CDD sites. Data collected for each member of the Dioxin 

Subregistry include demographic information, smoking and occupational histories, and self-reported 
CDDs 502 

5. POTENTIAL FOR HUMAN EXPOSURE 

responses to 25 general health status questions. The data files for the Subregistry are established at the 

time baseline data are collected. A follow-up survey is conducted 1 year after baseline data collection, and 

surveys are, in most cases, conducted at 2-year intervals after that to update the files. For the Dioxin 

Subregistry, all interviews are conducted by means of computer-assisted telephone interviewing. 

Subregistry members will be questioned yearly about their health over the previous year. This activity is 

carried out by ATSDR. The data will become part of public-user data files maintained by ATSDR. 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 the exposure to this compound. 

The Air Force maintains an exposure registry of about 1,200 personnel previously involved in the spraying 

of Agent Orange (USAF 1991). The Air Force Health Study (AFHS) is an epidemiologic investigation of 

the association between occupational exposure to Agent Orange (and its CDD contaminants) and longterm adverse health effects experienced by Air Force personnel who served in Operation Ranch Hand units 

in Vietnam from 1962 to 1971 and sprayed Agent Orange from fixed winged aircraft. A comparison 

group, which was formed from Air Force veterans, is used as the unexposed cohort. Evaluations were 

performed in 1982, 1985, and 1987. In the 1987 examination, 1,670 participants were involved. Health 

outcomes were evaluated with respect to serum CDD levels. Additional evaluations are planned for 1997 

and 2002. 

5.8.2 Ongoing Studies 

The EPA is currently conducting a reassessment of the risk from exposure to CDDs, and other chlorinated 

dioxin-like compounds such as CDFs and PCBs. This reassessment involves a literature reevaluation of 

existing studies and new laboratory studies addressing health and ecological risks from exposure to these 

compounds (LaFleur et al. 1990; Rappe 1992; Schecter et al. 1994d). Currently, the EPA dioxin 

reassessment document is undergoing final review. 

The National Institute of Environment Health Sciences and the Centers for Disease Control and Prevention 

are measuring levels of CDDs, CDFs, PCBs and other chemicals in blood of members of the general U.S. 

population as part of the NHANES program. 

In addition, an international “Dioxin” research conference meets annually to discuss developments 

regarding these environmental contaminants. The proceedings of this international symposium on CDDs 

and related compounds are published annually in extended abstract form and frequently in a proceedings 
CDDs 503 

5. POTENTIAL FOR HUMAN EXPOSURE 

issue of the journal Chemosphere and are an extensive source for papers on a wide variety of 

environmental and health issues related to CDDs and dioxin-like compounds. 

A search of Federal Research in Progress (FEDRIP 1998) identified numerous research studies that are 

currently being conducted that may fill some of the data needs discussed in Section 5.7.1 (see Table 5-15). 


.
 
CDDs 507 

6. ANALYTICAL METHODS
 

The purpose of this chapter is to describe the analytical methods that are available for detecting, and/or 

measuring, and/or monitoring CDDs, its metabolites, and other biomarkers of exposure and effect to 

CDDs. 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. 

6.1 BIOLOGICAL SAMPLES 

The primary method of determining CDDs in biological samples is gas chromatography (GC) with mass 

spectrometry (MS). Sample preparation is critical, and extensive extraction and sample clean-up are 

required to separate the CDD homologues/congeners from fatty material and other organic contaminants. 

Extreme care must also be used to ensure that all reagents and equipment used in analysis are free of CDD 

contamination. Losses of CDDs can occur as a result of adsorption onto the surfaces of glassware used in 

sample preparation (EPA 1994c). The routine baking of glassware as a part of the cleaning process should 

be avoided because this may cause active sites on the glass that will irreversibly adsorb CDDs. The lack of 

interferences must be demonstrated under the conditions of analysis. Analysts should avoid polyvinyl 

chloride (PVC) gloves (EPA 1994c). The basic steps of sample preparation include extraction of the 

sample with a lipophilic organic solvent (e.g., hexane) followed by several evaporation and column 

chromatography steps to concentrate, clean up, and fractionate the CDDs. 

Methods of measuring CDDs in biological samples are very sensitive, generally having method (sample 

matrix) detection limits in the low- or sub-parts per trillion (ppt) level. If rigorous sample preparation 

methods are meticulously followed, sensitivity, accuracy, selectivity, and precision can be good. These 

parameters will vary with the analytical method used, the experience level of the technician, the nature of 

the sample matrix, the concentrations of the analyte(s) and possible interfering substances, and the specific 
CDDs 508 

6. ANALYTICAL METHODS 

homologue/congener being measured. High-resolution gas chromatography (HRGC) is used almost 

exclusively. The MS method may be low resolution (LRMS), high resolution (HRMS), or tandem LRMS 

(MS/MS). Individual ionization techniques that have been commonly used with MS to determine CDDs 

include electron impact ionization (EI), chemical ionization (CI), and negative chemical ionization (NCI). 

Electron impact ionization instruments are the most common although the least sensitive. The use of CI 

and NCI methods can improve instrumental sensitivity because less molecular fragmentation occurs, with 

the resulting ion current concentrated in fewer ions compared to EI. NCI is very selective for those 

compounds that tend to capture electrons and form negative ions. Both CI and NCI can greatly increase 

selectivity and sensitivity in complicated matrices. Selected ion monitoring (SIM) is most frequently used 

for quantitation; however, multiple ion monitoring (MIM), also called multiple ion detection (MID), has 

also been employed. Isotopically labeled internal standards (such as 13C- or 37Cl-labeled CDDs) are 

needed both for quantitation and to monitor method performance. Table 6-1 is a summary of some of the 

most commonly used methods for detecting CDDs in biological samples. Many of the methods for food 

and wildlife (Table 6-2) could have applicability to CDDs in human samples of similar composition. 

HRGC combined with HRMS has been used to determine parts per quadrillion (ppq) levels of CDDs in 

blood, serum, and plasma (Chang et al. 1993; Nygren et al. 1988; Patterson et al. 1987a, 1989b). Method 

8290 (EPA 1994c) is applicable to adipose tissue with a limit of detection of 1 ppt. Method 8290 has also 

been used to determine CDDs in blood and semen (Schecter 1996). The methods differ in the solvent 

system used to extract the dioxins and the types of columns used to clean up and fractionate the samples. 

The method of Chang et al. (1993) used solid phase extraction for the initial step of the isolation. 

Detection limits were comparable for CDD, but the method used by Patterson et al. (1987a) gave better 

recovery of the analyte. Precision was similar, with a coefficient of variation (CV) that ranged from 2 to 

22% for TCDD. 

2,3,7,8-TCDD has been detected (sub-ppt) in human feces using HRGC/LRMS (Wendling et al. 1990). In 

rodent metabolism studies both parent compound and metabolite were detected in feces and metabolites 

were detected in urine using GC/LRMS. HRGC/LRMS has also been used successfully in determination 

of CDDs in rat feces (Abraham et al. 1989a). Adequate comparisons of sensitivity, accuracy, and 

precision cannot be made because of the lack of these data for several methods and the differences in the 

media and analytes for the available data. 













CDDs 522 

6. ANALYTICAL METHODS 

HRGC has been combined with LRMS, HRMS, and MS/MS for the detection of CDDs in tissues. 

Sensitivity is generally in the ppt range with the best sensitivity (2 ppt) reported with MS/MS using CI 

(Ryan et al. 1987a). The limit of detection was higher for MS than for MS/MS (Schecter et al. 1985b; 

Stanley 1986; Takizawa and Muto 1987). No recovery data were given for HRMS (Nygren et al. 1988). 

Precision for these methods is usually <20% (Takizawa and Muto 1987; Van den Berg et al. 1989). 

CDDs have been measured in breast milk using HRGC/MS in the SIM mode. Reported detection limits 

are in the low- to sub-ppt (Van den Berg et al. 1986b), and recovery (75–89%) is good (Noren and Sjoevall 

1987). 

An additional screening test for TCDD-like (aryl hydrocarbon receptor, AhR, active) chemicals has been 

developed (Garrison et al. 1996) and is available commercially (Anonymous 1997). Dubbed the CALUX 

(for chemically activated luciferase gene expression) system, the assay is based on recombinant cell lines 

into which researchers have inserted a firefly luciferase gene. When exposed to dioxin-like compounds, 

the recombinant cells luminesce. The method is sensitive to ppt levels of 2,3,7,8-TCDD equivalents in 

blood, serum, and milk (Anonymous 1997). Samples testing positive can be subjected to more definitive 

and specific analytical testing. 

6.2 ENVIRONMENTAL SAMPLES 

As with biological samples, the most common method of determining CDDs in environmental samples is 

HRGC/HRMS. Other methods, including enzyme bioassays, and monoclonal antibody-based enzymelinked immunosorbent assays (ELISAs) have also been used or are under development. Even in relatively 

simple matrices, such as air and water, detection and quantitation of CDDs require rigorous sample 

preparation procedures. Methods used to prepare environmental samples are similar to those used for 

biological samples: organic solvent extraction of CDDs from the sample and concentration, clean up, and 

fractionation of the dioxins using evaporative and column chromatography techniques. The same MS 

techniques described for biological samples are available for environmental samples, with essentially the 

same results and limitations. Table 6-2 describes some of the most common methods that have been used 

to determine CDDs in environmental samples, with specific MS techniques listed when known. The 

following section describes the methods available for the different types of environmental samples. 
CDDs 523 

6. ANALYTICAL METHODS 

HRGC/LRMS and HRGC/HRMS have been used to analyze for CDDs in ambient and hazardous waste 

site air, cigarette smoke, car exhaust, and gaseous waste emissions. Sample preparation steps for gaseous 

samples are very similar for these two analytical methods. The steps consist of collection of sample 

contaminants on a filter/trapping cartridge apparatus, organic solvent extraction of the cartridge, and clean 

up and fractionation of the extract using column chromatography (Bingham et al. 1989; Cooke et al. 1988; 

Fairless et al. 1987; Harless et al. 1992; Muto and Takizawa 1989; Oehme et al. 1986; Rappe et al. 1988; 

Smith et al. 1986). A quartz fiber filter and polyurethane foam plug are commonly used to collect air 

samples (EPA 1988g; Harless et al. 1992; Kuwata et al. 1993), although XAD-2 has also been used 

(Hippelein et al. 1993). The sensitivity of these methods is in the low- to sub-pg/m3 range. Reported 

recovery and precision were generally good for measurements in air and gaseous waste emissions (Cooke 

et al. 1988; Fairless et al. 1987; Oehme et al. 1986), but severe sample loss can occur (Bingham et al. 

1989; Rappe et al. 1988). Electron capture, negative ionization, low resolution MS has also been used to 

quantify CDDs in ambient air; however, 2,3,7,8-TCDD is difficult to detect using this method and results 

must be confirmed with HRGC (Koester et al. 1992). 

Methods have been developed for detecting CDDs in liquid samples including drinking water (McCurvin 

et al. 1989; O'Keefe et al. 1986), groundwater (EPA 1986k, 1994a, 1994c; Pereira et al. 1985), fog 

(Czuczwa et al. 1989), liquid waste effluents (Cooke et al. 1988), an oil extract of landfill leachate (Först et 

al. 1988), pentachlorophenol (Singh et al. 1985), fuel oils, still bottoms, and reactor residues (EPA 1986k, 

1994a), and pyrolyzed transformer oil (Hardin et al. 1989). HRGC was combined with either LRMS, 

HRMS, or MS/MS in these methods. Not all methods reported on recovery, precision, and sensitivity, so 

it is difficult to compare these parameters. Based on the data available, sensitivities range from sub-ppq 

(O'Keefe et al. 1986) to low-ppt levels (Först et al. 1988). Recoveries were usually >60% (Först et al. 

1988; O'Keefe et al. 1986), although some lower values were reported (Hardin et al. 1989). 

HRGC/LRMS, HRGC/HRMS, HRGC/MS/MS, and HRGC/ECD have been used to analyze for CDDs in 

soils and/or sediments (Bobbie et al. 1989; Creaser and Al-Haddad 1989; Donnelly et al. 1986; EPA 

1986k, 1994a, 1994c; Eschenroeder et al. 1986; Jasinski 1989; Pereira et al. 1985; Simon et al. 1989; 

Stalling et al. 1986), solid wastes (Donnelly et al. 1986; Först et al. 1988; Popp et al. 1997), and other 

solid materials (Donnelly et al. 1986; Hardin et al. 1989; Korfmacher et al. 1985; Muto and Takizawa 

1989). Detection limits for the MS methods range from low-ppt to low-ppb levels. The sensitivity cannot 

be compared to ECD because no detection limits were reported for the ECD methods. For soil/sediments, 

recovery seemed to be better for GC/ECD (92–100%) (Jasinski 1989) than for the HRGC/MS methods 
CDDs 524 

6. ANALYTICAL METHODS 

(40–102%) (Creaser and Al-Haddad 1989; Donnelly et al. 1986; Simon et al. 1989). Polychlorinated 

biphenyls, polychlorinated diphenyl ethers, polychlorinated naphthalenes, and polychlorinated 

alkydibenzofurans may be found at concentrations several orders of magnitude higher than the analytes of 

interest (EPA 1994a) and could thus interfere with the CDDs. Retention times must be verified using 

reference standards. 

A method for determining CDDs in municipal incinerator fly ash has been reported (Alexandrou and 

Pawliszyn 1990). The method uses supercritical fluid extraction (SFE) to recover CDDs from fly ash 

samples prior to GC. Supercritical fluid extraction is faster and less expensive than the typically used 

Soxhlet extraction and gives quantitative removal of CDDs and CDFs from fly ash. Extracts obtained 

using SFE will still require additional clean-up steps prior to analysis. Supercritical CO2 has also been 

used to assist solvent-based extraction of CDDs from soils (Friedrich and Kleiböhmer 1997). In this case, 

the supercritical fluid was combined with accelerated solvent extraction (liquid extractions conducted 

under elevated temperature and pressure) to provide good recoveries relative to Soxhlet extractions. 

TCDD and other CDDs have been measured in foods (Jasinski 1989; Schecter et al. 1994; Takizawa and 

Muto 1987) and wildlife (birds and bird eggs, fish, and seals) (Bobbie et al. 1989; Buser et al. 1985; EPA 

1994a; Stalling et al. 1986) using HRGC/ECD or HRGC/LRMS. Schecter et al. (1994) reported data as 

TCDD toxic equivalents with detection limits of approximately 0.01 ppt. Ferrario et al. (1996) reported a 

new modification of EPA Method 1613 (EPA 1994a) for use in measuring CDDs and CDFs in beef fat; an 

LOD of 0.05 ppt was shown. A comparison of HRGC/LRMS methods conducted using samples from 

fish, birds, and seals showed that NCI was substantially more sensitive than EI for some, but not all, 

congeners (Buser et al. 1985). A within-lab comparison of fish tissue analysis using HRGC combined 

with either LRMS, HRMS, or MS/MS showed HRMS to be the most sensitive of the three methods 

(Bobbie et al. 1989). However, the large variations in recovery obtained with these methods also 

demonstrated the significance of the problems of sample loss and sample contamination that can occur in 

the analyses of CDDs. The data were not sufficient to permit a comparison of methods among different 

laboratories. 

Bioassays using induction of the enzymes ethoxyresorufin o-deethylase (EROD) and/or arylhydrocarbon 

hydroxylase (AHH) in rat hepatoma H-4-IIE cells (Zacharewski et al. 1989) and modified mouse liver 

cells (Schuman and Hunter 1988) have been developed and tested on water, soil, and fish samples. The 

bioassays are based on induction of AHH or EROD enzymatic activity in the cell cultures. Since the cells 

used in the bioassays are most sensitive to induction by 2,3,7,8-TCDD, this dioxin is used to generate a 
CDDs 525 

6. ANALYTICAL METHODS 

standard curve for the bioassays, and induction of activity is expressed as TCDD equivalents. These 

bioassays are highly sensitive to concentrations of Ah receptor-mediated cytochrome P-450 inducers 

(Holcomb et al. 1988; Zacharewski et al. 1989), and could be used to rapidly pre-screen environmental 

samples for 2,3,7,8-TCDD toxicity equivalents. A major drawback to these assays is that they are not 

highly selective. A number of halogenated aromatics other than CDDs can induce AHH and EROD 

activity (e.g., chlorinated dibenzofurans, polychlorinated biphenyls, and polychlorinated phenols), 

although none to the extent of TCDD induction. There is also a question about the possible effects of 

chemical mixtures, such as might be found in contaminated soil or fish, on the assay results (Zacharewski 

et al. 1989). An ELISA based on derivation of monoclonal antibodies specific to CDDs has also been 

investigated as a means of screening environmental samples for chlorinated dioxins (Stanker et al. 1987). 

Monoclonal antibodies (MAbs) developed using 1-amino-substituted 3,7,8-TrCDD derivatives could 

detect sub-ng levels of TCDD standards. The derived antibodies had a stronger affinity for CDDs 

substituted at the 1 position and for CDFs substituted at the 2, 3, 7, and 8 positions than for other CDDs 

including 2,3,7,8-TCDD. However, development of MAbs more specific for CDDs, especially 

2,3,7,8-TCDD would provide a rapid, inexpensive, sensitive, and reasonably selective method for 

screening samples for CDD contamination. Sugawara and coworkers (Sugawara et al. 1998) have recently 

described an ELISA-based method for polychlorinated dibenzo-p-dioxins that can detect as little as 0.5 

pg/well of 2,3,7,8-TCDD and shows great promise as a screening tool. The cross reactivity for octachlorodibenzo-p-dioxin is very low (<0.1%), but it is much higher for compounds with three, four, or five 

chlorine atoms in a substitution pattern similar to the of 2,3,7,8-TCDD. As with all screening approaches, 

more accurate chemical analysis would be needed to confirm the compounds present. 

The CALUX assay described in Section 6.1 has been applied to Ah receptor-active compounds (not 

limited to dioxins) in sediments and pore waters (Murk et al. 1996) and to blood with mixed results. 

Sensitivities as low as 0.5 fmol of 2,3,7,8-TCDD were reported. Two polychlorinated terphenyl mixtures, 

the PCB-substituted Ugilec 141, polybrominated diphenyl ethers, and the PCB mixture Clophen 150 were 

tested in the CALUX assay and had induction potencies that were 10-4 to 10-7 compared to TCDD. Thus, 

this assay is more selective than earlier, induction-based assays, although clearly not as selective as 

GC/MS. 

6.3 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 
 

CDDs 526 

6. ANALYTICAL METHODS 

information on the health effects of CDDs is available. Where adequate information is not available, 

ATSDR, in conjunction with the 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 

CDDs. 

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. 

6.3.1 Identification of Data Needs 

Methods for Determining Biomarkers of Exposure and Effect. Methods exist for determining 

CDDs in human serum and plasma, feces, biological tissues, and milk (Abraham et al. 1989a; Anonymous 

1997; Chang et al. 1993; EPA 1994a, 1994c; Noren and Sjoevall 1987; Nygren et al. 1988; Patterson et al. 

1987a, 1987b; Ryan et al. 1987a; Schecter et al. 1985b; Stanley 1986; Takizawa and Muto 1987; Van den 

Berg et al. 1989; Wendling et al. 1990). These methods have been used to determine ppq to ppt levels of 

CDDs in biological samples. The commonly used methods are sensitive enough to detect background 

levels of CDDs in most media, especially adipose tissue. The background concentration for nonoccupationally-exposed people has been reported to be on the order of 4 ppt in lipid (Michalek et al. 1998). 

Improved clean-up and instrument sensitivity could make blood a more useful monitoring medium, 

although it is usually reagent and background contamination that is most problematic; CDD concentrations 

in blood tend be quite low. Improvements in current methods or development of new methods to increase 

sensitivity and selectivity would help to decrease the time involved in sample preparation, and would 

reduce the high cost ($800–$1,000 per sample) and possible errors associated with current methods of 

determining exposure to CDDs. 

Several effects such as chloracne and alterations in hepatic metabolism have been associated with exposure 

to 2,3,7,8-TCDD in humans. However, these effects are not specific for 2,3,7,8-TCDD or other CDDs, but 

may be induced by numerous other chlorinated hydrocarbons. Determination of specific biomarkers of 

effect for CDD and development of reliable methods to quantify these effects would be useful in assessing 

the effects associated with exposure to CDDs. 
 

CDDs 527 

6. ANALYTICAL METHODS 

Methods for Determining Parent Compounds and Degradation Products in Environmental 

Media. Methods exist for measuring CDDs in a variety of environmental media, including air, water, 

sediment, soil, chemical waste, foods, fish, and other solid matrices (Bingham et al. 1989; Bobbie et al. 

1989; Buser et al. 1985; Cai et al. 1994; Cooke et al. 1988; Creaser and Al-Haddad 1989; Donnelly et al. 

1986; EPA 1986k, 1988g, 1994a, 1994c; Fairless et al. 1987; Jasinski 1989; Marquis et al. 1994; 

McCurvin et al. 1989; Muto and Takizawa 1989; Oehme et al. 1986; O'Keefe et al. 1986; Pereira et al. 

1985; Rappe et al. 1988; Smith et al. 1986a). Of the EPA methods, Method 8280 (EPA 1986k) and 8290 

(EPA 1994a) are both commonly used; Method 8290 is approximately three orders of magnitude more 

sensitive. Assuming an acute oral MRL of 20 pg/kg/day, an intermediate oral MRL of 7 pg/kg/day, and a 

70-kg individual, the limit of detection needed for water (2 L/day consumption) is 770 ppq for acute and 

245 ppq for intermediate exposure. The methods of O'Keefe et al. (1986) (LOD reported to be 

0.5–1.1 ppq) and EPA (1994a, 1994c) (LODs reported to be 4 ppq to 10 ppq) are adequate for detecting 

CDDs in drinking water. If a 2 kg/day consumption of food is assumed, the needed method LODs will be 

700 ppq for acute and 245 ppq for intermediate exposure. Of those method reporting LODs in foods, the 

methods of Bobbie et al. (1989) and of Ferrario et al. (1996) have the required LODs. Since CDDs are 

typically determined on a fat weight basis, the method of Ferrario et al. (1996) should be suitable for most 

food types once the fat is extracted. The sensitivity of the HRGC/MS methods is excellent, but because of 

the very low levels of these chemicals in the environment, increased sensitivity may be desirable in order 

to obtain detectable values. Increased accuracy and selectivity would help make analyses more reliable 

and possibly reduce the costly and time-consuming sample preparation steps that are currently required. 

Additional development of bioassays to detect CDDs could provide screening methods with sufficient 

sensitivity to detect the very low concentrations of toxicological importance. 

6.3.2 Ongoing Studies 

A collaborative study was identified in which researchers at CDC, NIEHS, University of Mainz in 

Germany and the German Cancer research Center in Heidelberg are studying biochemical markers of 

exposure and susceptibility to dioxin in human peripheral blood lymphocytes (Yang et al. 1997). 

The following information was obtained from a search of Federal Research in Progress (FEDRIP 1998). 

Under an SBIR (Small Business Innovative Research) grant, Xeonobiotic Detection Systems, Inc. of 

Durham, NC, is marketing the CALUX assay (Anonymous 1997) described in Section 6.1. Hybrizyme 
CDDs 528 

6. ANALYTICAL METHODS 

Corp, of Raleigh, NC, is working on a new test method for dioxins in human and animal samples. This 

work is also being performed under an SBIR. No other details were available. Antibody-based methods 

for 2,3,7,8-TCDD analysis is the subject of a project lead by R. Carlson of Ecochem Research, Inc. 

(another SBIR) during which methods for gases will be developed. Finally, G. Wheelock, Paracelsian, 

Inc., Ithaca, NY, is using SBIR funding to develop an Ah receptor-based assay for the determination of 

toxic equivalency factors. 
CDDs 529 

7. REGULATIONS AND ADVISORIES
 

The international, national, and state regulations and guidelines regarding CDDs in air, water, and other 

media are summarized in Table 7-1. 

It is ATSDR’s policy (see Appendix B) to use health guidance values (i.e., MRLs and EMEGs) derived for 

2,3,7,8-TCDD for other dioxin-like compounds expressed as total TEQs. 

ATSDR has derived an acute-duration oral MRL of 0.0002 μg/day (2×10-4 μg/kg/day) for 2,3,7,8-TCDD 

based on its ability to suppress serum total hemolytic complement activity in B6C3F1 mice (White et al. 

1986). 

An intermediate-duration oral MRL of 0.00002 μg/day (2×10-5 μg/kg/day) was derived based on observed 

decreases in thymus weight in guinea pigs (Decaprio et al. 1986). 

A chronic-duration oral MRL of 0.000001 μg/day (1×10-6 μg/kg/day) was derived for 2,3,7,8-TCDD based 

on altered social interactions with peers in monkeys exposed to 2,3,7,8-TCDD prenatally and during 

lactation (Schantz et al. 1992). 

Neither a reference concentration (RfC) nor a reference dose (RfD) is listed in IRIS (IRIS 1998) for any of 

the CDDs. 

The IRIS database does not contain a weight-of evidence cancer classification for 2,3,7,8-TCDD. The 

EPA is currently in the final stages of re-evaluating the risks and hazards from exposures to CDDs and 

CDD-like-like compounds. In its proposed rule to add a chemical category that includes dioxin and 

dioxin-like compounds to the list of toxic chemicals subject to release reporting requirements, EPA 

acknowledged that existing data shows “2,3,7,8-TCDD is a potent toxicant in animals and has the potential 

to a produce a wide spectrum of toxic effects in humans” (EPA 1997c). In the preamble of the rule EPA 

further states that “Available human data cannot clearly demonstrate whether a cause and effect 

relationship exists between 2,3,7,8-TCDD exposure and increased incidence of cancer. However, there are 

a number of epidemiological studies associating exposure to 2,3,7,8,-TCDD with cancer mortality” (EPA 

1997c). Making reference to the 1985 cancer slope factor (1.56 × 105

 [mg/kg/day]-1) (EPA 1985d) for 
CDDs 530 

7. REGULATIONS AND ADVISORIES 

2,3,7,8-TCDD and considering its own weight-of-evidence classification criteria, EPA states that “there is 

sufficient evidence to conclude that the compound is a probable human carcinogen” (EPA 1997c). In 

February, 1997, the International Agency for Research on Cancer (IARC) evaluated the evidence for 

CDDs being risk factors for human cancer (IARC 1997). Consequently, IARC currently identifies 2,3,7,8­

TCDD as being carcinogenic to humans; Group 1 carcinogen (IARC 1997). IARC concluded that there is 

limited evidence in humans for the carcinogenicity of 2,3,7,8-TCDD; however, data from studies involving 

experimental animals provided sufficient evidence of carcinogenicity. Giving consideration to supporting 

evidence such as 2,3,7,8-TCDD being a multi-site carcinogen in experimental animals; its acting through a 

mechanism involving Ah receptor which functions the same way in humans as in experimental animals; 

and similar tissue concentrations both in heavily exposed human populations and rats exposed to 

carcinogenic dosages, IARC’s overall evaluation for 2,3,7,8-TCDD is that it is carcinogenic to humans 

(IARC 1997). The Department of Health and Human Services (DHHS), National Toxicology Program 

(NTP) considers it to be a substance that is “reasonably anticipated to be a carcinogen.” Again, the 

supporting data indicate that the evidence of 2,3,7,8-TCDD carcinogenicity in humans is limited, but that 

there is sufficient evidence of carcinogenic effects in studies involving experimental animals (NTP 1998). 

NTP is currently considering a reclassification of 2,3,7,8-TCDD and the decision is pending. 

EPA regulates dioxins as hazardous air pollutants (HAPs) in accordance with the provisions of the Clean 

Air Act (CAA). EPA has promulgated guidelines and performance standards limiting dioxin and other 

HAP emissions from various sources (i.e., major, stationary, and area). A wide variety of health effects 

(e.g., cancer, respiratory problems, developmental and/or reproductive effects) have been associated with 

exposure to HAP emissions (EPA 1998c). Some of the sources for which EPA has most recently 

promulgated or proposed guidelines and standards under the authority of the CAA are municipal waste 

combustors (MWCs), hospital/medical/infectious/waste combustors (HMIWI), and process operations in 

the Portland Cement industry (EPA 1997a, 1997b, 1998c). 

Owners and operators of facilities that have chemicals subject to "The Emergency Planning and 

Community Right-to-Know Act (EPCRA) of 1986" on their sites in amounts exceeding a designated 

“reporting threshold level” are required to annually report releases of such chemicals to any environmental 

media (U.S. Congress 1986). On May 7, 1997, EPA proposed adding a chemical category that includes 

dioxin and 27 dioxin-like compounds to the list of toxic chemicals subject to the EPCRA reporting 

requirements (EPA 1997c). 
CDDs 531 

7. REGULATIONS AND ADVISORIES 

2,3,7,8-TCDD is also regulated as a drinking water contaminant. As an impurity in the production of some 

pesticides, 2,3,7,8-TCDD may get into drinking water by industrial discharge of waste. EPA has set a 

drinking water standard (Maximum Contaminant Level [MCL]) for dioxin at 3×10-8 ppm (EPA 1994d). 

People who drink water containing dioxin in excess of the MCL over many years could experience 

problems with their reproductive systems and may have an increased risk of cancer (EPA 1998d). There is 

little to no risk associated with drinking-water that meets the MCL (EPA 1994d). In February 1998, as 

mandated by the Safe Drinking Water Act (SDWA), EPA issued a proposed rule that would require 

community water systems to inform the public as to the quality of the water delivered by the system (EPA 

1998). The community right-to-know provisions of the SDWA mandate a reporting which informs the 

public of where their water comes from, shows them the process by which safe drinking water is delivered 

to their homes, provide access to information concerning source water assessments, and allows them to 

make informed decisions about their drinking water (EPA 1998d). The SDWA provisions also include 

requirements for timely notifications of violations. Within 24 hours, people served by public water 

systems must be notified of any violations of the national drinking water standard that have the potential to 

have serious adverse health effects (EPA 1998d). The SDWA amendments of 1996 required the FDA to 

issue monitoring requirements for nine allowable contaminants in bottled water (FDA 1998a, 1998b). 

2,3,7,8-TCDD is included among these nine chemicals.

 2,3,7,8-TCDD is regulated by the Clean Water Effluent Guidelines for the following industrial point 

sources: electroplating (EPA 1981a), steam electric power generating (EPA 1982a), and metal finishing 

(EPA 1986g). Limitations depend on the type of industry and plant. If waters and their sediments become 

contaminated from sources such as atmospheric deposition and discharges from industrial, municipal, or 

agricultural operations, toxic substances could concentrate in the tissue of fish and wildlife. Sixty-six 

advisories have been issued by 21 states recommending consumers limit their consumption of fish and 

shellfish (EPA 1998b). For 14 states (Wisconsin, Rhode Island, New Hampshire, West Virginia, 

Louisiana, Arkansas, Virginia, Michigan, Mississippi, Florida, Massachusetts, Oregon, Tennessee, and 

Delaware) advisories were issued for freshwater fish (EPA 1998b). Only two states (Texas and California) 

issued advisories for marine waters only (EPA 1998b). An advisory for woodduck (migratory fowl) was 

issued for the state of Arkansas. This information is current as of December 1997, based on the EPA Fish 

and Wildlife Advisory Database searched October 1998 at: http://www.epa.gov/OST/fishadvice/. More 

detailed information can be obtained from the state Public Health Department or the state Department of 

Natural Resources. A fish or wildlife advisory will specify the bodies of water or hunting areas with 

restrictions. The advisory will indicate what species and size of fish or game are of concern. The advisory 
CDDs 532 

7. REGULATIONS AND ADVISORIES 

may completely ban consumption or recommend limiting meals of a certain fish or wildlife species to a 

particular frequency. For example, an advisory may recommend that a person eat a certain type of fish no 

more than once a month. The advisory may indicate that only certain parts of the fish or game should be 

consumed and recommend preparation methods that minimize exposure. Fish and wildlife advisories may 

also provide restrictions specifically targeting pregnant women, nursing mothers, and young children. 

Each state, Native American Tribe, or U.S. Territory chooses its own criteria for issuing fish and wildlife 

advisories. 

2,3,7,8-TCDD is regulated as a hazardous waste constituent under the requirements of the Resource 

Conservation and Recovery Act (RCRA) (EPA 1988d). Non-specific sources of 2,3,7,8-TCDD-containing 

waste are wastes from the production or manufacturing use of tri-, tetra-, or pentachlorophenols and their 

pesticide derivatives, discarded or unused formulations containing these compounds, and residues from 

incineration or thermal treatment of soil contaminated with these compounds. RCRA prohibits land 

disposal of hazardous waste unless it meets treatment standards established by the EPA. On May 12, 

1997, the EPA promulgated universal treatment standards (UTSs) for hazardous constituents in wood 

preserving waste. These wastes have been assigned EPA hazardous waste codes F032, F034, and F035 

(EPA 1997d). The final rule also promulgated a compliance alternative for dioxin constituents in 

nonwastewater and wastewater forms of F032 waste which allowed combustion to be used as a method of 

treatment (EPA 1997d). 

The Toxic Substance Control Act (TSCA) authorizes the EPA to determine whether the use of a chemical 

substance is a “significant new use” (EPA 1998e). Once it has been determined that a use of a chemical is 

a significant new use, it must be reported to the EPA prior to manufacturing, importing, or processing for 

the new use. The required notice will provide the EPA with an opportunity to evaluate the intended use, 

and if necessary, to prohibit or limit the activity before it occurs (EPA 1998e). For example, brominated 

phthalate ester was included among the 163 chemical substances for which the EPA promulgated 

significant new use rules (SNURs). The toxicity concern for the new use was that when similar chemicals 

have been incinerated under combustion conditions of municipal incinerators, dibenzodioxins and 

dibenzofurans were formed (EPA 1998e). Persons providing notice of this new use would need to 

characterize, through an incineration simulation, the potential for dioxin and furan formation and agree not 

to exceed the production volume limit without performing the characterization (EPA 1998e). 










CDDs 543 

8. REFERENCES
 

*Abbott BD, Birnbaum LS. 1989a. Cellular alterations and enhanced induction of cleft palate after 

coadministration of retinoic acid and TCDD. Toxicol Appl Pharmacol 99:287-301. 

*Abbott BD, Birnbaum LS. 1989b. TCDD alters medial epithelial cell differentiation during 

palatogenesis. Toxicol Appl Pharmacol 99:276-286. 

*Abbott BD, Birnbaum LS. 1990a. TCDD-induced altered expression of growth factors may have a role 

in producing cleft palate and enhancing the incidence of clefts after coadministration of retinoic acid and 

TCDD. Toxicol Appl Pharmacol 106:418-432. 

*Abbott BD, Birnbaum LS. 1990b. Effects of TCDD on embryonic ureteric epithelial EGF receptor 

expression and cell proliferation. Teratol 41:71-84. 

*Abbott BD, Birnbaum LS, Dilberto JJ. 1996. Rapid distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) to embryonic tissues in C57BL/6N mice and correlation with palatal uptake in vitro. Toxicol 

Appl Pharmacol 141:256-263. 

*Abbott BD, Birnbaum LS, Perdew GH. 1995. Developmental expression of two members of a new class 

of transcription factors: I. Expression of aryl hydrocarbon receptor in the C57BL/6N mouse embryo. 

Develop Dynam 204:133-143. 

*Abbott BD, Birnbaum LS, Pratt RM. 1987a. TCDD-induced hyperplasia of the ureteral epithelium 

produces hydronephrosis in murine fetuses. Teratology 35:329-334. 

*Abbott BD, Diliberto JJ, Birnbaum LS. 1989. 2,3,7,8-Tetrachlorodibenzo-p-dioxin alters embryonic 

palatal medial epithelial cell differentiation in vitro. Toxicol Appl Pharmacol 100:119-131. 

*Abbott BD, Harris MW, Birnbaum LS. 1992. Comparisons of the effects of TCDD and hydrocortisone 

on growth factor expression provide insight into their interaction in the embryonic mouse palate. Teratol 

45:35-53. 

*Abbott BD, Morgan KS, Birnbaum LS, et al. 1987b. TCDD alters the extracellular matrix and basal 

lamina of the fetal mouse kidney. Teratol 35:335-344. 

*Abbott BD, Perdew GH, Birnbaum LS. 1994b. Ah receptor in embryonic mouse palate and effects of 

TCDD on receptor expression. Toxicol Appl Pharmacol 126:16-25. 

*Abbott BD, Perdew GH, Buckalew AR, et al. 1994a. Interactive regulation of Ah and glucocorticoid 

receptors in the synergistic induction of cleft palate by 2,3,7,8-tetrachlorodibenzo-p-dioxin and 

hydrocortisone. Toxicol Appl Pharmacol 128:138-150. 

*Abbott BD, Probst MR. 1995. Developmental expression of two members of a new class of transcription 

factors: II. Expression of aryl hydrocarbon receptor nuclear translocator in the C57BL/6N mouse embryo. 

Develop Dynam 204:144-155 

*Cited in text 
CDDs 544 

8. REFERENCES 

Abernathy DJ, Boreiko CJ. 1987. Promotion of C3H/10T1/2 morphological transformation by 

polychlorinated dibenzo-p-dioxin isomers. Carcinogenesis 8:1485-1490. 

Abernathy DJ, Greenlee WF, Hubard JC, et al. 1985. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) 

promotes the transformation of C3H/10T1/2 cells. Develop Dynam 204:144-155 

*Abraham K, Hille A, Ende M, et al. 1994. Intake and fecal excretion of PCDDs, PCDFs, HCB and PCBs 

(138,153,180) in a breast-fed and a formula-fed infant. Chemosphere 29:2279-2286. 

*Abraham K, Knoll A, Ende M, et al. 1996. Intake, fecal excretion, and body burden of polychlorinated 

dibenzo-p-dioxins and dibenzofurans in breast-fed and formula-fed infants. Pediatr Res 40:671-679. 

*Abraham K, Krowke R, Neubert D. 1988. Pharmacokinetics and biological activity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin: 1. Dose-dependent tissue distribution and induction of hepatic 

ethoxyresorufin O-deethylase in rats following a single injection. Arch Toxicol 62:359-368. 

*Abraham K, Weberrub U, Wiesmuller T, et al. 1989c. Comparative studies on absorption and 

distribution in the liver and adipose tissue of PCDDs and PCDFs in rats and marmoset monkeys. 

Chemosphere 19:887-892. 

*Abraham K, Wiesmueller T, Brunner H, et al. 1989a. Elimination of various polychlorinated 

dibenzo-p-dioxins and dibenzofurans (PCDDs and PCDFs) in rat feces. Arch Toxicol 63:75-78. 

*Abraham K, Wiesmuller T, Brunner H, et al. 1989b. Absorption and tissue distribution of various 

polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDDs and PCDFs) in the rat. Arch Toxicol 

63:193-202. 

*ACGIH. 1990. Threshold limit values and biological exposure indices for 1990û1991. American 

Conference of Governmental Industrial Hygienists. Cincinnati, OH. 

*Ackermann MF, Gasiewicz TA, Lamm KR, et al. 1989. Selective inhibition of polymorphonuclear 

neutrophil activity by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 101:470-480. 

Adams WJ, Blaine KM. 1986. A water solubility determination of 2,3,7,8-TCDD. Chemosphere 

15:1397-1400. 

*Adams WJ, Degraeve GM, Sabourin TD, et al. 1986. Toxicity and bioconcentration of 2,3,7,8,-TCDD 

to fathead minnows (Pimephales promelas). Chemosphere 15:1503-1511. 

*Adinolfi M. 1985. The Development of the Human Blood-CSF-Brain Barrier. Developmental Medicine 

& Child Neurology 27:532-537. 

*Adriaens P, Grbic-Galic D. 1994. Reductive dechlorination of PCDD/F by anaerobic cultures and 

sediments. Chemosphere 29(9-11):2253-2253. 

Afghan BK, Carron J, Goulden PD, et al. 1987. Recent advances in ultratrace analysis of dioxins and 

related halogenated hydrocarbons. Can J Chem 65:1086-1097. 
CDDs 545 

8. REFERENCES 

*Ahlborg UG, Becking GC, Birnbaum LS, et al. 1994. Toxic equivalency factors for dioxin-like PCBs. 

Chemosphere 28(6):1049-1067. 

*Ahlborg UG, Brouwer A, Fingerhut MA, et al. 1992. Impact of polychlorinated dibenzo-p-dioxins, 

dibenzofurans, and biphenyls on human and environmental health, with special emphasis on application of 

the toxic equivalency factor concept. Europ J Pharmacol 228:179-199. 

Ahlborg UG, Hanberg A. 1992. Toxicokinetics of PCDDs and PCDFs of importance to the development 

of human risk assessment. Toxic Substances Journal 12:197-211. 

Ahlborg UG, Waern F, Håkansson H. 1987. Interactive effects of PCDDs and PCDFs occurring in human 

mother's milk. Chemosphere 16:1701-1706. 

Ahling B. 1979. Destruction of chlorinated hydrocarbons in cement kiln. Environ Sci Technol 

13:1377-1379. 

Ahling B, Liondskog A, Jansson B, et al. 1977. Formation of polychlorinated dibenzo-p-dioxins and 

dibenzofurans during combustion of a 2,4,5-T formulation. Chemosphere 6:461-468. 

*Al-Bayati ZA, Wahba ZZ, Stohs SJ. 1988. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)-induced 

alterations in lipid peroxidation, enzymes, and divalent cations in rat testis. Xenobiotica 18:1281-1289. 

*Alaluusua S, Lukinmaa P-L, Vartiainen T, et al. 1996. Polychlorinated dibenzo-p-dioxins and 

dibenzofurans via mother's milk may cause developmental defects in the child's teeth. Environ Toxicol 

Pharmacol 1:193-197. 

Albro PW, Crummett WB, Dupuy AE Jr, et al. 1985. Methods for the quantitative determination of 

multiple, specific polychlorinated dibenzo-p-dioxin and dibenzofuran isomer in human adipose tissue in 

the parts-per-trillion range: An interlaboratory study. Anal Chem 57:2717-2725. 

Albro PW, Lister MI, Cahe K, et al. 1979. A radioimmunoassay for chlorinated dibenzo-p-dioxins. 

Toxicol Appl Pharmacol 50:137-146. 

Aldred JE. 1978. Report of the Consultative Council on congenital abnormalities in the Yarrom District. 

Minister of Health. Melbourne, Victoria, Australia. 

Alexander LR, Maggio VL, Gill JB, et al. 1989. Performance of the high sensitivity ion source for 

PCDD/PCDF analyses in human serum and adipose tissue. Chemosphere 19:241-248. 

Alexandrou N, Miao Z, Colquhoun M, et al. 1992. Supercritical fluid extraction and cleanup with 

capillary GC-ion trap mass spectrometry for determination of polychlorinated dibenzo-p-dioxins and 

dibenzofurans in environemental samples. J Chromatogr Sci 30(9):335-382. 

*Alexandrou N, Pawliszyn JB. 1990. Supercritical fluid extraction for the rapid determination of 

polychlorinated dibenzo-p-dioxins and dibenzofurans in municipal incinerator fly ash [Abstract]. New 

York, NY: 41st Pittsburgh Conference and Exposition on Analytical Chemistry and applied chemistry. 

Allen JR, Barsotti DA, Lambrecht LK, et al. 1979. Reproductive effects of halogenated aromatic 

hydrocarbons on nonhuman primates. Ann NY Acad Sci 320:419-425. 

*Allen JR, Barsotti DA, Van Miller JP, et al. 1977. Morphological changes in monkeys consuming a diet 

containing low levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Food Cosmet Toxicol 15:401-410. 
CDDs 546 

8. REFERENCES 

*Allen JR, Van Miller JP, Norback DH. 1975. Tissue distribution, excretion and biological effects of 

[14C]tetrachlorodibenzo-p-dioxin in rats. Food Cosmet Toxicol 13:501-505. 

*Altman PK, Dittmer DS. 1974. In: Biological handbooks: Biology data book, Volume III, second 

edition. Bethesda, MD: Federation of American Societies for Experimental Biology, 1987-2008, 2041. 

*Amendola G, Barna D, Blosser R, et al. 1989. The occurrence and fate of PCDDs and PCDFs in five 

bleached kraft pulp and paper mills. Chemosphere 18:1181-1188. 

*Anaizi NH, Cohen JJ. 1978. The effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on the renal tubular 

secretion of phenolsulfonphthalein. J Pharmacol Exp Ther 207:748-755. 

Andersen ME. 1985. Adjusting exposure limits for long and short exposure periods using a physiological 

pharmacokinetic model. Am ind Hyg Assoc J 48(4):335-343. 

*Andersen ME, Birnbaum LS, Barton HA, et al. 1997b. Regional hepatic CYP1A1 and CYP1A2 

induction with 2,3,7,8-tetrachlorodibenzo-p-dioxin evaluated with a multicompartment geometric model of 

hepatic zonation. Toxicol Appl Pharmacol 144:145-155. 

*Andersen ME, Eklund CR, Mills JJ, et al. 1997a. A multicompartment geometric model of the liver in 

relation to regional induction of cytochrome P450s. Toxicol Appl Pharmacol 144:135-144. 

*Andersen ME, Krishman K. 1994. Relating in vitro to in vivo exposures with physiologically-based 

tissue dosimetry and tissue response models. In: H. Salem, ed. Current concepts and approaches on animal 

test alternatves. U.S. Army Chemical Research Development and Engineering Center, Aberdeen Proving 

Ground, Maryland. 

*Andersen ME, Krishnan 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. U.S. Army 

Chemical Research Development and Engineering Center, Aberdeen Proving Ground, Maryland. 

*Andersen ME, MacNaughton MG, Clewell HJ, et al. 1987. Adjusting exposure limits for long and short 

exposure periods using a physiological pharmacokinetic model. Am Ind Hyg Assoc J 48(4):335-343. 

*Andersen ME, Mills JJ, Gargas ML, et al. 1993. Modeling receptor-mediated processes with dioxin: 

Implications for pharmacokinetics and risk assessment. Risk Analysis 13:25-36. 

*Anderson HA, Falk C, Hanrahan L, et al. 1998. Profiles of Great Lakes critical pollutants: A sentinel 

analysis of human blood and urine. Environmental Health Perspectives 106(5):279-289. 

*Andrews JS Jr, Garret WA, Patterson DG, et al. 1989. 2,3,7,8-Tetrachlorodibenzo-p-dioxin levels in 

adipose tissue of persons with no known exposure and in exposed persons. Chemosphere 18:499-506. 

Aniline O. 1973. Preparation of chlorodibenzo-p-dioxins for toxicological evaluation. In: Blair EH, ed. 

Advances in Chemistry Series No. 120: Chlorodioxins-origin and fate. Proceedings of the 162nd 

Meeting of the American Chemical Society. Washington, DC, 126-135. 
CDDs 547 

8. REFERENCES 

*Ankley GT, Niemi GJ, Lodge KB, et al. 1993. Uptake of planar polychlorinated biphenyls and 

2,3,7,8-substituted polychlorinated dibenozfurans and dibenzo-p-dioxins by birds nesting in the Lower Fox 

River and Green Bay, Wisconsin, USA. Arch Environ Contam Toxicol 24(3):332-344. 

Anonymous. 1985. Dioxins in the environment: No consensus on human hazard. Chem Eng News May 

27, 41-44. 

*Anonymous. 1997. An enlightened approach to screening for dioxins. Environmental Health 

Perspectives 105(11):1176-1177. 

*Antila S, Hietanen E, Vainio H, et al. 1991. Smoking and peripheral type of cancer are related to high 

levels of pulmonary cytochrome P450IA in lung cancer patients. Int J Cancer 47:681-685. 

*Arthur MR, Frea JL. 1989. 2,3,7,8-Tetrachlorodibenzo-p-dioxin: Aspects of its important properties 

and its potential biodegradation in soils. J Environ Qual 18:1-11. 

*Aschengrau A, Monson RR. 1989. Paternal military service in Vietnam and risk of spontaneous 

abortion. J Occup Med 31:618-623. 

*Aschengrau A, Monson RR. 1990. Paternal military service in Vietnam and the risk of late adverse 

pregnancy outcomes. Am J Publ Health 80:1218-1223. 

*Assennato G, Cervino D, Emmett EA, et al. 1989. Follow-up of subjects who developed chloracne 

following TCDD exposure at Seveso. Am J Ind Med 16:119-125. 

*ASTER. 1995. ASTER (Assessment Tools for the Evaluation of Risk) ecotoxicity profile. Duluth, MN: 

Environmental Research Laboratory, U.S. Environmental Protection Agency, 

Astle JW, Gobas FAPC, Siu WY, et al. 1987. Lake sediments as historic records of atmospheric 

contamination by organic chemicals. Adv Chem Ser 216:57-77. 

*Astroff B, Eldridge B, Safe S. 1991. Inhibition of the 17'-estradiol-induced and constitutive expression 

of the cellular proto-oncogene c-fos by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the female rat 

uterus. Toxicol Lett 56:305-315. 

*Astroff B, Safe S. 1990. 2,3,7,8-Tetrachlorodibenzo-p-dioxin as an antiestrogen: Effect on rat uterine 

peroxidase activity. Biochem Pharmacol 39:485-488. 

Atkinson R. 1987. Estimation of OH radical reaction rate constants and atmospheric lifetimes for 

polychlorobiphenyls, dibenzo-p-dioxins, and dibenzofurans. Environ Sci Technol 21:305-307. 

*Atkinson R. 1991. Atmospheric lifetimes of dibenzo-p-dioxins and dibenzofurans. Sci Total Environ 

104(1-2):17-33. 

*ATSDR. 1989. Decision guide for identifying substance-specific data needs related to toxicologicol 

profiles. Agency for Toxic Substances and Disease Registry, Division of Toxicology, Atlanta, GA. 

*ATSDR. 1994. Toxicological profile for chlorinated dibenzofurans (CDFs). Agency for Toxic 

Substances and Disease Registry, Atlanta, GA. 
CDDs 548 

8. REFERENCES 

*ATSDR. 1996. National exposure registry: Dioxin subregistry. Baseline and followups 1 and 2 

technical report. U.S. Department of Health & Human Services, Public Health Service, Agency for Toxic 

Substances and Disease Registry (Atlanta, Georgia). 

*ATSDR. 1997. Toxicological profile for polychlorinated biphenyls (update). Agency for Toxic 

Substances and Disease Registry, Atlanta, GA. 

*ATSDR/CDC. 1990. Subcommittee report on biological indicators of organ damage. Agency for Toxic 

Substances and Disease Registry, Centers for Disease Control and Prevention, Atlanta, GA. 

Aulerich RJ, Bursian SJ, Napolitano AC. 1988. Biological effects of epidermal growth factor and 

2,3,7,8-tetrachlorodibenzo-p-dioxin on developmental parameters of neonatal mink. Arch Environ 

Contam Toxicol 17:27-31. 

*Aust SD. 1984. On the mechanism of anorexia and toxicity of TCDD and related compounds. In: 

Poland A, Kimbrough RD, eds. Banbury report 18. Cold Spring Harbor Laboratory, 309-319. 

Axegard P, Renberg L. 1989. The influence of bleaching chemicals and lignin content on the formation 

of polychlorinated dioxins and dibenzofurans. Chemosphere 19:661-668. 

*Axelson O, Sundell L, Anderson K, et al. 1980. Herbicide exposure and tumor mortality: An updated 

epidemiologic investigation on Swedish railroad workers. Scand J Work Environ Health 6:73-79. 

*Ayotte P, Carrier G, Dewailly E. 1996. Health risk assessement for Inuit newborns exposed to 

dioxin-like compounds through breast milk. Chemosphere 32:531-542. 

*Ayotte P, Dewailly E, Ryan JJ, et al. 1997. PCBs and dioxin-like compounds in plasma of adult inuit 

living in Nunavik (Arctic Quebec). Chemosphere 34(5-7):1459-1468. 

Ayres DC. 1981. Destruction of polychlorinated-p-dioxins. Nature 290:323-324. 

Ayres SM, Webb KB, Evans RG, et al. 1985. Is 2,3,7,8-TCDD (dioxin) a carcinogen for humans? 

Environ Health Perspect 62:329-335. 

Baars AJ, Jansen M, Breimer DD. 1978. The influence of phenobarbital, 3-methylcholanthrene and 

2,3,7,8-tetra chlorodibenzo-p-dioxin on glutathione S-transferase activity of rat liver cytosol. Biochem 

Pharmacol 27:2487-2494. 

*Badesha JS, Maliji G, Flaks B. 1995. Immunotoxic effects of prolonged dietary exposure of male rats to 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Eur J Pharmacol 293:429-437. 

Balarajan R, Acheson ED. 1984. Soft tissue sarcoma in agriculture and forestry workers. J Epidemiol 

Community Health 38:113-116. 

*Balk JL, Piper WN. 1984. Altered blood levels of corticosteroids in the rat after exposure of 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem Pharmacol 33:2531-2534. 

*Ball LM, Chhabra RS. 1981. Intestinal absorption of nutrients in rats treated with 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). J Toxicol Environ Health 8:629-638. 
CDDs 549 

8. REFERENCES 

Ballschmiter K, Budhert H, Neimczyk R, et al. 1986. Automobile exhausts versus municipal-waste 

incineration as sources of the polychloro-dibenzodioxins (PCDD) and furans (PCDF) found in the 

environment. Chemosphere 15:901-915. 

Ballschmiter K, Kraemer W, Magg H, et al. 1984. Distribution of polychlorodibenzodioxin and furan 

emissions between particulates, flue gas condensate, and ipenger adsorption in stack gas sampling. 

Chemosphere 13:1139-1142. 

*Banks YB, Birnbaum LS. 1991. Absorption of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) after low 

dose dermal exposure. Toxicol Appl Pharmacol 107:302-310. 

*Banks YB, Brewster DW, Birnbaum LS. 1990. Age-related changes in dermal absorption of 

2,3,7,8-tetrachlorodibenzo-p-dioxin and 2,3,4,7,8-pentachlorodibenzofuran. Fund Appl Toxicol 

15:163-173. 

*Banks YB, Jackson JA, Birnbaum LS. 1993. Maturational changes in dermal absorption of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in Fischer 344 rats. Toxicol Appl Pharmacol. 119:214-220. 

*Bannister R, Davis D, Zacharewski Z, et al. 1987. Aroclor 1254 as a 2,3,7,8-tetrachlorodibenzo-p-dioxin 

antagonist: Effects on enzyme induction and immunotoxicity. Toxicology 40:29-42. 

Barber MC, Suarez LA, Lassiter RR. 1988. Modeling bioconcentration of non polar organic pollutants by 

fish. Environ Toxicol Chem 7:545-558. 

*Barbieri S, Pirovano C, Scarlato G, et al. 1988. Long-term effects on 

2,3,7,8-tetrachlorodibenzo-p-dioxin on the peripheral nervous system: Clinical and neurophysiological 

controlled study on subjects with chloracne from the Seveso area. Neuroepidemiology 7:29-37. 

*Barnes D, Alford-Stevens A, Birnbaum L, et al. 1991. Toxicity equivalency factors for PCBs? Quality 

assurance: Good Practice, Regulation, and Law. 1:70-81. 

*Barnes D, Dourson M. 1988. Reference dose (RfD): Description and use in health risk assessments. 

Regul Toxicol Pharmacol 8:471-486. 

Barnes DG. 1983. Chlorinated dioxins work group. Washington, DC: U.S. Environmental Protection 

Agency. 12:645-655. 

*Barnes DG. 1991. Toxicity equivalents and EPA's risk assessment of 2,3,7,8-TCDD. Sci Total Environ 

104:73-86. 

*Barnes DG, Dourson M. 1988. Reference dose (RfD): Description and use in health risk assessments. 

U.S. Environmental Protection Agency. Regul Toxicol Pharmacol 8:471-486. 

Barnes JM. 1975. Assessing hazards from prolonged and repeated exposure to low doses of toxic 

substances. Brit Med Bull 31:196-200. 

*Barthel E. 1981. Increased risk of lung cancer in pesticide-exposed male agricultural workers. J Toxicol 

Environ Health 8:1027-1040. 
CDDs 550 

8. REFERENCES 

*Bartsch H, Hietanen E, Petruzzelli S, et al. 1990. Possible prognostic value of pulmonary 

AH-locus-linked enzymes in patients with tobacco-related lung cancer. Int J Cancer 46:185-188. 

*Basharova GR. 1996. Reproduction in families of workers exposed to 2,4,5-T intoxication. 

Organohalogen Compounds 315-318. 

*Bastomsky CH. 1977. Enhanced thyroxine metabolism and high uptake goiters in rats after a single dose 

of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Endocrinology 101:292-296. 

Batterman AR, Cook PM, Lodge KB, et al. 1989. Methodology used for a laboratory determination of 

relative contributions of water, sediment and food chain routes of uptake for 2,3,7,8-TCDD 

bioaccumulation by lake trout in Lake Ontario. Chemosphere 19:451-458. 

*Baughman RW, Meselson M. 1973. An analytical method for detecting TCDD (dioxin): Levels of 

TCDD in samples from Vietnam. Environ Health Perspect 5:27-35. 

*Becher H, Flesch-Janys D, Kauppinen T, et al. 1996. Cancer mortality in German male workers exposed 

to phenoxy herbicides and dioxins. Cancer Causes Control 7:312-321. 

*Beck H, Dro A, Eckart K, et al. 1989d. PCDDs, PCDFs and related compounds in paper products. 

Chemosphere 19:655-660. 

Beck H, Dro A, Mathar W. 1990. Interlaboratory study on PCDDs and PCDFs in coffee filter paper. 

Chemosphere 20:1307-1312. 

Beck H, Eckart K, Kellert M, et al. 1987. Levels of PCDD's and PCDF's in samples of human origin and 

food in the Federal Republic of Germany. Chemosphere 16:1977-1982. 

Beck H, Eckart K, Mathar W, et al. 1988a. Isomer-specific determination of 

2,3,7,8-tetrachlorodibenzo-p-dioxin and related compounds in human fat and food. Biomed Environ Mass 

Spectrom 16:161-165. 

*Beck H, Eckart K, Mathar W, et al. 1988b. Occurrence of PCDD and PCDF in different kinds of paper. 

Chemosphere 17:51-57. 

*Beck H, Eckart K, Mathar W, et al. 1989a. PCDD and PCDF body burden from food intake in the 

Federal Republic of Germany. Chemosphere 18:417-424. 

Beck H, Eckart K, Mathar W, et al. 1989b. Dependence of PCDD and PCDF levels in human milk on 

various parameters in the Federal Republic of Germany. Chemosphere 18:1063-1066. 

*Beck H, Eckart K, Mathar W, et al. 1989c. Levels of PCDDs and PCDFs in adipose tissue of 

occupationally exposed workers. Chemosphere 18:507-516. 

Beebe LE. 1990. Adenylate cyclase regulation in the spermatogenic cell plasma membrane: Modulating 

effects of TPA and TCDD [Abstract]. Diss Abstr Int [B] 50:2860 

Benthe C, Heinzow B, Jessen H, et al. 1992. Polychlorinated biphenyls indoor air contamination due to 

thiokol-rubber sealants in an office building. Eleventh international symposium on chlorinated dioxins and 
CDDs 551 

8. REFERENCES 

related compounds, Part 2. Research Triangle Park, North Carolina, USA. Chemosphere 

25(7-10):1481-1486. 

*Berghard A, Gradin K, Pongratz I, et al. 1993. Cross-coupling of signal transduction pathways: the dioxin 

receptor mediates induction of cytochrome P-450IA1 expression via a protein kinase C-dependent 

mechanism. Mol Cell Biol 13:67-689. 

*Berry DL, DiGiovanni J, Juchau MR, et al. 1978. Lack of tumor-promoting ability of certain 

environmental chemicals in a two-stage mouse skin tumorigenesis assay. Res Commun Chem Pathol 

Pharmacol 20:101-108. 

*Berry DL, Slaga TJ, DiGiovanni J, et al. 1979. Studies with chlorinated dibenzo-p-dioxins, 

polybrominated biphenyls and polychlorinated biphenyls in a two-stage system of mouse skin tumorigenesis: 

Potent anticarcinogenic effects. Ann NY Acad Sci 320:405-414. 

*Berry RM, Luthe CE, Voss RH. 1993. Ubiquitous nature of dioxins: A comparison of the dioxins content 

of common everyday materials with that of pulps and papers. Environ Sci Technol 27(6):1164-1168. 

*Bertazzi PA, Pesatori AC, Consonni D, et al. 1993. Cancer incidence in a population accidentally exposed 

to 2,3,7,8-tetrachlorodibenzo-para-dioxin. Epidemiol 4:398-406. 

*Bertazzi PA, Zocchetti C, Guercilena S, et al. 1997. Dioxin exposure and cancer risk: A 15-year mortality 

study after the "Seveso Accident". Epidemiology 8:646-652. 

*Bertazzi PA, Zocchetti C, Pesatori AC, et al. 1989a. Mortality in an area contaminated by TCDD 

following an industrial incident. Med Lav 80:316-329. 

*Bertazzi PA, Zocchetti C, Pesatori AC, et al. 1989b. Ten-year mortality study of the population involved 

in the Seveso incident in 1976. Am J Epidemiol 129:1187-1200. 

*Bestervelt LL, Cai Y, Piper DW, et al. 1993. TCDD alters pituitary-adrenal function I: Adrenal 

responsiveness to exogenous ACTH. Neurotoxicol Teratol 15:365-370. 

Bestervelt LL, Nolan CJ, Cai Y, et al. 1991. Altered rat brain mu opiate receptor number after exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) [Abstract 1192]. The Toxicologist 11:08. 

*Bestervelt LL, Pitt JA, Nolan CJ, et al. 1993. TCDD alters pituitary-adrenal function. II. Evidence for 

decreased bioactivity of ACTH. Neurotoxicol Teratol 15(6):371-376. 

*Beurskens JEM, Toussaint M, De Wolf J, et al. 1995. Dehalogenation of chlorinated dioxins by an 

anaerobic microbial consortium from sediment. Environmental Toxicology and Chemistry 14(6):939-943. 

Beychok MR. 1987. A data base of dioxin and furan emissions from municipal refuse incinerators. Atmos 

Environ 21:29-36. 

Bickel MH. 1982. Polychlorinated persistent compounds. Experientia 38:879. 
CDDs 552 

8. REFERENCES 

Biddinger GR, Gloss SP. 1984. The importance of trophic transfer in the bioaccumulation of chemical 

contaminants in aquatic ecosystems. Res Rev 91:103-146. 

*Bingham AG, Edmunds CJ, Graham BWL, et al. 1989. Determination of PCDDs and PCDFs in car 

exhaust. Chemosphere 19:669-673. 

Birmingham B. 1990. Analysis of PCDD and PCDF patterns in soil samples: Use in the estimation of the 

risk of exposure. Chemosphere 20(7-9):807-814. 

*Birmingham B, Thorpe B, Frank R, et al. 1989. Dietary intake of PCDD and PCDF from food in Ontario, 

Canada. Chemosphere 19:507-512. 

*Birnbaum LS. 1985. The role of structure in the disposition of halogenated aromatic xenobiotics. Environ 

Health Perspect 61:11-20. 

Birnbaum LS. 1986. Distribution and excretion of 2,3,7,8-tetrachlorodibenzo-p-dioxin in cogenic strains of 

mice which differ at the Ah locus. Drug Metab Dispos 14:34-40. 

*Birnbaum LS. 1994a. The mechanism of dioxin toxicity: Relationship to risk assessment. Environ Health 

Perspect 102:157-167. 

*Birnbaum LS. 1994b. Evidence for the role of the Ah receptor in response to dioxin. In: 

Receptor-mediated biological process: Implications for evaluating carcinogenesis, 139-154. 

*Birnbaum LS, Couture LA. 1988. Disposition of octachlorodibenzo-p-dioxin (OCDD) in male rats. 

Toxicol Appl Pharmacol. 93:22-30. 

*Birnbaum LS, Couture LA, Elwell MR. 1989. Subchronic effects of exposure to octachlorodibenzodioxin 

(OCDD). Chemosphere 18:389-390. 

*Birnbaum LS, Harris MW, Miller CP, et al. 1986. Synergistic interaction of 

2,3,7,8-tetrachlorodibenzo-p-dioxin and hydrocortisone in the induction of cleft palate in mice. Teratol 

33:29-35. 

*Birnbaum LS, Harris MW, Stocking LM, et al. 1989. Retinoic acid and 

2,3,7,8-tetrachlorodibenzo-p-dioxin selectively enhance teratogenesis in C57BL/6N mice. Toxicol Appl 

Pharmacol 98:487-500. 

*Birnbaum LS, McDonald MM, Blair PC, et al. 1990. Differential toxicity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in C57Bl/6J mice congenic at the Ah locus. Fundam Appl 

Toxicol 15:186-197. 

*Birnbaum LS, Weber H, Harris MW, et al. 1985. Toxic interaction of specific polychlorinated biphenyls 

and 2,3,7,8-tetrachlorodibenzo-p-dioxin: Increased incidence of cleft palate in mice. Toxicol Appl 

Pharmacol 77:292-302. 

*Bisanti L, Bonetti F, Caramaschi F, et al. 1980. Experiences from the accident of Seveso. Acta 

Morphologica Acad Sci Hung 28:139-157. 
CDDs 553 

8. REFERENCES 

Bishop CM, Jones AH. 1981. Non-Hodgkin's lymphoma of the scalp in workers exposed to dioxins. 

Lancet 2:369. 

*Bjerke DL, Brown TJ, MacLusky NJ, et al. 1994b. Partial demasculinization and feminization of sex 

behavior in male rats by in utero and lactational exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin is not 

associated with alterations in estrogen receptor binding or volumes of sexually differentiated brain nuclei. 

Toxicol Appl Pharmacol 127:258-267. 

*Bjerke DL, Peterson RE. 1994. Reproductive toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in male rats: 

different effects of in utero versus lactational exposure. Toxicol Appl Pharmacol 127:241-249. 

*Bjerke DL,Sommer RJ, Moore RW et al. 1994a. Effects of in utero and lactational 

2,3,7,8-tetrachlorodibenzo-p-dioxin exposure on responsiveness of the male rat reproductive system to 

testosterone stimulation in adulthood. Toxicol Appl Pharmacol 127:250-257. 

*Blank JA, Tucker AN, Sweatlock J, et al. 1987. Naphthoflavone antagonism of 

2,3,7,8-tetrachlorodibenzo-p-dioxin-induced murine lymphocyte ethoxyresorufin-O-deethylase activity and 

immunosuppression. Mol Pharmacol 32:168-172. 

*Blankenship AL, Suffia MC, Matsumura F, et al. 1993. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCCD) 

accelerated differentiation of murine preimplantation embryos in vitro. Reprod Toxicol 7(3):255-261. 

*Blaylock BL, Holladay SD, Comment CE, et al. 1992. Exposure to tetrachlorodibenzo-p-dioxin (TCDD) 

alters fetal thymocyte maturation. Toxicol Appl Pharmacol 112:207-213. 

*Bobbie BA, Clement RE, Taguchi VY. 1989. Determination of chlorinated dibenzo-p-dioxins (CDD) and 

dibenzofurans (CDF) in environmental samples by tandem mass spectrometry. Chemosphere 18:155-162. 

*Bobet E, Berard MF, Dann T. 1990. The measurement of PCDD and PCDF in ambient air in southwestern 

Ontario. Chemosphere 20(10-12):1439-1445. 

*Bock KW. 1991. Roles of UDP-glucuronyltransferases in chemical carcinogenesis. Crit Rev Biochem 

Mol Biol 26:129-150. 

Bogen G. 1979. Symptoms of Vietnam veterans exposed to Agent Orange. JAMA 242:2391. 

*Bombick DW, Jankun J, Tullis K, et al. 1988. 2,3,7,8-Tetrachlorodibenzo-p-dioxin causes increases in 

expression of c-erb-A and levels of protein-tyrosine kinases in selected tissues of responsive mouse strains. 

Proc Natl Acad Sci 85:4128-4132. 

Bombick DW, Matsumura F. 1987. TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin) causes an increase in 

protein tryosine kinase activities at an early stage of poisoning in vivo in rat hepatocyte membranes. Life Sci 

41:429-436. 

Bonaccorsi A, Fanelli R, Tognoni G. 1978. In the wake of Seveso. Ambio 7:234-239. 

*Bond GG, McLaren EA, Brenner FE, et al. 1989a. Incidence of chloracne among chemical workers 

potentially exposed to chlorinated dioxins. J Occup Med 31:771-774. 
CDDs 554 

8. REFERENCES 

*Bond GG, McLaren EA, Lipps TE, et al. 1989b. Update of mortality among chemical workers with 

potential exposure to the higher chlorinated dioxins. J Occup Med 31:121-123. 

*Bond GG, Ott MG, Brenner FE, et al. 1983. Medical and morbidity surveillance findings among 

employees potentially exposed to TCDD. Br J Ind Med 40:318-324. 

Bookstaff RC, Kamel F, Moore RW, et al. 1990b. Altered regulation of pituitary gonadotropin-releasing 

hormone (GnRH) receptor number and pituitary responsiveness to GnRH in 2,3,7,8-tetrachlorodibenzo-pdioxin-treated male rats. Toxicol Appl Pharmacol 105:78-92. 

*Bookstaff RC, Moore RW, Peterson RE. 1990a. 2,3,7,8-Tetrachlorodibenzo-p-dioxin increases the 

potency of androgens and estrogens as feedback inhibitors of luteinizing hormone secretion in male rats. 

Toxicol Appl Pharmacol 104:212-224. 

*Bopp RF, Gross ML, Tong H, et al. 1991. A major incident of dioxin contamination: Sediments of New 

Jersey Estuaries. Environ Sci Technol 25:951-956. 

Botre C, Memoli A, Alhaigue F. 1978. TCDD solubilization and photodecomposition in aqueous solutions. 

Environ Sci Tech 124:335. 

Boule P, Hutzinger O. 1987. Photoreactivity of dibenzo-para-dioxins - can chlorodibenzo-para-dioxins be 

nitrated in the environment. Toxicol Environ Chem 13:229-256. 

*Bovey RW. 1980. Uses of phenoxy herbicides and their methods of application. In: Bovey RW, Young 

AL, eds. The science of 2,4,5-T and associated phenoxy herbicides. New York, NY: John Wiley and Sons, 

Inc., 49-69. 

Bowes GW, Simoneit BR, Burlingame AL, et al. 1973. The search for chlorinated dibenzofurans and 

chlorinated dibenzodioxins in wildlife populations showing elevated levels of embryonic death. Environ 

Health Perspect 5:191-198. 

*Bowman RE, Schantz SL, Gross ML, et al. 1989a. Behavioral effects in monkeys exposed to 

2,3,7,8-TCDD transmitted maternally during gestation and for four months of nursing. Chemosphere 

18:235-242. 

*Bowman RE, Schantz SL, Weerasinghe NCA, et al. 1989b. Chronic dietary intake of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) at 5 or 25 parts per trillion in the monkey: TCDD kinetics and 

dose-effect estimate of reproductive toxicity. Chemosphere 18:243-252. 

*Bowman RE, Tong HY, Gross ML, et al. 1990. Controlled exposure of female Rhesus monkeys to 

2,3,7,8,-TCDD: concentrations of TCDD in fat of offspring, and its decline over time. Chemosphere 

20:1199-1202. 

*Boyd JA, Clark GC, Walmer DK, et al. 1995. Endometriosis and the environment biomarkers of toxin 

exposure. Abstract of Endometriosis 2000 Workshop, May 15-17. 

Bradlaw JA, Casterline JL Jr. 1979. Induction of enzyme activity in cell culture: A rapid screen for 

detection of planar polychlorinated organic compounds. J Assoc Off Anal Chem 62:904-916. 
CDDs 555 

8. REFERENCES 

Bradlaw JA, Garthoff LH, Hurley NE, et al. 1980. Comparative induction of aryl hydrocarbon hydroxylase 

activity in vitro by analogs of dibenzo-p-dioxin. Food Cosmet Toxicol 18:627-635. 

*Branson D, Takahashi T, Parker W, et al. 1985. Bioconcentration kinetics of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in rainbow trout. Environ Toxicol Chem 4:779-788. 

Brenner KS. 1986. Pu-foam-plug technique and extractive co-distillation (Bleidner apparatus), versatile 

tools for stack emission sampling and sample preparation. Chemosphere 15:1917-1922. 

*Brewster DW, Matsumura F. 1988. Reduction of adipose tissue lipoprotein lipase activity as a result of in 

vivo administration of 2,3,7,8-tetrachlorodibenzo-p-dioxin to the guinea pig. Biochem Pharmacol 

37:2247-2253. 

*Brewster DW, Matsumura F, Akera T. 1987. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on guinea pig 

heart muscle. Toxicol Appl Pharmacol 89:408-417. 

Bridle TR, Afghan BK, Wilkinson RJ, et al. 1984. The formation and fate of PCDDs and PCDFs during 

chlorophenol combustion. Proc - APCA Annu Meet 77:1-15. 

Broman D, Naef C, Zebuehr Y, et al. 1989. The composition, distribution, and flux of PCDDs and PCDFs 

in settling particulate matter (SPM): A sediment trap study in the northern Baltic. Chemosphere 

19:445-450. 

*Bronzetti G, Bauer C, Corsi C, et al. 1983. Mutagenicity study of TCDD and ashes from urban incinerator 

in vitro and in vivo using yeast D7 strain. Chemosphere 12:549-553. 

Brooks GT. 1980. Perspectives on the chemical fate and toxicity of pesticides. J Environ Sci Health B 

15:755-793. 

*Brown NM, Lamartiniere CA. 1995. Xenoestrogens alter mammary gland differentiation and cell 

proliferation in the rat. Environ Health Perspect 103:708-713. 

*Brown NM, Manzolillo PA, Zhang J-X, et al. 1998. Prenatal TCDD and predisposition to mammary 

cancer in the rat. Carcinogenesis 19:1623-1629. 

*Brown RP, Cooper KR, Cristini A. 1994. Polychlorinated dibenzo-p-dioxins and dibenzofurans in Mya 

Arenaria in the Newark/Raritan Bay Estuary. Environ Toxicol Chem 13(3):523-528. 

Brown RS, Pettit K, Jones PW. 1989. A GC/GC-MS study of 2,3,7,8-TCDD using a DB-225 column and 

selected decomposition monitoring a quantitative tool. Chemosphere 19:171-176. 

Brownlee LJ, Evans CH, Hollebone BR. 1986. The relative induction of mixed-function oxidase specific 

activity to carbon-hydrogen and carbon-carbon 1 bond strengths in polychlorinated derivatives of 

dibenzo-p-dioxin. J Appl Toxicol 6:67-72. 

*Brugge JS, Yonemoto W, Darrow D. 1983. Interaction between Rous Sarcoma virus transforming protein 

and two cellular phosphoproteins: Analysis of the turnover and distribution of this complex. Mol Cell Biol 

3: 9-19. 
CDDs 556 

8. REFERENCES 

*Bruggeman WA, Opperhuizen A, Wijbenga A, et al. 1984. Bioaccumulation of super-lipophilic chemicals 

in fish. Toxicol Environ Chem 7:173-189. 

Bryant CJ, Hartle RW, Crandall MS, et al. 1989. Polychlorinated biphenyl, polychlorinated 

dibenzo-p-dioxin, and polychlorinated dibenzofuran contamination in PCB disposal facilities. Chemosphere 

18:569-576. 

Buchert H, Ballschmiter K. 1986. Polychlorodibenzofurans (PCDF) and dioxins (PCDD) as part of the 

general pollution in environmental samples of urban areas. Chemosphere 15:1923-1926. 

Buck M, Kirschmer P. 1988. Concentrations of polychlorinated dibenzo-p-dioxins and dibenzofurans in 

outdoor air of North-Rhine, Westphalia, Germany. Toxicol Environ Chem 16:173-202. 

*Bueno de Mesquita HB, Doornbos G, van der Kuip DAM, et al. 1993. Occupational exposure to phenoxy 

herbicides and chlorophenols and cancer mortality in the Netherlands. Am J Industr Med 23:289-300. 

*Bumb RR, Crummett WB, Cutie SS, et al. 1980. Trace chemistries of fire: A source of chlorinated 

dioxins. Science 210:385-390. 

*Bumpus JA, Tien M, Wright D, et al. 1985. Oxidation of persistent environmental pollutants by a white 

rot fungus. Science 228:1434-1436. 

Burbach KM, Poland A, Bradfield CA. 1992. Cloning of the Ah-receptor cDNA reveals a novel 

ligand-activated transcription factor. Proc Natl Acad Sci USA 89:8185-8189. 

Bureau of Water Protection. 1988. Final 880607 groundwater contaminant cleanup target concentrations, 

final 880606 table of chemical references (WQ). Bureau of Water Protection, 6/6/88. 

Burkhard LP, Kuehl DW. 1986. N-octanol/water partition coefficients by reverse phase liquid 

chromatography/mass spectrometry for eight tetrachlorinated planar molecules. Chemosphere 15:163-167. 

*Burleson GR, Lebrec H, Yang YG, et al. 1996. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on 

influenza virus host resistance in mice. Fund Appl Toxicol 29:40-47. 

*Burton JE, Michalek JE, Rahe AJ. 1998. Serum dioxin, chloracne, and acne in veterans of Operation 

Ranch Hand. Arch Environ Health 53:199-204. 

Buser HR. 1975. Polychlorinated dibenzo-p-dioxins: Separation and identification of isomers by gas 

chromatography-mass spectrometry. J Chromatogr 114:95-108. 

Buser HR. 1979. Formation and identification of tetra- and pentachlorodibenzo-p-dioxins from photolysis 

of two isomeric hexachlorodibenzo-p-dioxins. Chemosphere 4:251-257. 

*Buser HR. 1987. Brominated and brominated/chlorinated dibenzodioxins and dibenzofurans: Potential 

environmental contaminants. Chemosphere 16:713-732. 

Buser HR. 1988. Rapid photolytic decomposition of brominated and brominated/chlorinated 

dibenzodioxins and dibenzofurans. Chemosphere 17:889-903. 
 

CDDs 557 

8. REFERENCES 

Buser HR, Bosshardt HP. 1976. Determination of polychlorinated dibenzo-p-dioxins and dibenzofurans in 

commercial pentachlorophenols by combined gas chromatography-mass spectrometry. J Assoc Off Anal 

Chem 59:562-569. 

Buser HR, Rappe C. 1978. Identification of substitution patterns in polychlorinated dibenzo-p-dioxins 

(PCDDs) by mass spectrometry. Chemosphere 7:199-211. 

Buser HR, Rappe C. 1980. High-resolution gas chromatography of the 22 tetrachlorodibenzo-p-dioxin 

isomers. Anal Chem 52:2257-2262. 

Buser HR, Rappe C. 1984. Isomer-specific separation of 2,3,7,8-substituted polychlorinated 

dibenzo-p-dioxins by high-resolution gas chromatography/mass spectrometry. Anal Chem 56:442-448. 

*Buser HR, Rappe C, Bergqvist P. 1985. Analysis of polychlorinated dibenzofurans, dioxins, and related 

compounds in environmental samples. Environ Health Perspect 60:293-302. 

Buu-Hoi NP, Pham-Huu-Chanh, Sesque G, et al. 1972. [Organs as targets of dioxin (2,3,7,8-tetrachlorodibenzo-p-dioxin) intoxication.] Naturwissenschaften 59:174-175. (German) 

*Bysshe SE. 1990. Bioconcentration factors in aquatic organisms. In: Lyman WJ, Reehl WF, Rosenblatt 

DF, eds. Handbook of chemical properties estimation methods. Environmental behavior of organic 

compounds. Washington, DC: American Chemical Society, 5-1 to 5-30. 

*Cai Z, Giblin DE, Ramanujam VMS, et al. 1994. Mass profile monitoring in trace analysis: Identification 

of polychlorodibenzothiophenes in crab tissue collected from the Newark/Raritan Bay system. Environ Sci 

Technol 28(8):1535-1538. 

*Calvert GM, Hornung RW, Sweeney MH, et al. 1992. Hepatic and gastrointestinal effects in an 

occupational cohort exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. J Am Med Assoc 267:2209-2214. 

*Calvert GM, Sweeney MH, Morris JA, et al. 1991. Evaluation of chronic bronchitis, chronic obstructive 

pulmonary disease, and ventilatory function among workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Am Rev Resp Dis 144:1302-1306. 

*Calvert GM, Wall DK, Sweeney MH, et al. 1998. Evaluation of cardiovascular outcomes among U.S. 

workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ Health Perspect 106 (Suppl 2):635-643. 

*Calvert GM, Wille KK, Sweeney MH, et al. 1996. Evaluation of serum lipid concentrations among U.S. 

workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Arch Environ Health 51:100-107. 

*Cambridge Isotope Laboratories. 1995. Environmental Contaminant Standards. Andover, Massachusetts. 

Campbell AD, Firestone D, Westly B. 1971. Chick edema factor--toxic dioxins. In: Westley B, ed. 

International Symposium on identification and measurement of environmental pollutants, Ottawa, Ontario, 

Canada, June 14-17, 1971. Ottawa, Canada: The Symposium, 105-108. 
CDDs 558 

8. REFERENCES 

*Canga L, Levi R, Rifkind AB. 1988. Heart as a target organ in 2,3,7,8-tetrachlorodibenzo-p-dioxin 

toxicity: Decreased β-adrenergic responsiveness and evidence of increased intracellular calcium. Proc Natl 

Acad Sci USA 85:905-909. 

*Cantoni L, Salmona M, Rizzardini M. 1981. Porphyrogenic effect of chronic treatment with 

2,3,7,8-tetrachlorodibenzo-p-dioxin in female rats. Dose-effect relationship following urinary excretion of 

porphyrins. Toxicol Appl Pharmacol 57:156-163 

Cantor DS, Holder G, Cantor W, et al. 1994. In utero and postnatal exposure to 2,3,7,8-TCDD in Times 

Beach, Missouri: 2. Impact on neurophysiological functioning. (unpublished literature) 

*Cantor KP. 1982. Farming and mortality from non-Hodgkin's lymphoma: A case-control study. Int J 

Cancer 29:239-247. 

*Caputo R, Monti M, Ermacora E,et al. 1988. Cutaneous mainfestations of tetrachlorodebenzo-p-dioxin in 

children and adolescents: Follow-up 10 years after the Seveso, Italy, accident. J Am Acad Dermatol 

19:812-819. 

*Caramaschi F, Del Corno G, Favaretti C, et al. 1981. Chloracne following environmental contamination 

by TCDD in Seveso, Italy. Int J Epidemiol 10:135-143. 

*Carrier G, Brunet RC, Brodeur J. 1995a. Modeling of the toxicokinetics of polychlorinated 

dibenzo-p-dioxins and dibenzofurans in mammalians, including humans. I. Nonlinear distribution of 

PCDD/PCDF body burden between liver and adipose tissues. Toxicol Appl Pharmacol 131:253-266. 

*Carrier G, Brunet RC, Brodeur J. 1995b. Modeling of the toxicokinetics of polychlorinated 

dibenzo-p-dioxins and dibenzofurans in mammalians, including humans. II. Kinetics of absorption and 

disposition of PCDDs/PCDFs. Toxicol Appl Pharmacol 131:267-276. 

Carsch S, Thoma H, Hutzinger O. 1986. Leaching of polychlorinated dibenzo-p-dioxins and 

polychlorinated dibenzofurans from municipal waste incinerator fly ash by water and organic solvents. 

Chemosphere 15:1927-1930. 

Carter CD, Kimbrough RD, Liddled JA, et al. 1975. Tetrachlorodibenzodioxin: An accidental poisoning 

episode in horse arenas. Science 188:738-740. 

*Carver LA, Hogenesch JB, Bradfield CA. 1994. Tissue specific expression of the rat Ah receptor and 

ARNT mRNAs. Nuc Acids Res 22:3038-3044. 

Cavallaro A, Bandi G, Invernizzi G, et al. 1982. Negative ion chemical ionization MS as a structure tool in 

the determination of small amounts of PCDD and PCDF. Pergamon Ser Environ Sci 5:55-65. 

CDC. 1986. Centers for Disease Control statement on breastfeeding. Centers for Disease Control. Vet 

Hum Toxicol 28:582-583. 

*CDC. 1987. Comparison of serum levels of 2,3,7,8-TCDD with indirect estimates of Agent Orange 

exposure in Vietnam veterans: Final Report Agent Orange Projects. Atlanta, Ga: Centers for Disease 

Control, Public Health Service, U.S. Department of Health and Human Services. 
CDDs 559 

8. REFERENCES 

*CDC. 1988. Serum 2,3,7,8-tetrachlorodibenzo-p-dioxin levels in US Army Vietnam-era veterans. Centers 

for Disease Control. JAMA 260:1249-1254. 

*CELDS. 1994. Computer-environmental legislative data systems database. University of Illinois, Urbana, 

IL. 

*Cerlisi S, DiDomenico A, Ratti S. 1989. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) persistence in the 

Seveso (Milan, Italy) soil. Toxicol Environ Safe 18:149-164. 

*Chaffin CL, Peterson RE, Hutz RJ. 1996. In utero and lactational exposure of female Holtzman rats to 

2,3,7,8-tetrachlorodibenzo-p-dioxin: Modulation of the estrogen signal. Biol Reprod 55:62-67. 

*Chaffin CL, Trewin AL, Watanabe G, et al. 1997. Alterations to the pituitary-gonadal axis in the 

peripubertal female rat exposed in utero and through lactation to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Biol 

Reprod 56:1498-1502. 

*Chahoud I, Krowke R, Schimmel A. 1989. Reproductive toxicity and pharmacokinetics of 

2,3,7,8-tetrachlorodibenzo-p-dioxin: 1. Effects of high doses on the fertility of male rats. Arch Toxicol 

63:432-439. 

*Chang RR, Jarman WM, Hennings JA. 1993. Sample cleanup by solid phase extraction for the ultratrace 

determination of polychlorinated dibenzo-p-dioxins and dibenzofurans in biological samples. Anal Chem 

65(18):2420-2427. 

Chang RR, Jarman WM, King CC, et al. 1990. Bioaccumulation of PCDDs and PCDFs in food animals: 

III. A rapid cleanup of biological materials using reverse phase adsorbent columns. Chemosphere 

20:881-886. 

*Chemical Regulation Reporter. 1997. (7/25/97), page 455. 

*Chemical Regulation Reporter. 1997. (8/22/97), page 533. 

*Chen H-S, Perdew G. 1994. Subunit composition of the heteromeric cytosolic aryl hydrocarbon receptor 

complex. J Biol Chem 269: 27554-27558. 

Chittim BG, Bunce NJ, HU K, et al. 1994. Comparison of GC-MS with an in vitro bioassy for PCDDs and 

related compounds in environmental samples. Chemospere 29(9-11):1783-1788. 

Chiu C, Halman R, Lanoy M, et al. 1987. Refined analytical procedures to detect and measure 

PCDD/PCDF, PAH, PCB, CB, and CP in environmental samples. Chemosphere 16:1619-1624. 

*Chiu C, Thomas RS, Lockwood J, et al. 1983. Polychlorinated hydrocarbons from power plants, wood 

burning, and municipal incinerators. Chemosphere 12:607-616. 

*Choi EJ, Toscano DJ, Ryan JA, et al. 1991. Dioxin induces transforming growth factor-alpha in human 

keratinocytes. J Biol Chem 266:9591-5997. 

*Choudhry GG, Hutzinger O. 1982. Photochemical formation and degradation of polychlorinated 

dibenzofurans and dibenzo-p-dioxins. Residue Reviews 84:113-161. 
CDDs 560 

8. REFERENCES 

Choudhry GG, Webster GRB. 1986. Photochemical quantum yields and sunlight half-lives of 

polychlorodibenzo-p-dioxins in aquatic systems. Chemosphere 15:1935-1940. 

*Choudhry GG, Webster GRB. 1987. Quantum yields of polychlorodibenzo-p-dioxins in water-acetonitrile 

mixtures and their environmental phototransformation rates. In: Zika RG, Cooper WJ, eds. Photochemistry 

of environmental aquatic systems. Washington, DC: American Chemical Society, 327:61-74. 

*Choudhry GG, Webster GRB. 1989. Environmental photochemistry of PCDDs: 2. Quantum yields of the 

direct phototransformation of 1,2,3,7-tetra-, 1,3,6,8-tetra-, 1,2,3,4,6,7,8-hepta-, and 

1,2,3,4,6,7,8,9-octachlorodibenzo-p-dioxin in aqueous acetonitrile and their sunlight half-lives. J Agric 

Food Chem 37:254-261. 

*Christian BJ, Inhorn SL, Peterson RE. 1986a. Relationship of the wasting syndrome to lethality in rats 

treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 82:239-255. 

*Christian BJ, Menahan LA, Peterson RE. 1986b. Intermediary metabolism of the mature rat following 

2,3,7,8-tetrachlorodibenzo-p-dioxin treatment. Toxicol Appl Pharmacol 83:360-378. 

Christmann W, Kloeppel KD, Partscht H, et al. 1989a. Determination of PCDD/PCDF in ambient air. 

Chemosphere 19:521-526. 

*Christmann W, Kloeppel KD, Partscht H, et al. 1989b. PCDD/PCDF and chlorinated phenols in wood 

preserving formulations for household use. Chemosphere 18:861-865. 

Christmann W, Kloeppel KD, Partscht H, et al. 1989c. Tetrachlorobenzoquinones, a source of 

PCDD/PCDF. Chemosphere 18:789-792. 

*Cirnies-Ross C, Stanmore B, Millar G. 1996. Dioxins in diesel exhaust. Nature 381:379. 

*Clark DA, Sweeney G, Safe S, et al. 1983. Cellular and genetic basis for suppression of cytotoxic T cell 

generation by haloaromatic hydrocarbons. Immunopharmacology 6:143-153. 

*Clark GC, Blank JA, Germolec DR, et al. 1991a. 2,3,7,8-Tetrachlorodibenzo-p-dioxin stimulation of 

tyrosine phosphorylation in B lymphocytes: potential role in immunosuppression. Mol Pharmacol 

39:495-501. 

*Clark GC, Taylor MJ, Tritscher AM, et al. 1991b. Tumor necrosis factor involvement in 

2,3,7,8-tetrachlorodibenzo-p-dioxin-mediated endotoxin hypersensitivity in C57BL/6 mice congenic at the 

Ah locus. Toxicol Appl Pharmacol 111:422-431. 

Clement RE. 1988. The analysis and environmental distribution of chlorinated dibenzo-p-dioxins and 

dibenzofurans [Abstract]. Abstracts of Papers of the American Chemical Society 195:199. 

Clement RE, Bolelie B, Taguchi V. 1986. Comparison of instrumental methods for chlorinated 

dibenzo-p-dioxin (CDD) determination: Interim results of a round-robin study involving GC/MS, MS/MS 

and high resolution MS. Chemosphere 15:1147-1156. 
CDDs 561 

8. REFERENCES 

Clement RE, Suter SA, Reiner E, et al. 1989a. Concentrations of chlorinated dibenzo-p-dioxins and 

dibenzofurans in effluents and centrifuged particles from Ontario pulp and paper mills. Chemosphere 

19:649-954. 

Clement RE, Tashiro C, Reiner E, et al. 1987a. Chlorinated dibenzofurans (CDFs) in effluents and sludges 

from pulp and paper mills. In: Proceedings of the Seventh International Symposium on Chlorinated Dioxins 

and Related Compounds. Las Vegas, NV: University of Nevada. 

*Clement RE, Tashiro C, Suter S, et al. 1989b. Chlorinated dibenzo-p-dioxins (CDDs) and dibenzofurans 

(CDFs) in effluents and sludges from pulp and paper mills. Chemosphere 18:1189-1197. 

Clement RE, Tosine HM. 1988. The gas-chromatography mass-spectrometry determination of 

chlorodibenzo-p-dioxins and dibenzofurans. Mass Spectrometry Reviews 7:593-636. 

*Clement RE, Tosine HM, Ali B. 1985. Levels of polychlorinated dibenzo-p-dioxin and dibenzofuran in 

woodburning stoves, fireplaces. Chemosphere 14:815. 

Clement RE, Tosine HM, Osborne J, et al. 1988. Gas chromatographic/mass spectrometric determination of 

chlorinated dibenzo-p-dioxin and dibenzofurans in incinerator stack emissions and fly-ash: A 13-test study. 

Biomed Environ Mass Spectrom 17:81-96. 

Clement RE, Tosine HM, Taguchi V, et al. 1987b. Investigation of American lobster, HomarusAmericanus, for the presence of chlorinated dibenzo-para-dioxins and dibenzofurans. Bull Environ Contam 

Toxicol 39:1069-1075. 

*Clewell HJ III, Andersen M. 1985. Risk assessment extrapolations and physiological modeling. Toxicol 

Ind Health 1(4):111-131. 

*Clewell HJ III, Andersen ME. 1985. Risk assessment extrapolations and physiological modeling. Toxicol 

Ind Health 1:111-113. 

*Cohen GM, Bracken WM, Iyer RP, et al. 1979. Anticarcinogenic effects of 2,3,7,8-tetrachlorodibenzop-dioxin on benzo(a)pyrene and 7,12-dimethylbenz(a)anthracene tumor initiation and its relationship to 

DNA binding. Cancer Res 39:4027-4033. 

Cole RH, Frederick RE, Healy RP, et al. 1984. Preliminary findings of the priority pollutant monitoring 

project of the nationwide urban runoff program. J Water Pollut Control Fed 56:898-908. 

*Conacher HBS, Page BD, Ryan JJ. 1993. Industrial chemical contamination of foods. Food Additives and 

Contaminants 10(1):129-143. 

*Connett P, Webster T. 1987. An estimation of the relative human exposure to 2,3,7,8-TCDD emissions via 

inhalation and ingestion of cow's milk. Chemosphere 16:2079-2084. 

Conso F, Dally S, Cordier S, et al. 1986. Accidental human exposure to polychlorinated dibenzofurans and 

polychlorinated dioxins following a transformer fire. Chemosphere 15:1953-1955. 

*Cook JC, Dold KM, Greenlee WF. 1987a. An in vitro model for studying the toxicity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin to human thymus. Toxicol Appl Pharmacol 89:256-268. 
CDDs 562 

8. REFERENCES 

*Cook JC, Greenlee WF. 1989. Characterization of a specific binding protein for 

2,3,7,8-tetrachlorodibenzo-p-dioxin in human thymic epithelial cells. Mol Pharmacol 35:713-719. 

*Cook PM, Kuehl DW ,Walker MK, et al. 1991. Bioaccumulation and toxicity of TCDD and related 

compounds in aquatic ecosystems. In: Gallo MA, Scheuplein JR, Van Der Heijden KA, eds. Banbury 

Report 35: Biological basis for risk assessment of dioxins and related compounds. Plainview, NY: Banbury 

Center Meeting Cold Spring Harbor Laboratory Press. 

*Cook RR. 1981. Dioxin, chloracne and soft-tissue sarcoma. Lancet 1:681-619. 

*Cook RR, Bond GG, Olson RA, et al. 1986. Evaluation of the mortality experience of workers exposed to 

the chlorinated dioxins. Chemosphere 15:1769-1776. 

*Cook RR, Bond GG, Olson RA, et al. 1987b. Update of the mortality experience of workers exposed to 

chlorinated dioxins. Chemosphere 16:2111-2116. 

*Cooke M, Hatchel JA, Czuczwa JM, et al. 1988. Analytical and sampling techniques for the determination 

of ultratrace quantities of polychlorinated dibenzofurans and dibenzodioxins. In: Proceedings 1987 EPRI 

PCB Seminar, Kansas City, MO, October 6-9, 1987. Palo Alto, CA: Electric Power Research Institute, 7-1 

-7-8. 

*Corbet RL, Muir D CG, Webster G RB. 1983. Fate of carbon-14 labelled 1,3,6,8-tetrachlorodibenzo-p-dioxin in an outdoor aquatic system. Chemosphere 12 (4-5):523-528. 

Corbet RL, Worobey BL, Webster GRB. 1981. Preliminary analysis and confirmation of the presence of 

polychlorinated dibenzo-p-dioxin (PCDD) and polychlorinated biphenyls (PCB) in lake sediment. In: Muir 

DC, ed. Sixteenth Annual Workshop on pesticide residue analysis and ninth annual workshop on the 

chemistry and biochemistry of herbicides, Winnepeg, Manitoba, Canada: Department of National Health 

and Welfare, Health Protection Branch, 111-116. 

*Coumailleau P, Poellinger L, Gustafsson J-A, et al. 1995. Definition of a minimal domain of the dioxin 

receptor that is associated with hsp90 and maintains wild type ligand binding affinity and specificity. J Biol 

Chem 270: 25291-25300. 

*Courtney KD. 1976. Mouse teratology studies with chlorodibenzo-p-dioxins. Bull Environ Contam 

Toxicol 16:674-681. 

*Courtney KD, Putnam JP, Andrews JE. 1978. Metabolic studies with TCDD (dioxin) treated rats. Arch 

Environ Contam Toxicol 7:385-396. 

*Couture LA, Abbott BD, Birnbaum LS. 1990. A critical review of the developmental toxicity and 

teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin: Recent advances toward understanding the 

mechanism. Teratol 42:619-627. 

*Couture LA, Elwell MR, Birnbaum LS. 1988. Dioxin-like effects observed in male rats following 

exposure to octachlorodibenzo-p-dioxin (OCDD) during a 13-week study. Toxicol Appl Pharmacol 

93:31-46. 
CDDs 563 

8. REFERENCES 

Couture-Haws L, Harris MW, McDonald MM, et al. 1991a. Evaluation of persistence of hydronephrosis 

induced in mice following in utero and/or lactational exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Toxicol Appl Pharmacol 107:402-412. 

*Couture-Haws L, Harris MW, McDonald MM, et al. 1991b. Hydronephrosis in mice exposed to 

TCDD-contaminated breast milk: Identification of the peak period of sensitivity and assessment of potential 

recovery. Toxicol Appl Pharmacol 107:413-428. 

*Cramer G, Bolder M, Henry S, et al. 1991. USFDA assessment of exposure to 2,3,7,8-TCDD and 

2,3,7,8-TCDF from foods contacting bleached paper products. Chemosphere 23(8-10):1537-1550. 

Cramer PH, Ayling RE, Thornburg KR, et al. 1990. Evaluation of an analytical method for the 

determination of polybrominated dibenzo-p-dioxins/dibenzofurans (PBDD/PBDF) in animal adipose. 

Chemosphere 20:821-828. 

*Creaser CS, Al-Haddad A. 1989. Fractionation of polychlorinated biphenyls, polychlorinated 

dibenzo-p-dioxins, and polychlorinated dibenzofurans on porous graphitic carbon. Anal Chem 

61:1300-1302. 

*CRITFC. 1994. A fish consumption survey of the Umatilla, Nez Perce, Yakama, and Warm Springs tribes 

of the Columbia River basin. Columbia River Inter-Tribal Fish Commission. CRITFC technical report no. 

94-3. 

*Crosby DG, Moilanen KW, Wong AS. 1973. Environmental generation and degradation of 

dibenzodioxins and dibenzofurans. Environ Health Perspect 5:259-266. 

*Crosby DG, Wong AS. 1976. Photochemical generation of chlorinated dioxins. Chemosphere 5:327-332. 

*Crow KD. 1978. Chloracne - an up to date assessment. Ann Occup Hyg 21:297-298. 

Crummett WB. 1982. Environmental chlorinated dioxins from combustion - the trace chemistries of fire 

hypothesis. Pergamon Ser Environ Sci 5:253-263. 

*Crunkilton RL, Smith LM, Petty JD, et al. 1987. Residues of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the 

Spring River, Missouri. Water Air Soil Pollut 32:219-231. 

Cuiu C, Halman R, Li K, et al. 1986. Analytical procedures to assay environmental samples for 

PCDD/PCDF, PAH, PCB, chlorobenzenes (CB) and chlorophenols (CP). Chemosphere 15:1091-1098. 

Cull MR, Dobbs AJ, Goudot M, et al. 1984. Polychlorodibenzo-p-dioxins and dibenzofurans in technical 

pentachlorophenol - results of a collaborative analytical exercise. Chemosphere 13:1157-1165. 

Cull MR, Dobbs AJ, Williams N. 1983. Polychlorinated dibenzo-p-dioxins (PCDDs) in commercial 

pentachlorophenol (PCP) used in wood preservation. Chemosphere 12:483-485. 

*Cummings AM, Metcalf JL, Birnbaum L. 1996. Promotion of endometriosis by 2,3,7,8-tetrachlorodibenzo-p-dioxin in rats and mice: Time-dose dependence and species comparison. Toxicol Appl Pharmacol 

138:131-139. 
CDDs 564 

8. REFERENCES 

*Curtis LR, Kerkvliet NI, Baecher-Steppan L, et al. 1990. 2,3,7,8-Tetrachlorodibenzo-p-dioxin 

pretreatment of female mice altered tissue distribution but not hepatic metabolism of a subsequent dose. 

Fundam Appl Toxicol 14:523-531. 

Czeizel E, Kiraly J. 1976. Chromosome examinations in workers producing Klorinal and Buminol. In: 

Banki L, ed. The development of a pesticide as a complex scientific task. Medicina, Budapest: 239-256. 

*Czuczwa J, Katona V, Pitts G, et al. 1989. Analysis of fog samples for PCDD and PCDF. Chemosphere 

18:847-850. 

*Czuczwa JM, Hites RA. 1984. Environmental fate of combusion-generated polychlorinated dioxins and 

furans. Environ Sci Technol 18:444-450. 

*Czuczwa JM, Hites RA. 1986. Airborne dioxins and dibenzofurans: Sources and fates. Environmental 

Science and Technology 20:195-200. 

*Czuczwa JM, Hites RA. 1986a. Airborne dioxins and dibenzofurans: Sources and fates. Environ Sci 

Technol 20:195-200. 

*Czuczwa JM, Hites RA. 1986b. Sources and fates of PCDD and PCDF. Chemosphere 15:1417-1420. 

*Dahl P, Lindstrom G, Wiberg K, et al. 1995. Absorption of polychlorinated biphenyls, dibenzo-p-dioxins 

and dibenzofurans by breast-fed infants. Chemosphere 30:2297-2306. 

*Dasenbrock C, Bittmann H, Creutzenberg O, et al. 1992. Cleft palate and liver weight data of two different 

mixtures of PCDDs in pregnant mice related to organ dosage. Chemosphere 25:1153-1157. 

*Davidson MD and Fujimoto JM. 1987. Increased permeability of the rat biliary tree by 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) treatment and protection by hepatoactive agents. Toxicol Appl 

Pharmacol 87:57-66. 

Davies K. 1988. Concentrations and dietary intake of selected organochlorines, including PCB, PCDDs, 

and PCDFs in fresh food composites grown in Ontario, Canada. Chemosphere 17:263-276. 

*Davis D, Safe S. 1989. Dose-response immunotoxicities of commercial polychlorinated biphenyls (PCBs) 

and their interaction with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Lett 48:35-43. 

*Davis D, Safe S. 1991. Halogenated aryl hydrocarbon-induced suppression of the in vitro plaque-forming 

cell response to sheep red blood cells is not dependent on the Ah receptor. Immunopharmacology 

21:183-190. 

*De Heer C, Schuuman H-J, Vos JG, et al. 1994b. Lymphodepletion of the thymus cortex in rats after a 

single oral intubation of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Chemosphere 29:2295-2299. 

*De Heer C, Schuurman H-J, Liem AKD, et al. 1995. Toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) to the human thymus after implantation in SCID mice. Toxicol Appl Pharmacol 134:296-304. 

*De Heer C, Verlaan APJ, Penninks AH, et al. 1994a. Time course of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD)-induced thymic atrophy in the Wistar rat. Toxicol Appl Pharmacol 128:97-104. 
CDDs 565 

8. REFERENCES 

De Jong APJ, Liem AKD, Den Boer AC, et al. 1989. Analysis of polychlorinated dibenzofurans and 

dibenzo-p-dioxins in human milk by tandem hybrid mass spectrometry. Chemosphere 19:59-66. 

*De Krey GK, Kerkvliet NI. 1995. Suppression of cytotoxic T lymphocyte activity by 2,3,7,8-tetrachlorodibenzo-p-dioxin occurs in vivo, but not in vitro, and is independent of corticosterone elevation. Toxicology 

97:105-112. 

*De Vault D, Dunn W, Bergqvist P, et al. 1989. Polychlorinated dibenzofurans and polychlorinated 

dibenzo-p-dioxins in Great Lakes fish: A baseline and interlake comparison. Environ Toxicol Chem 

8:1013-1022. 

*De Wall EJ, Schuurman H-J, Loeber JG, et al. 1992. Alterations in the cortical thymic epithelium of rats 

after in vivo exposure to 2,3,7,8-tetrachlorodibenxo-p-dioxin (TCDD): An (immuno)histological study. 

Toxicol Appl Pharmacol 115(1):80-88. 

*DeCaprio AP, McMartin DM, O'Keefe PW, et al. 1986. Subchronic oral toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the guinea pig: Comparisons with a PCB-containing transformer fluid pyrolysate. 

Fundam Appl Toxicol 6:454-463. 

*Della Porta G, Dragani TA, Sozzi G. 1987. Carcinogenic effects of infantile and long-term 

2,3,7,8-tetrachlorodibenzo-p-dioxin treatment in the mouse. Tumori 73:99-107. 

*Dencker L, Hassoun E, D'Argy R, et al. 1985. Fetal thymus organ culture as an in vitro model for the 

toxicity of 2,3,7,8-TCDD and its congeners. Mol Pharmacol 27:133-140. 

Dencker L, Pratt RM. 1981. Association between the presence of the Ah receptor in embryonic murine 

tissues and sensitivity to TCDD-induced cleft palate. Teratogenesis Carcinog Mutagen 1:399-406. 

Denison MS. 1991. Inactivation and dissociation of rat hepatic Ah receptor: evidence for distinct receptor 

forms? [Abstract 989]. The Toxicologist 11:260. 

*Denison MS, Fisher JM, Whitlock JP Jr. 1989. Protein-DNA interactions at recognition sites for the 

dioxin-Ah receptor complex. J Biol Chem 264:16478-16482. 

Denison MS, Vella LM. 1990. The hepatic Ah receptor for 2,3,7,8-tetrachlorodibenzo-p-dioxin: Species 

differences in subunit dissociation. Arch Biochem Biophys 277:382-388. 

Denison MS, Vella LM, Okey AB. 1986b. Structure and function of the Ah receptor for 

2,3,7,8-tetrachlorodibenzo-p-dioxin: species difference in molecular properties of the receptors from mouse 

and rat hepatic cytosols. J Biol Chem 61:3987-3995. 

*Denison MS, Wilkinson CF, Okey AB. 1986a. Ah receptor for 2,3,7,8-tetrachlorodibenzo-p-dioxin: 

Comparative studies in mammalian and nonmammalian species. Chemosphere 15:1665-1672. 

*Denomme MA, Homonoko K, Fujita T, et al. 1985. Effects of substituents on the cytosolic 

receptor-binding affinities of aryl hydrocarbon hydroxylase induction potencies of 7-substituted 

2,3-dichlorodibenzo-p-dioxins. Mol Pharmacol 27:656-661. 
CDDs 566 

8. REFERENCES 

Department of Health, New Zealand. 1980. Report to the Minister of Health of an investigation into 

allegations of an association between human congenital defects and 2,4,5-T spraying in and around teKuiti. 

New Zealand Med J 314-315. 

DeRoos FL, Bicking MKL. 1990. Supercritical fluid extraction for the determination of PCDDs and PCDFs 

in soil. Chemosphere 20:1355-1362. 

*Des Rosiers PE. 1986. Methodologies for materials contaminated with PCDDs and related compounds. 

Chemosphere 15:1513-1528. 

*Des Rosiers PE. 1987. National dioxin study. ACS Sympos Ser 338:34-53. 

Des Rosiers PE, Lee A. 1986. PCB fires: Correlation of chlorobenzene isomer and PCB homolog contents 

of PCB fluids with PCDD and PCDF contents of soot. Chemosphere 15:1313-1323. 

Deutsch DG, Goldfarb TD. 1988. PCDD/PCDF contamination following a plastics fire in a university 

lecture hall building. Chemosphere 17:2423-2431. 

*DeVito M, Thomas T, Meeker R, et al. 1991. The effects of TCDD on estrogen receptor levels in intact, 

ovariectomized and hypophysectomized mice [Abstract 995]. The Toxicologist 11:262. 

*DeVito MJ, Birnbaum LS. 1995. The importance of pharmacokinetics in determining the relative potency 

of 2,3,7,8-tetrachlorodibenzo-p-dioxin and 2,3,7,8-tetrachlorodibenzofuran. Fund Appl Toxicol 24:145-148. 

*DeVito MJ, Birnbaum LS, Farland WH, et al. 1995. Comparisons of estimated human body burdens of 

dioxin-like chemicals and TCDD body burdens in experimentally exposed animals. Environ Health Perspect 

103:820-831. 

*DeVito MJ, Diliberto JJ, Ross DG, et al. 1997. Dose-response relationships for polyhalogenated dioxins 

and dibenzofurans following subchronic treatment in mice. I. CYP1A1 and CYP1A2 enzyme activity in 

liver, lung, and skin. Toxicol Appl Pharmacol 147:267-280. 

*DeVito MJ, Ma X, Babish JG, et al. 1994. Dose-response relationships in mice following subchronic 

exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin: CYP1A1, CYP1A2, estrogen receptor, and protein tyrosine 

phosphorylation. Toxicol Appl Pharmacol 124(1):82-90. 

DeVito MJ, Maier WE, Diliberto JJ, et al. 1993. Comparative ability of various PCBs, PCDFs, and TCDD 

to induced cytochrome P450 1A1 and 1A2 activity following 4 weeks of treatment. Fundam Appl Toxicol 

20(1):125-130. 

*DeVito MJ, Thomas T, Martin E, et al. 1992. Antiestrogenic action of 2,3,7,8-tetrachlorodibenzop-dioxin: Tissue specific regulation of estrogen receptor in CD1 mice. Toxicol Appl Pharmacol 

113(2):284-292. 

*Dewailly E, Bruneau S, Ayotte P, et al. 1993. Health status at birth of Inuit newborn prenatally exposed to 

organochlorines. Chemosphere 27:359-366. 
CDDs 567 

8. REFERENCES 

*Dewailly E, Nantel A , Bruneau S, et al. 1992. Breast milk contamination of PCDDs, PCDFs and PCBs in 

Arctic Quebec: A preliminary assessment. Chemosphere 25(7-10):1245-1249. 

*Dewailly E, Ryan J, Laliberte C, et al. 1994. Exposure of remote maritime populations to coplanar PCBs. 

Environ Health Perspectives S102(1):205-209. 

*Dewailly E, Weber J-P, Gingras S, et al. 1991. Coplanar PCBs in human milk in the Province of Quebec, 

Canada: Are they more toxic than dioxin for breast fed infants? Bull Environ Contam Toxicol 47:491-498. 

*DiBartolomeis MJ, Moore RW, Peterson RE, et al. 1986. Hypercholesterolemia and the regulation of 

adrenal steroidogenesis in 2,3,7,8-tetrachlorodibenzo-p-dioxin-treated rats. Toxicol Appl Pharmacol 

85:313-323. 

*DiBartolomeis MJ, Moore RW, Peterson RE, et al. 1987. Altered regulation of adrenal steroidogenesis in 

2,3,7,8-tetrachlorodibenzo-p-dioxin-treated rats. Biochem Pharmacol 36:59-67. 

Dickson LC, Hutzinger M, Karasek F, et al. 1986. Determination of desorption losses of polychlorinated 

dibenzo-p-dioxins from municipal incinerator fly ash by gas chromatography/mass spectrometry. Int J 

Environ Anal Chem 2:55-74. 

Dickson LC, Karasek FW. 1987. Mechanism of formation of polychlorinated dibenzo-p-dioxins produced 

on municipal incinerator fly ash from reactions of chlorinated phenols. J Chromatogr 389:127-137. 

*DiGiovanni J, Viaje A, Berry DL, et al. 1977. Tumor-initiating ability of 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD) and Arochlor 1254 in the two-stage system of mouse skin carcinogenesis. Bull Environ 

Contam Toxicol 18:552-556. 

DiLernia R, Crimaudo C, Pacchetti G. 1982. The study of X-rays and TCDD effects on satellite 

associations may suggest a simple model for application in environmental mutagenesis. Hum Genet 

61:42-47. 

*Diliberto JJ, Akubue PI, Luebke RW, et al. 1995. Dose-response relationships of tissue distribution and 

induction of CYP1A1 and CYP1A2 enzymatic activities following acute exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in mice. Toxicol Appl Pharmacol 130:197-208. 

*Diliberto JJ, Burgin D, Birnbaum LS. 1997. Role of CYP1A2 in hepatic sequestration of dioxin: Studies 

using CYP1A2 knock-out mice. Biochem Biophys Res Commun 236:431-433. 

*Diliberto JJ, Jackson JA, Birnbaum LS. 1996. Comparison of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) 

disposition following pulmonary, oral, dermal and parenteral exposures to rats. Toxicol Appl Pharmacol. 

138:158-168. 

*Dimich-Ward H, Hertzman C, Teschke K, et al. 1996. Reproductive effects of paternal exposure to 

chlorophenate wood preservatives in the sawmill industry. Scand J Work Environ Health 22:267-273. 

Djien Liem AK, Van Zorge JA. 1995. Dioxins and related compounds: Status and regulatory aspects. 

Environmental Science and Pollution Research International 2(1):46-56. 
CDDs 568 

8. REFERENCES 

Dobbs AJ, Cull MR. 1982. Volatilization of chemicals - relative loss rates and the estimation of vapor 

pressures. Environ Pollut Ser [B] 3:289-298. 

*Dobbs AJ, Grant C. 1979. Photolysis of highly chlorinated dibenzo-p-dioxins by sunlight. Nature 

278:163-165. 

Dobbs AJ, Grant C. 1981. Octachlorodibenzo-p-dioxin in wood treatment materials and treated wood. 

Chemosphere 10:1185-1193. 

*Dolwick KM, Schmidt JV, Carver LA, et al. 1993. Cloning and expression of a human Ah receptor 

cDNA. Mol Pharmacol 44:911-917. 

*Donnelly JR, Vonnahme TL, Hedin CM, et al. 1986. Evaluation of RCRA method 8280 for analysis of 

dioxins and dibenzofurans. In: Rappe C, Choudhary G, Keith LH, eds. Chlorinated dioxins and 

dibenzofurans in perspective. Chelsea, MI: Lewis Publishers, Inc., 399-435. 

*Dooley RK, Morris DL, Holsapple MP. 1990. Elucidation of cellular targets responsible for 

tetrachlorodibenzo-p-dioxin (TCDD)-induced suppression of antibody responses: II. Role of the T 

lymphocyte. Immunopharmacology 19:47-58. 

DOT. 1989b. List of hazardous substance and reportable quantities. Department of Transportation. Code 

of Federal Regulations. 49 CFR 172.101. 

Doucette WJ, Andren AW. 1987. Correlation of octanol/water partition coefficients and total molecular 

surface area for highly hydrophobic aromatic compounds. Environ Sci Technol 21:821-824. 

*Doucette WJ, Andren AW. 1988. Aqueous solubility of selected biphenyl, furan and dioxin congeners. 

Chemosphere 17:243-252. 

*Dougherty EJ, McPeters AL, Overcash MR, et al. 1993. Theoretical analysis of a method for in situ 

decontamination of soil contaning 2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ Sci Technol 27(3):505-515. 

DOW Chemical Company. 1978. Assessment of risk and toxicological data. EPA/OTS public files 

document number 8EHQ-0778-0209. 

*Dragan YP, Xu X, Goldsworthy TL, et al. 1992. Characterization of the promotion of altered hepatic foci 

by 2,3,7,8-tetrachlorodibenzo-p-dioxin in the female rat. Carcinogenesis (London) 13(8):1389-1395. 

*Duarte-Davidson R, Harrad SJ, Allen SC, et al. 1992. The relative contribution of individual PCBs, 

PCDDs, and PCDFs to toxic equivalent values derived from bulked human breast milk samples from the 

UK. Chemosphere 25:1653-1663. 

Dulin D, Drossman H, Mill T. 1986. Products and quantum yields for photolysis of chloroaromatics in 

water. Environ Sci Technol 20:72-77. 

*Durrin LK, Whitlock JP Jr. 1989. 2,3,7,8-Tetrachlorodibenzo-p-dioxin: Ah receptor-mediated change in 

cytochrome P-450 chromatin structure occurs independent of transcription. Mol Cell Biol 9:5733-5737. 
CDDs 569 

8. REFERENCES 

*Eadon G, Kaminsky L, Silkworth J, et al. 1986. Calculation of 2,3,7,8-equivalent concentrations of 

complex environmental contaminant mixtures. Environ Health Perspect 70:221-227. 

*Ebert E, Price P, Keenan R. 1996. Estimating exposures to dioxin-like compounds for subsistence anglers 

in North American. In: Human exposure, human health risk assessment, and organohalogen compounds, 

16th Symposium on Chlorinated Dioxins and Related Compounds, Amsterdam Holland, volume 30, 66-69. 

*ECETOC. 1992. Exposure of man to dioxins: A perspective on industrial waste incineration. European 

Centre for Ecotoxicology and Toxicology of Chemicals. Tech Rep - ECETOC 49:1-88. 

Edgerton SA, Czuczwa JM, Rench JD, et al. 1988. Determination of polychlorinated dibenzo-p-dioxins and 

dibenzofurans and associated health risks in ambient air in Ohio. Proc APCA Annu Meet 81:16. 

*Edgerton SA, Czuczwa JM, Rench JD, et al. 1989. Ambient air concentrations of polychlorinated 

dibenzo-p-dioxins and dibenzofurans in Ohio: Sources and health risk assessment. Chemosphere 

18:1713-1730. 

*Eduljee GH. 1987. Volatility of TCDD and PCB from soil. Chemosphere 16:907-920. 

Edwards BH, Paullin JN, Coghlan-Jordan K. 1985. Emerging technologies for the control of hazardous 

wastes. J Hazard Mater 12:201-205. 

*Egeland GM, Sweeney MG, Fingerhut MA, et al. 1994. Total serum testosterone and gonadotropins in 

workers exposed to dioxins. Am J Epidemiol 139:272-281. 

*Ehrlich R, Wenning RJ, Johnson GW, et al. 1994. A mixing model for polychlorinated dibenzo-p-dioxins 

and dibenzofurans in surface sediment from Newark Bay, New Jersey, using polytopic vector analysis. Arch 

Environ Contam Toxicol 27(4):486-500. 

Eiceman GA, Clement RE, Karasek FW. 1979. Analysis of fly ash from municipal incinerators for trace 

organic compounds. Anal Chem 51:2343-2350. 

Eiceman GA, Clement RE, Karasek FW. 1981. Variations in concentrations of organic compounds 

including PCDDs and polynuclear aromatic hydrocarbons in fly ash from a municipal incinerator. Anal 

Chem 53:955-959. 

*Eisenreich SJ, Looney BB, Thornton JD. 1981. Airborne organic contaminants in the Great Lakes 

ecosystem. Environ Sci Technol 15(1):30-38. 

*Eitzer BD. 1993. Comparison of point and nonpoint sources of polychlorinated dibenzo-p-dioxins and 

polychlorinated dibenzofurans to sediments of the Housatonic River. Environ Sci Technol 27(8):1632-1637. 

Eitzer BD, Hites RA. 1986. Concentrations of dioxins and dibenzofurans in the atmosphere. Int J Environ 

Anal Chem 2:215-229. 

Eitzer BD, Hites RA. 1988. Vapor pressures of chlorinated dioxins and dibenzofurans. Environ Sci 

Technol 22:1362-1364. 
CDDs 570 

8. REFERENCES 

Eitzer BD, Hites RA. 1989a. Dioxins and furans in the ambient atmosphere: A baseline study. 

Chemosphere 18:593-598. 

*Eitzer BD, Hites RA. 1989b. Polychlorinated dibenzo-p-dioxins and dibenzofurans in the ambient 

atmosphere of Bloomington, Indiana. Environ Sci Technol 23:1389-1395. 

Eklund G, Pedersen JR, Stromberg B. 1986. Phenol and HCl at 550 degrees yield a large variety of 

chlorinated toxic compounds. Nature 320:155-156. 

*Elferink CJ, Whitlock JP Jr. 1990. 2,3,7,8-Tetrachlorodibenzo-p-dioxin: Ah receptor-mediated bending of 

enhancer DNA. J Biol Chem 265:5718-5721. 

Ema M, Sogawa K, Watanabe N, et al. 1992. cDNA cloning and structure of mouse putative Ah receptor. 

Biochem Biophys Res Comm 184:246-253. 

*Enan E, Matsumura F. 1995. Regulation by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) of the DNA 

binding activity via nuclear protein phosphorylation in guinea pig adipose tissue. Biochem Pharmacol 

50:1199-1206. 

*Enan E, Matsumura F. 1996. Identification of c-src as the integral component of the cytosolic Ah receptor 

complex, transducing the signal of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) through the protein 

phosphorylation pathway. Biochem Pharmacol 52:1599-1612. 

Enfield CG, Bengtsson G. 1988. Macromolecular transport of hydrophobic contaminants in aqueous 

environments. Groundwater 26:64-70. 

Engblom JF, Norback DH. 1973. Chlorinated dibenzo-p-dioxin distribution with rat tissues and 

subfractions of the liver. Fed Proc Fed Amer Soc Exp Biol 32:236. 

Engesser KH, Fischer P. 1991. Degradation of haloaromatic compounds. In: Betts WB, ed. Springer 

series in applied biology, Biodegradation: Natural and synthetic materials. Berlin, Germany, New York, 

NY: Springer-Verlag, 15-54. 

*EPA. 1971a. Hexachlorophene: Tolerance for residues. U.S. Environmental Protection Agency. Federal 

Register 36:22569. 

*EPA. 1971b. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 180.302. 

*EPA. 1973. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 136, 

Appendix A. 

*EPA. 1977. Ocean dumping: Constituents prohibited as other than trace contaminants. U.S. 

Environmental Protection Agency. Code of Federal Regulations. 40 CFR 227.6. 

*EPA. 1978. Pesticide chemicals. U.S. Environmental Protection Agency. Code of Federal Regulations. 

40 CFR 455. 
CDDs 571 

8. REFERENCES 

EPA. 1979a. Report of assessment of a field investigation of six-year spontaneous abortion rates in three 

Oregon areas in relation to forest 2,4,5-T spray practice. Washington, DC: U.S. Environmental Protection 

Agency, Office of Toxic Substances. 

EPA. 1979b. 2,4,5-T and silvex: Introduction to suspensions and notice of intent to cavell. U.S. 

Environmental Protection Agency. Federal Register 44:15874. 

*EPA. 1979c. Toxic pollutants. U.S. Environmental Protection Agency. Code of Federal Regulations. 40 

CFR 401.15. 

*EPA. 1981a. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 413.02. 

*EPA. 1981b. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 264.343 

*EPA. 1982a. Steam electric power generating point source category. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 423.17. 

*EPA. 1982b. Steam electric power generating point source category. U.S. Environmental Protection 

Agency. Federal Register 47:52304, Appendix A. 

EPA. 1982c. Test method: 2,3,7,8-Tetrachlorodibenzo-p-dioxin--method 613. Cincinnati, OH: U.S. 

Environmental Protection Agency, Environmental Monitoring and Support Laboratory. 

*EPA. 1983a. Underground injection control program. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 144.1. 

*EPA. 1983b. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 122.21. 

*EPA. 1984a. Ambient water quality criteria document for 2,3,7,8-tetrachlorodibenzo-p-dioxin. Cincinnati, 

OH: U.S. Environmental Protection Agency, Environmental Criteria and Assessment Office. EPA 

440/5-84/194/1. 

EPA. 1984b. Health and environmental effects profile for tetra-, penta-, and hexachlorodibenzo-p-dioxins. 

Cincinnati, OH: U.S. Environmental Protection Agency, Office of Research and Development, 

Environmental Criteria and Assessment Office. EPA/600/22. 

EPA. 1984c. Health assessment document for polychlorinated dibenzo-p-dioxins. U.S. Environmental 

Protection Agency. Federal Register 49:19408. 

*EPA. 1984d. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 261, 

Appendix IX. 

*EPA. 1985a. Drinking water criteria document for 2,3,7,8-tetrachlorodibenzo-p-dioxin. Cincinnati, OH: 

U.S. Environmental Protection Agency, Environmental Criteria and Assessment Office. EPA Report No. 

600/x-84-194-I. 

EPA. 1985b. Hazardous waste management system: Dioxin-containing wastes. U.S. Environmental 

Protection Agency. Federal Register 50:1978. 
CDDs 572 

8. REFERENCES 

EPA. 1985c. Health and environmental effects profile for brominated dibenzo-p-dioxins. Cincinnati, OH: 

U.S. Environmental Protection Agency, Office of Research and Development, Environmental Criteria and 

Assessment Office. EPA/600/22. 

EPA. 1985d. Health assessment document for polychlorinated dibenzo-p-dioxins. Washington, DC: U.S. 

Environmental Protection Agency, Office of Health and Environmental Assessment. EPA report no. 

600/8-84/014. 

EPA. 1985e. Notification requirements: Reportable quantity adjustments. U.S. Environmental Protection 

Agency. Federal Register 50:13457. 

*EPA. 1985f. Pesticide chemicals category effluent limitations guidelines, pretreatment standards and new 

source performance standards. U.S. Environmental Protection Agency. Federal Register 50:40672-40708. 

*EPA. 1985g. U. S. Environmental Protection Agency. Code of Federal Regulations 40 CFR 302.4. 

*EPA. 1985h. U. S. Environmental Protection Agency. Code of Federal Regulations CFR 266.23. 

*EPA. 1985i. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 266.112. 

*EPA. 1985j. Health assessment document for polychlorinated dibenzo-p-dioxins. U.S. Environmental 

Protection Agency. Office of Health and Environmental Assessment. Washington, DC 20460. 

EPA/800/8-84/014F. September 1985. Final Report. Pages 1.1, 2.1-2.8, 11.19, 11.122, 11.125, 11.126, and 

11.30-11.40 

EPA. 1986a. Broad scan analysis of the FY82 national human adipose tissue survey specimens. Vol. IV. 

Polychlorinated dibenzo-p-dioxins (PCDD) and polychlorinated dibenzofurans (PCDF). Washington, DC: 

U.S. Environmental Protection Agency, Office of Toxic Substances: EPA report no. 560/5-86/038. 

*EPA. 1986b. Electroplating point source category. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 413. 

*EPA. 1986c. Electroplating point source category. U.S. Environmental Protection Agency. Federal 

Register 51:40421. 

*EPA. 1986d. General pretreatment regulations for existing and new sources of pollution. U.S. 

Environmental Protection Agency. Code of Federal Regulations. 40 CFR 403.2 Appendix B. 

*EPA. 1986e. General pretreatment regulations for existing and new sources of pollution. U.S. 

Environmental Protection Agency. Federal Register 51:20429-20431. 

*EPA. 1986f. Hazardous waste management system: Land disposal restriction. U.S. Environmental 

Protection Agency. Federal Register 51:40572-40623. 

*EPA. 1986g. Metal finishing point source category. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 433. 

*EPA. 1986h. Metal finishing point source category. U.S. Environmental Protection Agency. Federal 

Register 51:40421. 
CDDs 573 

8. REFERENCES 

EPA. 1986i. Testing and reporting requirements for polyhalogenated dibenzo-p-dioxins/dibenzofurans: 

Addition of chlorinated and brominated benzenes to list of precursor chemicals. U.S. Environmental 

Protection Agency. Federal Register 51:37612-37613. 

EPA. 1986j. 2,3,7,8-Tetrachlorodibenzo-p-dioxins: Health advisory. Draft. Washington, DC: U.S. 

Environmental Protection Agency, Office of Drinking Water, October 16, 1986. 

*EPA. 1986k. Test methods for evaluating solid waste (SW 846) volume IB: Laboratory manual physical 

chemical methods. Method 8280, the analysis of polychlorinated dibenzo-p-dioxins and polychlorinated 

dibenzofurans. U. S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, 

Washington, DC. 

*EPA. 1986l. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 268.50. 

EPA. 1987a. EPA contract laboratory program: Statement of work for organic analysis. Washington, DC: 

U.S. Environmental Protection Agency. 

*EPA. 1987b. Groundwater monitoring list. U.S. Environmental Protection Agency. Code of Federal 

Regulations. 40 CFR 264, Appendix IX. 

*EPA. 1987c. Hazardous materials table and hazardous materials communications regulations. U.S. 

Environmental Protection Agency. Federal Register 52:4825-4843. 

*EPA. 1987d. Hexachlorophene: tolerance for residues. U.S. Environmental Protection Agency. Code of 

Federal Regulations. 40 CFR 180.302. 

*EPA. 1987e. Interim procedures for estimating risks associated with exposure to mixtures of chlorinated 

dibenzo-p-dioxins and dibenzo furans (CDDs and CDFs). Risk Assessment Forum. Washington, DC: U.S. 

Environmental Protection Agency: EPA report no. 625/3-87/012. 

*EPA. 1987f. Listing of hazardous substance and reportable quantities. U.S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 302.4. 

*EPA. 1987g. List (phase I) of hazardous constituents for ground-water monitoring. U.S. Environmental 

Protection Agency. Federal Register 52:25942-25952. 

EPA. 1987h. Polychlorinated dibenzo-p-dioxin/dibenzofurans: Testing and reporting requirements. U.S. 

Environmental Protection Agency. Federal Register 52:21412-21442. 

*EPA. 1987i. Reportable quantity adjustments. U.S. Environmental Protection Agency. Federal Register 

52:8140. 

*EPA. 1987j. 2,3,7,8-TCDD: Health advisory. Washington, DC: U.S. Environmental Protection Agency, 

Office of Drinking Water. 1987. 

*EPA. 1987k. 2,3,7,8-TCDD: Environmental chemistry. Athens, GA: Environmental Protection Agency, 

Office of Research and Development, Environmental Research Laboratory. EPA600/D-87/086. 

EPA. 1987l. 2,3,7,8-Tetrachlorodibenzo-p-dioxin: Environmental chemistry. ACS Symp Ser 338:82-93. 
CDDs 574 

8. REFERENCES 

*EPA. 1987m. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 268 

Appendix IIII. 

*EPA. 1987n. The National Dioxin Study; Tiers 3,5,6, and 7. Office of Water Regulations and Standards. 

Washington DC EPA 440/4-87-003. 

EPA. 1988a. A cancer risk-specific dose estimate of 2,3,7,8-TCDD. Washington, DC: U.S. Environmental 

Protection Agency, Office of Research and Development, Office of Health Effects Assessment. EPA report 

no. 600/6-88/007Aa. 

EPA. 1988b. A cancer risk-specific dose estimate of 2,3,7,8-TCDD: Appendices A through F. 

Washington, DC: U.S. Environmental Protection Agency, Office of Research and Development, Office of 

Health Effects Assessment. EPA report no. 600/6-88/007Ab. 

EPA. 1988c. Estimating exposure to 2,3,7,8-TCDD. Washington, DC: U.S. Environmental Protection 

Agency, Office of Research and Development, Office of Health Effects Assessment. EPA report no. 

600/6-88/005A. 

*EPA. 1988d. Hazardous constituents. U.S. Environmental Protection Agency. Code of Federal 

Regulations. 40 CFR 261, Appendix VIII. 

*EPA. 1988e. Hazardous waste management system: Identification and listing of hazardous waste. U.S. 

Environmental Protection Agency. Federal Register 53:13382-13393. 

*EPA. 1988f. Land disposal: Waste specified prohibitions--dioxin-containing wastes. U.S. Environmental 

Protection Agency. Code of Federal Regulations. 40 CFR 268.31. 

*EPA. 1988g. Compendium on methods for the determination of toxic organic compounds in ambient air, 

Method TO-9, method for the determination of polychlornated dibenzo-p-dioxins(PCDDs) in ambient air 

using high resolution gas chromatography high resolution mass spectrometry. (HRGC/HRMS). EPA 

600/4/89/017. 

*EPA. 1988h. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 148.11. 

EPA. 1989a. Assessment of the potential for transport of dioxins and codisposed materials to groundwater. 

Washington, DC: U.S. Environmental Protection Agency. EPA 600/S6-89/002. 

*EPA. 1989b. Drinking water standards and health advisory table. Washington, DC: U.S. Environmental 

Protection Agency, Office of Drinking Water. 

*EPA. 1989c. Interim methods for development of inhalation reference doses. Washington, DC: U.S. 

Environmental Protection Agency, Office of Health and Environmental Assessment. EPA 600/8-88-066F. 

*EPA. 1989d. Notes of 2/01/89 CRAVE meeting. Memorandum to CRAVE Workgroup, Office Directors 

from Rita Schoeny. Cincinnati, OH: U.S. Environmental Protection Agency, Office of Research and 

Development, Environmental Criteria and Assessment Office. 
CDDs 575 

8. REFERENCES 

*EPA. 1989e. Interim procedures for estimating risks associated with exposure to mixtures of chlorinated 

dibenzo-p-dioxins and dibenzofurans (CDDs and CDFs) and 1989 update. Risk Assessment Forum. 

Washington, DC: U.S. Environmental Protection Agency. EPA 625/3-89/016. NTIS PB90-145756. 

*EPA. 1989f. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 266, 

Appendix IX. 

*EPA. 1990. Standards of performance for volatile organic compounds (VOC) emissions from synthetic 

organic chemical manufacturing industry (SOCMI) distillation operation. U. S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 60.667. 

*EPA. 1990a. Characterization of municipal waste combustion ash, ash extracts, and leachates. Coalition 

on Resource Recovery and the Environment. Washington, DC: U.S. Environmental Protection Agency, 

Office of Solid Waste and Emergency Response. Contract No. 68-01-7310. EPA 530-SW-90-029A. 

*EPA. 1990b. Standards for owners and operators of hazardous waste treatment, storage, and disposal 

facilities. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 264.342. 

*EPA. 1990c. Dibenzo-para-dioxins/dibenzofurans in bleached wood pulp and paper product referral for 

action. U. S. Environmental Protection Agency. 55 Federal Register 248:53047-53049. 

*EPA. 1990d. Summary report on USEPA/ Industry Cooperative Dioxin Study. "The 104 mills study" 

Washington, DC Office of Water Regulations and Standards July 1990. 

*EPA. 1991a. Health effects assessment summary tables. U.S. Environmental Protection Agency. 

Prepared by Office of Health and Environmental Assessment, Environmental Assessment and Criteria 

Office, Cincinnati, OH, for the Office of Solid Waste and Emergency Response, Office of Emergency and 

Remedial Response, Washington, DC, January 1991. OERR 9200.6-303(91-1). 

*EPA. 1991b. Proposed regulation of land application of sludge from pulp and paper mills using chlorine 

and chlorine derivative bleaching processes. U.S. Environmental Protection Agency. Federal Register 

56:21802. 

*EPA. 1991c. Public meeting on EPA's scientific reassessment of dioxin. U.S. Environmental Protection 

Agency. Federal Register 56:50903-50904. 

*EPA. 1991d. Atmospheric transport and deposition of polychlorinated dibenzo-p-dioxins and 

dibenzofurans. Washington, DC: U.S. Environmental Protection Agency. EPA600/3-91/002 Order No. 

PB91-144667. 

*EPA. 1991e. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 60.51a. 

*EPA. 1991f. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 744. 

*EPA. 1991g. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 266.103. 

*EPA. 1991h. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 266.104. 
CDDs 576 

8. REFERENCES 

*EPA. 1991i. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 268.48. 

*EPA. 1992. National Study of Chemical Residues in Fish . Volume 1. Office of Science and Technology, 

Washington D.C. EPA 823-R-92-008a. September 1992. 

*EPA. 1993a. Development document for proposed effluent limitation guidelines and standards for the 

pulp, paper and paperboard point source category. Environmental Protection Agency, Office of Water, 

Washinton, DC. EPA-821-R-93-019. 

*EPA. 1993b. U. S. Environmental Protection Agency. Federal Register. 58 FR 408892. Code of Federal 

Regulations. 40 CFR 148. 

*EPA. 1993c. U. S. Environmental Protection Agency. Federal Register. 58 FR 25706. Code of Federal 

Regulations. 40 CFR 261.31. 

*EPA. 1994a. Method 1613: tetra-through octa chlorinated dioxins and furans by isotope dilution 

HRGC/HRMS. United States Environmental Protection Agency, Office of Water, Engingering and Analysis 

Division, Washington DC. EPA 821-B-94-005. 

*EPA. 1994b. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 268.48. 

*EPA. 1994c. Polychlorinated dibenzodioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) by 

high-resolution gas chromatography/high-resolution mass spectrometry (HRGC/HRMS). Test methods for 

evaluating solid waste, method 8290. Environmental Protection Agency, Washington, DC. 

*EPA. 1994d. National primary drinking water regulations. U.S. Environmental Protection Agency. Code 

of Federal Regulations. 40 CFR 141. 

*EPA. 1994e. Variances and exeptions from the maximum contaminant levels for organic and inorganic 

chemicals. U. S. Environmental Protection Agency. Code of Federal Regulations. 40 CFR 142.62. 

*EPA. 1995a. Drinking water regulations and health advisories. Office of Water, U. S. Environmental 

Protection Agency. Washington DC. 

*EPA. 1995b. Emissions guidelines and compliance times for municipal waste combustors. U. S. 

Environmental Protection Agency. Code of Federal Regulations. 40 CFR 60, Subpart Cb. 

*EPA. 1995c. Guidance for assessing chemical contaminant data for use in fish advisories. Volume 1. 

Sampling and analysis second edition. Office of Water Washington DC EPA 823-R-95-007 September 1995 

*EPA. 1995d. Standards of performance for municipal waste combustors for which construction is 

commenced after septermer 20, 1994. U.S. Environmental Protection Agency. Code of Federal Regulations. 

40 CFR 60, Subpart Eb. 

*EPA. 1995e. National emission standards for wood furniture manufacturing operations. U. S. 

Environmental Protection Agency. Code of Federal Regulations. 40 CFR 63, Subpart JJ. 
 

 

 

 

 

 

CDDs 577 

8. REFERENCES 

*EPA. 1995f. Standards of performance for municipal waste combustors for which contruction is 

commenced after december 20, 1989 and on or before september 20, 1994. U. S. Environmental Protection 

Agency. Code of Federal Regulations. 40 CFR 60, Subpart Ea. 

*EPA. 1995g. Toxic Chemical Release Inventory Reporting Form R and Instructions (revised 1994 version). 

Office of Pollution Prevention and Toxics, Washington DC EPA 745-K-95-051. 

*EPA. 1996a. Drinking water regulations and health advisories. U. S. Environmental Protection Agency. 

Office of Water. EPA 822-B-96-002. October 1996. 

*EPA. 1996b. National listing of fish and wildlife consumption advisories. Office of Science and 

Technology, Washington DC EPA 823-C-96-011. 

*EPA. 1997a. Large municipal waste combustion units; emission guidelines; final rule. U. S. 

Environmental Protection Agency. Federal Register. 62 FR 45116. August 25, 1997. 

*EPA. 1997b. Standards of perfomance for new stationary sources and emission guidelines for existing 

sources: hospital/medical/infectious waste incinerators; final rule. U. S. Environmental Protection Agency. 

Federal Register. 62 FR 48348. September 15, 1997. 

*EPA. 1997c. Addition of dioxin and dioxin-like compounds; modification of polychlorinated biphenyls 

(PCBs) listing; toxic chemical release reporting; community right-to-know. U. S. Environmental Protection 

Agency. Federal Register. 62 FR 24887. 

*EPA. 1997d. Land disposal restrictions-phas IV: treatment standards for wood preserving wstes, 

paperwork reduction and streamlining, exemptions from RCRA for certain processed materials; and 

miscellaneous hazardous waste provisions; final rule. U. S. Environmental Protection Agency. Federal 

Register. 62 FR 25998. May 12, 1997. 

*EPA. 1997e. Water quality standards; establishment of numeric criteria for priority toxic pollutants for the 

state of california; proposed rule. U. S. Environmental Protection Agency. Federal Register. 62 FR 42160. 

August 5, 1997. 

*EPA. 1997f. Emissions guidelines for existing sources and standards of performance for new stationary 

sources: large municipal waste combustion units; final rule. U. S. Environmental Protection Agency. 

Federal Register. 62 FR 45124. August 25, 1997. 

*EPA. 1997g. Guidelines establishing test procedures for the analysis of pollutants; epa method 1613;l final 

rule. U. S. Environmental Protection Agency. Federal Register. 62 FR 48394. 

*EPA. 1998a. Update: listing of fish and wildlife advisories. U.S. Environmental Protection Agency. 

Office of Water. Fact Sheet. EPA-823-F-98-009. March 1998. 

*EPA. 1998b. Listing of fish and wildlife advisories - 1997. U.S. Environmental Protection Agency Office 

of Water, Washington, DC. 

*EPA. 1998c. National emission standards for hazardous air pollutants; proposed standards for hazardous 

air pollutants emissions for portlancd cement manufacturing industry; proposed rule. U. S. Environmental 

Protection Agency. Federal Register. 63 FR 14182. March 24, 1998. 
 

 

 

 

 

CDDs 578 

8. REFERENCES 

*EPA. 1998d. National primary drinking water regulations: consumer confidence; proposed rule. U. S. 

Environmental Protection Agency. Federal Register. 63 FR 7606. February 13, 1998. 

*EPA. 1998e. Significant new uses of certain chemical substances; final rule. U. S. Environmental 

Protection Agency. Federal Register. 63 FR 3394. Janyary 22, 1998. 

*EPA. 1998f. Effluent limitiaitions guidelines, pretreatment standrds, and new soruce performance 

standards for the transportaion equipment cleaning point source category; proposed rule. U. S. 

Environmental Protection Agency. Federal Register. 63 FR 34686. June 25, 1998. 

*EPA. 1998g. Effluent limitations guidelines, pretreatment standards, and new source performance 

standards for the industrial waste combustor subcategory of the waste combustors point source category; 

proposed rule. U. S. Environmental Protection Agency. Federal Register. 63 FR 6392. February 6, 1998. 

*EPA. 1998h. Effluent limitations guideline, pretreatment standards, and new source performance standrds 

for the landfills point source category; proposed rule. U. S. Environmental Protection Agency. Federal 

Register. 63 FR 6426. February 6, 1998. 

*EPA. 1998i. Hazardous waste combustors; revised standards; final rule. U. S. Environmental Protection 

Agency. Federal Register. 63 FR 33782. June 19, 1998. 

*EPA. 1998j. Source category listing for section 112 (d) (2) rulemaking pursuant to section 112 (c) (6) 

requirements. U.S. Environmental Protection Agency (EPA) Federal Register 63 FR 17838-17851 (April 

10, 1998). 

Epling GA, Qiu QW, Kumar A. 1989. Hydride enhanced photoreaction of chlorodibenzodioxins. 

Chemosphere 18:329-332. 

*Erickson JD, Mulinare J, McClain PW, et al. 1984. Vietnam veterans' risks for fathering babies with birth 

defects. JAMA 252:903-912. 

Erickson MD, Swanson SE, Flora JD, et al. 1989. Polychlorinated dibenzofurans and other thermal 

combustion products from dielectric fluids containing polychlorinated biphenyls. Environ Sci Technol 

23:462-470. 

*Eriksson M, Hardell L, Adami HO. 1990. Exposure to dioxins as a risk factor for soft tissue sarcoma: A 

population-based case-control study. J Natl Cancer Inst 82:486-490. 

Eriksson M, Hardell L, Berg NO, et al. 1979. Case-control study on malignant mesenchymal tumors of the 

soft tissue and exposure to chemical substances. Lakartidningen 76:3872-3875. 

*Eriksson M, Hardell L, Berg NO, et al. 1981. Soft-tissue sarcomas and exposure to chemical substances: 

A case-referent study. Br J Ind Med 38:27-33. 

*Ernst M, Flesch-Janys D, Morgenstern I, et al. 1998. Immune cell functions in industrial workers after 

exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin: Dissociation of antigen-specific T-cell responses in 

cultures of diluted whole blood and of isolated peripheral blood mononuclear cells. Environ Health Perspect 

106(Suppl 2):701-705. 
CDDs 579 

8. REFERENCES 

*Eschenroeder A, Jaeger RJ, Ospital JJ, et al. 1986. Health risk analysis of human exposures to soil 

amended with sewage sludge contaminated with polychlorinated dibenzodioxins and dibenzofurans. Vet 

Hum Toxicol 28:435-442. 

*Evans RG, Webb KB, Knutsen AP, et al. 1988. A medical follow-up of the health effects of long-term 

exposure to 2,3,7,8-TCDD. Arch Environ Health 43:273-278. 

Evers EHG, Ree KCM, Olie K. 1988. Spatial variations and correlations in the distribution of PCDDs, 

PCDFs and related compounds in sediments for the River Rhine-Western Europe. Chemosphere 

17:2271-2288. 

*Facchetti S, Balasso A, Fichtner C, et al. 1986. Studies on the absorption of TCDD by plant species. In: 

Rappe C, Choudhary G, Keith LH, eds. Chlorinated dioxins and dibenzofurans in perspective. Chelsea, MI: 

Lewis Publishers, Inc., 225-235. 

*Fairless BJ, Bates DI, Hudson J, et al. 1987. Procedures used to measure the amount of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in the ambient air near a Superfund site cleanup operation. Environ Sci 

Technol 21:550-555. 

*Faith RE, Luster MI, Moore JA. 1978. Chemical separation of helper cell function and delayed type 

hypersensitivity. Cell Immunol 40:275-284. 

*Faith RE, Moore JA. 1977. Impairment of thymus-dependent immune functions by exposure of 

developing immune system to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). J Toxicol Environ Health 

3:451-464. 

Falk H, Stehr PA, Stein GF, et al. 1984. A pilot epidemiology study of health effects due to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) contamination in Missouri. Banbury Rep. Issue Biol Mech 

Dioxin Action 18:447-460. 

*Fan F, Rozman KK. 1995. Short- and long-term biochemical effects of 2,3,7,8-tetrachorodibenzo-p-dioxin 

in female Long-Evans rats. Toxicol Letters 75:209-216. 

*Fan F, Wierda D, Rozman KK. 1996. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on humoral and 

cell-mediated immunity in Sprague-Dawley rats. Toxicology 106:221-228. 

*Faqi AS, Dalsenter PR, Merker H-J, et al. 1998. Reproductive toxicity and tissue concentrations of low 

doses of 2,3,7,8-tetrachlorodibenzo-p-dioxin in male offspring rats exposed throughout pregnancy and 

lactation. Tox Appl Pharmacol 150:383-392. 

Favreau LV, Pickett CB. 1993. Transcriptional regulation of the rat NAD(P)H:quinone reductase gene. 

Characterization of a DNA-protein interaction at the antioxidant responsive element and induction by 

12-O-tetradecoylphorbol 13-acetate. J Biol Chem 268:19875-19881. 

FDA. 1983. Statement by SA Miller, Director of Bureau of Foods, FDA: Testimony before the 

Subcommittee on natural resources, agriculture, research and environment. Washington, DC: U.S. House of 

Representatives, Food and Drug Administration, June 30. 
 

 

 

CDDs 580 

8. REFERENCES 

*FDA. 1995. Requirements for Specifice Standardized Beverages. U. S. Food and Drug Administration. 

Code of Federal Regulations. 21 CFR165.110. 

*FDA. 1998a. Beverages: bottled water; direct final rule. U. S. Food and Drug Administration. Federal 

Register. 63 FR 25764. May 11, 1998. 

*FDA. 1998b. Beverages: bottled water; corrections. U. S. Food and Drug Administration. Federal 

Register. 63 FR 30620. June 5, 1998. 

*FDA. 1998c. Beverage: bottled water; direct final rule; confirmaiton. U. S. Food and Drug 

Administration. Federal Register. 63 FR 42198. August 6, 1998. 

*FEDRIP. 1998. Federal Research In Progress. Dialog Information Service, Inc., Palo Alto, CA. February 

1998. 

Fehringer NV, Walter SM, Kozara RJ, et al. 1985. Survey of 2,3,7,8-tetrachlorodibenzo-para-dioxin in fish 

from the Great Lakes and selected Michigan rivers. J Agric Food Chem 33:626-630. 

*Fernandez-Salguero PM, Hilbert DM, Rudikoff S, et al. 1996. Aryl-hydrocarbon receptor-deficient mice 

are resistant to 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced toxicity. Toxicol Appl Pharmacol 140:173-179. 

*Ferrario J, Byrne C, Lorber M, et al. 1997. A statistical survey of dioxin-like compounds in United States 

poultry fat. Organohalogen Compounds 32:64-70. 

*Ferrario J, Byrne C, McDaniel D, et al. 1996. Determination of 2,3,7,8,-chlorine substituted 

dibenzo-p-dioxins and furans at the part per trillion level in United States beef fat using high resolution gas 

chromatography/high resolution mass spectrometry. Anal Chem 68:647-652. 

Fett MJ, Dunn M, Adena MA, et al. 1984. Australian veterans health studies: The mortality report. Part I: 

A retrospective cohort study of mortality among Australian national servicemen of the Vietnam conflict era, 

and an executive summary of the mortality report. Canberra, Australia: Australian Government Publishing 

Service. 

*Fiedler H, Cooper KR, Bergek S, et al. 1997. Polychlorinated dibenzo-p-dioxins and polychlorinated 

dibenzofurans (PCDD/PCDF) in food samples collected in southern Mississippi, USA. Chemosphere 

34(5-7):1411-1419. 

Field B, Kerr C. 1979. Herbicide use and incidence of neural tube defects. Lancet 1:1341-1342. 

*Filippini G, Bordo B, Crenna P, et al. 1981. Relationship between clinical and electrophysiological 

findings and indicators of heavy exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Scand J Work Environ 

Health 7:257-262. 

*Fine JS, Gasiewicz TA, Fiore NC, et al. 1990a. Prothymocyte activity is reduced by perinatal 

2,3,7,8-tetrachlorodibenzo-p-dioxin exposure. J Pharmacol Exp Ther 255:128-132. 

*Fine JS, Gasiewicz TA, Silverstone AE. 1989. Lymphocyte stem cell alterations following perinatal 

exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Mol Pharmacol 35:18-25. 
CDDs 581 

8. REFERENCES 

*Fine JS, Silverstone AE, Gasiewicz TA. 1990b. Impairment of prothymocyte activity by 

2,3,7,8-tetrachlorodibenzo-p-dioxin. J Immunol 144:1169-1176. 

Fingerhut MA, Halperin WE, Honchar PA, et al. 1984. An evaluation of reports of dioxin exposure and soft 

tissue sarcoma pathology among chemical workers in the United States. Scand J Work Environ Health 

10:299-303. 

*Fingerhut MA, Halperin WE, Marlow DA, et al. 1991. Cancer mortality in workers exposed to 

2,3,7,8-tetrachlorodibenzo-p-dioxin. N Engl J Med 324:212-218. 

*Fingerhut MA, Haring-Sweeney M, Patterson DG Jr, et al. 1989. Levels of 2,3,7,8-tetrachlorodibenzop-dioxin in the serum of U.S. chemical workers exposed to dioxin contaminated products: Interim results. 

Chemosphere 19:835-840. 

Firestone D, Clower M Jr, Borsetti AP, et al. 1979. Polychlorodibenzo-p-dioxin and pentachlorophenol 

residues in milk and blood of cows fed technical pentachlorophenol. J Agric Food Chem 27:1171-1177. 

*Firestone D, Niemann RA, Schneider LF, et al. 1986. Dioxin residues in fish and other foods. In: Rappe 

C, Choudhary G, Keith LH, eds. Chlorinated dioxins and dibenzofurans in perspective. Chelsea, MI: Lewis 

Publishers, Inc., 355-365. 

Firestone D, Ress J, Brown NL, et al. 1972. Determination of polychlordibenzo-p-dioxins and related 

compounds in commercial chlorophenols. J Assoc Off Anal Chem 55:85-92. 

Fishbein L. 1992. Exposure from occupational versus other sources. Scand J Work Environ Health 

18(1):5-16. 

*Fisher JM, Jones KW, Whitlock JP Jr. 1989. Activation of transcription as a general mechanism of 

2,3,7,8-tetrachlorodibenzo-p-dioxin action. Mol Carcinogenesis 1:216-221. 

*Fisher JM, Wu L, Denison MS, et al. 1990. Organization and function of a dioxin-responsive enhancer. J 

Biol Chem 265:9676-9681. 

*Fitzgerald EF, Hwang S-A, Brix KA, et al. 1995. Fish PCB concentrations and comsumption patterns 

among Mohawk women at Akwesasne. Journal of Exposure Analysis and Environmental Epidemiology 

5(1):1-19. 

*Flaws JA, Sommer RJ, Silbergeld EK, et al. 1997. In utero and lactational exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) induces genital dysmorphogenesis in the female rat. Toxicol 

Appl Pharmacol 147:351-362. 

*Flesch-Janys D, Becher H, Gurn P, et al. 1996. Elimination of polychlorinated dibenzo-p-dioxins and 

dibenzofurans in occupationally exposed persons. J Toxicol Environ Health 47:363-378. 

*Flesch-Janys D, Berger J, Gurn P, et al. 1995. Exposure to polychlorinated dioxins and furans (PCDD/F) 

and mortality in a cohort of workers from a herbicide-producing plant in Hamburg, Federal Republic of 

Germany. Am J Epidemiol 142:1165-1175. 
CDDs 582 

8. REFERENCES 

*Flesch-Janys D, Steindorf K, Gurn P, et al. 1998. Estimation of the cumulated exposure to polychlorinated 

dibenzo-p-dioxins/furans and standardized mortality ratio analysis of cancer mortality by dose in an 

occupationally exposed cohort. Environ Health Perspect 106(Suppl2):655-662. 

*Fletcher Cl, McKay WA. 1993. Polychlorinated dibenzo-p-dioxins PCDDs and dibenzofurans PCDFs in 

the aquatic environment a literature review. Chemosphere 26(6):1041-1069. 

Flick DF, Firestone D, Higginbotham GR. 1972. Chick edema disease: 9. Response of chicks fed or singly 

administered synthetic edema-producing compounds. Poult Sci 51:2026-2034. 

*Flodstrom S, Ahlborg UG. 1989. Tumour promoting effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD): Effects of exposure duration, administration schedule and type of diet. Chemosphere 19:779-783. 

*Flodstrom S, Ahlborg UG. 1992. Relative tumor promoting activity of some polychlorinated 

dibenzo-p-dioxin, dibenzofuran-, and biphenyl congeners in female rats. Chemosphere 25:169-172. 

*Flodstrom S, Busk L, Kronevi T, et al. 1991. Modulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin and 

phenobarbital-induced promotion of hepatocarcinogenesis in rats by the type of diet and vitamin A 

deficiency. Fund Appl Toxicol 16:375-391. 

*Foman 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. 

*Foman, SJ, Haschke F, Ziegler EE et al. 1982. Body composition of reference children from birth to 10 

years. American Journal of Clinical Nutrition 35:1169-1175. 

*Forsberg B, Nordstrom S. 1985. Miscarriages around a herbicide manufacturing company in Sweden. 

Ambio 14:110-111. 

*Först C, Stieglitz L, Zwick G. 1988. Isomer-specific determination of PCDD/PCDF in oil extracts from 

water leachate of a waste landfill. Chemosphere 17:1935-1944. 

Fossa AJ, Kerr RS, Columbus AF, et al. 1987. Air emission characterization of municipal waste combustors 

in New York State. Proc APCA Annu Meet 80:2-14. 

Foster MG, Huneault H, Cosby O, et al. 1990. Assessment of reference materials for application to 

PCDD/PCDF analysis. Chemosphere 20:1285-1290. 

*Foster WG, Ruka MP, Gareau P, et al. 1997. Morphologic characteristics of endometriosis in the mouse 

model: Application to toxicology. Can J Physiol Pharmacol 75:1188-1196. 

*Fowler BA, Hook GER, Lucier GW. 1977. Tetrachlorodibenzo-p-dioxin induction of renal microsomal 

enzyme systems: Ultrastructural effects on pars recta (S3) proximal tubule cells of the rat kidney. J 

Pharmacol Exp Ther 203:712-721. 

Frakes RA, Zeeman C QT, Mower B. 1993. Bioaccumulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) by fish downstream of pulp and paper mills in Maine. Ecotoxicol Environ Saf 25(2):244-252. 
CDDs 583 

8. REFERENCES 

*Frazier DE, Silverstone AE, Gasiewicz TA. 1994. 2,3,7,8-Tetrachlorodibenzo-p-dioxin-induced thymic 

atrophy and lymphocyte stem cell alterations by mechanisms independent of the estrogen receptor. Biochem 

Pharmacol 47:2039-2048. 

Freeman RA, Hileman FD, Noble RW, et al. 1987. Experiments on the mobility of 

2,3,7,8-tetrachlorodibenzo-p-dioxin at Times Beach, Missouri. ACS Symp Ser 338:114-130. 

Freeman RA, Schroy JM. 1985. Environmental mobility of TCDD. Chemosphere 14:873-876. 

*Freeman RA, Schroy JM. 1986. Modeling the transport of 2,3,7,8-TCDD and other low volatility 

chemicals in soils. Environ Progress 5:28-33. 

Freeman RA, Schroy JM. 1988. Modeling of chemical movement in soil matrices. Presented at the 195th 

ACS National Meeting, Toronto, Ontario, Canada, June 5-10, 8. 

*Freeman RA, Schroy JM, Hileman FD, et al. 1986. Environmental mobility of 2,3,7,8-TCDD and 

companion chemicals in a roadway soil matrix. In: Rappe C, Choudhary G, Keith LH, eds. Chlorinated 

dioxins and dibenzofurans in perspective. Chelsea, MI: Lewis Publishers, Inc., 171-183. 

Friedman JM. 1984. Does Agent Orange cause birth defects? Teratology 29:193-221. 

*Friedrich C, Kleiboehmer W. 1997. Supercritical co2-assisted liquid extraction of polycyclic aromatic 

hydrocarbons and polychlorinated dibenzo-p-dioxins and -furans from solid matrices. Journal of 

Chromatography A 777(2):289-294. 

*Fries GF, Marrow GS. 1975. Retention and excretion of 2,3,7,8-tetrachlorodibenzo-p-dioxin by rats. J 

Agric Food Chem 23:265-269. 

*Fries GF, Paustenbach DJ. 1990. Evaluation of potential transmission of 2,3,7,8-tetrachlorodibenzop-dioxin-contaminated incinerator emissions to human via foods. Journal of Toxicology and Environmental 

Health 29:1-43. 

Friesen KJ, Muir DCG, Webster GRB. 1990. Evidence of sensitized photolysis of polychlorinated 

dibenzo-p-dioxins in natural waters under sunlight conditions. Environ Sci Technol 24:1739-1744. 

*Friesen KJ, Sarna LP, Webster GRB. 1985. Aqueous solubility of polychlorinated dibenzo-p-dioxins 

determined by high pressure liquid chromatography. Chemosphere 14:1267-1274. 

Friesen MD, Hass Jr, Harvan D, et al. 1980. Negative ion chemical ionization mass spectrometry: 

Determination on chlorinated dibenzo-p-dioxins. Pract Spectrosc 3:324-375. 

Frommberger R. 1990. Polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans in cumulative 

samples of human milk from Baden-Wuerttemberg, FRG. Chemosphere 20:333-342. 

*FSTRAC. 1990. Summary of state and federal drinking water standards and guidelines. Federal-State 

Toxicology and Regulatory Alliance Committee, Chemical Communication Subcommittee. 

*Fukunaga BN, Probst MR, Reisz-Porszasz S, et al. 1995. Identification of functional domains of the aryl 

hydrocarbon receptor. J Biol Chem 270: 29270-29278. 
CDDs 584 

8. REFERENCES 

*Fürst P, Fürst C, Wilmers K. 1992. PCDDs and PCDFs in human milk statistical evaluation of a 6-years 

survey. 1992. Chemosphere 25(7-10):1029-1038. 

*Fürst P, Fürst C, Wilmers K. 1994. Human milk as a bioindicator for body burden of PCDDs, PCDFs, 

organochlorine pesticides, and PCBs. Environ Health Perspect 102(1):187-193. 

*Fürst P, Krause GHM, Hein D, et al. 1993. PCDD/PCDF in cow's milk in relation to their levels in grass 

and soil. Chemosphere 27(8):1349-1357. 

Fürst P, Krüger C, Meemken HA, et al. 1989a. Interaction between sample and packaging material - a 

potential source of contamination with PCDDs and PCDFs. Chemosphere 18:891-896. 

*Fürst P, Krüger C, Meemken HA, et al. 1989b. PCDD and PCDF levels in human milk - dependence on 

the period of lactation. Chemosphere 18:439-444. 

Fürst P, Meemken HA, Groebel W. 1986. Determination of polychlorinated dibenzodioxins and 

dibenzofurans in human milk. Chemosphere 15:1977-1980. 

Fürst P, Meemken HA, Krüger C, et al. 1987. Polychlorinated dibenzodioxins and dibenzofurans in human 

milk samples from Western Germany. Chemosphere 16:1983-1988. 

*Gaido KW, Maness SC. 1994. Regulation of gene expression and acceleration of differentiation in human 

keratinocytes by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 127:199-208. 

*Garrison PM, Tullis K, Aarts JMMJG, et al. 1996. Species-specific recombinant cell lines as bioassay 

systems for the detection of 2,3,7,8-tetrachlorodibenzo-p-dioxin-like chemicals. Fund Appl Toxicol 

30:194-203. 

Garten CT Jr, Trabalka JR. 1983. Evaluation of models for predicting terrestrial food chain behavior of 

xenobiotics. Environ Sci Technol 17:590-595. 

Gasiewicz TA. 1991. Nitro compounds and related phenolic pesticides. In: Hayes WJ, Laws ER, eds. 

Handbook of pesticide toxicology. Volume 3: Classes of pesticides. New York, NY: Academic Press, Inc., 

1191-1269. 

*Gasiewicz TA, Elferink CJ, Henry EC. 1991. Characterization of multiple forms of the Ah receptor: 

Recognition of a dioxin-responsive enhancer involves heteromer formation. Biochemistry 30:2909-2916. 

*Gasiewicz TA, Geiger LE, Rucci G, et al. 1983a. Distribution, excretion, and metabolism of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in C57BL/6J, DBA/2J, and B6D2F1/J mice. Drug Metab Dispos 

11:397-403. 

*Gasiewicz TA, Henry EC, Baggs RB, et al. 1986. Temporal- and dose-related characteristics of 

biochemical and morphological alterations in the hamster induced by 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Chemosphere 15:1749-1752. 

*Gasiewicz TA, Holscher MA, Neal RA. 1980. The effect of total parenteral nutrition on the toxicity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Toxicol Appl Pharmacol 54:469-488. 
CDDs 585 

8. REFERENCES 

*Gasiewicz TA, Neal RA. 1979. 2,3,7,8-Tetrachlorodibenzo-p-dioxin tissue distribution, excretion, and 

effects on clinical parameters in guinea pigs. Toxicol Appl Pharmacol 51:329-339. 

*Gasiewicz TA, Olson RJ, Geiger LE, et al. 1983b. Absorption, distribution, and metabolism of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in experimental animals. In: Tucker RE, Young AL, Gray AP, 

eds. Human and environmental risk of chlorinated dioxins and related compounds. New York, London: 

Plenum Press, 495-525. 

*Gasiewicz TA, Rucci G. 1984. Cytosolic receptor for 2,3,7,8-tetrachlorodibenzo-p-dioxin. Evidence for a 

homologous nature among various mammalian species. Mol Pharmacol 26:90-98. 

*Gasiewicz TA, Rucci G. 1991. Alpha-naphthoflavone acts as an antagonist of 2,3,7,8-tetrachlorodibenzop-dioxin by forming an inactive complex with the Ah receptor. Mol Pharmacol 40:607-612. 

*Gehrs BC, Riddle MM, Williams WC, et al. 1997a. Alterations in the developing immune system of the 

F344 rat after perinatal exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. II. Effects on the fetus and the 

neonate. Toxicology 122:219-228. 

*Gehrs BC, Riddle MM, Williams WC, et al. 1997b. Alterations in the developing immune system of the 

F344 rat after perinatal exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. II. Effects on the pup and the adult. 

Toxicology 122:229-240. 

*Geiger LE, Neal RA. 1981. Mutagenicity testing of 2,3,7,8-tetrachlorodibenzo-p-dioxin in histidine 

auxotrophs of Salmonella typhimurium. Toxicol Appl Pharmacol 59:125-129. 

Geyer HJ, Scheunert I, Fiser JG, et al. 1986a. Bioconcentration potential (BCP) of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) in terrestrial organisms including humans. 

Chemosphere 15:1495-1502. 

Geyer HJ, Scheunert I, Korte F. 1986b. Bioconcentration potential of organic environmental chemicals in 

humans. Regul Toxicol Pharmacol 6:313-347. 

Geyer HJ, Scheunert I, Korte F. 1987. Correlation between the bioconcentration potential of organic 

environmental chemicals in humans and their n-octanol/water partition coefficients. Chemosphere 

16:239-252. 

*Geyer HJ, Scheunert I, Rapp K, et al. 1990. Correlation between acute toxicity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and total body fat content in mammals. Toxicology 65:97-104. 

*Giavini E, Prati M, Vismara C. 1983. Embryotoxic effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

administered to female rats before mating. Environmental Research 31:105-110. 

*Giavini E, Prati M, Vismaro C. 1982. Rabbit teratology study: 2,3,7,8-Tetrachlorodibenzo-p-dioxin. 

Environ Res 27:74-78. 

*Gierthy JF, Lincoln DW, Gillespie MB, et al. 1987. Suppression of estrogen-regulated extracellular tissue 

plasminogen activator activity of MCF-7 cells by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Cancer Res 

47:6198-6203. 
CDDs 586 

8. REFERENCES 

*Gierthy JF, Lincoln DW, Kampick SJ, et al. 1988. Enhancement of 2- and 16-estradiol hydroxylation in 

MCF-7 human breast cancer cells by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem Biophys Res Comm 

157:515-520. 

*Gierthy JF, Silkworth JB, Tassinari M, et al. 1994. 2,3,7,8-Tetrachlorodibenzo-p-dioxin inhibits 

differentiation of normal diploid rat osteoblasts in vitro. J Cell Biochem 54:231-238. 

*Gilbert P, Saint-Ruf G, Poncelet F, et al. 1980. Genetic effects of chlorinated anilines and azobenzenes on 

Salmonella typhimurium. Arch Environ Contam Toxicol 9:533-541. 

Gilner M, Brittebo EB, Brandt I, et al. 1987. Uptake and specific binding of 2,3,7,8-tetrachlorodibenzop-dioxin in the olfactory mucosa of mice and rats. Cancer Res 47:4150-4159. 

Giri AK. 1986. Mutagenic and genotoxic effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin: A review. Mutat 

Res 168:241-242. 

*Gizzi F, Reginato R, Benfenati E, et al. 1982. PCDDs and PCDFs in emissions from an urban incinerator: 

I. Average peaks and values. Chemosphere 11:577-583. 

Gobas FAPC, Muir DCG, Mackay D. 1988. Dynamics of dietary bioaccumulation and faecal elimination of 

hydrophobic organic chemicals in fish. Chemosphere 17:943-962. 

Gochfeld M, Nygren M, Hansson M, et al. 1989. Correlation of adipose and blood levels of several dioxin 

and dibenzofuran congeners in Agent Orange exposed Vietnam veterans. Chemosphere 18:517-524. 

*Goetz CG, Bolla KI, Rogers SM. 1994. Neurologic health outcomes and Agent Orange: Institute of 

Medicine Report. Neurology 44:801-809. 

*Goldman LR, Hayward DG, Flattery J, et al. 1989. Serum, adipose, and autopsy tissue PCDD and PCDF 

levels in people eating dioxin-contaminated beef and chicken eggs. Chemosphere 19:841-848. 

*Goldman PJ. 1973. [Severest acute chloracne. A mass poisoning by 2,3,7,8-tetrachlorodibenzo-p-dioxin.] 

Hautarzt 24:149-152. (German) 

Goldstein JA, Friesen M, Linder RE, et al. 1977. Effects of pentachlorophenol on hepatic 

drug-metabolizing enzymes and porphyria related to contamination with chlorinated dibenzo-p-dioxins and 

dibenzofurans. Biochem Pharmacol 26:1549-1557. 

Goldstein JA, Friesen M, Scotti TM, et al. 1978. Assessment of the contribution of chlorinated 

dibenzo-p-dioxin and dibenzofurans to hexachlorobenzene toxicity, porphyria, changes in mixed function 

oxygenases, and histopathological changes. Toxicol Appl Pharmacol 46:633-649. 

*Goldstein JA, Hickman P, Bergman H, et al. 1973. Hepatic porphyria induced by 

2,3,7,8-tetrachlorodibenzo-p-dioxin in the mouse. Res Commun Chem Pathol Pharmacol 6:919-928. 

*Goldstein JA, Linko P, Bergman H. 1982. Induction of porphyria in the rat by chronic versus acute 

exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem Pharmacol 31:1607-1613. 
CDDs 587 

8. REFERENCES 

*Goldstein JA, Safe S. 1989. Mechanism of action and structure-activity relationships for the chlorinated 

dibenzo-p-dioxins and related compounds. In: Kimbrough RD, Jensen AA, eds. Halogenated biphenyls, 

naphthalenes, dibenzodioxins and related compounds, 2nd ed. Amsterdam: Elsevier Science Publishers B.V., 

239-293. 

Golor G, Neubert D. 1989. EROD (ethoxyresorufin-o-deethylase) activity in rat liver tissue after 

application of TCDD and OCDD alone and in combination [Abstract]. Naunyn Schmiedeberg's Arch 

Pharmacol 339:10. 

Goodman DG, Sauer RM. 1992. Hepatotoxicity and carcinogenicity in female Sprague-Dawley rats treated 

with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD): A pathology working group reevaluation. Regul Toxicol 

Pharmacol 15(3):245-252. 

*Gordon CJ, Gray LE, Jr., Monteiro-Riviere NA, et al. 1995. Temperature regulation and metabolism in 

rats exposed perinatally to dioxin: Permanent change in regulated body temperature? Toxicol Appl 

Pharmacol 133:172-176. 

*Gorski JR, Muzi G, Weber LW, et al. 1988a. Elevated plasma corticosterone levels and histopathology of 

the adrenals and thymuses in 2,3,7,8-tetrachlorodibenzo-p-dioxin-treated rats. Toxicology 53:19-32. 

*Gorski JR, Muzi G, Weber LWD, et al. 1988b. Some endocrine and morphological aspects of the acute 

toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicologic Pathology 16:313-320. 

*Gorski JR, Rozman K. 1987. Dose-response and time course of hypothyroxinemia and hypoinsulinemia 

and characterization of insulin hypersensitivity in 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-treated rats. 

Toxicology 44:297-307. 

*Gorski JR, Rozman T, Greim H, et al. 1988c. Corticosterone modulates acute toxicity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in male Sprague-Dawley rats. Fundam Appl Toxicol 

11:494-502. 

*Gorski JR, Weber LWD, Rozman K. 1990. Reduced gluconeogenesis in 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD)-treated rats. Arch Toxicol 64:66-71. 

Gorski T. 1981. Presence of polychlorinated dibenzo-p-dioxins in latex nipples. Bull Environ Contam 

Toxicol 27:68-71. 

Gorski T, Konopka L, Brodzki M. 1984. Persistence of some polychlorinated dibenzo-p-dioxins and 

polychlorinated dibenzofurans of pentachlorophenol in human adipose tissue. Rocz Panstw Zakl Hig 

35:297-301. 

Gotz R. 1986. Chlorinated dioxin and dibenzofurans in leachate and sediments of the sanitary landfill in 

Hamburg-Georgswerder. Chemosphere 15:1981-1984. 

*Gough M. 1991. Human exposure from dioxin in soil-a meeting report. J Toxicol Environ Health 

32:205-245. 

Graham M, Hileman FD, Orth RG, et al. 1986. Chlorocarbons in adipose tissue from a Missouri population. 

Chemosphere 15:1595-1600. 
CDDs 588 

8. REFERENCES 

Grahmann R, Claus D, Grehl H, et al. 1993. Electrophysiologic evidence for a toxic polyneuropathy in rats 

after exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). J Neurol Sci 115(1):71-75. 

Grainger J, Reddy VV, Patterson DG Jr. 1988. Analysis of tetra-through octachlorinated dibenzo-p-dioxins 

by gas chromatography/Fourier transform infrared spectroscopy. Chemosphere 19:249-254. 

Gray AP, Cepa SP, Dipinto VM, et al. 1975. Chlorinated dibenzodioxins and dibenzofurans. In: Hemphill 

DD, ed. Trace substances in environmental health-IX. Columbia, MO: University of Missouri, 255-263. 

*Gray J LE, Ostby JS, Kelce WR. 1997b. A dose-response analysis of the reproductive effects of a single 

gestational dose of 2,3,7,8-tetrachlorodibenzo-p-dioxin in male Long Evans Hooded rat offspring. Toxicol 

Appl Pharmacol 146:11-20. 

*Gray J LE, Wolf C, Mann P, et al. 1997a. In utero exposure to low doses of 2,3,7,8-tetrachlorodibenzop-dioxin alters reproductive development of female Long Evans Hooded rat offspring. Toxicol Appl 

Pharmacol 146:237-244. 

*Gray LE, Jr., Kelce WR, Monosson E, et al. 1995. Exposure to TCDD during development permanently 

alters reproductive function in male Long Evans rats and hamsters: Reduced ejaculated and epididymal 

sperm numbers and sex accessory gland weights in offspring with normal androgenic status. Toxicol Appl 

Pharmacol 131:108-118. 

*Gray LE, Jr., Ostby JS. 1995. In utero 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) alters reproductive 

morphology and function in female rate offspring. Tox Appl Pharm 133:285-294. 

Green S, Moreland FS. 1975. Cytogenetic evaluation of several dioxins in the rat. Toxicol Appl Pharmacol 

33:161. 

*Green S, Moreland, Sheu C. 1977. Cytogenic effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on rat bone 

marrow cells. Washington, DC: Food and Drug Administration. FDA by-lines 6:292. 

*Greenlee WF, Dold KM, Irons RD, et al. 1985. Evidence for direct action of 2,3,7,8-TCDD on thymic 

epithelium. Toxicol Appl Pharmacol 79:112-120. 

Greenlee WF, Osborne R, Hudson LG, et al. 1984. Studies on the mechanisms of toxicity of TCDD to 

human epidermis. Banbury Rep. Biological Mechanisms Dioxin Action 18:365-372. 

Greenwald P, Kovasznay B, Collins D, et al. 1984. Sarcomas of soft tissues after Vietnam service. J Natl 

Cancer Inst 73:1107-1109. 

*Grehl H, Grahmann F, Claus D, et al. 1993. Histologic evidence for a toxic polyneuropathy due to 

exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in rats. Acta Neurol Scand 88:354-357. 

Greig JB. 1979. The toxicology of 2,3,7,8-tetrachlorodibenzo-p-dioxin and its structural analogues. Ann 

Occup Hyg 22:411-420. 
CDDs 589 

8. REFERENCES 

*Greig JB. 1984. Differences between skin and liver toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

mice. Banbury Rep Biological Mechanisms Dioxin Action 18:391-397. 

Greig JB, Francis JE, Kay SJ, et al. 1984. Incomplete correlation of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

hepatotoxicity with Ah phenotype in mice. Toxicol Appl Pharmacol 74:17-25. 

*Greig JB, Francis JE, Kay SJ, et al. 1987. Pleiotropic effect of the gene hairless on hepatotoxicity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in mice. Arch Toxicol 60:350-354. 

Greim H. 1990. Toxicological evaluation of emission from modern municipal waste incinerators. 

Chemosphere 20:317-331. 

*Gross ML, Lay JO Jr, Lyon PA, et al. 1984. 2,3,7,8-Tetrachlorodibenzo-p-dioxin levels in adipose tissue 

of Vietnam veterans. Environ Res 33:261-268. 

*Guo YL, Lin CJ, Yao WJ, et al. 1994. Musculoskeletal changes in children prenatally exposed to 

polychlorinated biphenyls and related compounds (Yu-Cheng children). J Toxicol Environ Health 41:83-93. 

*Guzelian PS, Henry CJ, Olin SS. 1992. Similarities and differences between children and adults: 

implications for risk assessment. International Life Sciences Institute Press, Washington, D.C. 

*Ha M-C, Cordier S, Bard D, et al. 1996. Agent orange and the risk of gestational trophoblastic disease in 

Vietnam. Arch Environ Health 51:368-374. 

*Haake JM, Safe S, Mayura K, et al. 1987. Aroclor 1254 as an antagonist of the teratogenicity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Lett 38:299-306. 

Hagenmaier H. 1986. Determination of 2,3,7,8-TCDD in commercial chlorophenols and related products. 

Fresenius Z Anal Chem 325:603-606. 

Hagenmaier H, Brunner H. 1987. Isomer specific analysis of pentachlorophenol and sodium 

pentachlorophenate on 2,3,7,8-substituted PCDD and PCDF at sub-ppb levels. Chemosphere 16:1759-1764. 

Hagenmaier H, Kraft M, Brunner H, et al. 1987. Catalytic effects of fly ash from waste incineration 

facilities on the formation and decomposition of polychlorinated dibenzo-p-dioxins and polychlorinated 

dibenzofurans. Environ Sci Technol 21:1080-1084. 

Hagenmaier H, Kraft M, Jager W, et al. 1986. Comparison of various sampling methods for PCDDs and 

PCDFs in stack gas. Chemosphere 15:1187-1192. 

*Hagenmaier H, Wiesmueller T, Golor G, et al. 1990. Transfer of various polychlorinated 

dibenzo-p-dioxins and dibenzofurans (PCDDs and PCDFs) via placenta and through milk in a marmoset 

monkey. Arch Toxicol 64:601-615. 

Hahn JL, Von Dem Fange HP, Velzy CO. 1986. Effect of boiler operation and RDF feedstock on emissions 

of dioxins and furans from an RDF fired spreader-stoker system in Albany, NY. Chemosphere 

15:1239-1246. 
CDDs 590 

8. REFERENCES 

Hahn JL, Von Dem Fange HP, Westerman G. 1989a. A comparison of ambient and workplace dioxin levels 

from testing in and around modern resource recovery facilities with predicted ground level concentrations of 

dioxins from emission testing with corresponding workplace health risks. Chemosphere 19:629-636. 

*Hahn ME, Goldstein JA, Linko P, et al. 1989b. Interaction of hexachlorobenzene with the receptor for 

2,3,7,8-tetrachlorodibenzo-p-dioxin in vitro and in vivo: Evidence that hexachlorobenzene is a weak Ah 

receptor agonist. Arch Biochem Biophys 270:344-355. 

*Håkansson H, Hanberg A. 1989. The distribution of [14C]-2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) 

and its effect of the vitamin A content in parenchymal and stellate cells of rat liver. J Nutr 119:573-580. 

*Håkansson H, Johansson L, Ahlborg UG, et al. 1989a. Hepatic vitamin A storage in relation to paired 

restriction and to TCDD-treatment. Chemosphere 19:919-920. 

*Håkansson H, Johansson L, Manzoor E, et al. 1989b. 2,3,7,8-Tetrachlorodibenzo-p-dioxin 

(TCDD)-induced alterations in the vitamin A homeostasis and in the 7-ethoxyresorufin O-deethylase 

(EROD)-activity in Sprague-Dawley rats and Hartley guinea pigs. Chemosphere 18:299-305. 

*Håkansson H, Waern F, Ahlborg UG. 1987. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the 

lactating rat on maternal and neonatal vitamin A status. J Nutr 117:580-586. 

*Hallett DJ, Brooksbank MG. 1986. Trends of TCDD and related compounds in the Great Lakes: The 

Lake Ontario ecosystem. Chemosphere 15:1405-1416. 

Hammel KE, Kalyanaraman B, Kirk TK. 1986. Oxidation of polycyclic aromatic hydrocarbons and 

dibenzo(p)dioxins by Phanerochaete Chrysosporium Ligninase. J Biol Chem 261:16948-16952. 

*Hanberg A, Håkansson H, Ahlborg UG. 1989. "ED50" values for TCDD-induced reduction of body 

weight gain, liver enlargement, and thymic atrophy in Hartley guinea pigs, Sprague-Dawley rats, C57BL/6 

mice, and golden Syrian hamsters. Chemosphere 19:813-816. 

*Hanify JA, Metcalf P, Nobbs, CL, et al. 1981. Aerial spraying of 2,4,5-T and human birth malformations: 

An epidemiological investigation. Science 212:349-351. 

*Hankinson O. 1979. Single-step selection of clones of a mouse hepatoma cell line deficient in aryl 

hydrocarbon hydroxylase. Proc Natl Acad Sci USA 76:373-376. 

*Hankinson O. 1995. The aryl hydrocarbon receptor complex. Ann Rev Pharmacol Toxicol 35:307-340. 

Hanson DJ. 1987. Science failing to back up veteran concerns about Agent Orange. Chem Eng News. 

November 9, 7-11, 14. 

Hansson M, Rappe C, Gochfeld M, et al. 1989. Effects of fasting on the blood levels of 2,3,7,8-TCDD and 

related compounds. Chemosphere 18:525-530. 

Haraguchi K, Yamashita T, Shigemori N. 1985. [Sampling and analysis of phosphoric acid triesters in 

ambient air.] Taiki Osen Gakkaishi 20:407-415. (Japanese). 
CDDs 591 

8. REFERENCES 

*Hardell L. 1986. Some aspects of exposure to polychlorinated dibenzo-p-dioxins and dibenzofurans in the 

ecological system. In: Morris CR, Cabral JRP, eds. Hexachlorobenzene: Proceedings of an International 

Symposium 77:243-247. 

*Hardell L, Erikksson M, Degereman A. 1995. Meta-analysis of four Swedish case-control studies on 

exposure to pesticides as risk-factor for soft-tissue sarcoma including the relation to tumour localization and 

histopathological type. Int J Oncol 6:847-851. 

*Hardell L, Eriksson M. 1988. The association between soft tissue sarcomas and exposure to phenoxyacetic 

acids, a new case referent study. Cancer 62:652-656. 

*Hardell L, Eriksson M, Axelson O, et al. 1994. Cancer epidemiology. In: Schecter A, ed. Dioxins and 

health. Plenum Press, NY, 525-547. 

*Hardell L, Eriksson M, Lenner P, et al. 1981. Malignant lymphoma and exposure to chemicals, especially 

organic solvents, chlorophenols and phenoxy acids: A case-control study. Br J Cancer 43:169-176. 

*Hardell L, Sandstrom A. 1979. Case-control study: Soft-tissue sarcomas and exposure to phenoxyacetic 

acids and chlorophenols. Brit J Cancer 39:711-717. 

*Hardin ED, Hayes DA, Shore FL, et al. 1989. A comparison of analytical methods for the analysis of 

chlorinated dioxin/dibenzofuran in pyrolyzed transformer oil samples. Chemosphere 18:1093-1100. 

Harless RL, Lewis RG. 1989. Identification of bromo/chloro dibenzo-p-dioxins and dibenzofurans in ash 

samples. Chemosphere 18:201-208. 

Harless RL, Lewis RG, McDaniel DD, et al. 1992. Evaluation of a sampling and analysis method for 

determination of polyhalogenated dibenzo-p-dioxins and dibenzofurans in ambient air. Chemosphere 

25(7-10):1317-1322. 

*Harless RL, Oswald EO, Lewis RG, et al. 1982. Determination of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

freshwater fish. Chemosphere 2:193-198. 

Harms H, Wittich RM, Sinnwell V, et al. 1990. Transformation of dibenzo-p-dioxin by Pseudomonas sp. 

strain HH69. Appl Environ Microbiol 56:1157-1159. 

*Harper PA, Prokipcak RD, Bush LE, et al. 1991. Detection and characterization of the Ah receptor in the 

human colon adenocarcinoma cell line LS 180. Arch Biochem Biophys 290:27-36. 

*Harris M, Piskorska-Pliszczynska J, Zacharewski T, et al. 1989a. Structure-dependent induction of aryl 

hydrocarbon hydroxylase in human breast cancer cell lines and characterization of the Ah receptor. Cancer 

Res 49:4531-4545. 

*Harris M, Zacharewski T, Astroff B, et al. 1989b. Characterization of 6-methyl-1,3,8-trichlorodibenzofuran (MCDF) as a 2,3,7,8-TCDD antagonist in male rats: Induction of monooxygenases. 

Chemosphere 19:769-772. 
CDDs 592 

8. REFERENCES 

*Harris M, Zacharewski T, Piskorska-Pliszczynska J, et al. 1990. Structure-dependent induction of aryl 

hydrocarbon hydroxylase activity in C57BL/6 mice by 2,3,7,8-tetrachlorodibenzo-p-dioxin and related 

congeners: Mechanistic studies. Toxicol Appl Pharmacol 105:243-253. 

*Hashimoto S, ItoH, Morita M. 1992. Elution of polychlorinated dibenzo-p-dioxins and dibenzofurans 

from coffee filter papers. Chemosphere 25(3):297-305. 

Hass JR, Friesen MD, Parker CE, et al. 1977. Use of negative-ion chemical ionization mass spectrometry in 

the determination of polychlorodibenzo-p-dioxins at the part-per-trillion level in tissue samples. Environ 

Health Perspect 20:246. 

*Hassan MQ, Mohammadpour H, Hermansky SJ, et al. 1987. Comparitive effects of BHA and ascorbic 

acid on the toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in rats. Gen Pharmacol 5:547-550. 

*Hassoun E, D'Argy R, Dencker L, et al. 1984a. Teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

BXD recombinant inbred strains. Toxicol Lett 23:37-42. 

*Hassoun EA. 1987. In vivo and in vitro interactions of TCDD and other ligands of the Ah-receptor: Effect 

on embryonic and fetal tissues. Arch Toxicol 61:145-149. 

*Hattemer-Frey HA, Travis CC. 1989. Comparison of human exposure to dioxin from municipal waste 

incineration and background environmental contamination. Chemosphere 18:643-649. 

*Hauge PM, Belton TJ, Ruppel BE, et al. 1994. 2,3,7,8-TCDD and 2,3,7,8-TCDF in Blue crabs and 

American lobsters from the Hudson-Raritan estuary and the New York Bight. Bull Environ Contam Toxicol 

52(5):734-741. 

Hawker D. 1990. Description of fish bioconcentration factors in terms of solvatochromic parameters. 

Chemosphere 20:467-477. 

Hay A. 1982. Toxicology of dioxins. In: The chemical scythe: Lessons of 2,4,5-T and dioxin. New York, 

London: Plenum Press, 41-47. 

*Hayashi S, Watanabe J, Nakachi K, et al. 1994. Interindividual difference in expression of human Ah 

receptor and related P450 genes. Carcinogenesis 15:801-806. 

*HazDat. 1998. Database. Agency for Toxic Substances and Disease Registry (ATSDR), Atlanta, GA. 

Heath RG, Harless RL, Gross ML, et al. 1986. Determination of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

human milk at the 0.1-10 parts-per-trillion level: Method validation and survey results. Anal Chem 

58:463-468. 

*Hebert C, Gamberg M, Mychasiw L, et al. 1996. Polychlorinated dibenzodioxins, dibenzofurans and 

non-ortho polychlorinated biphenyls in caribou (Rangifer tarandus) from the Canadian Arctic. Sci Total 

Environment 185:195-204. 

*Hebert CD, Birnbaum LS. 1987. The influence of aging on intestinal absorption of TCDD in rats. Toxicol 

Lett 37:47-55. 
CDDs 593 

8. REFERENCES 

*Hebert CD, Harris MW, Elwell MR, et al. 1990. Relative toxicity and tumor-promoting ability of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), 2,3,4,7,8-pentachlorodibenzofuran (PCDF), and 

1,2,3,4,7,8-hexachlorodibenzofuran (HDCF) in hairless mice. Toxicol Appl Pharmacol 102:362-377. 

*Hebert CE, Norstrom RJ, Simon M, et al. 1994. Temporal trends and sources of PCDDs and PCDFs in the 

Great Lakes: Herring gull egg monitoring, 1981-1991. Environmental Science & Technology 

28(7):1268-1277. 

*Heimler I, Trewin AL, Chaffin CL, et al. 1998. Modulation of ovarian follicle maturation and effects on 

apoptotic cell death in Holtzman rats exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in utero and 

lactationally. Reprod Toxicol 12:69-73. 

Heindl A, Hutzinger O. 1986. Search for industrial sources of PCDD/PCDFs: II. Metal chlorides. 

Chemosphere 15:653-658. 

Heindl A, Hutzinger O. 1987. Search for industrial sources of PCDD/PCDF: III. Short-chain chlorinated 

hydrocarbons. Chemosphere 16:1949-1957. 

*Hellou J, Payne JF. 1993. Polychlorinated dibenzo-p-dioxins and dibenzofurans in cod (Gadus morhua) 

from the Northwest Atlantic. Mar Environ Res 36(2):117-128. 

*Henck JW, New MA, Kociba RJ, et al. 1981. 2,3,7,8-tetrachlorodibenzo-p-dioxin: Acute oral toxicity in 

hamsters. Toxicol Appl Pharmacol 59:405-407. 

*Henderson LO, Patterson DG Jr. 1988. Distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin in human 

whole blood and its association with, and extractability from lipoproteins. Bull Environ Contam Toxicol 

40:604-611. 

*Henriksen GL, Ketchum NS, Michalek JE, et al. 1997. Serum dioxin and diabetes mellitus in veterans of 

operation ranch hand. Epidemiology 8:252-258. 

*Henriksen GL, Michalek JE, Swaby JA, et al. 1996. Serum dioxin, testosterone, and gonadotropins in 

veterans of operation ranch hand. Epidemiology 7:352-357. 

*Henry EC, Gasiewicz TA. 1986. Effects of thyroidectomy on the Ah receptor and enzyme inducibility by 

2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat liver. Chem-Biol Interact 59:29-42. 

*Henry EC, Gasiewicz TA. 1987. Changes in thyroid hormones and thyroxine glucuronidation in hamsters 

compared with rats following treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 

89:165-174. 

*Henry EC, Gasiewicz TA. 1993. Transformation of the aryl hydrocarbon receptor to a DNA-binding form 

is accompanied by release of the 90 kDA heat-shock protein and increased affinity for 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem J 294: 95-101. 

*Henry EC, Rucci G, Gasiewicz TA. 1994. Purification to homogeneity of the heteromeric DNA-binding 

form of the Ah receptor from rat liver. Mol Pharmacol 46:1022-1027. 
CDDs 594 

8. REFERENCES 

Henry S, Cramer G, Bolger M, et al. 1991. Exposures and risks of dioxin in the U.S. food supply. 

Proceedings of the 11th International Symposium on Chlorinated Dioxins and Related Compounds, Research 

Triangle Park, NC, 1 -18. 

*Hermansky SJ, Holczlaw TL, Murray WJ, et al. 1988. Biochemical and functional effects of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the heart of female rats. Toxicol Appl Pharmacol 

95:175-184. 

*Hermansky SJ, Mohammadpour HA, Murray WJ, et al. 1987. Effect of thyroidectomy on 

2,3,7,8-tetrachlorodibenzo-p-dioxin-induced lipid peroxidation. Pharmacology 35(6):301-7. 

*Hesso A, Hameila M, Tornaeus J. 1992. Poly chlorinated dioxins, furans and non-ortho polychlorinated 

biphenyls in blood of exposed laboratory personnel. Chemosphere 25:1053-1059. 

*Hilardes RJ, Gray KA, Guzzetta J, et al. 1994. Radiolytic degradation of 2,3,7,8,-TCDD in artificially 

contaminated soils. Environ Sci Techol 28(13):2249-2258. 

*Hippelein M, Kaupp H, Doerr G, et al. 1993. Testing of a sampling system and analytical method for 

determination of semivolatile organic compounds in ambient air. Chemosphere 26(12):2255-2263. 

Hiraoka M, Takizawa Y, Masuda Y, et al. 1987. Investigation of generation of dioxins and related 

compounds from municipal incinerators in Japan. Chemosphere 16:1901-1906. 

Hiremath C, Bayliss D, Bayard S. 1986. Qualitative and quantitative cancer risk assessment of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD). Chemosphere 15:1815-1823. 

*Hirose K, Morita M, Ema M, et al. 1996. cDNA cloning and tissue-specific expression of a novel basic 

helix-loop-helix/PAS factor (Arnt2) with close sequence similarity to the aryl hydrocarbon receptor nuclear 

translocator (Arnt). Mol Cell Biol 16:1706-1713. 

*Hites R, Harless R. 1991. Atmospheric transport and deposition of polychlorinated dibenzo-p-dioxins and 

dibenzofurans. Office of Research and Development, US Environmental Protection Agency, Research 

Triangle Park, NC. EPA/600/3-91/002. 

Hoang TQ, Dai LC, Hong Thom LT. 1989. Effects of geographical conditions, soil movement and other 

variables on the distribution of 2,3,7,8-TCDD levels in adipose tissues from Vietnam: Preliminary 

observations. Chemosphere 18:967-974. 

*Hochstein JR, Aulerich RJ, Bursian SJ. 1988. Acute toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin to 

mink. Arch Environ Contam Toxicol 17:33-37. 

*Hodson PV, McWhirter M, Ralph K, et al. 1992. Effects of bleached draft mill effluent on fish in the St. 

Maurice River, Quebec. Environ Toxicol Chem 11(11):1635-1651. 

*Hoffman R, Stehr-Green PA, Webb KB, et al. 1986. Health effects of long-term exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin. JAMA 255 (15):2031-2038. 
CDDs 595 

8. REFERENCES 

*Holcomb M, Yao C, Safe S. 1988. Biologic and toxic effects of polychlorinated dibenzo-p-dioxin and 

dibenzofuran congeners in the guinea pig: Quantitative structure-activity relationships. Biochem Pharmacol 

37:1535-1539. 

Holden C. 1979. Agent Orange furor continues to build. Science 205:77-772. 

*Holder JW, Menzel HM. 1989. Analysis of 2,3,7,8-TCDD tumor promotion activity and its relationship to 

cancer. Chemosphere 19:861-868. 

*Holladay SD, Lindstrom P, Blaylock BL, et al. 1991. Perinatal thymocyte antigen expression and 

postnatal immune development altered by gestational esposure to tetrachlorodibenzo-p-dioxin (TCDD). 

Teratology 44(4):385-393. 

*Holmstedt B. 1980. Prolegomena to Seveso, Ecclesiastes 1:18. Arch Toxicol 44:211-230. 

*Holsapple MP, Dooley RK, McNerney PJ, et al. 1986a. Direct suppression of antibody-responses by 

chlorinated dibenzodioxins in cultured spleen-cells from (C57BL/6XC3H)F1 and DBA/2 mice. 

Immunopharmacology 12:175-186. 

*Holsapple MP, McCay JA, Barnes DW. 1986b. Immunosuppression without liver induction by subchronic 

exposure to 2,7-dichlorodibenzo-p-dioxin in adult female B6C3F1 mice. Toxicol Appl Pharmacol 

83:445-455. 

Holsapple MP, Snyder NK, Wood SC, et al. 1991. A review of 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced 

changes in immunocompetence: 1991 Update. Toxicology 69(3):219-256. 

*Hong LH, McKinney JD, Luster MI. 1987. Modulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD)-mediated myelotoxicity by thyroid hormones. Biochem Pharmacol 36:1361-1365. 

*Hong R, Taylor K, Abonour R. 1989. Immune abnormalities associated with chronic TCDD exposure in 

rhesus. Chemosphere 18:313-320. 

*Hooiveld M, Heederil DJJ, Kogevinas M, et al. 1998. Second follow-up of a Dutch cohort occupationally 

exposed to phenoxy herbicides, chlorophenols, and contaminants. Am J Epidemiol 147:891-901. 

Horaguchi M, Ogawa H, Ose K, et al. 1989. PCDDs and PCDFs from the MSW incinerator. Chemosphere 

18:1785-1797. 

*Hortsmann M, McLachlan M. 1994. Textiles as a source of CDD/CDF in human skin and sewage sludge. 

Environ Science and Pollution Research 1(1):15-20. 

*House RV, Lauer LD, Murray MJ. 1990. Examination of immune parameters and host resistance 

mechanisms in B6C3F1 mice following adult exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. J Toxicol 

Environ Health 31:203-215. 

Hryhorczuk DO, Orris P, Kominsky JR, et al. 1986. PCB, PCDF, and PCDD exposure following a 

transformer fire: Chicago. Chemosphere 15:1297-1303. 
CDDs 596 

8. REFERENCES 

*HSDB. 1995. Hazardous Substance Databank. National Library of Medicine, Bethesda, MD. March 13, 

1993. 

*Hsu C-C, Yu M-LM, Chen Y-CJ, et al. 1994. The Yu-Cheng rice oil poisoning incident. In: Schecter A, 

ed. Dioxins and health, New York , NY: Plenum Press, 661-684. 

*Huang ZJ, Edery I, Rosbash M. 1993. PAS is a dimerization domain common to Drosophila Period and 

several transcription factors. Nature 364:256-262. 

*Hudson LG, Toscano WA Jr, Greenlee WF. 1985. Regulation of epidermal growth factor binding in 

human keratinocyte cell line by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 77:251-259. 

Huetter R, Philippi M. 1982. Studies on microbial metabolism of TCDD under laboratory conditions. 

Pergamon Ser Environ Sci 5:87-93. 

Huff JE, Moore JA, Saracci R, et al. 1980. Long-term hazards of polychlorinated dibenzodioxins and 

polychlorinated dibenzofurans. Environ Health Perspect 36:221-240. 

Huff JE, Salmon AG, Hooper NK, et al. 1991. Long-term carcinogenesis studies on 

2,3,7,8-tetrachlorodibenzo-p-dioxin and hexachlorodibenzo-p-dioxins. Cell Biol Toxicol 7(1):67-94. 

Huff JE, Wassom JS. 1973. Perspective on chlorinated dibenzodioxins and dibenzofurans. Environ Health 

Perspect 5:283-313. 

*Huisman M, Koopman-Esseboom C, Fidler V, et al. 1995a. Perinatal exposure to polychlorinated 

biphenyls and dioixins and its effect on neaonatal neurological developement. Early Hum Develop 

41:111-127. 

*Huisman M, Koopman-Esseboom C, Lanting CI, et al. 1995b. Neurological condition in 18-month-old 

children perinatally exposed to polychlorinated biphenyls and dioxins. Early Hum Dev 43:165-176. 

*Hulster A, Marschner H. 1993. Transfer of PCDD-PCDF from contaminated soils to food and fodder crop 

plants. Twelfth international symposium on chlorinated dioxins and related compounds. Tampere, Finland. 

Chemosphere 27(1-3):439-446. 

*Hulster A, Muller J, Marschner H. 1994. Soil-Plant trasfer of Polychlorinated dibenzo-p-diozins to 

vegetables of the cucumber family (Curcurbitaceae) Envir Sci Technol 28(6):1110-1115. 

Hung LS, Ingram LL Jr. 1990. Effects of solvents on the photodegradation rates of 

octachlorodibenzo-p-dioxin. Bull Environ Contam Toxicol 44:380-386. 

Hungerford CM. 1988. Comments on behalf of Syntex Agribusiness Inc., in response to Draft 

Toxicological Profile for 2,3,7,8-Tetrachlorodibenzo-p-dioxin. February 11, 1988. 

*Hunt GT, Maisel BE. 1992. Atmospheric concentrations of PCDDs/PCDFs in southern California. J Air 

Wast Manage Assoc 42(5):672-680. 
CDDs 597 

8. REFERENCES 

*Huntley SL, Iannuzzi TJ, Avantaggio JD, et al. 1997. Combined sewer overflows (CSOs) as sources of 

sediment contamination in the lower Passaic River, New Jersey : Polychlorinated dibenzo-p-dioxins, 

polychlorinated dibenzofurans, and polychlorinated biphenyls. Chemosphere 34(2):233-250. 

*Hussain S, Ehrenberg L, Lofroth G, et al. 1972. Mutagenic effects of TCDD on bacterial systems. Ambio 

1:32-33. 

*Hutzinger O, Blumich MJ, Berg M, et al. 1985. Sources and fate of PCDDs and PCDFs: An overview. 

Chemosphere 14:581-600. 

Hutzinger O, Fiedler H. 1993. Form source to exposure: Some open questions. Chemosphere 27(ISS 

1-3):121-129. 

Hutzinger O, Olie K, Lustenhouwer JWA, et al. 1981. Polychlorinated dibenzo-p-dioxins and 

dibenzofurans: A bioanalytical approach. Chemosphere 10:19-25. 

*Huuskonen H, Undila M, Pohjanvirta R, et al. 1994. Developmental toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the most TCDD-resistant and -susceptible rat strains. Toxicol Appl Pharmacol 

124(2):174-80. 

*IARC. 1977. IARC monographs on the evaluation of the carcinogenic risk of chemicals to man: Some 

fumigants, the herbicides 2,4-D and 2,4,5-T, chlorinated dibenzodioxins and miscellaneous industrial 

chemicals. Lyon, France: World Health Organization, International Agency for Research on Cancer. IARC 

Monogr 15:41-102. 

IARC. 1982. IARC monographs on the evaluation of the carcinogenic risk of chemicals to humans: 

Chemicals, industrial processes, and industries associated with cancer in humans. Lyon, France: World 

Health Organization, International Agency for Research on Cancer. IARC Monogr Suppl 4:238-243. 

*IARC. 1986a. IARC mongraphs on the evaluation of carcingenic risks to humans. Occupational 

exposures to chlorophenols. Lyon, France: World Health Organization, International Agency for Research 

on Cancer. IARC Monogr 41:319-337. 

*IARC. 1986b. IARC mongraphs on the evaluation of carcingenic risks to humans. Occupational 

exposures to chlorophenoxy herbicides. Lyon, France: World Health Organization, International Agency 

for Research on Cancer. IARC Monogr 41:357-377. 

*IARC. 1987. IARC mongraphs on the evaluation of carcingenic risks to humans. Overall evaluations of 

carcinogenicity: an updating of IARC monographs volumes 1 to 42. Lyon, France: World Health 

Organization, International Agnecy for Research on Cancer. IARC Monogr Suppl F. 

*IARC. 1987. IInternational Agency for Research on Cancer. IARC monographs on the evaluation of 

carcinogenic risks to humans. Overall evaluations of carcinogenicity: an updating of IARC monographs 

volumes 1 to 41. Supplement 7. Lyone, France. 1987. International Agency for Research on Cancer. 

World Health Organization. 

*IARC. 1997. International Agency for Research on Cancer. IARC Monographs on the Evaluation of 

Carcinogenic Risks to Humans: Polychorinated Dibenzo-para-dioxins and Polychlorinated Dibenzofurans 

(vol. 69). Summary Statements. http://www.iarc.fr/publications/vol69.htm. May 8, 1997. 
 

CDDs 598 

8. REFERENCES 

*IARC. 1997. International Agency for Research on Cancer. Monographs database. Polychlorinated 

Dibenzo-para-Dioxins, htttp://www.iarc.fr. and htttp://193.51.164.11/htdocs/monographs/Vol 

69/dioxin.html. International Agency for Research on Cancer. World Health Organization. 

ICRP (International Commission on Radiological Protection). 1981. Report of the Task Group on Reference 

Man. ICRP Publication No. 23. Pergamon Press. 

*Ideo G, Bellati G, Bellobuono A, et al. 1982. Increased urinary D-glucaric acid excretion by children 

living in an area polluted with 2,3,7,8-tetrachlorodibenzo-para-dioxin (TCDD). Clin Chim Acta 

120:273-283. 

*Iida T, Hirakawa H, Matsueda T, et al. 1991. [Therapeutic trial for promotion of fecal excretion of PCDFs 

and PCBs by the adminstration of cholestyramine in Yusho patients.] Fukuoka Ishi 82:317-325. (Japanese) 

*Ilsen A, Briet JM, Koppe JG, et al. 1996. Signs of enhanced neuromotor maturation in children due to 

perinatal load with background levels of dioxins. Chemosphere 33:1317-1326. 

*Imamura M, Tung T-C. 1984. A trial of fasting cure for PC-poisoned patients in Taiwan. Am J Ind Med 

5:147-153. 

*Ireland RC, Li S-Y, Dougherty JJ. 1995. The DNA binding of purified Ah receptor heterodimer is 

regulated by redox conditions. Arch Biochem Biophys 319:470-480. 

*IRIS. 1997. Integrated risk informaiton system. Hexachloradibenzo-p-dioxin, mixture (HxCDD). 

CASRN 1940-74-3. U. S. Environmental Protection Agency. Integrated risk informaiton system. 

http://www.epa.gov/ngispgm3/iris/subst/0166.htm. March 1, 1997. 

*Isensee AR. 1978. Bioaccumulation of 2,3,7,8-tetrachlorodibenzo-para-dioxin. Ecol Bull. (Stockholm) 

27:255-262. 

Isensee AR, Jones GE. 1971. Absorption and translocation of root and foliage applied 2,4-dichlorophenol, 

2,7-dichlorodibenzo-p-dioxin, and 2,3,7,8-tetrachlorodibenzo-p-dioxin. J Agr Food Chem 19:1210-1214. 

Isensee AR, Jones GE. 1975. Distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in aquatic model 

ecosystems. Environ Sci Technol 9:668-672. 

Isnard P, Lambert S. 1988. Estimating bioconcentration factors from octanol-water partition coefficient and 

aqueous solubility. Chemosphere 17:21-34. 

Isnard P, Lambert S. 1989. Aqueous solubility and n-octanol/water partition coefficient correlations. 

Chemosphere 18:1837-1853. 

*Israel DI, Estolano MG, Galeazzi DR, et al. 1985. Superinduction of cytochrome Pi-450 gene transcription 

by inhibition of protein synthesis in wild type and variant mouse hepatoma cells. J Biol Chem 

260:5648-5653. 

*James WH. 1997. The sex ratio of offspring sired by men exposed to wood preservatives contaminated by 

dioxin. Scand J Work Environ Health 23:69. 
CDDs 599 

8. REFERENCES 

*Jansing P-J, Korff R. 1994. Blood levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin and -globulins in a 

follow-up investigation of employees with chloracne. J Dermatol Sci 8:91-95. 

*Jarman WM, Burns SA, Chang RR, et al. 1993. Determination of PCDDS, PCDFS, and PCBS in 

California Peregrine falcons (FALCO PEREGRINUS) and their eggs. Environ Toxic Chem 12(1):105-114. 

*Jasinski JS. 1989. Multiresidue procedures for the determination of chlorinated dibenzodioxins and 

dibenzofurans in a variety of foods using capillary gas chromatography-electron-capture detection. J 

Chromatogr 478:349-367. 

*Jennings AM, Wild G, Ward JD, et al. 1988. Immunological abnormalities 17 years after accidental 

exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Br J Ind Med 45:701-704. 

Jensen AA. 1987. Polychlorobiphenyls (PCBs), polychlorodibenzo-p-dioxins (PCDDs) and 

polychlorodibenzofurans (PCDFs) in human milk, blood and adipose tissue. Sci Total Environ 64:259-293. 

*Jensen DJ, Hummel RA, Mahle NH, et al. 1981. A residue study on beef cattle consuming 

2,3,7,8-tetrachlorodibenzo-p-dioxin. J Agric Food Chem 29:265-268. 

*Jirasek L, Kalensky J, Kubec K, et al. 1976. [Chloracne, porphyria cutanea tarda and other intoxication by 

herbicides.] Hautarzt 27:328-333. (German) 

*Jobb B, Uza M, Hunsinger R, et al. 1990. A survey of drinking water supplies in the Province of Ontario 

for dioxins and furans. Chemosphere 20(10-12):1553-1558. 

*Johanson CE. 1980. Permeability and vascularity of the developing brain: Cerebellum vs cerebral cortex. 

Brain Research 190:3-16. 

*Johnson ES. 1993. Important aspects of the evidence for TCDD carcinogenicity in man. Environ Health 

Perspect 99:383-390. 

*Johnson KL, Cummings AM, Birnbaum LS. 1997. Promotion of endometriosis in mice by polychlorinated 

dibenzo-p-dioxins, dibenzofurans, and biphenyls. Environ Health Perspect 105:750-755. 

*Johnson L, Dickerson R, Safe SH, et al. 1992. Reduced Leydig cell volume and function in adult rats 

exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin without a significant effect on spermatogenesis. Toxicology 

76(2):103-118. 

Johnson RL, Gehring PJ, Kociba RJ, et al. 1973. Chlorinated dibenzodioxins and pentachlorophenol. 

Environ Health Perspect 5:171-175. 

Jones D, Safe E, Morcum E, et al. 1989. Bioavailability of grain and soil-borne tritiated 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) administered to lactating Holstein cows. Chemosphere 

18:1257-1263. 

*Jones EL, Krizek HA. 1962. A technique for testing acnegenic potency in rabbits applied to the potent 

acnegen 2,3,7,8-tetrachlorodibenzo-p-dioxin. J Invest Dermatol 39:511-517. 
CDDs 600 

8. REFERENCES 

*Jones K. 1995. Diesel engine emissions and the link to human dioxin exposure. Dioxin 95 Secretariat, 

Edmonton AB Canada Organohalogen Compounds 24:69-74. 

Jones KC, Bennett BG. 1989. Human exposure to environmental polychlorinated dibenzo-p-dioxins and 

dibenzofurans: An exposure commitment assessment for 2,3,7,8-TCDD. Sci Total Environ 78:99-116. 

*Jones KG, Sweeney GD. 1977. Association between induction of aryl hydrocarbon hydroxylase and 

depression of uroporphyrinogen decarboxylase activity. Res Commun Chem Pathol Pharmacol 17:631-637. 

*Jones KG, Sweeney GD. 1980. Dependence of the porphyrogenic effect of 2,3,7,8-tetrachlorodibenzop-dioxin upon inheritance of aryl hydrocarbon hydroxylase responsiveness. Toxicol Appl Pharmacol 

54:42-49. 

*Jones KW, Whitlock JP Jr. 1990. Functional analysis of the transcriptional promoter for the CYP1A1 

gene. Mol Cell Biol 10:5098-5105. 

*Jones PBC, Durrin LK, Galeazzi DR, et al. 1986. Control of cytochrome Pi-450 gene expression: analysis 

of a dioxin-responsive enhancer system. Proc Natl Acad Sci USA 83:2802-2806. 

*Jones PD, Giesy JP, Newsted JL, et al. 1994. Accumulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

equivalents by double-crested cormorant (phalacrorax auritus, Pelicaniformes) chicks in the North American 

Great Lakes. Ecotoxicol Environ Safety 27:192-209. 

Jones RE, Chelsky M. 1986. Further discussion concerning porphyria cutanea tarda and TCDD exposure. 

Arch Environ Health 41:100-103. 

Jonsson P, Rappe C, Kjeller L-O. 1993. Pulp-mill related polychlorinated organic compounds in Baltic Sea 

sediments. Ambio 22(1):37-43. 

*Josephson J. 1983. Chlorinated dioxins and furans in the environment. Environ Sci Technol 

17:124A-128A. 

*Jung D, Berg PA, Edler L, et al. 1998. Immunologic findings in workers formerly exposed to 

2,3,7,8-tetrachlorodibenzo-p-dioxin and its congeners. Environ Health Perspect 106(Suppl2):689-695. 

*Kahn PC, Gochfeld M, Nygren M, et al. 1988. Dioxins and dibenzofurans in blood and adipose tissue of 

Agent Orange-exposed Vietnam veterans and matched controls. JAMA 259:1661-1667. 

*Kahr G, Pauliny I, Fercher E. 1990. Destruction of dioxins and furans adsorbed on fly ash in a rotary kiln 

furnace. Chemosphere 20:1855-1858. 

*Kalliokoski P, Kauppinen T. 1990. Complex chlorinated hydrocarbons: Occupational exposure in the 

sawmill. IARC Sci Publ 104:390-396. 

Kaminsky LS, DeCaprio AP, Gierthy JF, et al. 1985. The role of environmental matrices and experimental 

vehicles in chlorinated dibenzodioxin and dibenzofuran toxicity. Chemosphere 14:685-695. 
CDDs 601 

8. REFERENCES 

*Kang HK, Weatherbee L, Breslin P, et al. 1987. Soft tissue sarcoma and military service in Vietnam: A 

case comparison group analysis of hospital patients. J Occup Med 28:1215-1218. 

*Kapila S, Yanders AF, Orazio CE, et al. 1989. Field and laboratory studies on the movement and fate of 

tetrachlorodibenzo-p-dioxin in soil. Chemosphere 18:1297-1304. 

Karasek FW, Dickson LC. 1987. Model studies of polychlorinated dibenzo-p-dioxin formation during 

municipal refuse incineration. Science 237:754-756. 

Karasek FW, Tong HY. 1985. Semi-preparative high-performance liquid chromatographic analysis of 

complex organic mixtures. J Chromatogr 332:169-179. 

*Kawajiri K, Nakachi K, Imai K, et al. 1990. Identification of genetically high risk individuals to lung 

cancer by DNA polymorphisms of the cytochrome P450IA1 gene. FEBS Lett 263:131-133. 

*Kaye CI, Rao S, Simpson SJ, et al. 1985. Evaluation of chromosomal damage in males exposed to Agent 

Orange and their families. J Craniofac Genet Dev Bio 1:259-265. 

*KBWP. 1988. Kansas Board of Water Protection. Topeka, KS. 

*Kearney PC, Isensee AR, Helling CS, et al. 1971. Environmental significance of chlorodioxins. In: 

Chlorodioxins - origin and fate. Advances in Chemistry Series 120:105-111. 

Kedderis LB, Andersen ME, Birnbaum LS. 1993b. Effect of dose, time and pretreatment on the biliary 

excretion and tissue distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Fundam Appl Toxicol 

21(4):405-411. 

Kedderis LB, Diliberto JJ, Jackson JA, et al. 1992. Effects of dose and route of exposure on dioxin 

disposition. Chemosphere 25(1-2):7-10. 

Kedderis LB, Diliberto JJ, Linko P, et al. 1991a. Disposition of 2,3,7,8-tetrabromodibenzo-p-dioxin and 

2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat: Biliary excretion and induction of cytochromes CYP1A1 and 

CYP1A2. Toxicol Appl Pharmacol 111:163-172. 

Kedderis LB, Diliberto JJ, Linko P, et al. 1991b. Disposition and excretion in intravenous 

2,3,7,8-tetrabromodibenzo-p-dioxin (TBDD) in rats. Toxicol Appl Pharmacol 108:397-406. 

Kedderis LB, Mills JJ, Anderson ME, et al. 1993a. A physiologically-based pharmacokinetic model for 

2,3,7,8-tetrabromo-dibenzo-p-dioxin (TBDD) in the rat: Tissue distribution and CYP1A induction. Toxicol 

Appl Pharmacol 121:87-98. 

Keenan RE, Sullivan MJ. 1989. Assessing potential health risks of dioxin in paper products. Environ Sci 

Technol 23:643-644. 

*Kelling CK, Christian BJ, Inhorn SL, et al. 1985. Hypophagia-induced weight loss in mice, rats, and 

guinea pigs treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Fundam Appl Toxicol 5:700-712. 

*Kelling CK, Menahan LA, Peterson RE. 1987. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin treatment on 

mechanical function of the rat heart. Toxicol Appl Pharmacol 91:497-501. 
CDDs 602 

8. REFERENCES 

*Kelly TJ, Czuczwa JM, Sticksel PR, et al. 1991. Atmospheric and tributary inputs of toxic substances to 

Lake Erie. J Gt Lakes Res 17(4):504-516. 

*Kennedy SW, Lorenzen A, Norstrom RJ. 1996. Chicken embryo hepatocyte bioassay for measuring 

cytochrome P4501A-based 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalent concentrations in environmental 

samples. Environ Sci Technol 30:706-715. 

*Kerger B, Corbett G, El-Sururi S, et al. 1995. Validating dermal exposure assessment techniques for 

dioxin using body burden data and pharmacokinetic modeling. Organohalogen Compounds 25:137-141. 

*Kerkvliet NI. 1994. Immunotoxicology of dioxins and related chemicals. In: Schecter A, ed. Dioxins and 

health. New York, NY: Plenum Press, 199-225. 

*Kerkvliet NI. 1995. Immunological effects of chlorinated dibenzo-p-dioxins. Environ Health Perspect 

103:47-53. 

*Kerkvliet NI, Baecher-Steppan L, Shepherd DM, et al. 1996. Inhibition of TC-1 cytokine production, 

effector cytotoxic T lymphocyte development and alloantibody production by 

2,3,7,8-tetrachlorodibenzo-p-dioxin. J Immunol 157:2310-2319. 

*Kerkvliet NI, Baecher-Steppan L, Smith BB, et al. 1990b. Role of the Ah locus in suppression of 

cytotoxic T lymphocyte activity by halogenated aromatic hydrocarbons (PCBs and TCDD): 

Structure-activity relationships and effects in C57Bl/6 mice congenic at the Ah locus. Fundam Appl Toxicol 

14:532-541. 

Kerkvliet NI, Brauner JA. 1984. Humoral immune suppression by contaminants of technical grade 

pentachlorophenol with special reference to 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin. J Leukoc Biol 

36:392. 

Kerkvliet NI, Brauner JA. 1987. Mechanisms of 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin 

(HpCDD)-induced humoral immune suppression: Evidence of primary defect in T-cell regulation. Toxicol 

Appl Pharmacol 87:18-31. 

Kerkvliet NI, Brauner JA, Matlock JP. 1985. Humoral immunotoxicity of polychlorinated diphenyl ethers, 

phenoxyphenols, dioxins and furans present as contaminants of technical grade pentachlorophenol. 

Toxicology 36:307-324. 

*Kerkvliet NI, Steppan LB, Brauner JA, et al. 1990a. Influence of the Ah locus on the humoral 

immunotoxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD): evidence for Ah receptor-dependent and 

Ah receptor independent mechanisms of immunosuppression. Toxicol Appl Pharmacol 105:26-36. 

Kew GA, Schaum JL, White P, et al. 1989. Review of plant uptake of 2,3,7,8-TCDD from soil and 

potential influence of bioavailability. Chemosphere 18:1313-1318. 

*Khera KS, Ruddick JA. 1973. Polychlorodibenzo-p-dioxins: Perinatal effects and the dominant lethal test 

in Wistar rats. Advances in Chemistry Series 12C:70-84. 

*Kieatiwong S, Nguyen LV, Herbert VR, et al. 1990. Photolysis of chlorinated dioxins in organic solvents 

and on soils. Environ Sci Technol 24:1575-1580. 
CDDs 603 

8. REFERENCES 

Kimbrough RD. 1972. Toxicity of chlorinated hydrocarbons and related compounds. A review including 

chlorinated dibenzodioxins and chlorinated dibenzofurans. Arch Environ Health 25:125-131. 

Kimbrough RD. 1974. The toxicity of polychlorinated polycyclic compounds and related chemicals. CRC 

Crit Rev Toxicol 2:445-498. 

*Kimbrough RD, Carter CD, Liddle JA, et al. 1977. Epidemiology and pathology of a tetrachlorodibenzodioxin poisoning episode. Arch Environ Health 77-85. 

*Kimbrough RD, Falk H, Stehr P, et al. 1984. Health implications of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) contamination of residential soil. J Toxicol Environ Health 14:47-93. 

Kimbrough RD, Houk VN. 1987. Effects of chlorinated dibenzodioxins. ACS Symposium Series 

338:68-79. 

King ME, Shefner AM, Bates RR. 1973. Carcinogenesis bioassay of chlorinated dibenzodioxins and related 

chemicals. Environ Health Perspect 5:163-170. 

King TL, Uthe JF, Musial CJ. 1993. Rapid screening of fish tissue for polychlorinated dibenzo-p-dioxins 

and dibenzofurans. Analyst 118(10):1269-1275. 

*Kissel JC, Robarge GM. 1988. Assessing the elimination of 2,3,7,8-TCDD from humans with a 

physiologically based pharmacokinetic model. Chemosphere 17:2017-2027. 

Kitunen VH, Salkinoja-Salonen MS. 1989. Occurrence of PCDDs and PCDFs in pulp and board products. 

Chemosphere 19:721-726. 

*Klawans HL. 1987. Dystonia and tremor following exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Movement Disorders 2:255-261. 

*Klee AJ. 1988. New approaches to the decontamination of hazardous wastes. In: Abbou R, ed. 

Hazardous waste: Detection, control, treatment. Amsterdam, The Netherlands: Elsevier Science Publishers 

B.V., 721-729. 

*Kleeman JM, Moore RW, Peterson RE. 1990. Inhibition of testicular steroidogenesis in 

2,3,7,8-tetrachlorodibenzo-p-dioxin-treated rats: Evidence that the key lesion occurs prior to or during 

pregnenolone formation. Toxicol Appl Pharmacol 106:112-125. 

*Kleeman JM, Olson JR, Chen SM, et al. 1986a. Metabolism and disposition of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in rainbow trout. Toxicol Appl Pharmacol 83:391-401. 

*Kleeman JM, Olson JR, Chen SM, et al. 1986b. 2,3,7,8-Tetrachlorodibenzo-p-dioxin metabolism and 

disposition in yellow perch. Toxicol Appl Pharmacol 83:402-411. 

Kleinveld AJ, Verhoeve P, Nielen MWF. 1989. Evaluation of clean-up methods for the determination of 

chlorinated aromatic hydrocarbons in fly ash by capillary gas chromatography and mass selective detection. 

Chemosphere 18:1401-1412. 

Kleopfer RD. 1985. 2,3,7,8-TCDD contamination in Missouri. Chemosphere 14:739-744. 
CDDs 604 

8. REFERENCES 

Klicius R, Finkelstein A, Hay D. 1988. The national incinerator testing and evaluation program: Air 

pollution control technology assessment results. Hazard Waste Hazard Mater 5:53-64. 

Knutsen AP. 1984. Immunologic effects of TCDD exposure in humans. Bull Environ Contam Toxicol 

33:673-681. 

*Knutsen AP, Roodman ST, Evans RG, et al. 1987. Immune studies in dioxin-exposed Missouri residents: 

Quail Run. Bull Environ Contam Toxicol 39:481-489. 

Knutson JC, Poland A. 1980. Keratinization of mouse teratoma cell line XB produced by 

2,3,7,8-tetrachlorodibenzo-p-dioxin: an in vitro model of toxicity. Cell 22:27-36. 

*Knutson JC, Poland A. 1982. Response of murine epidermis to 2,3,7,8-tetrachlorodibenzo-p-dioxin: 

interaction of the Ah and hr loci. Cell 30:225-234. 

*Ko HP, Okino ST, Ma Q, et al. 1996. Dioxin-induced CYP1A1 transcription in vivo: The aromatic 

hydrocarbon receptor mediates transactivation, enhancer-promoter communication, and changes in 

chromatin structure. Mol Cell Biol 16:430-436. 

*Kociba RJ, Keyes DG, Beyer JE, et al. 1978a. Results of a two-year chronic toxicity and oncogenicity 

study of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rats. Toxicol Appl Pharmacol 46:279-303. 

Kociba RJ, Keyes DG, Beyer JE, et al. 1978b. Toxicologic studies of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) in rats. Toxicol Occup Med (De Toxicol Environ Sci) 4:281-287. 

Kociba RJ, Keyes DG, Beyer JE, et al. 1979. Long-term toxicologic studies of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in laboratory animals. Ann NY Acad Sci 320:397-404. 

Kociba RJ, Schwetz BA. 1982. Toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Drug Metab Rev 

13:387-406. 

*Koester CJ, Harless RL, Hites RA. 1992. Comparative analysis of dioxins and furans in ambient air by 

high resolution and electron capture mass spectrometry. Chemosphere 24(4):421-426. 

Koester CJ, Hites RA. 1988. Calculated physical properties of polychlorinated dibenzo-p-dioxins and 

dibenzofurans. Chemosphere 17:2355-2362. 

*Koester CJ, Hites RA. 1992. Photo degradation of polychlorinated dioxins and dibenzofurans adsorbed to 

fly ash. Environ Sci Technol 26(3):502-507. 

Kogan MD, Clapp RW. 1985. Mortality among Vietnam veterans in Massachusetts, 1972-1983. Boston, 

MA: Massachusetts Office of the Commissioner of Veterans Services, Agent Orange Program, 

Massachusetts Department of Public Health, Division of Health Statistics and Research. 

*Kogevinas M, Becher H, Benn T, et al. 1997. Cancer mortality in workers exposed to phenoxy herbicides, 

chlorophenols, and dioxins. Am J Epidemiol 145:1061-1075. 
CDDs 605 

8. REFERENCES 

*Kogevinas M, Kauppinen T, Winkelmann R, et al. 1995. Soft tissue sarcoma and non-Hodgkin's 

lymphoma in workers exposed to phenoxy herbicides, chlorophenols, and dioxins: two nested case-control 

studies. Epidemiol 6:396-402. 

*Kogevinas M, Saracci R, Winkelmann R, et al. 1993. Cancer incidence and mortality in women 

occupationally exposed to chlorophenoxy herbicides, chlorophenols, and dioxins. Cancer Causes and 

Control 4:547-553. 

*Kohn MC, Lucier GW, Clark GC, et al. 1993. A mechanistic model of effects of dioxin on gene expression 

in the rat liver. Toxicol Appl Pharmacol 120:138-154. 

*Kohn MC, Sewall CH, Lucier GW, et al. 1996. A mechanistic model of effects of dioxin on thyroid 

hormones in the rat. Toxicol Appl Pharmacol 165:29-48. 

*Kolenda J, Gass H, Wilken M, et al. 1994. Determination and reduction of PCDD/F emissions from 

wood burning facilities. Chemosphere 29(9-11):1927-1938. 

Kominsky JR, Kwoka CD. 1989. Background concentrations of polychlorinated dibenzofurans (PCDFs) 

and polychlorinated dibenzo-p-dioxins (PCDDs) in office buildings in Boston, Massachusetts. Chemosphere 

18:599-608. 

Kominsky JR, Slade WT, Mooring R. 1988. Background concentrations of polychlorinated biphenyls, 

polychlorinated dibenzo-p-dioxins, and polychlorinated dibenzofurans in office buildings in Santa Fe, New 

Mexico. Electr Power Res Inst (Rep), 7-15 - 7-26. 

*Komori M, Nishio K, Kitada M et al. 1990. Fetus-specific expression of a form of cytochrome P-450 in 

human liver. Biochemistry 29:4430-4433. 

*Koope JG, Pluim HJ, Olie K. 1991. Breast milk, dioxins and the possible effects on health of newborn 

infants. Sci Tot Environ 106:33-41. 

*Koopman-Esseboom C, Morse DC, Weisglas-Kuperus N, et al. 1994. Effects of dioxins and 

polychlorinated biphenyls on thyroid hormaone status of pregant women and their infants. Pediatr Res 

36:468-473. 

*Koppe JG. 1995. Nutrition and breast-feeding. European Journal of Obstetrics & Gynecology and 

Reproductive 61(1):73-78. 

Korcher CW, Mahle NH, Hummels RA, et al. 1978. A search for the presence of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in beef fat. Bull Environ Contam Toxicol 19:229-236. 

Korfmacher WA, Hansen EBJR, Rowland KL. 1986. Use of bullfrogs (Rana catesbeiana) as biological 

markers for 2,3,7,8-tetrachlorodibenzo-p-dioxin contamination in the environment. Sci Total Environ 

57:257-262. 

*Korfmacher WA, Rushing LG, Nestorick DM, et al. 1985. Analysis of dust and surface swab samples for 

octachlorodibenzo-p-dioxin and heptachlorodibenzo-p-dioxins by fused silica capillary GC with EC 

detection. Journal of High Resolution Chromatography and Chromatography Communications 8:12-19. 
CDDs 606 

8. REFERENCES 

Korte M, Renscher C, Luebbe U. 1989. TCDD liver concentrations and biological activity in cross-fostered 

rat offspring after maternal treatment. Naunyn-Schmiedeberg's Abstract R10. Arch Pharmacol 339 Suppl. 

*Koshakji RP, Harbison RD, Bush MT. 1984. Studies on the metabolic fate of [14C]2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the mouse. Toxicol Appl Pharmacol 73:69-77. 

Koshioka M, Kanazawa J, Lizuka H, et al. 1987. Photodegradation of decachlorobiphenyl. Bull Environ 

Contam Toxicol 38:409-415. 

Koshioka M, Yamada T, Kanazawa J, et al. 1989a. Photodegradation products of 

tetrachlorodibenzo-p-dioxins under xenon lamp irradiation. J Pesticide Sci 14:175-179. 

*Koshioka M, Yamada T, Kanazawa J, et al. 1989b. Photolysis of tetrachlorodibenzo-p-dioxins. 

Chemosphere 19:681-684. 

*Koshioka M, Yamada T, Kanazawa J, et al. 1989c. Wavelength-dependence of photodegradation of 

tetrachlorodibenzo-p-dioxins. Journal of Pesticide Science 14:181-187. 

*Kouri RE, Ratrie H, Whitmire CE. 1974. Genetic control of susceptibility to 3-methylcholanthreneinduced lung subcutaneous sarcomas. J Cancer 13:714-720. 

*Kouri RE, Rude TH, Joglekar R, et al. 1978. 2,3,7,8-Tetrachlorodibenzo-p-dioxin as cocarcinogen causing 

3-methylcholanthrene-initiated subcutaneous tumors in mice genetically "nonresponsive" at Ah locus. 

Cancer Res 38:2777-2783. 

*Kramer CM, Johnson KW, Dooley RK, et al. 1987. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) 

enhances antibody production and protein kinase activity in murine B cells. Biochem Biophys Res Comm 

145:25-32. 

*Kreuzer PE, Csanady GA, Baur C, et al. 1997. 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and 

congeners in infants. A toxicokinetic model of human lifetime body burden by TCDD with special emphasis 

on its uptake by nutrition. Arch Toxicol 71:383-400. 

*Krishnan K, Andersen ME. 1994. Physiologically-based pharmacokinetic modeling in toxicology. In: 

Hayes W, ed. Principles and methods of toxicology. 3rd edition. New York, NY: Raven Press, Ltd. 

*Krishnan K, Andersen ME. 1994. Physiologically-based pharmacokinetic modeling in toxicology. In: 

Hayes W, ed. Principles and methods of toxicology. 3rd edition. New York, NY: Raven Press, Ltd. 

*Krishnan K, Andersen ME, Clewell HJ III, et al. 1994. Physiologically-based pharmacokinetic modeling 

of chemical mixtures. In: Yang RSA, ed. Toxicology of chemical mixtures. New York, NY: Academic 

Press. 

*Krishnan K, Andersen ME, Clewell HJ III, et al. 1994. Physiologically-based pharmacokinetic modeling 

of chemical mixtures. In: Yang RSA, ed. Toxicology of chemical mixtures. New York, NY: Academic 

Press. 
CDDs 607 

8. REFERENCES 

*Krishnan V, Porter W, Santostefano M, et al. 1995. Molecular mechanism of inhibition of 

estrogen-induced cathepsin D gene expression by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in MCF-7 

cells. Mol Cell Biol 15: 6710-6719. 

*Krishnan V, Safe S. 1993. Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), and 

dibenzofurans (PCFDs) as antiestrogens in MCF-7 human breast cancer cells: Quantitative structure-activity 

relationships. Toxicol Appl Pharmacol 120:55-61. 

*Krowke R. 1986. Studies on distribution and embryotoxicity of different PCDD and PCDF in mice and 

marmosets. Chemosphere 15:2011-2022. 

*Krowke R, Chahoud I, Baumann-Wilschke I, et al. 1989. Pharmacokinetics and biological activity of 

2,3,7,8-tetrachlorodibenzo-p-dioxin: 2. Pharmacokinetics in rats using a loading-dose/maintenance-dose 

regime with high doses. Arch Toxicol 63:356-360. 

*Kuehl DW, Butterworth BC, DeVita WM, et al. 1987a. Environmental contamination by polychlorinated 

dibenzo-p-dioxins and dibenzofurans associated with pulp and paper mill discharge. Biomed Environ Mass 

Spectrom 14:443-447. 

*Kuehl DW, Butterworth BC, Marquis PJ. 1994. A national study of chemical residues in fish: III Study 

results. Chemosphere 29(3):523-535. 

*Kuehl DW, Butterworth BC, McBride A, et al. 1989. Contamination of fish by 

2,3,7,8-tetrachlorodibenzo-p-dioxin: A survey of fish from major watersheds in the United States. 

Chemosphere 18:1997-2014. 

*Kuehl DW, Cook PM, Batterman AR. 1985. Studies on the bioavailability of 2,3,7,8-TCDD from 

municipal incinerator fly ash to freshwater fish. Chemosphere 14:871-872. 

*Kuehl DW, Cook PM, Batterman AR, et al. 1987b. Bioavailability of polychlorinated dibenzo-p-dioxins 

and dibenzofurans from contaminated Wisconsin River sediment to carp. Chemosphere 16:667-679. 

*Kuehl DW, Cook PM, Batterman AR, et al. 1987c. Isomer dependent bioavailability of polychlorinated 

dibenzofurans from municipal incinerator fly ash to carp. Chemosphere 16:657-666. 

*Kuwata K, Yamashita Y, Nakashima S, et al. 1993. Practical method for monitoring polychlorinated 

dibenzo-p-dioxins and polychlorinated dibenzofurans in the atmosphere. J Chromatogr 643(1-2):25-33. 

Kuykendal WB, Lamason WH, Miles AJ, et al. 1989. Ash data from combustion sources: Results of tier 4 

of the national dioxin study. Chemosphere 18:1227-1234. 

*LaFleur L, Bousquet T, Ramage K, et al. 1990. Analysis of TCD and TCDF on the ppq-level in milk and 

food sources. Chemosphere 20(10-12):1657-1662. 

*LaGoy P. 1987. Estimated soil ingestion rates for use in risk assessment. Risk Analysis 7:355-359. 

*Lai Z-W, Pineau T, Esser C. 1996. Identification of dioxin-responsive elements (DREs) in the 5' regions 

of putative dioxin-inducible genes. Chem-Biol Interact 100:97-112. 
CDDs 608 

8. REFERENCES 

*Lakshmanan MR, Campbell BS, Chirtel SJ, et al. 1986. Studies on the mechanism of absorption and 

distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. J Pharmacol Exp Ther 239:673-677. 

Lamb JC IV, Harris MW, McKinney JD, et al. 1986. Effects of thyroid hormones on the induction of cleft 

palate by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in C57BL/6N mice. Toxicol Appl Pharmacol 

84:115-124. 

*Lamb JC IV, Moore JA. 1981. Development and viability of offspring of male mice treated with 

chlorinated phenoxy acids and 2,3,7,8-tetrachlorodibenzo-p-dioxin. J Toxicol Environ Health 8:835-844. 

Lamparski LL, Mahle NH, Shadoff LA. 1978. Determination of pentachlorophenol, 

hexachloro-dibenzo-p-dioxin, and octachloro-dibenzo-p-dioxin in bovine milk. J Agric Food Chem 

26:1113-1116. 

Lamparski LL, Nestrick TJ. 1980. Determination of tetra-, hexa-, hepta- and octachlorodibenzo-p-dioxin 

isomers in particle samples at parts-per-trillion range. Anal Chem 52:2045-2054. 

Lamparski LL, Nestrick TJ. 1988. Accuracy in the determination of chlorinated dibenzo-p-dioxins and 

dibenzofurans in environmental samples. Journal and Research of the National Bureau of Standards 

93:241-242. 

Lamparski LL, Nestrick TJ, Frawley NN, et al. 1986. Perspectives of a large scale environmental study for 

chlorinated dioxins: Water analysis. Chemosphere 15:1445-1452. 

Lamparski LL, Nestrick TJ, Stehl RH. 1979a. Determination of part per trillion concentrations of 

2,3,7,8-tetra chlorodibenzo-p-dioxin in fish. Anal Chem 51:1453-1458. 

Lamparski LL, Stehl RJ, Johnson RL. 1979b. Photolysis of pentachlorophenol-treated wood: Chlorinated 

dibenzo-p-dioxin formation. Environ Sci Technol 14:196-200. 

Lampi P, Vertianen T, Tuomisto J. 1990. Population exposure to chlorophenols, dibenzo-p-dioxins and 

dibenzofurans after a prolonged groundwater pollution by chlorophenols. Chemosphere 20:625-634. 

*Landers JP, Bunce NJ. 1991. The Ah receptor and the mechanism of dioxin toxicity. Biochem J 

276:273-287. 

*Landers JP, Piskorska-Pliszczynska J, Zacharewski T, et al. 1989. Photoaffinity labeling of the nuclear Ah 

receptor from mouse Hepa-1c1c7 cells using 2,3,7,8-[3H]tetrachlorodibenzo-p-dioxin. J Biol Chem 

264:18463-1841. 

*Lang DS, Becker S, Clark GC, et al. 1994. Lack of direct immunosuppressive effects of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on human peripheral blood lymphocyte subsets in vitro. Arch 

Toxicol 68:296-302. 

Langhorst ML, Shadoff LA. 1980. Determination of parts-per-trillion concentrations of tetra-, hexa-, 

hepta-, and octachlorodibenzo-p-dioxins in human milk samples. Anal Chem 52:2037-2044. 
CDDs 609 

8. REFERENCES 

*Lans MC, Klasson-Wehler E, Willemsen M, et al. 1993. Structure-dependent, competitive interaction of 

hydroxy-polychlorobiphenyls, -dibenzo-p-dioxins and -dibenzofurans with human transthyretin. Chem Biol 

Interactions 88:7-21. 

*Lans MC, Spiertz C, Brouwer A. 1994. Different competition of thyroxine binding to transthyretin and 

thyroxine-binding globulin by hydroxy-PCBs, PCDDs and PCDFs. Environ Toxicol Pharmacol 

270:129-136. 

Lapeza CR Jr, Patterson DG Jr, Liddle JA. 1986. Automated apparatus for the extraction and enrichment of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in human adipose tissue. Anal Chem 58:713-716. 

Lathrop GD, Wolfe WH, Albanese RA, et al. 1984. An epidemiological investigation of health effects in 

Air Force personnel following exposure to herbicides: Baseline morbidity study results. USAF School of 

Aerospace Medicine, Brooks Air Force Base, Texas. 

Lawn RE. 1988. Some aspects of the determination of polychlorodibenzo-p-dioxins and polychlorodibenzofurans in environmental and biological samples. Anal Proc 25:49-50. 

*Lawrence BP, Kerkvliet NI. 1997. Role of altered arachidonic acid metabolism in 2,3,7,8-tetrachlorodibenzo-p-dioxin-induced immune suppression in C57BI/6 mice. Toxicol Sci 42:13-22. 

Lawrence J, Onuska F, Wilinson R, et al. 1986. Methods in research: Determination of dioxins in fish and 

sediment. Chemosphere 15:1085-1090. 

*LeBel, GL, Williams DT, Benoit FM. 1992. Chlorinated dibenzodioxins and dibenzofurans in consumer 

paper products. Chemopshere 25(11):1683-1690. 

*Leeder JS, Kearns GL. 1997. Pharmacogenetics in pediatrics: implications for practice. Pediatric Clinics 

of North America 44:55-77. 

*Leung H. 1993. Physiologically-based pharmacokinetic modeling. In: Ballantine B, Marro T, Turner T, 

eds. General and applied toxicology. Vol. 1. New York, NY: Stockton Press, 153-164. 

*Leung H-W, Ku RH, Paustenbach DJ, et al. 1988. A physiologically based pharmacokinetic model for 

2,3,7,8-tetrachlorodibenzo-p-dioxin in C57BL/6J and DBA/2J mice. Toxicol Letters 42:15-28. 

*Leung H-W, Paustenbach DJ, Murray FJ, et al. 1990b. A physiological pharmacokinetic description of the 

tissue distribution and enzyme-inducing properties of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Toxicol 

Appl Pharmacol 103:399-410. 

*Leung H-W, Wendling JM, Orth R, et al. 1990a. Relative distribution of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in human hepatic and adipose tissues. Toxicol Lett 50:275-282. 

*Levy CJ. 1988. Agent Orange exposure and post-traumatic stress disorder. J Nerv Ment Dis 176:242-245. 

*Li X, Johnson DC, Rozman KK. 1995a. Reproductive effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) in female rats: Ovulation, hormonal regulation, and possible mechanisms. Toxicol Appl Pharmacol 

133:321-327. 
CDDs 610 

8. REFERENCES 

*Li X, Johnson DC, Rozman KK. 1995b. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on estrous 

cyclicity and ovulation in female Sprague-Dawley rats. Toxicol Lett 78:219-222. 

*Li X, Rozman KK. 1995. Subchronic effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and their 

reversibility in male Sprague-Dawley rats. Toxicology 97:133-140. 

*Li X, Weber LWD, Rozman KK. 1995c. Toxicokientics of 2,3,7,8-tetrachlorodibenzo-p-dioxin in female 

Sprague-Dawley rats including placental and lactational transfer to fetuses and neonates. Fund Appl Toxicol 

27:70-76. 

Liberti A, Ciccioli P, Brancaleoni E, et al. 1982. Determination of polychlorodibenzo-p-dioxins and 

polychlorodibenzofurans in environmental samples by gas chromatography-mass spectrometry. J 

Chromatogr 242:111-118. 

*Liem A, Hoogerbrugge R, Koostra PR, et al. 1991. Occurrence of dioxins in cow's milk in the vicinity of 

municipal waste incinerators and a metal reclamation plant in the Netherlands. Chemosphere 

23(11-12):1675-1684. 

Liem AKD, De Jong APJM, Marsman JA, et al. 1990. A rapid clean-up procedure for the analysis of 

polychlorinated dibenzo-p-dioxins and dibenzofurans in milk samples. Chemosphere 20:843-850. 

*Ligon WV Jr, Dorn SB, May RJ, et al. 1989. Chlorodibenzofuran and chlorodibenzodioxins levels in 

Chilean mummies date about 2800 years before the present. Environ Sci Technol 23(10):1286-1290. 

Ligon WV Jr, May RJ. 1986. Determination of selected chlorodibenzofurans and chlorodibenzodioxins 

using two-dimensional gas chromatography/mass spectrometry. Anal Chem 58:558-561. 

*Lim M, Jacobson-Kram D, Bowman RE, et al. 1987. Effect of chronic exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin on sister chromatid exchange levels in peripheral lymphocytes of the 

Rhesus monkey. Cell Biol Toxicol 3:279-284. 

*Lin FH, Clark G, Birnbaum LS, et al. 1991a. Influence of the Ah locus on the effects of 

2,3,7,8-tetrachlorodibenzo-p-dioxin on the hepatic epidermal growth factor receptor. Mol Pharmacol 

39:307-313. 

*Lin FH, Stohs SJ, Birnbaum LS, et al. 1991. The effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) 

on the hepatic estrogen and glucocorticoid receptors in congenic strains of Ah responsive and Ah 

nonresponsive C57Bl/6J mice. Toxicol Appl Pharmacol 108:129-139. 

*Lin W-Q, White KL. 1993. Production of complement component C3 in vivo following 

2,3,7,8-tetrachlorodibenzo-p-dioxin exposure. J Toxicol Environ Health 39:273-285. 

Lin WQ, White KL Jr. 1991. In vivo and in vitro suppression of C3 production by TCDD [Abstract 51]. 

The Toxicologist 11:37. 

Lindberg JJ. 1988. Polychlorinated dioxins and dibenzofurans in the pulp and paper industry. Papieri ja 

Puu-Paper and Timber 70:713-715, 717, 719. 
CDDs 611 

8. REFERENCES 

Lindstroem G, Rappe C. 1988. Analytical method for analysis of polychlorinated dibenzo-p-dioxins and 

dibenzofurans in milk. Chemosphere 17:921-935. 

Lindstroem G, Rappe C. 1990. Analytical procedure operating on 0.1-1 ppt level applied in toxicological 

studies on PCDFs/PCDDs. Chemosphere 20:851-856. 

Lindstroem GUM, Sjoestroem MS, Stephen E, et al. 1988. Multivariate statistical approach to a data set of 

dioxin and furan contaminations in human milk. Bull Environ Contam Toxicol 40:641-646. 

Lodge KB, Cook PM. 1989. Partitioning studies of dioxin between sediment and water: The measurement 

of Koc for Lake Ontario sediments. Chemosphere 19:439-444. 

*Lofroth G, Zebuhr Y. 1992. Polychlorinated dibenzo-p-dioxins (PCDDs and dibenzofurans (PCDFs) in 

mainstream and sidestream cigarette smoke. Bull Environ Contam Toxicol 48:789-794. 

*Loonen H, Tonkes M, Parsons JR, et al. 1993. Relative contributions of water and food to the 

bioaccumulation of a mixture of polychlorinated dibenzoi-p-dioxins and polychlorinated dibenzofurans in 

guppies. Sci Total Environ (S):491-498. 

*Loprieno N, Sbrana I, Rusciano D, et al. 1982. In vivo cytogenic studies on mice and rats exposed to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). In: Hutzinger O, Frei RW, Merian E, et al., eds. Chlorinated 

dioxins and related compounds: Impact on the environment. New York, NY: Pergamon Press, 419-428. 

*Lorber M, Cleverly D, Schaum J, et al. 1994. Development and validation of an air-to-beef food chain 

model for dioxin-like compounds. Sci Total Enviro 156:39-65. 

*Lorber M, Saunders P, Ferrario J, et al. 1997. A statistical survey of dioxin-like compounds in United 

States pork fat. Organohalogen Compounds 32:80-86. 

*Lorber MN, Winters DL, Griggs J, et al. 1998. A national survey of dioxin-like compounds in the United 

States milk supply. Organohalogen Compounds 38:87-97. 

*Lorenzen A, Okey AP. 1991. Detection and characterization of Ah receptor in tissue and cells from human 

tonsils. Toxicol Appl Pharmacol 107:203-214. 

Lower WR, Yanders AF, Orazio Ce, et al. 1989. A survey of 2,3,7,8-tetrachlorodibenzo-p-dioxin residues 

in selected species from Times Beach, Missouri. Chemosphere 18:1079-1088. 

*Lu CH, Baggs RB, Redmond D, et al. 1986. Toxicity and evidence for metabolic alterations in 

2,3,7,8-tetrachlorodibenzo-p-dioxin-treated guinea pigs fed by total parenteral nutrition. Toxicol Appl 

Pharmacol 84:439-453. 

Lucier G, Clark G, Hiermath C, et al. 1993a. Carcinogenicity of TCDD in laboratory animals: Implications 

for risk assessment. Toxicol Ind Health 9:631-668. 

*Lucier GW, Portier CJ, Gallo MA. 1993b. Receptor mechanisms and dose-response models for the effects 

of dioxins. Environ Health Perspect 101:36-44. 
CDDs 612 

8. REFERENCES 

*Lucier GW, Rumbaugh RC, McCoy Z, et al. 1986. Ingestion of soil contaminated with 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) alters hepatic enzyme activities in rats. Fundam Appl Toxicol 

6:364-371. 

*Lucier GW, Tritscher A, Goldsworthy T, et al. 1991. Ovarian hormones enhance 2,3,7,8-tetrachlorodibenzo-p-dioxin-mediated increases in cell proliferation and preneoplastic foci in a two-stage model for rat 

hepatocarcinogenesis. Cancer Res 51:1391-1397. 

*Lundberg K, Dencker L, Gronvik KO. 1990. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) 

treatment in vivo on thymocyte functions in mice after activation in vitro. Int J Immunopharmacol 

12:459-466. 

Lundgren K, Andries M, Thompson C, et al. 1986. Dioxin treatment of rats results in increased in vitro 

induction of sister chromatid exchanges by -naphthoflavone: An animal model for human exposure to 

halogenated aromatics. Toxicol Appl Pharmacol 85:189-195. 

*Lustenhouwer JWA, Olie K, Hutzinger O. 1980. Chlorinated dibenzo-p-dioxins and related compounds in 

incinerator effluents. Chemosphere 9:501-522. 

*Luster MI, Clark G, Lawson LD, et al. 1979. Effects of brief in vitro exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on mouse lymphocytes. J Environ Pathol Toxicol 2:965-977. 

Luster MI, Dean JH, Boorman GA. 1982. Altered immune functions in rodents treated with 

2,3,7,8-tetrachlorodibenzo-p-dioxin, phorbol-12-myristate-13-acetate, and benzo[a]pyrene. Banbury Rep. 

Iss Environmental Factors in Human Growth and Development 11:199-213. 

*Luster MI, Germolec DR, Clark G, et al. 1988. Selective effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

and corticosteroid on in vitro lymphocyte maturation. J Immunol 140:928-935. 

Luster MI, Hong LH, Boorman GA, et al. 1985. Acute myelotoxic responses in mice exposed to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicol Appl Pharmacol 81:156-165. 

*Luster, MI, Hong LH, Osborne R, et al. 1986. I-Amino-3,7,8-trichlorodibenzo-p-dioxin: A specific 

antagonist for TCDD-induced myelotoxicity. Biochem Biophys Res Commun 139:747-756. 

*Luster MJ, Boorman GA, Dean JH, et al. 1980. Examination of bone marrow, immunologic parameters 

and hose susceptibility following pre- and postnatal exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD). Int J Immunopharmacol 2:301-310. 

*Lyman WJ, Reehl WF, Rosenblatt DH. 1982. Handbook of chemical property estimation methods. New 

York, NY: McGraw-Hill Book Company, 4-9, 15-16. 

*Lynge E. 1985. A follow-up study of cancer incidence among workers in manufacture of phenoxy 

herbicides in Denmark. Br J Cancer 52:259-270. 

Ma X, Mufti NA, Babish JG. 1992. Protein tyrosine phosphorylation as an indicator of 

2,3,7,8-tetrachlorodibenzo-p-dioxin exposure in vivo and in vitro. Biochem Biophys Res Commun 

189(1):59-65. 
CDDs 613 

8. REFERENCES 

*Mabey WR, Smith JH, Podoll RT, et al. 1982. Aquatic fate process data for organic priority pollutants. 

Washington, DC: U.S. Environmental Protection Agency. EPA 440/4-81-014, 107-108. 

Mably TA, Moore RW, Peterson RE. 1991. Effects of perinatal exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on pre- and postnatal plasma testosterone (T) concentrations in 

male rats [Abstract 996]. The Toxicologist 11:262. 

*Mably TA, Moore RW, Peterson RE. 1992a. In utero and lactational exposure of male rats to 

2,3,7,8-tetrachlordibenzo-p-dioxin. 1. Effects on androgenic status. Toxicol Appl Pharmacol 114(1):97-107. 

*Mably TA, Moore RW, Peterson RE. 1992b. In utero and lactational exposure of male rats to 

2,3,7,8-tetrachlordibenzo-p-dioxin. 2. Effects on sexual behavior and the regulation of luteinizing hormone 

secretion in adulthood. Toxicol Appl Pjarmacol 114(1):108-117. 

*Mably TA, Moore RW, Peterson RE. 1992c. In utero and lactational exposure of male rats to 

2,3,7,8-tetrachlordibenzo-p-dioxin. 3. Effects on spermatogenesis and reproductive capability. Toxicol Appl 

Pharmacol 114(1):118-126. 

*Mably TA, Theobald HM, Ingall GB, et al. 1990. Hypergastrinemia is associated with decreased gastric 

acid secretion in 2,3,7,8-tetrachlorodibenzo-p-dioxin-treated rats. Toxicol Appl Pharmacol 106:518-528. 

Madge DS. 1977. Effects of trichlorphenoxyacetic acid and chlorodioxins on small intestinal function. Gen 

Pharmacol 8:319-324. 

*Madhukar BV, Brewster DW, Matsumura F. 1984. Effects of in vivo administered 

2,3,7,8-tetrachlorodibenzo-p-dioxin on receptor binding of epidermal growth factor in the hepatic plasma 

membrane of rat, guinea pig, mouse and hamster. Proc Natl Acad Sci USA 81:7404-7411. 

*Madsen C, Larsen JC. 1989. Relative toxicity of chlorinated dibenzo-para-dioxins, and dibenzofurans 

measured by thymus weight and liver-enzyme induction in perinatally dosed rats 2,3,7,8-TCDD, 

2,3,4,7,8-PCDF, 1,2,3,7,8-PCDD. Chemosphere 18:955-966. 

Mahle NJ, Whiting LF. 1980. The formation of chlorodibenzo-p-dioxins by air oxidation and chlorination 

of bituminous coal. Chemosphere 9:693-699. 

*Mahon MJ, Gasiewicz TA. 1995. Ah receptor phosphorylation: Localization of phosphorylation sites to 

the C-terminal half of the protein. Arch Biochem Biophys 318:166-174. 

*Malisch R. 1994. Determination of dioxins and furans in coloured candle wax. Chemosphere 

29(9-11):1957-1963. 

*Mamantov A. 1984. Linear correlation between photolysis rates and toxicity of polychlorinated 

dibenzo-p-dioxins. Environ Sci Technol 18:808-810. 

*Manara L, Coccia P, Croci T. 1984. Prevention of TCDD toxicity in laboratory rodents by addition of 

charcoal or cholic acids to chow. Food Chem Toxicol 22:815-818. 
CDDs 614 

8. REFERENCES 

*Mantovani A, Vecci A, Luini W, et al. 1980. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on macrophage 

and natural killer cell-mediated cytotoxicity in mice. Biomedicine 32:200-204. 

*Manz A, Berger J, Dwyer JH, et al. 1991. Cancer mortality among workers in chemical plant 

contaminated with dioxin. Lancet 338:959-964. 

Marcheterre L, Webster GR, Barrie MDC, et al. 1985. Fate of 14C-octachlorodibenzo-p-dioxin in artificial 

outdoor ponds. Chemosphere 14:835-838. 

Mariani G, Benfenati E, Fanelli R, et al. 1992. Incineration of agro-industrial wastes and macro- and 

micropollutants emission. Chemosphere 23:1359-64. 

*Marinovich M, Sirtori CR, Galli CL, et al. 1983. The binding of 2,3,7,8-tetrachlorodibenzo-p-dioxin to 

plasma lipoproteins may delay toxicity in experimental hyperlipidemia. Chem Biol Interact 45:393-399. 

Marklund S, Andersson R, Tysklind M, et al. 1990. Emissions of PCDDs and PCDFs in gasoline and diesel 

fueled cars. Chemosphere 20:553-561. 

Marklund S, Kjeller LO, Hansson M, et al. 1986. Determination of PCDDs and PCDFs in incineration 

samples and pyrolytic products. In: Rappe C, Choudhary G, Keith LH, eds. Chlorinated dioxins and 

dibenzofurans in perspective. Chelsea, MI: Lewis Publishers, Inc., 79-92. 

*Marklund S, Rappe C, Tyslind M, et al. 1987. Identification of polychlorinated dibenzofurans and dioxins 

in exhaust from cars run on leaded gasoline. Chemosphere 16:29-36. 

*Marklund S, Tysklind M, Andersson R, Ljung K, et al. 1991. Environmental deposition of PCDDs and 

PCDFs as determined by the analysis of snow samples from the northern Sweden. Chemosphere 

23:1359-1364. 

Marlow DA. 1986. Hexachlorobenzene exposure in the production of chlorophenols: Industrywide studies 

on branch. Cincinnati, OH: National Institute for Occupational Safety and Health, 161-169. 

Marple L, Berridge B, Throop L. 1986a. Measurement of the water-octanol partition coefficient of 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ Sci Technol 20:397-399. 

*Marple L, Brunck R, Throop L. 1986b. Water solubility of 2,3,7,8,-tetrachlorodibenzo-p-dioxin. Environ 

Sci Technol 20(2):180-182. 

*Marquis PJ, Hacket M, Holland LG, et al. 1994. Analytical methods for a national study of chemical 

residues in fish. 1: Polychlorinated dibenzo-p-dioxins, dibenzofurans. Chemosphere 29(3):495-508. 

Mason G, Denomme MA, Safe L, et al. 1987. Polybrominated and chlorinated dibenzo-para-dioxins 

structure-activity relationships. Chemosphere 16:1729-1731. 

Mason G, Farrell K, Keys B, et al. 1986. Polychlorinated dibenzo-p-dioxins: Quantitative in vitro and in 

vivo structure-activity relationships. Toxicology 41:21-31. 

*Mason G, Safe S. 1986a. Synthesis, biologic and toxic effects of the major 

2,3,7,8-tetrachlorodibenzo-p-dioxin metabolites in the rat. Toxicology 41:153-159. 
CDDs 615 

8. REFERENCES 

Mason G, Safe S. 1986b. Synthesis, biologic and toxic properties of 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Chemosphere 15:2081-2083. 

*Mastroiacovo P, Spagnolo A, Marni E, et al. 1988. Birth defects in the Seveso area after TCDD 

contamination. JAMA 259:1668-1672. 

*Matsumura F. 1994. How important is the protein phosphorylation pathway in the toxic expression of 

dioxin-type chemicals? Biochem Pharmacol 48:215-224. 

*Matsumura F, Benezet HJ. 1973. Studies on the bioaccumulation and microbial degradation of 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ Health Persp 5:253-258. 

Matsumura F, Quensen J, Tsushimoto G. 1983. Microbial degradation of TCDD in a model ecosystem. In: 

Tucker RE, et al. eds. Human and Environmental risks of chlorinated dioxins and related compounds. New 

York, London: Plenum Press, 191-219. 

*Max SR, Silbergeld EK. 1987. Skeletal muscle glucocorticoid receptor and glutamine synthetase activity 

in the wasting syndrome in rats-treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 

15:523-527. 

*May G. 1973. Chloracne from the accidental production of tetrachlorodibenzodioxin. Br J Med 

30:276-283. 

*McConkey DJ, Orrenius S. 1989. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) kills 

glucocorticoid-sensitive thymocytes in vivo. Biochem Biophys Res Commun 160:1003-1008. 

McConnell EE. 1985. Comparative toxicity of PCBs and related compounds in various species of animals. 

Environ Health Perspect 60:29-33. 

*McConnell EE, Lucier GW, Rumbaugh RC, et al. 1984. Dioxin in soil: Bioavailability after ingestion by 

rats and guinea pigs. Science 223:1077-1079. 

*McConnell EE, Moore JA. 1979. Toxicopathology characteristics of the halogenated aromatics. Ann. NY 

Acad. Sci. 320:138-147. 

*McConnell EE, Moore JA, Dalgard DW. 1978a. Toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rhesus 

monkeys (Macaca mulatta) following a single oral dose. Toxicol Appl Pharmacol 43:175-187. 

*McConnell EE, Moore JA, Haseman JK, et al. 1978b. The comparative toxicity of chlorinated 

dibenzo-p-dioxins in mice and guinea pigs. Toxicol Appl Pharmacol 44:335-356. 

*McCormack KM, Gibson JE, Hook JB. 1976. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on 

renal function in rats in vivo. Toxicol Appl Pharmacol 37:177. 

*McCrady JK, Maggard SP. 1993. Uptake and photodegradation of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

sorbed to grass foliage. Environ Sci Tech 27:343-350. 

*McCrady JK, McFarlane C, Gander LK. 1990. The transport and fate of 2,3,7,8- TCDD in soybeans and 

corn. Chemosphere 21: 359-376. 
CDDs 616 

8. REFERENCES 

*McCurvin DMA, Schellenberg DH, Clement RE, et al. 1989. A comparison of tandem quadrupole mass 

spectrometry versus single quadrupole capabilities for the determination of chlorinated dioxins and 

dibenzofurans in environmental samples. Chemosphere 19:201-204. 

*McKee P, Burt A, McCurvin D, et al. 1990. Levels of dioxins, furans and other organic contaminants in 

harbour sediments near a wood preserving plant using pentachlorophenol and creosote. Chemosphere 

20(10-12):1679-1685. 

*McKinney JD, Chae K, Oatley SJ, et al. 1985a. Molecular interactions of toxic chlorinated 

dibenzo-p-dioxins and dibenzofurans with thyroxine binding prealbumin. J Med Chem 28:375-381. 

*McKinney JD, Fawkes J, Jordan S, et al. 1985b. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) as a potent 

and persistent thyroxine agonist: A mechanistic model for toxicity based on molecular reactivity. Environ 

Health Perspect 61:41-53. 

McKone TE, Ryan PB. 1989. Human exposure to chemicals through food chains: An uncertainty analysis. 

Environ Sci Technol 23:1154-1163. 

McLachlan M, MacKay D, Jones PH. 1990. A conceptual model of organic chemical volatilization at 

waterfalls. Environ Sci Technol 24:252-257. 

*McLachlan MS. 1993. Digestive tract absorption of polychlorinated dibenzo-p-dioxins, dibenzofurans, 

and biphenyls in a nursing infant. Toxicol Appl Pharmacol 123:68-72. 

*McLachlan MS, Hinkel M, Reissinger M, et al. 1994. A study of the influence of sewage sludge 

fertilization on the concentrations of PCDD/F and PCB in soil and milk. Environmental Pollution 

85:337-343. 

*McLemore TL, Adelberg S, Liu MC, et al. 1990. Expression of CYP1A1 gene in patients with lung 

cancer: Evidence for cigarette smoke-induced gene expression in normal lung tissue and for altered gene 

regulation in primary pulmonary carcinomas. J Natl Cancer Inst 82:127-133. 

McMahon LW. 1983. Organic priority pollutants in wastewater. Proceedings of the 1982 UCC-ND/GAT 

Environ Protection Seminar, April 5-7, 1982, Glenstone Lodge, Gatlinburg TN. NTIS DE83010817. 

*McNulty W. 1984. Fetotoxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) for Rhesus Macaques 

(Macaca mulatta). Am J Primatol 6:41-47. 

*McNulty W. 1985. Toxicity and fetotoxicity of TCDD, TCDF, and PCB isomers in Rhesus Macaques 

(Macaca mulatta). Environ Health Perspect 60:77-88. 

*McNulty WP, Nielsen-Smith JO, Lay Jr JO, et al. 1982. Persistence of TCDD in monkey adipose tissue. 

Food Cosmet Toxicol 20:985-987. 

*McPeters AL, Overcash MR. 1993. Demonstrations of photodegradation by sunlight of 

2,3,7,8-tetrachlorodibenzo-p-dioxins in 6 cm soil columns. Chemosphere 27:1221-34. 

McQueen EG, Veale AMO, Alexander WS, et al. 1977. 2,4,5-T and human birth defects. Report of the 

Division of Public Health, New Zealand Department of Health. 
CDDs 617 

8. REFERENCES 

*MDNR. 1987. Clean water commission: Water quality standards. Missouri Department of Natural 

Resources. Washington, DC: Bureau of National Affairs, Inc., 10 CSR 20-7.031. 

*Mebus CA, Piper WN. 1986. Decreased rat adrenal 21-hydroxylase activity associated with decreased 

adrenal microsomal cytochrome P-450 after exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Biochem 

Pharmacol 35:4359-4262. 

*Mebus CA, Reddy VR, Piper WN. 1987. Depression of rat testicular 17-hydroxylase and 17,20-lyase after 

administration of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Biochem Pharmacol 36:727-731. 

Mehrle PM, Buckler DR, Little EE, et al. 1988. Toxicity and bioconcentration of 2,3,7,8-tetrachlorodibenzodioxin and 2,3,7,8-tetrachlorodibenzofuran in rainbow trout. Environ Toxicol Chem 1:47-62. 

Mes J. 1981. Human milk analysis for halogenated hydrocarbon residues. Int J Environ Anal Chem 

9:283-300. 

Mes J, Weber D. 1989. Non-orthochlorine substituted coplanar polychlorinated biphenyl congeners in 

Canadian adipose tissue, breast milk and fatty foods. Chemosphere 19:1357-1365. 

*Messerer P, Zober A, Becher H. 1998. Blood lipid concentrations of dioxins and furans in a sample of 

BASF employees included in the IARC registry of workers exposed to phenoxy acid herbicides and/or 

chlorophenols. Environ Health Perspect 106 (Suppl 2):733-735. 

Meyer C, O'Keefe P, Hilker D, et al. 1989. A survey of twenty community water systems in New York 

State for PCDDs and PCDFs. Chemosphere 19:21-26. 

*Meyne J, Allison DC, Bose K, et al. 1985. Hepatotoxic doses of dioxin do not damage mouse bone 

marrow chromosomes. Mutat Res 157:63-69. 

*Michalek JE, Pirkle JL, Caudill SP, et al. 1996. Pharmacokinetics of TCDD in veterans of Operation 

Ranch Hand: 10-year follow-up. J Toxicol Environ Health 47:209-220. 

*Michalek JE, Rahe AJ, Boyle CA. 1998. Paternal dioxin, preterm birth, intrauterine growth retardation, 

and infant death. Epidemiology 9:161-167. 

*Michalek JE, Rahe AJ, Kulkarni PM, et al. 1998. Levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 1,302 

unexposed air force Vietnam-era veterans. Journal of Exposure Analysis and Environmental Epidemiology 

8(1):59-64. 

*Miller AG, Whitlock JP Jr. 1981. Novel variant in benzo[a]pyrene metabolism. J Biol Chem 

256:2433-2437. 

*Miller GC, Herbert VR, Miille MJ, et al. 1989a. Photolysis of octachlorodibenzo-p-dioxin on soils: 

Production of 2,3,7,8-TCDD. Chemosphere 18:1265-1274. 

Miller GC, Herbert VR, Miller WW. 1989b. Effect of sunlight on organic contaminants at the 

atmosphere-soil interface. In: Reactions and movement of organic chemicals in soils. SSSA special 

publication no. 22. 99-110. 
CDDs 618 

8. REFERENCES 

*Miller GC, Herbert VR, Zepp RG. 1987. Chemistry and photochemistry of low-volatility organic 

chemicals on environmental surfaces. Environ Sci Technol 21:1164-1167. 

*Mimnaugh EG, Worland PJ, Whitesell L, et al. 1995. Possible role for serine/threonine phsphorylation in 

the regulation of the heteroprotein complex between the hsp90 stress protein and pp60v-src tyrosine kinase. 

J Biol Chem 270:28654-28659. 

Miyata H, Takayama K, Ogaki J, et al. 1989. Level of PCDDs, coplanar PCBs and PCDFs in patients with 

Yusho disease and in Yusho oil. Chemosphere 18:407-416. 

*MMWR. 1987. Epidemiologic notes and reports. Serum dioxin in Vietnam-era veterans--preliminary 

report. Atlanta, GA: Centers for Disease Control, Morbidity and Mortality Weekly Report 36:470-475. 

*MMWR. 1988. Preliminary report: 2,3,7,8-tetrachlorodibenzo -p-dioxin exposure to humans - Seveso, 

Italy. Morbidity and Mortality Weekly Report 37(48):733-736. 

*Mocarelli P, Brambilla P, Gerthoux PM, et al. 1996. Change in sex ratio with exposure to dioxin. Lancet 

348:409. 

*Mocarelli P, Marocchi A, Brambilla P, et al. 1986. Clinical laboratory manifestations of exposure to 

dioxin in children. JAMA 256:2687-2695. 

*Mocarelli P, Needham LL, Marocchi A, et al. 1991. Serum concentrations of 2,3,7,8-tetrachlorodibenzop-dioxin and test results from selected residents of Seveso, Italy. J Toxicol Environ Health 32:357-366. 

Mohamad AH, Zerezghi M, Caruso JA. 1986. Determination of polychlorinated dibenzo-p-dioxins using 

capillary gas chromatography with microwave-induced plasma detection. Anal Chem 58:469-471. 

*Moore JA, Gupta BN, Zinkl JG, et al. 1973. Postnatal effects of maternal exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Environ Health Perspect 5:81-85. 

Moore M, Narasimhan TR, Steinberg MA, et al. 1993. Potentiation of CYP1A1 gene expression in MCF-7 

human breast cancer cells cotreated with 2,3,7,8-tetrachlorodibenzo-p-dioxin and 12-O-tetradecanoylph 

orbol-13-acetate. Arch Biochem Biophys 305(2):483-488. 

*Moore RW, Jefcoate CR, Peterson RE. 1991. 2,3,7,8-Tetrachlorodibenzo-p-dioxin inhibits steroidogenesis 

in the rat testis by inhibiting the mobilization of cholesterol to cytochrome P450scc. Toxicol Appl 

Pharmacol 109:85-97. 

Moore RW, Mably TA, Bjerke KL, et al. 1992. In utero and lactational 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) exposure decreased androgenic responsiveness of male sex organs and permanently inhibits 

spermatogenesis and demasculinizes sexual behavior in rats. Toxicologist 12(1):81 

*Moore RW, Parsons JA, Bookstaff RC, et al. 1989. Plasma concentrations of pituitary hormones in 

2,3,7,8-tetrachlorodibenzo-p-dioxin-treated male rats. J Biochem Toxicol 4:165-172. 

*Moore RW, Peterson RE. 1988. Androgen catabolism and excretion in 2,3,7,8-tetrachlorodibenzo-pdioxin-treated rats. Biochem Pharmacol 37:560-562. 
CDDs 619 

8. REFERENCES 

*Moore RW, Potter CL, Theobald HM, et al. 1985. Androgenic deficiency in male rats treated with 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 79:99-111. 

*Moran RA, Lee CW, Fujimoto JM, et al. 1986. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on 

IgA serum and bile levels in rats. Immunopharmacology 12:245-250. 

*Morgan JE, Whitlock JP Jr. 1992. Transcription-dependent and transcription-independent nucleosome 

disruption induced by dioxin. Proc Natl Acad Sci USA 89:11622-11626. 

Morita M, Yashuhara A, Seki H, et al. 1987. Chlorodibenzo-para-dioxins in the feral pigeon. Chemosphere 

16:1749-1752. 

*Morris DL, Holsapple MP. 1991. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on humoral 

immunity: II. B-Cell activation. Immunopharmacology 21:171-182. 

*Morris DL, Jordan SD, Holsapple MP. 1991. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on 

humoral immunity: similarities to Staphylococcus aureus Cowan strain I (SAC) in the in vitro T-dependent 

antibody response. Immunopharmacology 21:159-170. 

*Morris DL, Snyder NK, Gokani V, et al. 1992. Enhanced suppression of humoral immunity in DBA/2 

mice following subchronic exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicol Appl 

Pharmacol 112:128-132. 

Morrissey RE, Harris MW, Diliberto JJ, et al. 1992. Limited PCB antagonism of TCDD-induced 

malformations in mice. Toxicol Lett 60(1):19-25. 

Morselli L, Zappoli S, Liberti A, et al. 1989. Evaluation and comparison of organic and inorganic 

compounds between emission and immission samples from municipal solid waste incinerator. Chemosphere 

18:2263-2273. 

*Morselli PL, Franco-Morselli R, Bossi L. 1980. Clinical Pharmacokinetics in Newborns and Infants. 

Clinical Pharmacokinetics 5:485-527. 

*Mortelmans K, Haworth S, Speck W, et al. 1984. Mutagenicity testing of Agent Orange and related 

compounds. Toxicol Appl Pharmacol 75:137-146. 

*Moses M, Lilis R, Crow KD, et al. 1984. Health status of workers with past exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin in the manufacture of 2,4,5-trichlorophenoxyacetic acid: Comparison of 

findings with and without chloracne. Am J Ind Med 5:161-182. 

*Moses M, Prioleau PG. 1985. Cutaneous histologic findings in chemical workers with and without 

chloracne with past exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. J Am Acad Dermatol 12:497-506. 

*Mott L. 1995. The disproportionate impact of environmental health threats on children of color. 

Environmental Health Perspectives 103(suppl6):33-35. 

Mottet NK. 1985. Morphogenic injury by chemical agents. In: Mottet NK, ed. Environmental pathology. 

New York, NY: Oxford University, 82-125. 
CDDs 620 

8. REFERENCES 

Mottura A, Zei G, Nuzzo F, et al. 1981. Evaluation of results of chromosome analyses on lymphocytes of 

TCDD exposed subjects after the Seveso accident. Mutat Res 85:238-239. 

*Muir DC G, Lawrence S, Holoka M, et al. 1992. Partitioning of polychlorinated dioxins and furans 

between water, sediments and biota in lake mesocosms. Chemosphere 25(1-2):119-24. 

*Muir DCG, Townsend BE, Webster GRB. 1983. Bioavailability of 

14C-1,3,6,8-tetrachlorodibenzo-p-dioxin and 14C-octachlorodibenzo-p-dioxin to aquatic insects in sediment 

and water. Fate and Distribution of PCDD and PCDF, 89-102. 

*Muir DCG, Yarachewski AL, Webster GRB. 1986b. Bioconcentration of four chlorinated dioxins by 

rainbow trout and fathead minnows. In: Bahner RC, Hansen DJ, eds. Aquatic toxicology and hazard 

assessment: Eighth Symposium, Fort Mitchell, KY, April 15-17, 1984. Philadelphia, PA: ASTM, 442-454. 

*Muir DCG, Yarechewski AL. 1988. Dietary accumulation of four chlorinated dioxin congeners by 

rainbow trout and fathead minnow. Environ Toxicol Chem 7:227-236. 

*Muir DCG, Yarechewski AL, Knoll A, et al. 1986a. Bioconcentration and disposition of 

1,3,6,8-tetrachlorodibenzo-p-dioxin and octachlorodibenzo-p-dioxin by rainbow trout and fathead minnows. 

Environ Toxicol Chem 5:261-272. 

*Mulcahy MT. 1980. Chromosome aberrations and "Agent Orange." Med J Aust 2:573-574. 

*Muller JF, Hulster A, Papke A, et al. 1993. Transfer pathways of PCDD/PCDF to fruits. Chemosphere 

27(1-3):195-201. 

*Muller JF, Hulster AA, Papke OC, et al. 1994. Transfer of PCDD/PCDF from contaminated soils into 

carrots, lettuce and peas. Chemosphere 29(9-11):2175-2181. 

*Murai K, Tsuji H, Fujishima F. 1991. [Treatment of Yusho PCB poisioning patients with cholestyramine.] 

Fukuoka Ishi 82:226-329. (Japanese) 

*Murk AJ, Legler J, Denison MS, et al. 1996. Chemical-activated luciferase gene expression (calux): A 

novel in vitro bioassay for ah receptor active compounds in sediments and pore water. Fundam Appl 

Toxicol 33(1):149-60. 

*Murray FJ, Smith FA, Nitschke KD, et al. 1979. Three-generation reproduction study of rats given 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the diet. Toxicol Appl Pharmacol 50:241-251. 

*Muto H, Takizawa Y. 1989. Dioxins in cigarette smoke. Arch Environ Health 44:171-174. 

*MWQS. 1988a. Minnesota recommended allowable limits for drinking water contaminants release no. 2, 

November. Minnesota Water Quality Standards. 

*MWQS. 1988b. Minnesota water quality standards and criteria for toxic substances, provisional data 

subject to change, 11/88. Minnesota Water Quality Standards. 
CDDs 621 

8. REFERENCES 

*Nagarkatti PS, Sweeney GD, Gauldie J, et al. 1984. Sensitivity to suppression of cytotoxic T-cell 

generation by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) is dependent on the Ah genotype of the murine 

host. Toxicol Appl Pharmacol 72:169-176. 

Nagayama J, Kiyohara C, Masuda Y, et al. 1985. Genetically mediated induction of aryl hydrocarbon 

hydroxylase activity in human lymphoblastoid cells by polychlorinated dibenzofuran isomers and 

2,3,7,8-TCDD. Arch Toxicol 56:230-235. 

*Nagayama J, Mohri N, Kiyohara C, et al. 1989. Comparative toxicologic study of 2,3,7,8-tetrachlorodibenzo-p-dioxin in Ah responsive and nonresponsive strains of mice. Chemosphere 19:927-932. 

Naikwadi KP, Karasek FW. 1986. Gas chromatographic separation of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

from polychlorinated biphenyls and tetrachlorodibenzo-p-dioxin isomers using a polymeric liquid crystal 

capillary column. J Chromatogr 369:203-207. 

Naikwadi KP, Karasek FW. 1990a. Development of GC capillary column for isomer-specific separation of 

toxic isomer of PCCDs and PCDFs in environmental samples. Chemosphere 20:1379-1384. 

Naikwadi KP, Karasek FW. 1990b. Use of C-13 labelled dioxins mixture for analysis of polychlorinated 

dibenzo-para-dioxins in environmental samples by GC. Int J Environ Anal Chem 38:329-342. 

Nakano T, Tsuji M, Okuno T. 1987. Levels of chlorinated organic compounds in the atmosphere. 

Chemosphere 16:1781-1786. 

Nakano T, Tsuji M, Okuno T. 1990. Distribution of PCDDs, PCDFs and PCBs in the atmosphere. Atmos 

Environ 24A:1361-1368. 

*NAS/NRC. 1977. Drinking water and health: Part II. Washington, DC: National Academy of Sciences, 

Safe Drinking Water Committee, National Academy Press, 508-513. 

*NAS/NRC. 1989. Biologic markers in reproductive toxicology. National Academy of Sciences/National 

Research Council. Washington, DC: National Academy Press, 15-35. 

*NAS/NRC. 1989. Biological markers in reproductive toxicology. National Research Council. Board of 

Environmental Studies and Toxicology. Committee on Biological Markers. 15-35. 

*NAS/NRC. 1989. Recommended Dietary Allowances, 10th ed. National Academy of Sciences, National 

Research Council, Food and Nutrition Board. National Academy Press, Washington DC. 

*Nash RC, Beall ML. 1980. Distribution of silvex, 2,4-D, and TCDD applied to turf in chambers and field 

plots. J Agri Food Chem 28(3):614-623. 

*NATICH. 1992. National Air Toxics Information and Clearinghouse. Database report on state, local, and 

EPA air toxics activities. Washington, DC: U.S. Environmental Protection Agency, Office of Air Quality 

Planning and Standards. 

*Nau H, Bass R. 1981. Transfer of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) to the mouse embryo and 

fetus. Toxicology 20:299-308. 
CDDs 622 

8. REFERENCES 

*Nau H, Bass R, Neubert D. 1986. Transfer of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) via placenta 

and milk, and postnatal toxicity in the mouse. Arch Toxicol 59:36-40. 

Nauman CH, Schaum JL. 1987. Human exposure estimation for 2,3,7,8-TCDD. Chemosphere 

16:1851-1856. 

*NCASI. 1993. Summary of data reflective of pulp and paper industry progress in reducing the 

TCDD/TCDF content of effluents, pulps and wastewater treatment sludges. National Council of the Paper 

Industry for Air and Stream Improvement, Inc. Special report no. 93-08. 

*NCI/NTP. 1979a. Bioassay of 2,7-dichlorodibenzo-p-dioxin (DCDD) for possible carcinogenicity. 

Bethesda MD: U.S. Department of Health, Education, and Welfare, Public Health Service, National Cancer 

Institutes/National Toxicology Program. (NIH) DHEW Publication no. 79-1378. 

*NCI/NTP. 1979b. Bioassay of dibenzo-p-dioxin for possible carcinogenicity. Bethesda MD: U.S. 

Department of Health, Education, and Welfare, Public Health Service, National Cancer Institutes/National 

Toxicology Program. (NIH) DHEW Publication no. 79-1377. 

*NCI/NTP. 1980. Bioassay of a mixture of 1,2,3,6,7,8-hexachlorodibenzo-p-dioxin and 

1,2,3,7,8,9-hexachlorodibenzo-p-dioxin (gavage) for possible carcinogenicity: CAS No. 57653-85-7, CAS 

No. 19408-74-3. Research Triangle Park, NC: U.S. Department of Health and Human Services, Public 

Health Service, National Institutes of Health/National Toxicology Program. NCI Carcinogen Tech Rep Ser 

no. 198. NTP no. 80-12. 

Nebert DW, Robinson JR, Niwa A, et al. 1975. Genetic expression of aryl hydrocarbon hydroxylase 

activity in the mouse. J Cell Physiol 85:393-414. 

*Needham LL, Gerthoux PM, Patterson DG, et al. 1994. Half-life of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

serum of Seveso adults: Interim report. Organohalogen Compounds 21:81-85. 

Needham LL, Patterson DG, Alley CC, et al. 1987. Polychlorinated dibenzo-p-dioxins and dibenzofurans 

levels in persons with high and normal levels of 2,3,7,8-tetrachlorodibenzo-p-dioxins. Chemosphere 

16:2027-2031. 

*Needham LL, Patterson DG Jr, Burse VW, et al. 1996. Reference range data for assessing exposure to 

selected environmental toxicants. Toxicology and Industrial Health 2(3/4):507-513. 

*Needham LL, Patterson DG Jr, Houk VN. 1991. Levels of TCDD in selected human populations and their 

relevance to human risk assessment. Banbury Report 35: Biological basis for risk assessment of dioxins and 

related compounds. Cold Spring Harbor Press, 229-247. 

Needham LL, Patterson DG Jr, Pirkle JL, et al. 1989. The basis of measuring 2,3,7,8-tetrachlorodibenzo-p-dioxin in serum. Chemosphere 18:425-430. 

*Nelson CJ, Holson JF, Green HG, et al. 1979. Retrospective study of the relationship between agricultural 

use of 2,4,5-T and cleft palate occurrence in Arkansas. Teratol 19:377-384. 
CDDs 623 

8. REFERENCES 

*Nessel CS, Amoruso MA, Umbreit TH, et al. 1990. Hepatic aryl hydrocarbon hydroxylase and 

cytochrome P450 induction following the transpulmonary absorption of TCDD from intratracheally instilled 

particles. Fundam Appl Toxicol 15:500-509. 

*Nessel CS, Amoruso MA, Umbreit TH, et al. 1992. Pulmonary bioavailability and fine particle enrichment 

of 2,3,7,8-tetrachlorodibenzo-p-dioxin in respirable soil particles. Fundam Appl Toxicol 19(2):279-85. 

*Nessel CS, Butler JP, Post GB, et al. 1991. Evaluation of the relative contribution of exposure routes in a 

health risk assessment of dioxin emissions from a municipal waste incinerator. J Exposure Anal Environ 

Epidemiol 1(3):283-307. 

Nestrick TJ, Lamparski LL. 1983. Assessment of chlorinated dibenzo-p-dioxin formation and potential 

emission to the environment from wood combustion. Chemosphere 12:617-626. 

*Nestrick TJ, Lamparski LL, Frawley NN, et al. 1986. Perspectives of a large scale environmental survey 

for chlorinated dioxins: Overview and soil data. Chemosphere 15:1453-1460. 

*Nestrick TJ, Lamparski LL, Townsend DI. 1980. Identification of tetrachlorodibenzo-p-dioxin isomers at 

the 1-ng level by photolytic degradation and pattern recognition techniques. Anal Chem 52:1865-1874. 

*Neubert D, Dillman I. 1972. Embryotoxic effects in mice treated with 2,4,5-trichlorophenoxyacetic acid 

and 2,3,7,8-tetrachlorodibenzo-p-dioxin. Arch Pharmacol 272:243-264. 

*Neubert D, Golor G, Neubert R. 1992c. TCDD-toxicity equivalencies for PCDD/PCDF congeners: 

prerequisites and limitations. Chemosphere 25:65-70. 

*Neubert D, Wiesmueller T, Abraham K, et al. 1990a. Persistence of various polychlorinated 

dibenzo-p-dioxins and dibenzofurans (PCDDs and PCDFs) in hepatic and adipose tissue of marmoset 

monkeys. Arch Toxicol 64:431-442. 

Neubert D, Zens P, Rothenwallner A, et al. 1973. A survey of the embryotoxic effects of TCDD in 

mammalian species. Environ Health Perspect 5:67-79. 

*Neubert R, Golor G, Stahlmann R, et al. 1992a. Dose-dependent TCDD-induced increase or decrease in 

T-lymphocyte subsets in the blood of New World monkeys (Callithrix jacchus). Chemosphere 

25(7-10):1201-1206. 

*Neubert R, Golor G, Stahlmann R, et al. 1992b. Polyhalogenated dibenzo-p-dioxins and dibenzofurans 

and the immune system. 4. Effects of multiple-dose treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) on peripheral lymphocyte subpopulations of a nonhuman primate (Callithrix jacchus). Arch 

Toxicol 66(4):250-259. 

*Neubert R, Jacob-M6ller U, Helge H, et al. 1991. Polyhalogenated dibenzo-p-dioxins and dibenzofurans 

and the immune system. Arch Toxicol 65:213-219. 
CDDs 624 

8. REFERENCES 

*Neubert R, Jacob-M6ller U, Stahlmann R, et al. 1990b. Polyhalogenated dibenzo-p-dioxins and 

dibenzofurans and the immune system. I. Effects on peripheral lymphocyte subpopulations of a non-human 

primate (Callithrix jacchus) after treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Arch Toxicol 

64:345-359. 

*Neubert R, Maskow L, Delgado I, et al. 1995. Chlorinated dibenzo-p-dioxins and dibenzofurans and the 

human immune system. 2. In vitro proliferation of lymphocytes from workers with quantified 

moderately-increased body burdens. Life Sci 56:421-436. 

*Neubert R, Maskow L, Webb J, et al. 1993. Chlorinated dibenzo-p-dioxins and dibenzofurans and the 

human immune system. 1. Blood cell receptors in volunteers with moderately increased body burdens. Life 

Sci 53:1995-2006. 

*Neuhold LA, Gonzalez FJ, Jaiswal AK, et al. 1986. Dioxin inducible enhancer region upstream from the 

mouse Pi-450 gene and interaction with a heterologous SV40 promoter. DNA 5:403-411. 

*Neuhold LA, Shirayoshi Y, Ozato K, et al. 1989. Regulation of mouse CYP1A1 gene expression by 

dioxin: requirements of two cis-acting elements during induction. Mol Cell Biol 9:2378-2386. 

Neumann CM, Brauner JA, Kerkvliet NI. 1991. T cell activation in vivo via the CD3 receptor: Effect of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) exposure [Abstract 48]. Toxicologist 11:36. 

*Neumann CM, Oughton JA, Kerkvliet NI. 1993. Anti-CD3-induced T-cell activation. II. Effect of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Int J Immunopharmacol 15(4):543-50. 

Neupert M, Grupe A, Weis H. 1988. Stability of polybrominated dibenzo-p-dioxins and dibenzofurans 

under laboratory handling conditions. Chemosphere 17:1089-1097. 

Newsome WH, Iverson F, Ryan JJ, et al. 1984. Chlorinated compounds in tissues of chickens raised on 

pentachlorophenol-contaminated litter. Food Addit Contam 1:3-9. 

Nguyen TNP, Bui SH, Dan QV, et al. 1989. Dioxin levels in adipose tissues of hospitalized women living 

in the south of Vietnam in 1984-85 with a brief review of their clinical histories. Chemosphere 19:933-936. 

Niemann RA. 1986. Surrogate-assisted determination of 2,3,7,8-tetrachlorodibenzo-para-dioxin in fish by 

electron-capture capillary gas-chromatography. J Assoc Off Anal Chem 69:976-980. 

Niimi AJ, Oliver BG. 1986. Biological half-lives of chlorinated dibenzo-p-dioxins and dibenzofurans in 

rainbow trout (Salmo gairdneri). Environ Toxicol Chem 5:49-53. 

*Niimi AJ, Oliver BG. 1989a. Distribution of polychlorinated biphenyl congeners and other halocarbons in 

whole fish and muscle among Lake Ontario salmonids. Environ Sci Technol 23:83-88. 

*Niimi AJ, Oliver BG. 1989b. Assessment of relative toxicity of chlorinated dibenzo-p-dioxins, and 

biphenyls in Lake Ontario salmonids to mammalian systems using toxic equivalent factors (TEF). 

Chemosphere 18(7-8):1413-1423. 
CDDs 625 

8. REFERENCES 

NIOSH. 1984. 2,3,7,8-Tetrachlorodibenzo-p-dioxin. Current intelligence bulletin 40. Cincinnati, OH: 

Department of Health and Human Services, National Institute for Occupational Safety and Health. 

NIOSH. 1992. NIOSH Recommendations for Occupational Safety. U. S. Department of Health and Human 

Services, Public Health Service, Centers for Disease Control, National Institute for Occupational Safety and 

Health, Division of Standards Development and Technology Transfer. Cincinnati, OH. 

*NIOSH. 1997. NIOSH pocket guide tochemical hazards. U.S. Depatment of Health and Human Services. 

Public Health Services. Centers for Disease Control and Prevention. National Institute for Occupational 

Safety and Health. 

Nirmalakhandan NN, Speece RE. 1989. Prediction of aqueous solubility of organic chemicals based on 

molecular structure: 2. Application to PNAs, PCBs, PCDDs, etc. Environ Sci Technol 23:708-713. 

Nisbet ICT, Paxton MB. 1982. Statistical aspects of three-generation studies of the reproductive toxicity of 

TCDD and 2,4,5-T. Am Stat 36:290-298. 

Noble R, Walker S, Sherman P, et al. 1991. Dioxins in the analytical laboratory. IARC Sci Publ 

108:181-196 

*Nobmann ED Byers T Lanier AP et al. 1992. The diet of Alaska native adults: 1987-1988. Am J Clin Nutr 

55:1024-1032. 

Nolan RJ, Smith FA, Hefner JG. 1979. Elimination and tissue distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in female guinea pigs following a single oral dose. Toxicol Appl Pharmacol 

48:A162. 

Norback DH, Allen JR. 1973. Biological responses of the nonhuman primate, chicken, and rat to 

chlorinated dibenzo-p-dioxin ingestion. Environ Health Perspect 5:233-240. 

*Norback DH, Engblom JF, Allen JR. 1975. Tissue distribution and extraction of octachlorodibenzopara-dioxin in the rat. Toxicol Appl Pharmacol 32:330-338. 

Noren K. 1988. Changes in the levels of organochlorine pesticides, polychlorinated biphenyls, 

dibenzo-p-dioxins and dibenzofurans in human milk from Stockholm, Sweden 1972-1985. Chemosphere 

17:39-50. 

*Noren K, Sj÷vall J. 1987. Analysis of organochlorine pesticides, polychlorinated biphenyls, 

dibenzo-p-dioxins and dibenzofurans in human milk by extraction with the lipophilic gel Lipidex« Noren K. 

J Chromator 422:103-115. 

Norris LA. 1981. The movement, persistance, and fate of the phenoxy herbicides and TCDD in forests. 

Residue Reviews 80:65-135. 

*Norstrom RJ, Simon M, Muir DCG. 1990. Polychlorinated dibenzo-p-dioxins and dibenzofurans in 

marine mammals in the Canadian north. Environ Pollut 66:1-19. 

*Norwood CB, Hackett M, Pruell RJ, et al. 1989. Polychlorinated dibenzo-p-dioxins and dibenzofurans in 

selected estuarine sediments. Chemosphere 18:553-560. 
CDDs 626 

8. REFERENCES 

*Nosek JA, Craven SR, Sullivan JR, et al. 1992. Metabolism and disposition of 2,3,7,8-tetrachlorodibenzop-dioxin in ring-necked pheasant hens, chicks and eggs. J Toxicol Environ Health 35(3):153-164. 

Nottrodt IA, Ballschmiter K. 1986. Causes for and reduction strategies against emissions of PCDD/PCDF 

from waste incineration plants-interpretations of recent measurements. Chemosphere 15:1225-1237. 

*NRC. 1993. Pesticides in the Diets of Infants and Children. National Research Council. Washington DC: 

National Academy Press. 

NRCC. 1981. Polychlorinated dibenzo-p-dioxins: Criteria for their effects on man and his environment. 

Ottawa, Canada: National Research Council of Canada, Associate Committee on Scientific Criteria for 

Environmental Quality. NRCC no. 18574. 

*NREPC. 1986. National Resources and Environmental Protection Cabinet. Department for Environmental 

Protection. Frankfort, KY: Proposed Regulation. 401 KAR 63:022. 

*NREPC. 1988. National Resources and Environmental Protection Cabinet. Kentucky Waste Management 

Regulations. Frankfort, KY: Department for Environmental Protection, Division of Waste Management. 

401 KAR 31:030, 040, and 160. 

*NTP. 1982a. Carcinogenesis bioassay of 2,3,7,8-tetrachlorodibenzo-p-dioxin (CAS no. 1746-01-6) in 

Swiss-Webster mice (dermal study). Bethesda, MD: Carcinogenesis Testing Program, National Cancer 

Institute, National Institute of Health. Research Triangle Park, NC: National Toxicology Program. (NIH) 

DHHS publication no 82-1757. 

*NTP. 1982b. Carcinogenesis bioassay of 2,3,7,8-tetrachlorodibenzo-p-dioxin (CAS no. 1746-01-6) in 

Osborne-Mendel rats and B6C3F1 mice (gavage study). Bethesda, MD: Carcinogenesis Testing Program, 

National Cancer Institute, National Institute of Health. Research Triangle Park, NC: National Toxicology 

Program. (NIH) DHHS publication no 82-1765. 

*NTP. 1989. Fifth annual report on carcinogens. Summary. U.S. Department of Health and Human 

Services, Public Health Services, Research Triangle Park, NC, by Technical Resources, Inc., Rockville, MD. 

*NTP. 1995. Long term technical reports and short term toxicity study reports. National Toxicology 

Program. National Institute of Occupations Safety and Health, Research Triangle Park, NC. 

*NTP. 1998. Eighth Annual Report on Carcinogens. U.S. Department of Health and Human Services, 

Public Health Services, Research Triangle Park, NC. 

http://www.ehis.niehs.nih.gov/roc/eighth/chemicals/tcdd.pdf. 

*NTP. 1998. Eighth report on carcinogens. 1998 Summary. U.S. Department of Health and Human 

Services. Public Health Services. National Toxicology Program. 

*Nygren M, Hansson M, Sj÷str÷m M, et al. 1988. Development and validation of a method for 

determination of PCDDs and PCDFs in human blood plasma: A multivariate comparison of blood and 

adipose tissue levels between Vietnam veterans and matched controls. Chemosphere 17:1663-1692. 

*Nygren M, Rappe C, Lindstrom G, et al. 1986. Identification of 2,3,7,8-substituted polychlorinated 

dioxins and dibenzofurans in environmental and human samples. In: Rappe C, Choudhary G, Keith LH, 

eds. Chlorinated dioxins and dibenzofurans in perspective. Chelsea, MI: Lewis Publishers, Inc., 17-34. 
CDDs 627 

8. REFERENCES 

*NYPDWS. 1989. New York Public Drinking Water Standards. Code revision, effective 1/9/89. 

*NYWQSGV. 1987. New York Water Ambient Quality Standards and Guidance Values, 4/1/87. 

O'Keefe P, Hilker D, Meyer C, et al. 1984. Tetrachlorodibenzo-p-dioxins and tetrachlorodibenzofurans in 

Atlantic Coast striped bass and in selected Hudson River fish, waterfowl and sediments. Chemosphere 

13:849-860. 

*O'Keefe P, Meyer C, Smith R, et al. 1986. Reverse-phase adsorbent cartridge for trapping dioxins in 

drinking water. Chemosphere 15:1127-1134. 

O'Keefe PW, Smith RM, Hilker DR, et al. 1989. Interlaboratory validation of PCDD and PCDF 

concentrations found in municipal incinerator emissions. Chemosphere 18:185-192. 

Oberg T, Bergstrom JGT. 1987. Emission and chlorination pattern of PCDD/PCDF from indicator 

parameters. Chemosphere 16:1221-1230. 

*Oehme M, Mano S, Bjerke B. 1989. Formation of polychlorinated dibenzofurans and dibenzo-p-dioxins 

by production processes for magnesium and refined nickel. Chemosphere 18:1379-1389. 

*Oehme M, Mano S, Mikalsen A, et al. 1986. Quantitative method for the determination of femtogram 

amounts of polychlorinated dibenzo-p-dioxins and dibenzofurans in outdoor air. Chemosphere 15:607-617. 

*Oehme M, Mano S, Mikalsen A, et al. 1987. Formation and presence of polyhalogenated and polycyclic 

compounds in the emissions of small and large scale municipal waste incinerators. Chemosphere 

16:143-153. 

Ogaki J, Takayama K, Miyata H, et al. 1987. Levels of PCDDs and PCDFs in human tissues and various 

foodstuffs in Japan. Chemosphere 16:2047-2056. 

*Okey AB, Riddick DS, Harper PA. 1994. The Ah receptor: Mediator of the toxicity of 

2.3.7.8-tetrachlorodibenzo-p-dioxin (TCDD) and related compounds. Toxicol Lett 70(1):1-22. 

*Okey AB, Vella LM, Harper PA. 1989. Detection and characterization of a low affinity form of cytosolic 

Ah receptor in livers of mice nonresponsive to induction of cytochrome P1-450 by 3-methylcholanthrene. 

Mol Pharmacol 35: 823-830. 

*Okino ST, Pendurthi UR, Tukey RH. 1992. Phorbol esters inhibit the dioxin receptor-mediated 

transcriptional activation of the mouse CYP1A1 and CYP1A2 genes by 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

J Biol Chem 267:6991-6998. 

*Oku A, Tomari K, Kamada T, et al. 1995. Destruction of PCCDs and PCDFs. A convenient method using 

alkali-metal hydroxide in 1,3,-dimethyl-2-imidazolidinone (DMI). Chemosphere 31(8):3873-3878. 
CDDs 628 

8. REFERENCES 

Olie K, Schecter A, Constable J, et al. 1989. Chlorinated dioxin and dibenzofuran levels in food and 

wildlife samples in North and South Vietnam. Chemosphere 19:493-496. 

Olie K, Slot PC, Wever H. 1989. Decomposition of octachlorodibenzofuran and formation of the hepta 

congeners during GLC analysis. Chemosphere 19:103-108. 

Olie K, Vermeulen PL, Hutzinger O. 1977. Chlorodibenzo-p-dioxins and chlorodibenzofurans are trace 

compounds in fly ash and flue gas of some municipal incinerators in the Netherlands. Chemosphere 

8:455-459. 

*Oliver RM. 1975. Toxic effects of 2,3,7,8-tetrachlorodibenzo-1,4-dioxin in laboratory workers. Br J Ind 

Med 32:49-53. 

*Olson JR. 1986. Metabolism and disposition of 2,3,7,8-tetrachlorodibenzo-p-dioxin in guinea pigs. 

Toxicol Appl Pharmacol 85:263-273. 

Olson JR, Bellin JS, Barnes DG. 1989. Reexamination of data used for establishing toxicity equivalence 

factors (TEFs) for chlorinated dibenzo-para-dioxins and dibenzofurans (CDDs and CDFs). Chemosphere 

18:371-381. 

Olson JR, Gasiewicz TA, Geiger LE, et al. 1983. The metabolism of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

mammalian systems. In: Coulston R, Pocchiari F, eds. Accidental exposure to dioxins: Human health 

aspects. New York, NY: Academic Press, 81-100. 

*Olson JR, Gasiewicz TA, Neal RA, et al. 1980b. Tissue distribution excretion, and metabolism of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the Golden Syrian Hamster. Toxicol Appl Pharmacol 

56:78-85. 

*Olson JR, Holscher MA, Neal RA. 1980a. Toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the Golden 

Syrian hamster. Toxicol Appl Pharmacol 55:67-78. 

*Olson JR, McGarrigle BP. 1992. Comparative developmental toxicity of 2,3,7,8-tetrachloro-dibenzop-dioxin (TCDD). Chemophere 25(1-2):71-74. 

*Olson JR, Wroblewski VJ. 1985. Metabolism of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in isolated 

hepatocytes from guinea pigs and rats. Chemosphere 14:979-982. 

*Ono M, Kashima Y, Wakimoto T, et al. 1987. Daily intake of PCDDs and PCDFs by Japanese through 

food. Chemosphere 16:1823-1828. 

Ono M, Wakimoto T, Tatsukawa R, et al. 1986. Polychlorinated dibenzo-p-dioxins and dibenzofurans in 

human adipose tissues of Japan. Chemosphere 15:1629-1634. 

Onuska FI. 1988. Trials, tribulations and trends in analysis of chlorinated dibenzo-para-dioxins [Abstract]. 

Abstracts of Papers of the American Chemical Society 195:203. 

*Onuska FI, Mudroch A, Terry KA. 1983. Identification and determination of trace organic substances in 

sediment cores from the western basin of Lake Ontario. J Great Lakes Res 9:169-182. 
CDDs 629 

8. REFERENCES 

*Oppelt ET. 1991. Air emissions from the incineration of hazardous waste. In: Mehlman MA, ed. 

Advances in modern environmental toxicology. Vol XIX. Health hazards and risks from exposure to 

complex mixtures and air toxic chemicals. Princeton, NJ: Princeton Scientific Publishing Co., Inc., 1-26. 

Opperhuizen A. 1986. Bioconcentration of hydrophobic chemicals in fish. ASTM Spec Tech Publ 

9:304-315. 

Opperhuizen A, Sijm DTHM. 1990. Bioaccumulation and biotransformation of polychlorinated 

dibenzo-p-dioxins and dibenzofurans in fish. Environ Toxicol Chem 9:175-186. 

*Orazio CE, Kapila S, Puri RK, et al. 1992. Persistence of chlorinated dioxins and furans in the soil 

environment. Chemosphere 25(7-10):1469-1474. 

*Orban JE, Stanley JS, Schwemberger JG, et al. 1994. Dioxins and dibenzofurans in adipose tissue of the 

general U.S. population and selected subpopulations. Am J Public Health 84:439-445. 

Orth RG, Ritchie C, Hileman F. 1989. Measurement of the photoinduced loss of vapor phase TCDD. 

Chemosphere 18:1275-1282. 

*OSHA. 1987. Access to employee exposure and medical records. U.S. Department of Labor, 

Occupational Safety and Health Administration. Code of Federal Regulations. 29 CFR 1910.20. 

*OTA. 1990. Neurotoxicology: Identifying and controlling poisons of the nervous system. Office of 

Technology Assessment, Washington, DC. OTA-BA-438. 

*Ott MG, Holder BB, Olson RO. 1980. A mortality analysis of employees engaged in the manufacture of 

2,4,5-trichlorophenoxyacetic acid. J Occup Med 22:47-50. 

*Ott MG, Messerer P, Zober A. 1993. Assessment of past occupational exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin using blood lipid analysis. Int Arch Environ Health 65:1-8. 

*Ott MG, Olson RA, Cook RR. 1987. Cohort mortality study of chemical workers with potential exposure 

to the higher chlorinated dioxins. J Occup Med 29:422-429. 

*Ott MG, Zober A. 1996. Cause specific mortality and cancer incidence among employees exposed to 

2,3,7,8-TCDD after a 1953 reactor accident. Occup Environ Med 53:606-612. 

*Ott MG, Zober A, Germann C. 1994. Laboratory results for selected target organs in 138 individuals 

occupationally exposed to TCDD. Chemosphere 29:9-11. 

*Oughton JA, Pereira CB, DeKrey GK, et al. 1995. Phenotypic analysis of spleen, thymus, and peripheral 

blood cells in aged C57BL/6 mice following long-term exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Fund Appl Toxicol 25:60-69. 

*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: Saunders, 222-238. 

Ozvacic V. 1986. A review of stack sampling methodology for PCDDs/PCDFs. Chemosphere 

15:1173-1178. 
CDDs 630 

8. REFERENCES 

*Päpke O, Ball M, Lis A. 1992. Various PCDD/PCDF patterns in human blood resulting from different 

occupational exposures. Chemosphere 25:1101-8. 

*Päpke O, Ball M, Lis A. 1993. Potential occupational exposure of municipal waste incinerator workers 

with PCDD/PCDF. Chemosphere 27:203-9. 

Päpke O, Ball M, Lis ZA, et al. 1989a. PCDD and PCDF in indoor air of kindergartens in northern W. 

Germany. Chemosphere 18:617-626. 

*Päpke O, Ball M, Lis ZA, et al. 1989b. PCDD/PCDF in whole blood samples of unexposed persons. 

Chemosphere 19:941-948. 

Päpke O, Ball M, Lis ZA, et al. 1990. Determination of PCDD/PCDF in whole blood from persons 

involved in fire incidents. Chemosphere 20:959-966. 

Palansky J, Kapila S, Manahan SE, et al. 1986. Studies on vapor phase transport and role of dispersing 

medium on mobility of 2,3,7,8-TCDD in soil. Chemosphere 15:1389-1396. 

*Park J-Y K, Shigenaga MK, Ames BN. 1996. Induction of cytochrome P4501A1 by 

2,3,7,8-tetrachlorodibenzo-p-dioxin or indolo(3,2-b) carbazole is associated with oxidative DNA damage. 

Proc Natl Acad Sci USA 93: 2322-2327. 

Parlar H. 1984. Geochemical induced degradation of environmental chemicals. Presenius Z Anal Chem 

319:114-118. 

*Parsons JR. 1992. Influence of suspended sediment on the biodegradation of chlorinated 

dibenzo-p-dioxins. Chemosphere 25(12):1973-1980 

Parsons JR, Storms MCM. 1989. Biodegradation of chlorinated dibenzo-p-dioxins in batch and continuous 

cultures of strain JB1. Chemosphere 19:1297-1308. 

*Patterson DG Jr, Fingerhut MA, Roberts DW, et al. 1989a. Levels of polychlorinated dibenzo-p-dioxins 

and dibenzofurans in workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Am J Ind Med 16:135-146. 

*Patterson DG Jr, Fürst P, Alexander LR, et al. 1989b. Analysis of human serum for PCDDs/PCDFs: A 

comparison of three extraction procedures. Chemosphere 19:89-96. 

*Patterson DG Jr, Fürst P, Henderson LO, et al. 1989d. Partitioning of in vivo bound PCDDs/PCDFs among 

various compartments in whole blood. Chemosphere 19:135-142. 

*Patterson DG Jr, Hampton L, Lapeza Cr Jr, et al. 1987a. High-resolution gas chromatographic/highresolution mass spectrometric analysis of human serum on a whole-weight and lipid basis for 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Anal Chem 59:2000-2005. 

*Patterson DG Jr, Hoffman RE, Needham LL, et al. 1986a. 2,3,7,8-Tetrachlorodibenzo-p-dioxin levels in 

adipose tissue of exposed and control persons in Missouri. JAMA 256:2683-2686. 

*Patterson DG Jr, Holler JS, Belser WT, et al. 1987b. Determination of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) in human adipose tissue on whole-weight and lipid basis. Chemosphere 16:935-936. 
CDDs 631 

8. REFERENCES 

Patterson DG Jr, Holler JS, Lapeza CR Jr, et al. 1986b. High-resolution gas chromatographic/highresolution mass spectrometric analysis of human adipose tissue for 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Anal Chem 58:705-713. 

Patterson DG Jr, Holler JS, Smith SJ, et al. 1986c. Human adipose data for 2,3,7,8-tetrachlorodibenzo-pdioxin in certain US samples. Chemosphere 15:2055-2060. 

Patterson DG Jr, Needham LL, Pirkle JL, et al. 1988. Correlation between serum and adipose tissue levels 

of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 50 persons from Missouri. Arch Environ Contam Toxicol 

17:139-143. 

Patterson DG Jr, Reddy W, Barnhart ER, et al. 1989c. Synthesis and analytical characterization of all tetra 

to octachlorodibenzo-p-dioxins. Chemosphere 19:233-240. 

Patterson DG Jr, Turner WE, Isaacs SG, et al. 1990. A method performance evaluation and lessons learned 

after analyzing more than 5,000 human adipose tissue, serum, and breast milk samples for polychlorinated 

dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs). Chemosphere 20:829-863. 

*Patterson Jr. DG, Todd GD, Turner WE, et al. 1994. Levels of non-ortho-substituted (coplanar), mono- and 

di-ortho-substituted polychlorinated biphenyls, dibenzo-p-dioxins, and dibenzofurans in human serum and 

adipose tissue. Environ Health Perspect 102(supp 1):195-204. 

*Paulson KE, Darnell J Jr, Rushmore T, et al. 1990. Analysis of the upstream elements of the xenobiotic 

compound-inducible and positionally regulated glutathione S-transferase Ya gene. Mol Cell Biol 

10:1841-1852. 

*Paustenbach D, Finley B, Lau V, et al. 1991. An evaluation of the inhalation hazard posed by 

diozin-contaminated soils. J Air Waste Manage Assn. 41:1334-1340. 

Paustenbach DJ, Shu HP, Murray FJ. 1986. A critical examination of assumptions used in risk assessments 

of dioxin contaminated soil. Regul Toxicol Pharmacol 6:284-307. 

*Paustenbach DJ, Wenning RJ, Lau V, et al. 1992. Recent developments on the hazards posed by 

2,3,7,8-tetrachlorodibenzo-p-dioxin in soil: Implications for setting risk-based cleanup levels at residential 

and industrial sites. J Toxicol Environ Health 36(2):103-150. 

*Pazdernik TL, Rozman KK. 1985. Effect of thyroidectomy and thyroxine on 2,3,7,8-tetrachlorodibenzop-dioxin-induced immunotoxicity. Life Sci 36:695-703. 

*Pazderova-Vejlupkova J, Nemcova M, Pickova J, et al. 1981. The development and prognosis of chronic 

intoxication by tetrachlorodibenzo-p-dioxin in men. Arch Environ Health 36:5-11. 

*Pearson RG, McLaughlin DL, McIlveen WD. 1990. Concentrations of PCDD and PCDF in Ontario soils 

from the vicinity of refuse and sewage sludge incinerators and remote rural and urban locations. 

Chemosphere 20(10-12):1543-1548. 

Pedersen LG, Darden TA, Oatley SJ, et al. 1986. A theoretical study of the binding of polychlorinated 

biphenyls (PCBs), dibenzodioxins, and dibenzofuran to human plasma prealbumin. Journal of Medicinal 

Chemistry 29:2451-2457. 
CDDs 632 

8. REFERENCES 

*Pegg DG, Hewett WR, McCormack K, et al. 1976. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on renal 

function in the rat. J Toxicol Environ Health 2:55-65. 

*Pegram RA, Diliberto JJ, Moore TC, et al. 1995. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) 

distribution and cytochrome P4501A induction in young adult and senescent male mice. Toxicol Letters 

76:119-126. 

Pelizzetti E, Borgareelo M, Minero C, et al. 1988. Photocatalytic degradation of polychlorinated dioxins 

and polychlorinated biphenyls in aqueous suspensions of semiconductors irradiated with simulated solar 

light. Chemosphere 17:499-510. 

*Peper M, Klett M, Frentzel-Beyme R, et al. 1993. Neuropsychological effects of chronic exposure to 

environmental dioxins and furans. Environ Res 60:124-135. 

Perdew GH. 1991. Chemical cross-linking of the cytosolic and nuclear forms of the Ah receptor: 

determination of protein composition [Abstract 12]. Toxicologist 11:27. 

Perdew GH. 1992. Chemical cross-linking of the cytosolic and nuclear form of the Ah receptor in hepatoma 

cell line 1c1c7. Biochem Biophys Res Comm 182:55-62. 

*Pereira WE, Rostad CE, Sisak ME. 1985. Geochemical investigation of polychlorinated dibenzo-p-dioxins 

in the subsurface environment at an abandoned wood-treatment facility. Environ Toxicol Chem 4:629-639. 

*Pesatori AC, Consonni S, Tironi A, et al. 1993. Cancer in a young population in a dioxin-contaminated 

area. Int J Epidemiol 22:1010-1013. 

*Pesatori AC, Zocchetti C, Guercilena S, et al. 1998. Dioxin exposure and non-malignant health effects: A 

mortality study. Occup Environ Med 55:126-131. 

Peterson RE, Seefeld MD, Christian BJ, et al. 1984. The wasting syndrome in 2,3,7,8-tetrachlorodibenzop-dioxin toxicity: Basic features and their interpretation. In: Poland A, Kimbrough RD, eds. Banbury Rep 

18. Cold Spring Harbor Laboratory, 291-308. 

*Peterson RE, Theobald HM, Kimmel GL. 1993. Developmental and reproductive toxicity of dioxins and 

related compounds: Cross-species comparisons. Crit Rev Toxicol 23:283-335. 

Petreas MX, Wiesmuller T, Palmer FH, et al. 1992. Aquatic life as biomonitors of dioxin/furan and 

coplanar polychlorinated biphenyls contamination in the Sacramento-San Joaquin River Delta. 

Chemosphere 25(4):621-631. 

Petty J, Rodgers P. 1985. Source, distribution and fate-workshop summary. In: Kamrin MA, Rodgers PW, 

eds. Dioxins in the environment. Washington, DC: Hemisphere Publishing Corporation, 299-306. 

Petty JD, Stalling DL, Smith LM, et al. 1983. Occurrence and potential impact of PCDFs and PCDDs in 

aquatic ecosystems. Trace Subst Environ Health 17:96-102. 

Pfeiffer Cd. 1976. Determination of chlorinated dibenzo-p-dioxins in pentachlorophenol by liquid 

chromatography. J Chromatogr Sci 14:386-391. 
CDDs 633 

8. REFERENCES 

*Phiet PH, Anh TK, Vu DQ, et al. 1989. Preliminary observations on the clinical histories, polychlorinated 

dibenzodioxin and dibenzofuran tissue levels and 2,3,7,8-TCDD toxic equivalents of potentially 

dioxin-exposed patients living in the south of Vietnam. Chemosphere 19:937-940. 

Philippi M, Krasnobage V, Zeyer J, et al. 1981. Fate of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in 

microbial cultures and soil under laboratory conditions. FEMS Symp 12:2210-2233. 

*Philippi M, Schmid J, Wipf HK, et al. 1982. A microbial metabolite of TCDD. Experientia 38:659-661. 

*Phillips LJ, Birchard GF. 1991. Regional variations in human toxics exposure in the USA: An analysis 

based on the national human adipose tissue survey. Arch Environ Contam Toxicol 21:159-168. 

*PHS. 1992. CCEHRP interim statement on dioxins. Committee to Coordinate Environmental Health and 

Related Programs, interagency committee. 

*Phuong NTN, Hung BS, Vu DQ, et al. 1989b. Dioxin levels in adipose tissues of hospitalized women 

living in south of Vietnam in 1984-85 with a brief review of their clinical histories. Chemosphere 

19:933-936. 

*Phuong NTN, Thuy TT, Phuong PK. 1989a. An estimate of reproductive abnormalities in women 

inhabiting herbicide sprayed and non-herbicide sprayed areas in the south of Vietnam, 1952-1981. 

Chemosphere 18:843-846. 

Piper WN, Rose RQ, Gehring PJ. 1973. Excretion and tissue distribution of 2,3,7,8-tetrachlorodibenzo-pdioxin in the rat. Environ Health Perspect 5:241-244. 

Piper WN, Ruangwises S, Bestervelt LL, et al. 1991. Leydig cell testosterone production and P45017 

(17-hydroxylase and C17-20lyase) activities after in vitro exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) [Abstract 997]. Toxicologist 11:262. 

*Pirkle J, Wolfe W, Patterson D, et al. 1989. Estimates of the half-life of 2,3,7,8-tetrachlorodibenzop-dioxin in Vietnam veterans of Operation Ranch Hand. J Toxicol Environ Health 27:165-171. 

*Piskorska-Pliszczynska J, Astroff B, Zacharewski T, et al. 1991. Mechanism of action of 

2,3,7,8-tetrachlorodibenzo-p-dioxin antagonists: Characterization of [125I]-6-methyl-8-iodo1,3-dichlorodibenzofuran-Ah receptor complexes [Abstract 1047]. Toxicologist 11: 272. 

Pitea D, Bonati L, Lasagni M, et al. 1989a. The combustion of municipal solid wastes and PCDF emissions: 

Part 1. PCDD and PCDF in MSW. Chemosphere 18:1457-1464. 

Pitea D, Cosentino U, Lasagni M, et al. 1989b. The combustion of municipal wastes and PCDD and PCDF 

emissions: Part 3. PCDD and PCDF in fly ash. Chemosphere 18:1475-1483. 

Pitea D, Lasagni M, Bonati L, et al. 1989c. The combustion of municipal solid wastes and PCDD and 

PCDF emissions: Part 2. PCDD and PCDF in stack gases. Chemosphere 18:1465-1474. 

*Pitot HC, Goldsworthy T, Campbell HA. 1980. Quantitative evaluation of the promotion by 

2,3,7,8-tetrachlorodibenzo-p-dioxin of hepatocarcinogenesis from diethylnitrosamine. Cancer Res 

40:3616-3620. 
CDDs 634 

8. REFERENCES 

*Plimmer JR, Klingebiel UI, Crosby DG, et al. 1973. Photochemistry of dibenzo-p-dioxins. In: Blair EH, 

ed. Chlorodioxins-origin and fate. Washington, DC: Advances in Chemistry Series, 44-54. 

Pluess N, Poiger H, Hohback C, et al. 1988. Subchronic toxicity of some chlorinated dibenzofurans 

(PCDFs) and a mixture of PCDFs and chlorinated dibenzodioxins (PCDDs) in rats. Chemosphere 

17:973-984. 

*Pluim HJ, Boersma ER, Kramer I, et al. 1994c. Influence of short-term dietary measures on dioxin 

concentrations in human milk. Environ Health Perspect 102:968-971. 

*Pluim HJ, deVijlder JJM, Olie K, et al. 1993c. Effects of pre and postnatal exposure to chlorinated dioxins 

and furans on human neonatal thyroid hormone concentrations. Environ Health Perspect 101:504-508. 

*Pluim HJ, Koope JG, Olie K, et al. 1994a. Clincal laboratory manifestations of exposure to background 

levels of dioxins in the perinatal period. Act Paediatr 83:583-587. 

*Pluim HJ, Koppe GJ, Oilie K, et al. 1992. Effect of dioxins on thyroid function in newborn babies. Lancet 

3339:1303. 

*Pluim HJ, van der Goot M, Olie K, et al. 1996. Missing effects of background dioxin exposure on 

development of breast-fed infants during the first half year of life. Chemosphere 33:1307-1315. 

*Pluim HJ, van der Slikke JW, Koppe JG, et al. 1993a. Levels of PCDDs and PCDFs in human milk: 

Dependence on several parameters and dietary habits. Chemosphere 26(10):1889-1894. 

*Pluim HJ, van der slikke JW, Olie K, et al. 1994b. Dioxins and vitamin K status of the newborn. J 

Environ Sci Health A29:793-802. 

*Pluim HJ, Wever J, Koppe JG, et al. 1993b. Intake and fecal excretion of chlorinated dioxins and 

dibenzofurans in breast-fed infants at different ages. Chemosphere 26:1947-1952. 

*Pocchiari F, Silano V, Zampieri A. 1979. Human health effects from accidental release of 

tetrachlorodibenzo-p-dioxin (TCDD) at Seveso, Italy. Ann NY Acad Sci 320:311-320. 

*Podoll RT, Jaber HM, Mill T. 1986. Tetrachlorodibenzodioxin: Rates of volatilization and photolysis in 

the environment. Environ Sci Technol 20:490-492. 

Poellinger L, Cuthill S, Wihelmsson A. 1991. The dioxin receptor: A signal-dependent gene regulatory 

protein [Abstract 13]. Toxicologist 11:27. 

Poellinger L, Lund J, Soderkvist P, et al. 1986. The receptor for 2,3,7,8-TCDD: Structure, possible 

functions and relationship to endocrine receptors: An overview. Chemosphere 15:9-12. 

*Pohjanvirta R, Tuomisto L, Tuomisto J. 1989. The central nervous system may be involved in TCDD 

toxicity. Toxicology 58:167-174. 
CDDs 635 

8. REFERENCES 

Pohjanvirta R, Unkila M, Tuomisto J. 1993a. Comparative acute lethality of 2,3,7,8-tetrachlorodibenzo-pdioxin (TCDD), 1,2,3,7,8-pentachlorodibenzo-p-dioxin and 1,2,3,4,7,8-hexachlorodibenzo-p-dioxin in the 

most TCDD-susceptible and the most TCDD-resistant rat strain. Pharmacol Toxicol 73(1):52-6. 

*Pohjanvirta R, Unkila M, Tuomisto JT, et al. 1993b. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) on plasma and tissue beta-endorphin-like immunoreactivity in the most TCDD-susceptible and the 

most TCDD-resistant rat strain. Life Sci 53(19):1479-1487. 

*Pohjanvirta R, Vartiainen T, Uusi-Rauva A, et al. 1990. Tissue distribution, metabolism, and excretion of 

[14C]-TCDD in a TCDD-susceptible and a TCDD-resistant rat strain. Pharmacol Toxicol 66:92-100. 

*Pohl H, DeRosa C, Holler J. 1995. Public health assessment for dioxins exposure from soil. Chemosphere 

1995 31(1):(2437-2454). 

*Pohl HR, Hibbs BF. 1996. Breast-feeding exposure of infants to environmental contaminants - a public 

health risk assessment viewpoint: chlorinated dibenzodioxins and chlorinated dibenzofurans. Toxicol 

Indust Health 12(5):105-114. 

*Poiger H, Buser HR. 1984. The metabolism of TCDD in the dog and rat. Banbury Rep 18:39-47. 

*Poiger H, Schlatter C. 1980. Influence of solvents and adsorbents on dermal and intestinal absorption of 

TCDD. Food Cosmet Toxicol 18:477-481. 

Poiger H, Schlatter C. 1983. Animal toxicology of chlorinated dibenzo-p-dioxins. Chemosphere 

12:453-462. 

*Poiger H, Schlatter C. 1985. Influence of phenobarbital and TCDD on the hepatic metabolism of TCDD in 

the dog. Experientia 41:376-378. 

*Poiger H, Schlatter C. 1986. Pharmacokinetics of 2,3,7,8-TCDD in man. Chemosphere 15:1489-1494. 

*Poiger H, Weber H, Schlatter CH. 1982. Special aspects of metabolism and kinetics of TCDD in dogs and 

rats: Assessment of toxicity of TCDD-metabolites(s) in guinea pigs. In: Hutzinger O, Frei RW, Merian E, 

et al., eds. Chlorinated dioxins and related compounds: Impact on the environment. New York, NY: 

Pergamon Press, 317-325. 

*Poland A, Glover E. 1973a. Chlorinated dibenzo-p-dioxins: Potent inducers of delta-aminolevulinic acid 

synthetase and aryl hydrocarbon hydroxylase: 2. Study of the structure-activity relationship. Mol 

Pharmacol 9:736-747. 

*Poland A, Glover E. 1973b. Mechanism of toxicity of the chlorinated dibenzo-p-dioxins. Environ Health 

Perspect 5:245-251. 

*Poland A, Glover E. 1979. An estimate of the maximum in vivo covalent binding of 

2,3,7,8-tetrachlorodibenzo-p-dioxin to rat liver protein, ribosomal RNA and DNA. Cancer Res 

39:3341-3344. 

*Poland A, Glover E. 1980. 2,3,7,8-Tetrachlorodibenzo-p-dioxin segregation of toxicity with the Ah locus. 

Mol Pharmacol 17:86-94. 
CDDs 636 

8. REFERENCES 

*Poland A, Glover E. 1987. Variation in the molecular mass of the Ah receptor among vertebrate species 

and strains of rats. Biochem Biophys Res Comm 146:1439-1449. 

*Poland A, Glover E. 1990. Characterization and strain distribution pattern of the murine Ah receptor 

specified by the Ahd and Ahb-3 alleles. Mol Pharmacol 38:306-312. 

*Poland A, Glover E, Bradfield CA. 1991. Characterization of polyclonal antibodies to the Ah receptor 

prepared by immunization with a synthetic peptide hapten. Mol Pharmacol 39:20-26. 

*Poland A, Glover E, Kende AS. 1976. Stereospecific, high affinity binding of 

2,3,7,8-tetrachlorodibenzo-p-dioxin by hepatic cytosol: Evidence that the binding species is a receptor for 

induction of aryl hydrocarbon hydroxylase. J Biol Chem 251:4936-4946. 

*Poland A, Greenlee WF, Kende AS. 1979. Studies on the mechanisms of action of the chlorinated 

dibenzo-p-dioxins and related compounds. Ann NY Acad Sci 320:214-230. 

*Poland A, Knutson JC. 1982. 2,3,7,8-Tetrachlorodibenzo-p-dioxin and related halogenated aromatic 

hydrocarbons: Examination of the mechanism of toxicity. Ann Rev Pharmacol Toxicol 22:517-554. 

*Poland A, Knutson JC, Glover E. 1984. Histologic changes produced by 2,3,7,8-tetrachlorodibenzop-dioxin in the skin of mice carrying mutations that affect the integument. J Invest Dermatol 83:454-459. 

*Poland A, Palen D, Glover E. 1982. Tumour promotion by TCDD in skin of HRS/J hairless mice. Nature 

300:271-273. 

*Poland A, Palen D, Glover E. 1994. Analysis of the four alleles of the murine aryl hydrocarbon receptor. 

Mol Pharmacol 46: 915-921. 

*Poland AP, Smith D, Metter G, et al. 1971. A health survey of workers in a 2,4-D and 2,4,5-T plant with 

special attention to chloracne, porphyria cutanea tarda, and psychologic parameters. Arch Environ Health 

22:316-327. 

*Pollenz RS, Sattler CA, Poland A. 1994. The aryl hydrocarbon receptor and aryl hydrocarbon receptor 

nuclear translocator protein show distinct subcellular localizations in Hepa 1c1c7 cells by 

immunofluorescence microscopy. Mol Pharmacol 45: 428-438. 

*Pongratz I, Mason GGF, Poellinger L. 1992. Dual roles of the 90 kDa heat shock protein hsp90 in 

modulating functional activities of the dioxin receptor. J Biol Chem 267:13728-13734. 

*Pongratz I, Stromstedt P-E, Mason GGF, et al. 1991. Inhibition of the specific DNA binding activity of 

the dioxin receptor by phosphatase treatment. J Biol Chem 266:16813-16817. 

*Popp P, Keil P, Moeder M, et al. 1997. Application of accelerated solvent extraction followed by gas 

chromatography, high-performance liquid chromatography and gas chromatography-mass spectrometry for 

the determination of polycyclic aromatic hydrocarbons, chlorinated pesticides and polychlorinated 

dibenzo-p-dioxins and dibenzofurans in solid wastes. Journal of Chromatography A 774(1-2):203-211. 

*Porte C, Albaiges J. 1993. Bioaccumulation patterns of hydrocarbons and polychlorinated biphenyls in 

bivalves, crustaceans, and fishes. Archives of Environmental Contamination and Toxicology 26:273-281. 
CDDs 637 

8. REFERENCES 

Portier C, Tritscher A, Kohn M, et al. 1993. Ligand/receptor binding for 2,3,7,8-TCDD: Implications for 

risk assessment. Fundam Appl Toxicol 20(1):48-56. 

*Potter CL, Menahan LA, Peterson RE. 1986a. Relationship of alterations in energy metabolism to 

hypophagia in rats treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Fundam Appl Toxicol 6:89-97. 

*Potter CL, Moore RW, Inhorn SL, et al. 1986b. Thyroid status and thermogenesis in rats treated with 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 84:45-55. 

*Potter CL, Sipes IG, Russell DH. 1983. Hypothyroxinemia and hypothermia in rats in response to 

2,3,7,8-tetrachlorodibenzo-p-dioxin administration. Toxicol Appl Pharmacol 69:89-95. 

*Pratt RM, Dencker L, Diewert VM. 1984. TCDD-induced cleft palate in the mouse: Evidence for 

alterations in palatal shelf fusion. Teratol Carcinogen Mutagen 4:427-436. 

*Prokipcak RD, Okey AB. 1990. Equivalent molecular mass of cytosolic and nuclear forms of Ah receptor 

from Hepa-1 cells determined by photoaffinity labeling with 2,3,7,8-[3H]tetrachlorodibenzo-p-dioxin. 

Biochem Biophys Res Comm 172:698-704. 

*Pruell RJ, Rubinstein NI, Taplin BK, et al. 1993. Accumulation of polychlorinated organic contaminants 

from sediment by three benthic marine species. Arch Environ Contam Toxicol 24(3):290-297. 

Puhvel SM, Connor MJ, Sakamoto M. 1990. Vitamin A deficiency and dioxin-induced cutaneous toxicity 

in murine skin [Abstract]. In: Tricontinental meeting: European Society for Dermatological Research 

(ESDR), Japanese Society for Investigative Dermatology (JSID) and Society for Investigative Dermatology 

(SID), Washington, DC, May 2-5, 1990. J Invest Dermatol 94:568. 

*Puhvel SM, Sakamoto M. 1988. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on murine skin. J Invest 

Dermatol 90:354-358. 

*Puhvel SM, Sakamoto M, Ertl DC, et al. 1982. Hairless mice as models for chloracne: A study of 

cutaneous changes induced by topical application of established chloracnegens. Toxicol Appl Pharmacol 

64:492-503. 

Puntoni R, Merlo F, Fini A, et al. 1986. Soft tissue sarcoma in Seveso. Lancet 2:525. 

Puri RK, Clevenger TE, Kapila S, et al. 1989. Studies of parameters affecting translocation of 

tetrachlorodibenzo-p-dioxin in soil. Chemosphere 18:1291-1296. 

Puri RK, Kapila S, Lo YH, et al. 1990. Effect of co-contaminants on the distribution of polychlorinated 

dibenzo-p-dioxins and polychlorinated dibenzofurans in saturated soils. Chemosphere 20:1589-1596. 

*Quattrochi LC, Vu T, Tukey RH. 1994. The human CYP1A2 gene and induction by 

3-methylcholanthrene. J Biol Chem 269: 6949-6954. 

*Ramsey JC, Hefner JG, Karbowski RJ, et al. 1982. The in vivo biotransformation of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the rat. Toxicol Appl Pharmacol 65:180-184. 
CDDs 638 

8. REFERENCES 

*Randerath K, Putman KL, Randerath E, et al. 1988. Organ-specific effects of long-term feeding 

2,3,7,8-tetrachlorodibenzo-p-dioxin and 1,2,3,7,8-pentachlorodibenzo-p-dioxin on I-compounds in hepatic 

and renal DNA of female Sprague-Dawley rats. Carcinogenesis 9:2285-2289. 

*Randerath K, Putman KL, Randerath E, et al. 1989. Liver DNA alterations by 

2,3,7,8-tetrachlorodibenzo-p-dioxin and 1,2,3,7,8-pentachlorodibenzo-p-dioxin in the female rat. 

Chemosphere 18:681-685. 

*Randerath K, Putman KL, Randerath E, et al. 1990. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on 

1-compounds in hepatic DNA of Sprague-Dawley rats: Sex specific effects and structure-activity 

relationships. Toxicol Appl Pharmacol 103:271-280. 

*Rannug U, Sjogren M, Rannug A, et al. 1991. Use of artificial intelligence in structure-affinity correlations 

of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) receptor ligands. Carcinogenesis 12:2007-2015. 

*Rao MS, Subbarao V, Prasad JD, et al. 1988. Carcinogenesis of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the 

Syrian golden hamster. Carcinogenesis 9:1677-1679. 

Rappe C. 1984. Analysis of polychlorinated dioxins and furans: All 75 PCDDs and 135 PCDFs can be 

identified by isomer-specific techniques. Environ Sci Technol 18:78a-90a. 

*Rappe C. 1992. Dietary exposure and human levels of PCDDs and PCDFs. Chemosphere 25:231-34. 

*Rappe C. 1993. Sources of exposure, environmental concentrations and exposure assessment of PCDDs 

and PCDFs. Chemosphere 27(ISS 1-3):211-25. 

Rappe C, Bergovist PA, Hansson M, et al. 1984. Chemistry and analysis of polychlorinated dioxins and 

dibenzofurans in biological samples. Banbury Rep 18:17-25. 

*Rappe C, Bergqvist P-A, Kjeller L-O, et al. 1991. Levels and patterns of PCDD and PCDF contamination 

in fish, crab, and lobsters from Newark Bay and the New York Bight. Chemosphere 22(3-4):239-266. 

Rappe C, Kjeller L. 1987. PCDDs and PCDFs in environmental samples air, particulates, sediments and 

soil. Chemosphere 16:1775-1780. 

Rappe C, Kjeller LO, Andersson R. 1989a. Analyses of PCDDs and PCDFs in sludge and water samples. 

Chemosphere 19:13-20. 

*Rappe C, Kjeller LO, Bruckmann P, et al. 1988. Identification and quantification of PCDDs and PCDFs in 

urban air. Chemosphere 17:3-20. 

Rappe C, Kjeller LO, Marklund S, et al. 1986a. Electrical PCB accidents: An update. Chemosphere 

15:1291-1295. 

Rappe C, Lundstroem G, Glas B, et al. 1990. Levels of PCDDs and PCDFs in milk cartons in commercial 

milk. Chemosphere 20:1649-1656. 
CDDs 639 

8. REFERENCES 

Rappe C, Marklund S, Kjeller LO, et al. 1986b. PCDDs and PCDFs in emissions from various incinerators. 

Chemosphere 15:1213-1218. 

Rappe C, Marklund S, Kjeller LO, et al. 1989b. Long-range transport of PCDDs and PCDFs on airborne 

particles. Chemosphere 18:1283-1290. 

Rappe C, Nygren M. 1984. Chemical analysis of human samples identification and quantification of 

polychlorinated dioxins and dibenzofurans. Environ Sci Res 30:305-314. 

Rappe C, Nygren M, Buser H, et al. 1983. Identification of polychlorinated dioxins (PCDDs) and 

dibenzofurans (PCDFs) in human samples, occupational exposure and Yusho patients. Environ Sci Res 

26:241-253. 

Rappe C, Nygren M, Buser HR, et al. 1982. Occupational exposure to polychlorinated dioxins and 

dibenzofurans. Pergamon Ser Environ Sci 5:495-513. 

Rappe C, Nygren M, Lindstrom G, et al. 1986c. Dioxins and dibenzofurans in biological samples of 

European origin. Chemosphere 15:1635-1640. 

Rappe C, Nygren M, Lindstrom G, et al. 1987. Polychlorinated dibenzofurans and dibenzo-p-dioxins and 

other chlorinated contaminants in cows milk from various locations in Switzerland. Environ Sci Technol 

21:964-970. 

Rappe C, Nygren M, Marklund S, et al. 1985. Assessment of human exposure to polychlorinated 

dibenzofurans and dioxins. Environ Health Perspect 60:303-304. 

Rappe C, Swanson SE, Glas B. 1989c. Formation of PCDDs and PCDFs by the chlorination of water. 

Chemosphere 19:1875-1880. 

Rappe C, Tarkowski S, Yrjaenh E. 1989d. The WHO/EURO quality control study on PCDDs and PCDFs 

in human milk. Chemosphere 18:883-889. 

Ravo VR, Unger A. 1995. A novel application of a competitive binding model in dioxin risk assessment. 

Reg Toxicol Pharmacol 21:108-116. 

Rawls RL. 1983. Dioxin's human toxicity is most difficult problem: Dioxin is extremely toxic to certain 

animals, but scientists are only beginning to understand the nature of its toxicity and what it does to humans. 

Chemical and Engineering News 61:37-48. 

*Reed LW, Hunt GT, Maisel BE, et al. 1990. Baseline assessment of PCDDs/PCDFs in the vicinity of the 

Elk River, Minnesota generating station. Chemosphere 21:159-171. 

*Reggiani G. 1980. Acute human exposure to TCDD in Seveso, Italy. J Toxicol Environ Health 6:27-43. 

Reiner EJ, Schellenberg DH, Taguchi VY. 1991. Environmental applications for the analysis of chlorinated 

dibenzo-p-dioxins and dibenzofurans using mass-spectrometry. Environ Sci Technol 25:110-117 
CDDs 640 

8. REFERENCES 

Reiner EJ, Schellenberg DH, Taguchi VY, et al. 1990. Application of tandem quadrupole mass 

spectrometry for the ultra-trace determination of polychlorinated dibenzo-p-dioxins and dibenzofurans. 

Chemosphere 20:1385-1392. 

Reischl A, Reissinger M, Thoma H, et al. 1989. Uptake and accumulation of PCDD/F in terrestrial plants: 

Basic considerations. Chemosphere 19:467-474. 

*Reyes H, Peisz-Porszasz S, Hankinson O. 1992. Identification of the Ah receptor nuclear translocator 

(Arnt) as a component of the DNA binding form of the Ah receptor. Science 256:1193-1195. 

Rghei HO, Eiceman GA. 1984. Adsorption and chlorination of dibenzo-p-dioxin and 

1-chlorodibenzo-p-dioxin on fly ash from municipal incinerators. Chemosphere 13:421-426. 

Rghei HO, Eiceman GA. 1985. Effect of matrix on heterogeneous phase chlorine substitution reactions for 

dibenzo-p-dioxin and hydrochloric-acid in air. Chemosphere 14:167-172. 

*Rier SE, Martin DC, Bowman RE, et al. 1993. Endometriosis in Rhesus monkeys (Macaca mulatta) 

following chronic exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Fund Appl Toxicol 21:433-441. 

Riihimaki V, Sisko A, Hernberg S. 1982. Mortality of 2,4-dichlorophenoxyacetic acid and 

2,4,5-trichlorophenoxyacetic acid herbicide applicators in Finland. Scand J Work Environ Health 8:37-42. 

*Rippen G, Wesp H. 1993. Kale uptake of of PCDD/PCDF, PCB and PAH under field conditions: 

Importance of gaseous dry deposition. Presented at Dioxin 93. 13th International Symposium on Chlorinated 

Dioxins and Related Compounds, Vienna ,Austria, September 1993. 

*Riss A, Hagenmaier H, Weberruss U, et al. 1990. Comparison of PCDD/PCDF levels in soil, grass, cow's 

milk, human blood and spruce needles in an areas of PCDD/PCDF contamination through emissions from a 

metal reclamation plant. Chemosphere 21(12):1451-1456. 

*Roberts EA, Johnson KC, Harper PA, et al. 1990. Characterization of the Ah receptor mediating aryl 

hydrocarbon hydroxylase induction in the human liver cell line HepG2. Arch Biochem Biophys 

276:442-450. 

Roberts EA, Shear NH, Okey AB, et al. 1985. The Ah receptor and dioxin toxicity: From rodent to human 

tissues. Chemosphere 14:661-674. 

*Roffman A, Roffman HK. 1991. Air emissions from municipal waste combustion and their environmental 

effects. Sci Total Environ 104(1-2):87-96. 

*Rogers AM, Anderson ME, Back KC. 1982. Mutagenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin and 

perfluoro-n-decanioc acid in L5178y mouse lymphoma cells. Mutat Res 105:445-449. 

Rogers CJ, Kornel A, Peterson RL. 1987. Mobile KPEG destruction unit for PCBs, dioxins and furans in 

contaminated waste. Presented at the 13th Annual Research Symposium: Land disposal, remedial action: 

Incineration and treatment of hazardous wastes, Cincinnati, Ohio. May 6-8, 1987, 361-365. 

*Rohde S, Moser GA, PΣpke O, et al. 1997. Fecal clearance of PCDD/Fs in occupationally exposed 

persons. Organohalogen Compounds 33:408-413. 
CDDs 641 

8. REFERENCES 

*Roman BL, Peterson RE. 1998. In utero and lactational exposure of the male rat to 

2,3,7,8-tetrachlorodibenzo-p-dioxin impairs prostate development. 1. Effects on gene expression. Tox Appl 

Pharmacol 150:240-253. 

*Roman BL, Pollenz RS, Peterson RE. 1998a. Responsiveness of the adult male rat reproductive tract to 

2,3,7,8-tetrachlorodibenzo-p-dioxin exposure. Ah receptor and ARNT expression, CYP1A1 induction, and 

Ah receptor down-regulation. Tox Appl Pharmacol 150:228-239. 

*Roman BL, Sommer RJ, Shinomiya K, et al. 1995. In utero and lactational exposure of the male rat to 

2,3,7,8-tetrachlorodibenzo-p-dioxin: Impaired prostate growth and development without inhibited androgen 

production. Toxicol Appl Pharmacol 134:241-250. 

*Roman BL, Timms BG, Prins GS, et al. 1998b. In utero and lactational exposure of the male rat to 

2,3,7,8-tetrachlorodibenzo-p-dioxin impairs prostate development. 2. Effects on growth and 

cytodifferentiation. Tox Appl Pharmacol 150:254-270. 

*Romkes M, Piskorska-Pliszczynska J, Keys B, et al. 1987. Quantitative structure-activity relationships: 

analysis of interactions of 2,3,7,8-tetrachlorodibenzo-p-dioxin and 2-substituted analogues with rats, mouse, 

guinea pig, and hamster cytosolic receptor. Cancer Res 47:5108-5111. 

*Romkes M, Safe S. 1988. Comparative activities of 2,3,7,8-tetrachlorodibenzo-p-dioxin and progesterone 

as antiestrogens in the female rat uterus. Toxicol Appl Pharmacol 92:368-380. 

*Rordorf BF. 1986. Thermal properties of dioxins, furans and related compounds. Chemosphere 

15:1325-1332. 

Rordorf BF. 1987. Prediction of vapor pressures, boiling points and enthalpies of fusion for twenty-nine 

halogenated dibenzo-p-dioxins. Thermochim Acta 112:117-122. 

*Rordorf BF. 1989. Prediction of vapor pressures, boiling points and enthalpies of fusion for twenty-nine 

halogenated dibenzo-p-dioxins and fifty-five dibenzofurans by vapor pressure correlation method. 

Chemosphere 18:783-788. 

*Rose JQ, Ramsey JC, Wentzler TH, et al. 1976. The fate of 2,3,7,8-tetrachlorodibenzo-p-dioxin following 

single and repeated oral doses to the rat. Toxicol Appl Pharmacol 36:209-226. 

Rosengren R, Safe S, Bunce NJ. 1991. Structure-dependent interactions of ligands with the rat cytosolic Ah 

receptor: Kinetic analysis [Abstract 10]. Toxicologist 11:27. 

Rosengren R, Safe S, Bunce NJ. 1992. Kinetics of the association of several tritiated polychlorinated 

dibenzo-p-dioxin and dibenzofuran congeners with hepatic cytosolic Ah receptor from the Wistar rat. Chem 

Res Toxicol 5(3):376-82. 

*Rosenthal GJ, Lebetkin E, Thigpen JE, et al. 1989. Characteristics of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

induced endotoxin hypersensitivity: Association with hepatotoxicity. Toxicology 56:239-251. 

Rotard W, Christmann W, Lattner A, et al. 1987. Occurrence of PCDD and PCDF in motor oils, refined oils 

and contaminated soils. Chemosphere 16:1847-1849. 
CDDs 642 

8. REFERENCES 

*Roth W, Voorman R, Aust SD. 1988. Activity of thyroid hormone-inducible enzymes following treatment 

with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 92:65-74. 

*Rozman K. 1984. Separation of wasting syndrome and lethality caused by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Lett 22:279-285. 

Rozman K, Greim H. 1986. Toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in cold-adapted rats. Arch 

Toxicol 59:211-215. 

*Rozman K, Pfeifer B, Kerecsen L, et al. 1991. Is a serotonergic mechanism involved in 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-induced appetite suppression in the Sprague-Dawley rat? Arch 

Toxicol 65:124-128. 

Rozman K, Roth WL, Greim H, et al. 1993. Relative potency of chlorinated dibenzo-p-dioxins (CDDs) in 

acute, subchronic and chronic (carcinogenicity) toxicity studies: Implications for risk assessment of 

chemical mixtures. Toxicology 77(1-2):39-50. 

*Rozman K, Rozman T, Scheufler E, et al. 1985. Thyroid hormones modulate the toxicity of 

2,3,7,8-TCDD. J Toxicol Environ Health 16:481-491. 

RTECS. 1996. Registry of Toxic Effects of Chemical Substances. National Library of Medicine, Bethesda, 

MD. August 21. 1996. 

Ryan JJ. 1986. Variation of dioxins and furans in human tissues. Chemosphere 15:1585-1593. 

Ryan JJ. 1987. Personal telephone communication between DK Basu and JJ Ryan Food Research Division, 

Health Protection Branch, Ottawa, Canada. Syracuse Research Corporation, N.Y., September 2, 1987. 

*Ryan JJ, Beaudoin N, Mills P, et al. 1997. Dioxin-like compounds in total diet food, Canada 1992-93. 

Organohalogen Compounds 32:121-124. 

*Ryan JJ, Conacher HBS, Panopio LG, et al. 1991. Gas chromatographic separations of all 136 tetra- to 

octa- polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans on nine different stationary 

phases. Journal of Chromatography 541:131-183. 

*Ryan JJ, Gasiewicz TA, Brown Jr JR. 1990. Human body burden of polychlorinated dibenzofurans 

associated with toxicity based on the Yusho and Yucheng incidents. Fund Appl Toxicol 15:722-731. 

Ryan JJ, Lau BPY, Pilon JC, et al. 1984a. Incidence and levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

Lake Ontario and commercial fish. Environ Sci Technol 18:719-721. 

*Ryan JJ, Levesque D, Panopio LG, et al. 1993a. Elimination of polychlorinated dibenzofurans (PCDFs) 

and polychlorinated biphenyls (PCBs) from human blood in the Yusho and Yu-Cheng rice oil poisonings. 

Arch Environ Contam Toxicol 24:504-512. 

*Ryan JJ, Lizotte R, Lau BPY. 1985a. Chlorinated dibenzo-p-dioxins and chlorinated dibenzofurans in 

Canadian human adipose tissue. Chemosphere 14:697-706. 
CDDs 643 

8. REFERENCES 

*Ryan JJ, Lizotte R, Lewis D. 1987a. Human tissue levels of PCDDs and PCDFs from a fatal 

pentachlorophenol poisoning. Chemosphere 16:1989-1996. 

Ryan JJ, Lizotte R, Newsome WH. 1984b. Study of chlorinated diphenyl ethers and chlorinated 

2-phenoxyphenols as interferences in the determination of chlorinated dibenzo-p-dioxins and chlorinated 

dibenzofurans in biological samples. J Chromatogr 303:351-360. 

*Ryan JJ, Lizotte R, Panopio LG, et al. 1993b. Polychlorinated dibenzo-p-dioxins (PCDDs) and 

polychlorinated dibenzofurans (PCDFs) in human milk samples collected across Canada in 1986-87. Food 

Addit Contam 10(4):419-428. 

Ryan JJ, Lizotte R, Sakuma T, et al. 1985b. Chlorinated dibenzo-para-dioxins, chlorinated dibenzofurans 

and pentachlorophenol in Canadian chicken and pork samples. J Agric Food Chem 33:1021-1026. 

Ryan JJ, Pilon JC. 1980. High-performance liquid chromatography in the analysis of chlorinated 

dibenzodioxins and dibenzofurans in chicken liver and wood shaving samples. J Chromatogr 197:171-180. 

Ryan JJ, Pilon JC. 1982. Chlorinated-dibenzodioxins and dibenzofurans in chicken litter and livers arising 

from pentachlorophenol contamination of wood shavings. In: Hutzinger O., ed. Chlorinated dioxins and 

related compounds: Impact on the environment. New York, NY: Pergamon Press, 183-189. 

Ryan JJ, Schecter A, Lizotte R, et al. 1985c. Tissue distributions of dioxins and furans in humans from the 

general population. Chemosphere 14:929-932. 

Ryan JJ, Schecter A, Masuda Y, et al. 1987b. Comparison of PCDDs and PCDFs in the tissues of Yusho 

patients with those from the general population in Japan and China. Chemosphere 16:2017-2026. 

*Ryan JJ, Schecter A, Sun WF, et al. 1986. Distribution of chlorinated dibenzo-p-dioxins and chlorinated 

dibenzofurans in human tissues from the general population. In: Rappe C, Choudhary G, Keith LH, eds. 

Chlorinated dioxins and dibenzofurans in perspective. Chelsea, MI: Lewis Publishers, Inc., 3-16. 

*Ryan JJ, Shewchuk C, Lau P-Y, et al. 1992. Polychlorinated dibenzo-p-dioxins and polychlorinated 

dibenzofurans in Canadian bleached paperboard milk containers (1988-1989) and their transfer to fluid milk. 

J Agric Fd Chem 40:919-923. 

Ryan JJ, Williams DT. 1983. Analysis of human fat tissue from the Great Lakes area for 

2,3,7,8-tetrachlorodibenzo-p-dioxin and furan residues. Environ Chem 23:157-158. 

Sacchi GA, Vigano P, Fortunati G, et al. 1986. Accumulation of 2,3,7,8-tetrachlorodibenzo-p-dioxin from 

soil and nutrient solution by bean and maize plants. Experientia 42(5):586-588. 

*Safe S. 1986. Comparative toxicology and mechanism of action of polychlorinated dibenzo-p-dioxins and 

dibenzofurans. Annu Rev Pharmacol Toxicol 26:371-399. 

*Safe S. 1987. Determination of 2,3,7,8-TCDD toxic equivalent factors (TEFs): Support for the use of the 

in vitro induction assay. Chemosphere 16:791-802. 
CDDs 644 

8. REFERENCES 

*Safe S. 1990. Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), 

and related compounds: Enviornmental and mechanistic considerations which support the development of 

toxic equivalency factors. Toxicology 21:51-88. 

*Safe S. 1992. Development, validation and limitations of toxic equivalency factors. Chemosphere 

25:61-64. 

*Safe S. 1998a. Limitations of the toxic equivalency factor approach for risk assessment of TCDD and 

related compounds. Teratogenesis Carcinog Mutagen 17:285-304. 

*Safe S, Astroff B, Harris M, et al. 1991. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and related 

compounds as antiestrogens: Characterization and mechanism of action. Pharmacol Toxicol 69:400-409. 

Safe S, Brown KW, Donnelly KC, et al. 1992. Polychlorinated dibenzo-p-dioxins and dibenzofurans 

associated with wood-preserving chemical sites: Biomontoring with pine needles. Environ Sci Technol 

26:394-396. 

Safe S, Mason G, Keys B, et al. 1986. Polychlorinated dibenzo-p-dioxins and dibenzofurans: Correlation 

between in vitro and in vivo structure-activity relationships (SARs). Chemosphere 15:1725-1731. 

*Safe SH. 1995. Modulation of gene expression and endocrine response pathways by 

2,3,7,8-tetrachlorodibenzo-p-dioxin and related compounds. Pharmacol Ther 67:247-281. 

*Safe SH. 1998b. Development validation and problems with the toxic equivalency factor approach for risk 

assessment of dioxins and related compounds. J Anim Sci 76:134-141. 

*Sanderson JT, Elliott JE, Norstrom RJ, et al. 1994. Monitoring biological effects of polychlorinated 

dibenzo-p-dioxins, dibenzofurans, and bipehnyls in Great Blue Heron chicks in British Columbia. J Toxicol 

Environ Health 41(4):434-450. 

*Santostefano MJ, Johnson KL, Whisnant NA, et al. 1996. Subcellular localization of TCDD differs 

between the liver, lungs, and kidneys after acute and subchronic exposure: Species/dose comparisons and 

possible mechanism. Fundam Appl Toxicol 34:265-275. 

*Santostefano MJ, Wang X, Richardson VM, et al. 1998. A pharmacodynamic analysis of TCDD-induced 

cytochrome P450 gene expression in multiple tissues: dose- and time-dependent effects. Tox Appl 

Pharmacol 151:294-310. 

*Saracci R, Kogevinas M, Bertazzi PA, et al. 1991. Cancer mortality in workers exposed to chlorphenoxy 

herbicides and chlorophenols. Lancet 338:1027-1032. 

Sarna LP, Hodge PE, Webster GRB. 1984. Octanol-water partition coefficients of chlorinated dioxins and 

dibenzofurans by reversed-phase HPLC using several C18 columns. Chemosphere 13:975-983. 

*Sawahata T, Olson JR, Neal RA. 1982. Identification of metabolites of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) formed on incubation with isolated rat hepatocytes. Biochem 

Biphys Res Commun 105:341-346. 
CDDs 645 

8. REFERENCES 

*Sax NI, Lewis RJ, eds. 1987. Hawley's Condensed Chemical Dictionary. New York, NY: Van Nostrand 

Reinhold, 11:886. 

Schafer W, Ballschmiter K. 1986. Monobromo-polychloro-derivatives of benzene, biphenyl, dibenzofuran, 

and dibenzodioxin formed in chemical-waste burning. Chemosphere 15:755-763. 

*Schantz S, Bowman RE. 1989. Learning in monkeys exposed perinatally to 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD). Neurotoxicol Teratol 11:13-19. 

Schantz SL, Barsotti DA, Allen JR. 1979. Toxicological effects produced in nonhuman primates 

chronically exposed to fifty parts-per-trillion 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicol Appl 

Pharmacol 48:A180. 

*Schantz SL, Ferguson SA, Bowman RE. 1992. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on behavior 

of monkey in peer groups. Neurotoxicol Teratol 14:433-446. 

*Schantz SL, Laughlin NK, Van Valkenberg HC, et al. 1986. Maternal care by rhesus monkeys of infant 

monkey exposed to either lead or 2,3,7,8-tetrachlorodibenzo-p-dioxin. Neurotoxicol 2:637-650. 

*Schantz SL, Mably TA, Peterson RE. 1991. Effects of perinatal exposure to 2,3,7,8-tetrachlorodibenzop-dioxin (2,3,7,8-TCDD) on spatial learning and memory and locomotor activity in rats. Teratol 43:497. 

*Schaum J, Cleverly D, Lorber M, et al. 1994. Updated analysis of U.S. sources of dioxin-like compounds 

and background exposure levels. Organohalogen Compounds 20:178-184. 

Schecter A. 1983. Contamination of an office building in Binghamton, New York, by PCBs, dioxins, 

furans, and biphenylenes after an electrical panel and electrical transformer incident. Chemosphere 

12:669-680. 

*Schecter A. 1985. Medical surveillance of exposed persons after exposure to PCBs, chlorinated 

dibenzodioxins and dibenzofurans after PCB transformer or capacitor incidents. Environ Health Perspect 

60:333-338. 

*Schecter A. 1994. Elevated dioxin blood levels in British chemical workers. J Occup Med 

36(12):1283-1287. 

*Schecter A, Constable JD, Arghestani S, et al. 1987a. Elevated levels of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in adipose tissue of certain U.S. veterans of the Vietnam war. 

Chemosphere 16:1997-2001. 

*Schecter A, Constable JD, Bangert JV, et al. 1989a. Elevated body burdens of 2,3,7,8-tetrachlorodibenzodioxin in adipose tissue of United States Vietnam veterans. Chemosphere 18:431-438. 

Schecter A, Constable JD, Bangert JV, et al. 1989b. Isomer specific measurement of polychlorinated 

dibenzodioxin and dibenzofuran isomers in human blood from American Vietnam veterans two decades after 

exposure to Agent Orange. Chemosphere 18:531-538. 

*Schecter A, Cramer P, Boggess K, et al. 1996c. Dioxin intake from U.S. Food: Results from a new 

nationwide food survey. In: Organohalogen Compounds volume 28. 
CDDs 646 

8. REFERENCES 

*Schecter A, Dai LC, Thuy LTB, et al. 1995. Agent Orange and the Vietnamese: The persistence of 

elevated dioxin levels in human tissues. Am J Public Health 85:516-522. 

Schecter A, Dekin A, Weerasinghe NCA, et al. 1988. Sources of dioxins in the environment: A study of 

PCDDs and PCDFs in ancient, frozen Eskimo tissue. Chemosphere 17:627-631. 

*Schecter A, Fürst P, Fürst C, et al. 1989c. Levels of polychlorinated dibenzodioxins and dibenzofurans in 

cow's milk and in soybean derived infant formulas sold in the United Sates and other countries. 

Chemosphere 19:913-918. 

*Schecter A, Fürst P, Fürst C, et al. 1991a. Dioxins, dibenzofurans and selected chlorinated organic 

compounds in human milk and blood from Cambodia, Germany, Thailand, the USA, the USSR, and 

Vietnam. Chemosphere 23:1903-1912. 

*Schecter A, Fürst P, Fürst C, et al. 1994a. Chlorinated dioxins and dibenzofurans in human tissue from 

general populations: A selective review. Environ Health Perspectives Supplements 102(1):159-171. 

*Schecter A, Fürst P, Krüger C, et al. 1989d. Levels of polychlorinated dibenzofurans, dibenzodioxins, 

PCBs, DDT, and DDE, hexachlorobenzene, dieldrin, hexachlorocyclohexanes, and oxychlordane in human 

breast milk from the United States, Thailand, Vietnam, and Germany. Chemosphere 18:445-454. 

*Schecter A, Fürst P, Ryan JJ, et al. 1989e. Polychlorinated dioxins and dibenzofurans levels from human 

milk from several locations in the United States, Germany, and Vietnam. Chemosphere 19:979-984. 

*Schecter A, Gasiewicz TA. 1987a. Health hazard assessment of chlorinated dioxins and dibenzofurans 

contained in human milk. Chemosphere 16:2147-2154. 

Schecter A, Gasiewicz TA. 1987b. Human breast milk levels of dioxins and dibenzofurans: Significance 

with respect to current risk assessments. ACS Symp Ser 338:162-173. 

Schecter A, Kooke R, Olie K, et al. 1989f. Chlorinated dioxin and dibenzofuran levels in food samples 

collected between 1985-87 in north and south of Vietnam. Chemosphere 18:627-634. 

*Schecter A, Li L. 1997. Dioxins, dibenzofurans, dioxin-like PCBs, and DDE in U.S. fast food, 1995. 

Chemosphere 34(5-7):1449-1457. 

Schecter A, Malkin R, PΣpke O, et al. 1991b. Dioxin levels in blood of municipal incinerator workers. Med 

Sci Res 19:331-332. 

*Schecter A, McGee H, Stanley JS, et al. 1996b. Dioxins and dioxin like chemicals in blood and semen of 

American Vietnam veterans from the state of Michigan. Amer J Ind Med 30:647-654. 

*Schecter A, PΣpke O, Ball M. 1990a. Evidence for transplacental transfer of dioxins from mother to fetus: 

Chlorinated dioxinal dibenzofuran levels in the livers of stillborn infants. Chemosphere 21:1017-1022. 

*Schecter A, Papke O, Dellarco M, et al. 1996d. A comparison of dioxins and dibenzofurans in cooked and 

uncooked food. Organohalogen compounds, volume 28. 
CDDs 647 

8. REFERENCES 

*Schecter A, Papke O, Lis A, et al. 1996a. Decrease in milk and blood dioxin levels over two years in a 

mother nursing twins: Estimates of decreased maternal and increased infant dioxin body burden from 

nursing. Chemosphere 32:543-549. 

*Schecter A, Ryan J, Papke O. 1994b. Elevated dioxin blood levels in Russian chemical workers and their 

children following maternal exposure. Chemosphere 29 9-11):2361-2370. 

Schecter A, Ryan JJ. 1988. Polychlorinated dibenzo-p-dioxin and dibenzofuran levels in human adipose 

tissues from workers 32 years after occupational exposure to 2,3,7,8-TCDD. Chemosphere 17:915-920. 

Schecter A, Ryan JJ. 1989. Blood and adipose tissue levels of PCDDs/PCDFs over three years in a patient 

after exposure to polychlorinated dioxins and dibenzofurans. Chemosphere 18:635-642. 

*Schecter A, Ryan JJ. 1991. Brominated and chlorinated dioxin blood levels in a chemist 34 years after 

exposure to 2,3,7,8-tetrachlorodibenzodioxin and 2,3,7,8-tetrabromodibenzodioxin. Chemosphere 

23:1921-1924. 

*Schecter A, Ryan JJ, Constable JD. 1986a. Chlorinated dibenzo-p-dioxin and dibenzofuran levels in 

human adipose tissue and milk samples from the north and south of Vietnam. Chemosphere 15:1613-1620. 

*Schecter A, Ryan JJ, Constable JD. 1987b. Polychlorinated dibenzo-p-dioxin and polychlorinated 

dibenzofuran levels in human breast milk from Vietnam compared with cow's milk and human breast milk 

from the North American continent. Chemosphere 16:2003-2016. 

Schecter A, Ryan JJ, Constable JD. 1989g. Chlorinated dioxins and dibenzofurans in human milk from 

Japan, India, and the United States of America. Chemosphere 18:975-980. 

*Schecter A, Ryan JJ, Constable JD, et al. 1990b. Partioning of 2,3,7,8-chloroinated dibenzo-p-dioxins and 

dibenzofurans between adipose tissue and plasma lipid of 20 Massachusetts Vietnam veterans. 

Chemosphere 20:951-958. 

*Schecter A, Ryan JJ, Gitlitz G. 1986b. Chlorinated dioxin and dibenzofurans levels in human adipose 

tissues from exposed and control populations. In: Rappe C, Choudhary G, Keith LH, eds. Chlorinated 

dioxins and dibenzofurans in perspective. Chelsea, Michigan: Lewis Publishers, Inc., 51-56. 

*Schecter A, Ryan JJ, Lizotte R, et al. 1985a. Chlorinated dibenzodioxins and dibenzofurans in human 

adipose tissue from exposed and control New York State patients. Chemosphere 14:933-937. 

*Schecter A, Ryan JJ, Masuda Y, et al. 1994c. Chlorinated and brominated dioxins and dibenzofurans in 

human tissue following exposure. Environ Health Perspect 102(S1):134-147. 

*Schecter A, Ryan JJ, PΣpke O, et al. 1993. Elevated dioxin levels in the blood of male and female Russian 

workers with and without chloracne 25 years after phenoxyherbicide exposure: the UFA "Khimprom" 

incident. Chemosphere 27:253-258. 

*Schecter A, Startin J, Wright C, et al. 1994. Dioxins in U.S. food and estimated daily intake. 

Chemosphere 29(9-11):2261-2265. 
CDDs 648 

8. REFERENCES 

*Schecter A, Startin J, Wright C, et al. 1994d. Congener-specific levels of dioxins and dibenzofurans in 

U.S. food and estimated daily dioxin toxic equivalent intake. Environ Health Perspectives 102(11):962-966 

*Schecter A, Startin J, Wright C, et al. 1996e. Concentrations of polychlorinated dibenzo-p-dixions and 

dibenzofurans in human placental and fetal tissues from the U.S. and in placentas from Yu-Cheng exposed 

mothers. Chemosphere. 32:551-557. 

*Schecter A, Tiernan T. 1985. Occupational exposure to polychlorinated dioxins, polychlorinated furans, 

polychlorinated biphenyls, and biphenylenes after an electrical panel and transformer accident in an office 

building in Binghamton, NY. Environ Health Perspect 60:305-313. 

*Schecter A, Tiernan T, Schaffner F. 1985b. Patient fat biopsies for chemical analysis and liver biopsies for 

ultrastructural characterization after exposure to polychlorinated dioxins, furans and PCBs. Environ Health 

Perspect 60:241-254. 

Scheidl K, Kuna RP, Wurst F. 1985. Chlorinated dioxins and furans in emissions from municipal 

incineration. Chemosphere 14:913-917. 

*Schlaud M, Seidler A, Salje A, et al. 1995. Organochlorine residues in human breast milk: analysis 

through a sentinel practice network. Journal of Epidemiology & Community Health 49(suppl. 1):17-21. 

Schmid P, Suter-Hofmann ME, Schlatter C. 1989. A simplified clean-up procedure for the rapid 

determination of PCDDs and PCDFs based on retention on activated carbon AX-21. Chemosphere 

18:1741-1748. 

*Schrenk D, Buchmann A, Dietz K, et al. 1994. Promotion of preneoplastic foci in rat liver with 

2,3,7,8-tetrachlorodibenzo-p-dioxin, 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin and a defined mixture of 49 

polychlorinated dibenzo-p-dioxins. Carcinogenesis 15(3):509-515 

*Schroll R, Bierling B, Cao G, et al. 1994. Uptake pathways of organic chemicals from soil by agricultural 

plants. Chemosphere 28(2):297-303. 

*Schroll R, Scheunert I. 1993. Uptake pathways of octachlorodibenzo-p-dioxin from soil by carrots. 

Chemosphere 26(9):1631-1640. 

*Schroy JM, Hileman FD, Cheng SC. 1985. Physical/chemical properties of 2,3,7,8-TCDD 

(2,3,7,8-tetrachlorodibenzo-p-dioxin). Chemosphere 14:877-880. 

*Schulz-Schlage T, Golor G, Stahlmann R. 1992. Biological effect of high doses of TCDD in the marmoset 

monkey, Callithrix jacchus, Plateau of ethoxyresorufin O-deethylase activity in liver microsomes. 

Chemosphere 25(7-10):991-995. 

*Schuman RF, Hunter KW. 1988. Ultra sensitive bioassay for dioxin. Electric Power Research Institute 

Report, 7-11 - 7-14. 

*Schwetz BA, Norris JM, Sparschu GL, et al. 1973. Toxicology of chlorinated dibenzo-p-dioxins. Environ 

Health Perspect 5:87-99. 
CDDs 649 

8. REFERENCES 

SDGWQS. 1989. South Dakota Ground-Water Quality Standards, 2/89. South Dakota Chapter 74:03:15. 

*Secor CL, Mills EL, Harshbarger, et al. 1993. Bioaccumulation of toxicants, element and nutrient 

composition, and soft tissue histology of Zebra mussels (dreissena polymorpha) from New York state waters. 

Chemosphere 26(8):1559-1575. 

*Sedman RM, Esparza JR. 1991. Evaluation of the public health risks associated with semivolatile metal 

and dioxin emissions from hazardous waste incinerators. Environmental Health Perspectives 94:181-187. 

*Seefeld MS, Corbett SW, Keesey RE, et al. 1984a. Characterization of the wasting syndrome in rats 

treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 73:311-322. 

*Seefeld MS, Keesey RE, Peterson RE. 1984b. Body weight regulation in rats treated with 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 76:526-536. 

*Seefeld MS, Peterson RE. 1984. Digestible energy and efficiency of feed utilization in rats treated with 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 74:214-222. 

*Seiler JP. 1973. A survey on the mutagenicity of various pesticides. Experientia 29:622-623. 

*Seo B-W, Li M-H, Hansen LG et al. 1995. Effects of gestational and lactational exposure to coplanar 

polychlorinated biphenyl (PCB) congeners or 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on thyroid 

hormone concentrations in weanling rats. Toxicol Lett 78:253-262. 

*Servos M, Muir D, Whittle D, et al. 1989b. Bioavailability of octachlorodibenzo-p-dioxin in aquatic 

systems. Chemosphere 19(1-6):969-972. 

*Servos MR, Muir D CG, Webster G RB. 1992. Environmental fate of polychlorinated dibenzo-p-dioxins 

in lake enclosures. Can J Fish Aquat Sci 49(4):722-734. 

Servos MR, Muir DCG, Webster GRB. 1989a. The effect of dissolved organic matter on the bioavailability 

of polychlorinated dibenzo-p-dioxins. Aquat Toxicol 14:169-184. 

Seta J. 1991. Written communication (March 26) to Susan Press, ICF Incorporated, regarding 1990 NOES 

data for 2,3,7,8-TCDD. Cincinnati, OH: Department of Health and Human Services. 

*Setchell BP, Waites GMH. 1975. The blood testis barrier. In: Creep RO, Astwood EB, eds., Geiger SR, 

executive ed. Handbook of physiology: Endocrinology V. (Chapter 6). American Physiological Society, 

Washington DC. 

*Sewall CH, Flager N, Van den Heuvel JP, et al. 1995. Alterations in thyroid function in female 

Sprague-Dawley rats following chronic treatment with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl 

Pharmacol 132:237-244. 

*Sewall CH, Lucier GW, Tritscher AM, et al. 1993. TCDD-mediated changes in hepatic epidermal growth 

factor receptor may be a critical event in the hepatocarcinogenic action of TCDD. Carcinogenisis 

14(9):1885-1893. 
CDDs 650 

8. REFERENCES 

*Shane BS, Metosh-Dickey CA, Ebel JG Jr., et al. 1993. Toxicological evaluation of milk from lactating 

dairy cows following consumption of newsprint. J Agric Food Chem 41(2):240-245. 

Shara MA, Stohs SJ. 1987. Biochemical and toxicological effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) congeners in female rats. Arch Environ Contam Toxicol 16:599-606. 

*Shear N, Schmidt C, Huntley S, et al. 1996. Evaluation of the factors relating combined sewer overflows 

with sediment contamination of the Lower Passaic River. Mar Poll Bull 32(3):288-304. 

Sheffield A. 1985. Sources and releases of PCDDs and PCDFs to the Canadian environment. Chemosphere 

14:811. 

*Shen ES, Guengerich FP, Olson JR. 1989. Biphasic response for hepatic microsomal enzyme induction by 

2,3,7,8-tetrachlorodibenzo-p-dioxin in C57BL/6J and DBA/2J mice. Biochem Pharmacol 38:4075-4084. 

*Shen ES, Gutman SI, Olson JR. 1991. Comparison of 2,3,7,8-tetrachlorodibenzo-p-dioxin-mediated 

heptotoxicity in C57BL/6J and DBA/2J mice. J Toxicol Environ Health 32:367-381. 

*Shen ES, Whitlock JP Jr. 1992. The potential role of DNA methylation in the response to 

2,3,7,8-tetrachlorodibenzo-p-dioxin. J Biol Chem 264:17754-17758. 

Shepard BM, Young AL. 1981. Dioxins as contaminants of herbicides: The U.S. perspective. In: Tucker 

RE, Young AL, Gray AP, eds. Human and environmental risks of chlorinated dioxins and related 

compounds. New York, NY: Plenum Press, 3-11. 

*Sherman RK, Clement RE, Tashiro C. 1990. The distribution of polychlorinated dibenzo-p-dioxins and 

dibenzofurans in Jackfish Bay, Lake Superior, in relation to a Kraft pulp mill effluent. Chemosphere 

20(10-12):1641-1648. 

*Sherman WR, Keenan RE, Gunster DG, et al. 1992. Reevaluation of dioxin bioconcentration and 

bioaccumulation factors for regulatory purposes. J Toxicol Environ Health 37:211-229. 

Sherry JP, Apsimon JW, Collier TL, et al. 1990. Use of DMSO as solubilization agent in the detection of 

2,3,7,8-TCDD by radioimmunoassay. Chemosphere 20:1409-1416. 

Sherry JP, Tse H. 1990. A procedure for the determination of polychlorinated dibenzo-p-dioxins in fish. 

Chemosphere 20:865-872. 

*Shiu WY, Doucette W, Gobas FAPC, et al. 1988. Physical-chemical properties of chlorinated 

dibenzo-p-dioxins. Environ Sci Technol 22:651-658. 

Shiverick KT, Muther TF. 1983. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) effects on hepatic 

microsomal steroid metabolism and serum estradiol of pregnant rats. Biochem Pharmacol 32:991-995. 

Shu H, Teitelbaum P, Webb AS, et al. 1988. Bioavailability of soil-bound TCDD: Dermal bioavailability 

in the rat. Fundam Appl Toxicol 10:335-343. 
CDDs 651 

8. REFERENCES 

Shu HP, Paustenbach DJ, Murray FJ. 1987. A critical evaluation of the use of mutagenesis, carcinogenesis, 

and tumor promotion data in a cancer risk assessment of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Regul Toxicol 

Pharmacol 7:57-88. 

Sijm DTHM, Weaver H, Devries PJ, et al. 1989. Octanol/water partition coefficients of polychlorinated 

dibenzo-p-dioxins and dibenzofurans: Experimental values determined with a stirring method. 

Chemosphere 19:263-266. 

*Silbergeld EK, Gasiewicsz TA. 1989. Dioxins and the Ah receptor. Am J Ind Med 16:455-474. 

*Silkworth J, McMartin D, DeCaprio A, et al. 1982. Acute toxicity in guinea pigs and rabbits of soot from a 

polychlorinated biphenyl-containing transformer fire. Tox Appl Pharmacol 65:425-439. 

*Silkworth JB, Cutler DS, Antrim L, et al. 1989b. Teratology of 2,3,7,8-tetrachlorodibenzo-p-dioxin in a 

complex environmental mixture from the Love Canal. Fundam Appl Toxicol 13:1-15. 

*Silkworth JB, Cutler DS, Sack G. 1989a. Immunotoxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in a 

complex environmental mixture from the Love Canal. Fundam Appl Toxicol 12:303-312. 

*Silverstone AE, Frazier, DE, Fiore NC, et al. 1994. Dexamethasone, -estradiol, and 

2,3,7,8-tetrachlorodibenzo-p-dioxin elicit thymic atrophy through different cellular targets. Toxicol Appl 

Pharmacol 126: 248-259. 

*Simon N, Guzzetta J, Thielen D. 1989. GC/MS and GC/MS/MS: Complimentary techniques for the 

quantitation of tetrachlorodibenzo-p-dioxin in soils and sediments from industrial sites. Chemosphere 

18:163-168. 

*Sine C, ed. 1990. Farm chemicals handbook. Willoughby, OH: Meister Publishing Co., C87, C261. 

*Singer R, Moses M, Valciukas J, et al. 1982. Nerve conduction velocity studies of workers employed in 

the manufacture of phenoxy herbicides. Environ Res 29:297-311. 

*Singh J, Miles W, Barrette JP. 1985. Determination of polychlorinated dibenzo-p-dioxins in technical 

pentachlorophenol. J Assoc Off Anal Chem 68:583-586. 

Slaga TJ, Nesnow S. 1985. Sencar mouse skin tumorigenesis. In: Millmen HA, Weisburger EK, eds. 

Handbook of carcinogen testing. Park Ridge, N.J.: Noyes Publ, 230-250. 

*Slob W, Van Jaarsveld JA. 1993. A chain model for dioxins: From emission to cow's milk. Chemosphere 

27(ISS 1-3):509-16. 

*Sloop TC, Lucier GW. 1987. Dose-dependent elevation of Ah receptor binding by TCDD rat liver. 

Toxicol Appl Pharmacol 88:329-337. 

Smith AH. 1987. Infant exposure assessment for breast milk dioxins and furans derived from waste 

incineration emissions. Risk Anal 7:347-353. 

*Smith AH, Fisher DO, Dip NP, et al. 1982. Congenital defects and miscarriages among New Zealand 

2,4,5-T sprayers. Arch Environ Health 37:197-200. 
CDDs 652 

8. REFERENCES 

*Smith AH, Francis JE, Kay SJ, et al. 1981. Hepatic toxicity and uroporphyringen decarboxylase activity 

following a single dose of 2,3,7,8-tetrachlorodibenzo-p-dioxin to mice. Biochem Pharmacol 30:2825-2830. 

*Smith AH, Pearce NE, Fisher DO, et al. 1984a. Soft tissue sarcoma and exposure to phenoxyherbicides 

and chlorophenols in New Zealand. J Natl Cancer Inst 73:1111-1117. 

*Smith FA, Schwetz BA, Nitschke KD. 1976. Teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 

CF-1 mice. Toxicol Appl Pharmacol 38:517-523. 

Smith LM, Schwartz TR, Feltz K, et al. 1990a. Determination and occurrence of AHH-active 

polychlorinated-biphenyls, 2,3,7,8-tetrachloro-para-dioxin and 2,3,7,8-tetrachlorodibenzofuran in Lake 

Michigan sediment and biota - the question of their relative toxicological significance. Chemosphere 

21:1063-1085. 

Smith LM, Stalling DL, Johnson JL. 1984b. Determination of part-per-trillion levels of polychlorinated 

dibenzofurans and dioxins in environmental samples. Anal Chem 56:1830-1842. 

*Smith RM, O'Keefe P, Aldous K, et al. 1992. Measurement of PCDFs and PCDDs in air samples and lake 

sediments at several locations in upstate New York. Chemosphere 25(1-2):95-8. 

*Smith RM, O'Keefe PW, Aldous KM, et al. 1983. 2,3,7,8-Tetrachlorodibenzo-p-dioxin in sediment 

samples from Love Canal storm sewers and creeks. Environ Sci Technol 17:6-10. 

*Smith RM, O'Keefe PW, Aldous KM, et al. 1990b. Chlorinated dibenzofurans and dioxins in atmospheric 

samples from cities in New York. Environ Sci Technol 24:1502-1506. 

*Smith RM, O'Keefe PW, Briggs R, et al. 1990c. Continuous ambient air monitoring for CDFs and CDDs: 

Niagara Falls New York. Organohalogen Compd 4:233-236. 

*Smith RM, O'Keefe PW, Hilker DR, et al. 1986a. Determination of picogram per cubic meter 

concentrations of tetra- and pentachlorinated dibenzofurans and dibenzo-p-dioxins in indoor air by 

high-resolution gas chromatography/high-resolution mass spectrometry. Anal Chem 58:2414-2420. 

Smith RM, O'Keefe PW, Hilker DR, et al. 1986b. Sampling, analytical method and results for chlorinated 

dibenzo-p-dioxins and chlorinated dibenzofurans from incinerator stack effluent and contaminated building 

indoor air samples. In: Rappe C, Choudhary G, Keith LH, eds. Chlorinated dioxins and dibenzofurans in 

perspective. Chelsea, MI: Lewis Publishers, Inc., 93-108. 

Smith RM, O'Keefe PW, Hilker DR, et al. 1989. Ambient air and incinerator testing for chlorinated 

dibenzofurans and dioxins by low resolution mass spectrometry. Chemosphere 18:585-592. 

*Sommer RJ, Ippolito DL, Peterson RE. 1996. In utero and lactational exposure of the Male Holtzman rat 

to 2,3,7,8-tetrachlorodibenzo-p-dioxin: decreased epididymal and ejaculated sperm numbers without 

alterations in sperm transit rate. Toxicol Appl Pharmacol 140:146-153. 

Sovocool GW, Mitchum RK, Tondeur Y, et al. 1988. Bromo- and bromochloro-polynuclear aromatic 

hydrocarbons, dioxins and dibenzofurans in municipal incinerator fly ash. Biomed Environ Mass Spectrum 

15:669-676. 
CDDs 653 

8. REFERENCES 

Sparschu GL Jr, Dunn FL Jr, Lisowe RW, et al. 1971a. Effects of high levels of 

2,4,5-trichlorophenoxyacetic acid on fetal development in the rat. Food Cosmet Toxicol 9:527-530. 

Sparschu GL Jr, Dunn FL Jr, Row VK Jr. 1971b. Study of the teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Food Cosmet Toxicol 9:405-412. 

*SRI. 1991. 1991 Directory of Chemical Producers: United States of America. Menlo Park, CA: Stanford 

Research Institute International, 823, 1033. 

Srinivasan KR, Fogler HS. 1986. Binding of OCDD, 2,3,7,8-TCDD and HCB to clay-based sorbents. In: 

Rappe C, Choudhary G, Keith LH, eds. Chlorinated dioxins and dibenzofurans in perspective. Chelsea, MI: 

Lewis Publishers, Inc., 531-539. 

*Stahl BU, Kettrup A, Rozman K. 1992. Comparative toxicity of four chlorinated dibenzo-p-dioxins 

(CDDs) and their mixture. Part I: Acute toxicity and toxic equivalency factors (TEFs). Arch Toxicol 

66(7):471-7. 

*Stahl BU, Rozman K. 1990. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)-induced appetite suppression in 

the Sprague-Dawley rat is not a direct effect on feed intake regulation in the brain. Toxicol Appl Pharmacol 

106:158-162. 

*Stahlmann R, Schulz-Schalge T, Korte M, et al. 1992. Clinical chemistry and hematology in rats after 

short-term and after long-term exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Chemosphere 

25(7-10):1207-1214. 

Stalling DL, Norstrom RJ, Smith LM, et al. 1985. Patterns of PCDD, PCDF, and PCB contamination in 

Great Lakes, fish and birds and their characterization by principal components analysis. Chemosphere 

14:627-643. 

*Stalling DL, Peterman PH, Smith LM, et al. 1986. Use of pattern recognition in the evaluation of PCDD 

and PCDF residue data from GC/MS analysis. Chemosphere 15:1435-1443. 

Stalling DL, Smith LM, Petty JD, et al. 1983. Residues of polychlorinated dibenzo-p-dioxins and 

dibenzofurans in Laurentar Great Lakes fish. In: Tucker E, Young AL, Gray AP, eds. Human and 

environmental risks of chlorinated dioxins and related compounds. New York, London: Plenum Publishing 

Corp, 221-240. 

*Stanker LH, Watkins B, Rogers N, et al. 1987. Monoclonal antibodies for dioxin: Antibody 

characterization and assay development. Toxicology 45:229-243. 

*Stanley JS. 1986. Broad scan analysis of the FY82 National human adipose tissue survey specimens: 

Volume 1. Executive summary. Washington, D.C.: U.S. Environmental Protection Agency. EPA 

560/5-86-035. 

*Stanley JS, Boggess KE, Onstot J, et al. 1986. PCDDs and PCDFs in human adipose tissue from the EPA 

FY82 NHATS repository. Chemosphere 15:1605-1612. 

*Staples CA, Werner A, Hoogheem T. 1985. Assessment of priority pollutant concentrations in the United 

States using STORET database. Environ Toxicol Chem 4:131-142. 
CDDs 654 

8. REFERENCES 

Startin JR. 1994. Dioxins in food. In: Schecter A., ed, Dioxins and health. New York, Plenum Press, 

115-137. 

Startin JR, Rose M, Offen C. 1989. Analysis of PCDDs and PCDFs in human milk from the U.K. 

Chemosphere 19:985-988. 

*Stehl RH, Papenfuss RR, Bredeweg RA, et al. 1973. Stability of pentachlorophenol and chlorinated 

dioxins to sunlight, heat, and combustion. Advan Chem Ser 120:119-125. 

*Stehr PA, Stein G, Falk H, et al. 1986. A pilot epidemiologic study of possible health effects associated 

with 2,3,7,8-tetrachlorodibenzo-p-dioxin contaminations in Missouri. Arch Environ Health 41:16-22. 

*Stehr-Green PA, Naylor PH, Hoffman RE. 1989. Diminished thymosinalpha-1 levels in persons exposed 

to 2,3,7,8-tetrachlorodibenzo-p-dioxin. J Toxicol Environ Health 28:285-295. 

Steinwandter H. 1994. A simple and fast micromethod for the determination of 

1,2,3,4,6,7,8-Heptachlorodibenzo-p-dioxin and octachlorodibenzo-p-dioxin in sewage sludges by GC-ECD. 

Fresenius Journal of Analytical Chemistry 348(5-6):385-389. 

Steppan LB, Kirkvliet NI. 1991. Influence of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the 

production of inflammatory cytokine mRNA by C57Bl/6 macrophages [Abstract 45]. Toxicologist 11:35. 

Stevens KM. 1981. Agent Orange toxicity: A quantitative perspective. Hum Toxicol 1:31-39. 

Stieglitz L, Zwick G, Roth W. 1986. Investigation of different treatment techniques for PCDD/PCDF in fly 

ash. Chemosphere 15:1135-1140. 

Stockbauer JW, Hoffman RE, Schramm WF, et al. 1988. Reproductive outcomes of mothers with potential 

exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Am J Epidemiol 128:410-419. 

*Stohs SJ, Shara MA, Alsharif NZ, et al. 1990. 2,3,7,8-Tetrachlorodibenzo-p-dioxin-induced oxidative 

stress in female rats. Toxicol Appl Pharmacol 106:126-135. 

*Sugawara Y, Gee SJ, Sanborn JR, et al. 1998. Development of a highly sensitive enzyme-linked 

immunosorbent assay based on polyclonal antibodies for the detection of polychlorinated dibenzo-p-dioxins. 

Anal Chem 70(6):1092-9. 

*Sullivan MJ, Stanford JT. 1990. Risk associated with potential exposure to dioxin through consumption of 

tea brewed using tea bags containing bleached pulp. Chemosphere 20:1755-1762. 

*Sunahara GI, Nelson KG, Wong TK, et al. 1987. Decreased human birth weights after in utero exposure to 

PCBs and PCDFs are associated with decreased placental EGF-stimulated receptor autophosphorylation 

capacity. Mol Pharmacol 32: 572-578. 

Suskind RR. 1985. Chloracne, "the hallmark of dioxin intoxication." Scand J Work Environ Health 

11:165-171. 

*Suskind RR, Hertzberg VS. 1984. Human health effects of 2,4,5-T and its toxic contaminants. J Am Med 

Assoc 251:2372-2380. 
CDDs 655 

8. REFERENCES 

*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. 

Svensson B-G, Nilsson A, Hansson M, et al. 1991. Exposure to dioxins and dibenzofurans through the 

consumption of fish. N Engl J Med 324(1):8-12. 

*Swann RL, Laskowski DA, McCall PJ, et al. 1983. A rapid method for the estimation of the environmental 

parameters octanol/water partition coefficient, soil sorption constant, water to air ratio, and water solubility. 

Residue Rev 85:17-28. 

*Swanson HI, Bradfield CA. 1993. The Ah receptor: Genetics, structure and function. Pharmacogenetics 

3: 213-230. 

*Swanson HI, Chan WK, Bradfield CA. 1995. DNA binding specificities and pairing rules of the Ah 

receptor, ARNT, and SIM proteins. J Biol Chem 270: 26292-26302. 

*Swanson SE, Rappe C, Malmstr÷m J, et al. 1988. Emissions of PCDDs and PCDFs from the pulp 

industry. Chemosphere 17:681-691. 

*Sweeney A. 1994. Reproductive epidemiology of dioxins. In: Schecter A, ed. Dioxins and health. New 

York: Plenum Press, 549-585. 

Sweeney M, Fingerhut M, Arezzo J, et al. 1991. Neurologic changes in workers exposed to 2,3,7,8-TCDD. 

National Institute for Occupational Safety and Health, Cincinnati, OH, 56. 

*Sweeney MH, Fingerhut MA Arezzo J, et al. 1993. Peripheral neurophathy after occupational exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Am J Ind Med 23:845-858. 

*Sweeney MH, Fingerhut MA, Connally LB, et al. 1989. Progress of the NIOSH cross-sectional medical 

study of workers occupationally exposed to chemicals contaminated with 2,3,7,8-TCDD. Chemosphere 

19:973-977. 

Sweeney MH, Fingerhut MA, Patterson DG, et al. 1990. Comparison of serum levels of 2,3,7,8-TCDD in 

TCP production workers and in an unexposed comparison group. Chemosphere 20:993-1000. 

*Sweeney MP, Hornung RW, Wall DK, et al. 1992. Prevalence of diabetes and increased fasting serum 

glucose in workers with long-term exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. 12th International 

Symposium on Dioxins and Related Compounds, August 23-28 Tampere, Finland. 

Swerev M, Ballschmiter K. 1989. Pattern analysis of PCDDs and PCDFs in environmental samples as an 

approach to an occurrence/source correlation. Chemosphere 18:609-616. 

Takayama K, Miyata H, Mimura M, et al. 1988. Rapid analytical method for PCDDs, PCDFs, and 

polychlorobenzenes in samples from municipal waste incinerators. Eisei Kagaku 34:231-236. 

*Takimoto K, Lindahl R, Dunn TJ, et al. 1994. Structure of the 5' flanking region of class 3 aldehyde 

dehydrogenase in the rat. Arch Biochem Biophys 312: 539-546. 
CDDs 656 

8. REFERENCES 

*Takizawa Y, Muto H. 1987. PCDDs and PCDFs carried to the human body from the diet. Chemosphere 

16:1971-1975. 

Tanabe S, Kannan N, Subramanian A, et al. 1987. Highly toxic coplanar PCBs: Occurrence, source, 

persistency and toxic implications to wildlife and humans. Environ Pollut 47:147-163. 

Tanaka K, Hara S, Makino S. 1992. Development of new program for dioxin analysis using mass 

spectrometry. Kankyo Kagaku 2(2):428-9. 

Tarkowski S, Yrjanheikki E. 1986. Polychlorinated dibenzo-p-dioxins and dibenzofurans in human milk 

Reasons for concern. Chemosphere 15:1641-1648. 

Tarkowski S, Yrjanheikki E. 1989. Polychlorinated dibenzodioxins and dibenzofurans in human milk: An 

interim health risk evaluation. Chemosphere 18:1107-1113. 

*Tarkowski S, Yrjanheikki E. 1989. Polychlorinated dibenzodioxins and dibenzofurans in human milk: an 

interim health risk evaluation. Chemosphere 18(1-6):1107-1113. 

Tashiro C, Clement RE, Davies S, et al. 1990. Water round robin parts-per-quadrillion determination of 

PCDDs and PCDFs. Chemosphere 20:1313-1318. 

*Tashiro C, Clement RE, Lusis M, et al. 1989a. Monitoring dioxins and furans in precipitation samples. 

Chemosphere 18:777-782. 

*Tashiro C, Clement RE, Reid N, et al. 1989b. Determination of dioxins and furans in precipitation 

collected in urban and rural Ontario locations. Chemosphere 19:535-540. 

Tashiro C, Clement RE, Szakolcai A, et al. 1989c. Comparison of high volume sampling techniques for 

dioxins and furans in ambient air. Chemosphere 19:1-6. 

Taylor JS. 1979. Environmental chloracne: Update and overview. Ann NY Acad Sci 320:295-307. 

*Taylor MJ, Lucier GW, Mahler JF, et al. 1992. Inhibition of acute TCDD toxicity by treatment with 

anti-tumor necrosis factor antibody or dexamethasone. Toxicol Appl Pharmacol 117(1):126-132. 

*Tenchini ML, Crimaudo C, Pacchetti G, et al. 1983. A comparative cytogenetic study on cases of induced 

abortions in TCDD-exposed and nonexposed women. Environ Mutagen 5:73-85. 

*Tepper A, Burr S, Piacitelli L, et al. 1997. Serum levels of polychlorinated dibenzo-p-dioxins and 

dibenzofurans in pulp and paper mill workers. Chemosphere 34(5-7):1587-1603. 

Theisen J, Funcke W, Balfanz E, et al. 1989. Determination of PCDFs and PCDDs in fire accidents and 

laboratory combustion tests involving PVC-containing materials. Chemosphere 19:423-428. 

*Theobald HM, Ingall GB, Mably TA, et al. 1991. Response of the antral mucuso of the rat stomach to 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 108:167-179. 

Thibodeaux LJ, Lipsly D. 1985. A fate and transport model for 2,3,7,8-tetrachlorodibenzo-p-dioxin in fly 

ash on soil and urban surfaces. Hazard Waste Hazard Mater 2:225-235. 
CDDs 657 

8. REFERENCES 

Thiess AM, Frentzel-Beyme R. 1977. Mortality study of persons exposed to dioxin following an accident 

which occurred in the BASF on 13, November 1953. Working Papers, Joint NIEHS/IARC Working Group 

Report. Lyon, France, June. 

*Thiess AM, Frentzel-Beyme R, Link R. 1982. Mortality study of persons exposed to dioxin in a 

trichlorophenol-process accident that occurred in the BASF/AG on November 17, 1953. Am J Ind Med 

3:179-189. 

*Thigpen JE, Faith RE, McConnell EE, et al. 1975. Increased susceptibility of bacterial infection as a 

sequela of exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Infect Immun 12:1319-1324. 

*Thoma H. 1988. PCDD/F-concentrations in chimney soot from house heating systems. Chemosphere 

17:1369-1379. 

*Thoma H, Mucke W, Kauert G. 1990. Comparison of the polychlorinated dibenzo-p-dioxin and 

dibenzofuran in human tissue and human liver. Chemosphere 20:433-442. 

Thoma H, Muecke W, Kretschmer E. 1989. Concentrations of PCDD and PCDF in human fat and liver 

samples. Chemosphere 18:491-498. 

Thomann RV. 1989. Bioaccumulation model of organic chemical distribution in aquatic food chains. 

Environ Sci Technol 23:699-707. 

Thomas HF. 1980. Internal memo to P. Cohn, Washington, DC: U.S. Environmental Protection Agency, 

Office of Toxic Substances. 

*Thomas PT, Hinsdill RD. 1979. The effect of perinatal exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin on 

the immune response of young mice. Drug Chem Toxicol 2:77-98. 

*Thomas VM, Spiro TG. 1995. An estimation of dioxinemissiions in the United States. Toxicology and 

Environmental Chemistry 50:1-37. 

*Thomas VM, Spiro TG. 1996. The U.S. dioxin inventory: are there missing sources? Environmental 

Science & Technology 30 (2):82A-85A. 

Thompson HC Jr, Kendall DC, Korfmacher WA, et al. 1986a. Assessment of the contamination of a 

multibuilding facility by polychlorinated biphenyls, polychlorinated dibenzo-p-dioxins, and polychlorinated 

dibenzofuran. Environ Sci Technol 20:597-607. 

Thompson HC Jr, Kendall DC, Korfmacher WA, et al. 1986b. Polychlorinated dibenzo-p-dioxins, 

polychlorinated dibenzofurans, and polychlorinated biphenyls at a contaminated multibuilding facility. In: 

Rappe C, Choudhary G, Keith LH, eds. Chlorinated dioxins and dibenzofurans in perspective. Chelsea, MI: 

Lewis Publishers, Inc., 121-136. 

*Thompson TS, Clement RE, Thornton N, et al. 1990. Formation and emission of PCDDs/PCDFs in the 

petroleum refining industry. Chemosphere 20(10-12):1525-32. 

*Thunberg T. 1984. Effects of TCDD on vitamin A and its relation to TCDD toxicity. Banbury Rep 

38:333-344. 
CDDs 658 

8. REFERENCES 

*Thunberg T, Ahlborg UG, Hakansson H, et al. 1980. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on the 

hepatic storage of retinol in rats with different dietary supplies of vitamin A (retinol). Arch Toxicol 

45:273-285. 

*Thunberg T, Ahlborg UG, Johnsson H. 1979. Vitamin A (retinol) status in the rat after a single oral dose 

of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Arch Toxicol 42:265-274. 

*Thunberg T, Ahlborg UG, Wahlstr÷m B. 1984. Comparison between the effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin and six other compounds on the vitamin A storage, the UDP-glucuronosyltransferase and 

the aryl hydrocarbon hydroxylase activity in the rat liver. Arch Toxicol 55:16-19. 

*Thunberg T, Hakansson H. 1983. Vitamin A (retinol) status in the Gunn rat: The effect of 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Arch Toxicol 53:225-233. 

Tiernan TO, Taylor ML, Garrett JH, et al. 1983. PCDDs, PCDFs and related compounds in the effluents 

from combustion processes. Chemosphere 12:595-606. 

*Tiernan TO, Taylor ML, Garrett JH, et al. 1985. Sources and fate of polychlorinated dibenzodioxins, 

dibenzofurans and related compounds in human environments. Environ Health Perspect 59:145-158. 

Tiernan TO, Taylor ML, Vanness GF, et al. 1984. Analyses of human tissues for chlorinated 

dibenzo-p-dioxins and chlorinated dibenzofurans: The state of the art. In: Lowrance WW, ed. Public 

health risks of the dioxins. New York, NY: Rockerfeller University, 31-56. 

Tiernan TO, Wagel DJ, Vanness GF, et al. 1989a. Dechlorination of PCDD and PCDF sorbed on activated 

carbon using the KPEG reagent. Chemosphere 19:573-578. 

*Tiernan TO, Wagel DJ, Vanness GF, et al. 1989b. PCDD/PCDF in the ambient air of metropolitan area in 

the U.S. Chemosphere 19:541-546. 

TMN. 1987. Survey finds little dioxin at control sites: Industry finds trace amounts in paper products and 

highest levels at pesticide plants. Toxic Materials News. September 30, 301. 

Tognoni G, Bonaccorsi A. 1982. Epidemiological problems with TCDD: A critical view. Drug Metab Rev 

13:447-469. 

*Tomaszewski KE, Montgomery CA, Melnick RL. 1988. Modulation of 

2,3,7,8-tetrachlorodibenzo-p-dioxin toxicity in F344 rats by di(2-ethylhexyl)phthalate. Chem Biol Interact 

65:205-222. 

*Tong H , Gross M, Schecter A, et al. 1990. Sources of dioxins in the environment : Second stage study of 

PCDDs/PCDFs in ancient human tissue and environmental samples. Chemosphere 20:987-992. 

*Tong HY, Arghestani S, Gross ML, et al. 1989a. Polychlorodibenzodioxins and polychlorodibenzofurans 

in the dust of the ambient atmosphere in a municipal waste incinerator. Chemosphere 18:577-584. 
CDDs 659 

8. REFERENCES 

Tong HY, Karasek FW. 1986. Comparison of quantitation of polychlorinated dibenzodioxins and 

polychlorinated dibenzofurans in complex environmental samples by high resolution gas chromatography 

with flame ionization, electron capture and mass spectrometric detection. Chemosphere 15:1141-1146. 

*Tong HY, Monson SJ, Gross ML, et al. 1989b. Elevated levels of 2,3,7,8-TCDD in the tissue of an 

agricultural sprayer of herbicides: A single case study. Chemosphere 18:469-476. 

Tong HY, Shore DL, Karasek FW. 1984b. Isolation of polychlorinated dibenzodioxins and polychlorinated 

dibenzofurans for a complex organic mixture by two-step liquid chromatographic fractionation for 

quantitative analysis. Anal Chem 56:2442-2447. 

Tong HY, Shore DL, Karasek FW, et al. 1984a. Identification of organic wastes obtained from incineration 

of municipal waste by high-performance liquid chromatographic fractionation and gas chromatography-mass 

spectrometry. J Chromatogr 285:413-441. 

*Tonn T, Esser C, Schneider EM, et al. 1996. Persistence of decreased T-helper cell function in industrial 

workers 20 years after exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Environ Health Perspect 

104:422-426. 

*Toth K, Olah E, Somfai-Relle S, et al. 1984. Effect of herbicide 2,4,5-trichlorophenoxyethanol (TCPE) 

containing dioxin on mutation and induction of sister chromatid exchanges. Carcinogenesis 5:1725-1728. 

*Toth K, Somfai-Relle S, Sugar J, et al. 1979. Carcinogenicity testing of herbicide 2,4,5-trichlorophenoxyethanol containing dioxin and of pure dioxin in Swiss mice. Nature 278:548-549. 

*Towara J, Hiller B, Hutzinger O, et al. 1992. PCDD/F in distillation residues from dry cleaners. 

Chemosphere 25:1509-1516. 

Townsend DI. 1980. The use of dioxin isomer group ratios to identify sources and define background levels 

of dioxins in the environment. J Environ Sci Health 15:571-609. 

*Townsend JC, Bodner KM, Van Peenen PFD, et al. 1982. Survey of reproductive events of wives of 

employees exposed to chlorinated dioxins. Am J Epidemiol 115:695-713. 

Travis CC, Arms AD. 1988. Bioconcentration of organics in beef, milk, and vegetation. Environ Sci 

Technol 22:271-274. 

*Travis CC, Hattemer-Frey HA. 1987. Human exposure to 2,3,7,8-TCDD. Chemosphere 16:2331-2342. 

*Travis CC, Hattemer-Frey HA. 1989. Human exposure to dioxin from municipal solid waste incineration. 

Waste Management 9:151-156. 

*Travis CC, Hattemer-Frey HA. 1991. Human exposure to dioxin. The Science of the Total Environment 

104(1-2):97-127. 

*Tritscher AM, Goldstein JA, Portier CJ, et al. 1992. Dose-response relationships for chronic exposure to 

2,3,7,8-tetrachlorodibenzo-p-dioxin in a rat tumor promotion model: Quantification and immunolocalization 

of CYP1A1 and CYP1A2 in the liver. Cancer Res 52:3436-3442. 
CDDs 660 

8. REFERENCES 

Tsushimoto G, Matsumura F, Sago R. 1982. Fate of 2,3,7,8-TCDD in an outdoor pond and in model 

aquatic ecosystems. Environ Toxicol Chem 1:61-68. 

*Tucker AN, Vore SJ, Luster MI. 1986. Suppression on B cell differentiation by 2,3,7,8-tetrachlorodibenzo-p-dioxin. Mol Pharmacol 29:372-377. 

Tuinstra LGMT, Van Rhijn JA, Roos AH, et al. 1990. Determination of planar chlorobiphenyls in animal 

fat. Journal of High Resolution Chromatography 13:797-802. 

*Tulp MTM, Hutzinger O. 1978. Rat metabolism of polychlorinated dibenzo-p-dioxins. Chemosphere 

7:761-768. 

*Tuomisto J, Pohjanvirta R. 1991. Do we hypothese on the mechanism of action of dioxins help in risk 

assessment? Sci Tot Environ 106:21-31. 

*Tuomisto J, Pohjanvirta R, MacDonald E, et al. 1990. Changes in rat brain monoamines, monoamine 

metabolites and histamine after a single administration of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). 

Pharmacol Toxicol 67:260-265. 

Turkstra E, Pols HB. 1989. PCDDs and PCDFs in Dutch inland waters. Chemosphere 18:539-551. 

*Turner JN, Collins DN. 1983. Liver morphology in guinea pigs administered either pyrolysis products of a 

polychlorinated biphenyl transformer fluid or 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Appl Pharmacol 

67:417-429. 

*Turteltaub KW, Felton JS, Gledhill BL, et al. 1990. Accelerator mass spectrometry in biomedical 

dosimetry: Relationship between low-level exposure and covalent binding of heterocyclic amine carcinogens 

to DNA. Proc Natl Acad Sci USA 87:5288-5292. 

*Tysklind M, Faengmark I, Marklund S, et al. 1993. Atmospheric transport and transformation of 

polychlorinated dibenzo-p-dioxins and dibenzofurans. Environ Sci Technol 27(10):2190-7. 

*U.S. Congress. 1986. Superfund amendments and reauthorization act of 1986. Title III. Emergency 

planning and community right-to-know act. Ninety-ninth Congress of the United States of America. Second 

Sesstion. January 21, 1986. 

U.S. Congress 1991. Dioxin treatment technologies background paper. Office of Technology Assessment. 

OTA-BP-0-93. Washington, DC: U.S. Government Printing Office. November 1991. 

*Uematsu F, Kikuchi H, Motomiya M, et al. 1991. Association between restriction frgament length 

polymorphism of the human cytochrome P450IIE1 gene and susceptibility to lung cancer. Jpn J Cancer Res 

82:254-256. 

*Umbreit TH, Hesse EJ, Gallo MA. 1986a. Bioavailability of dioxin in soil from a 2,4,5-T manufacturing 

site. Science 232:497-499. 

*Umbreit TH, Hesse EJ, Gallo MA. 1986b. Comparative toxicity of TCDD contaminated soil from Times 

Beach, Missouri, and Newark, New Jersey. Chemosphere 15:2121-2124. 
CDDs 661 

8. REFERENCES 

*Umbreit TH, Hesse EJ, Gallo MA. 1987. Reproductive toxicity in female mice of dioxin-contaminated 

soils from a 2,4,5-trichlorophenoxyacetic acid manufacturing site. Arch Environ Contam Toxicol 

16:461-466. 

*Umbreit TH, Hesse EJ, Gallo MA. 1988. Reproductive studies of C57B/6 male mice treated with 

TCDD-contaminated soils from a 2,4,5-trichlorophenoxyacetic acid manufacturing site. Arch Environ 

Contam Toxicol 17:145-150. 

*Umbreit TH, Patel D, Gallo MA. 1985. Acute toxicity of TCDD contaminated soil from an industrial site. 

Chemosphere 14:945-947. 

*Unkila M, Pohjanvirta R, MacDonald E, et al. 1993a. Differential effect of TCDD on brain serotonin 

metabolism in a TCDD-susceptible and a TCDD-resistant rat strain. Chemosphere 27(1-3):401-6. 

*Unkila M, Pohjanvirta R, MacDonald E, et al. 1994. Dose response and time course of alterations in 

tryptophan metabolism by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the most TCDD-susceptible and 

the most TCDD-resistant rat strain: Relationship with TCDD lethality. Toxicol Appl Pharmacol 

128:280-292. 

*Unkila M, Tuomisto JT, Pohjanvirta R, et al. 1993b. Effect of a single lethal dose of TCDD on the levels 

of monoamines, their metabolites and tryptophan in discrete brain nuclei and peripheral tissues of 

Long-Evans rats. Pharmacol Toxicol 72(4-5):279-285. 

*USAF. 1991. Air Force health study: An epidemiological investigation of health effects in Air Force 

personnel following exposure to herbicides. Brooks Air Force Base, TX: U.S. Air Force, Chapters 1-5, 18, 

19. 

USDOC. 1984. Technical issues concerned with PCDD and PCDF formation and destruction in MSW fired 

incinerators. Gaithersburg, MD: National Bureau of Standards, U.S. Department of Commerce. 

Uza M, Hunsinger R, Thompson T, et al. 1990. Validation studies of an automated preconcentration water 

sampler APS for chlorinated dibenzo-p-dioxins and dibenzofurans. Chemosphere 20:1349-1354. 

Vaino H, Hesso A, Haeppinen P. 1989. Chlorinated dioxins and dibenzofurans in the environment. Scand J 

Work Environ Health 15:377-382. 

*Valli K, Wariishi H, Gold MH. 1992. Degradation of 2,7-dichlorodibenzo-p-dioxin by the 

lignin-degrading Basidiomycete Phanerochaete chrysosporium. Journal of Bacteriology 174(7):2131-2137. 

*Van Birgelen APJM, Van der Kolk J, Fase KM, et al. 1995. Subchronic dose-response study of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in female Sprague-Dawley rats. Toxicol Appl Pharmacol 132:1-13. 

*Van den Berg M, Birnbaum L, Bosveld ATC, et al. 1998. Toxic equivalency factors (TEFs) for PCBs, 

PCDDs, and PCDFs for humans and wildlife. Environ Health Perspect 106(12):775-792. 

Van den Berg M, Blank F, Heeremans C, et al. 1987a. Presence of polychlorinated dibenzo-p-dioxins and 

polychlorinated dibenzofurans in fish-eating birds and fish from the Netherlands. Arch Environ Contam 

Toxicol 165:149-158. 
CDDs 662 

8. REFERENCES 

*Van den Berg M, De Jongh J, Poiger H, et al. 1994. The toxicokinetics and metabolism of 

polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs) and their relevance to toxicity. Crit 

Rev. Toxicol 24:1-74. 

Van den Berg M, de Vroom E, van Greevenbroek M, et al. 1985. Bioavailability of PCDDs and PCDFs 

absorbed on fly ash in rat, guinea pig and Syrian golden hamster. Chemosphere 14:865-869. 

*Van den Berg M, Heeremans C, Veenhoven E, et al. 1987b. Transfer of polychlorinated dibenzo-p-dioxins 

and dibenzofurans to fetal and neonatal rats. Fundam Appl Toxicol 9:635-644. 

Van den Berg M, Meerman L, Olie K, et al. 1986a. Retention of PCDDs and PCDFs in the liver of the rat 

and hamster after oral administration of a municipal incinerator fly ash extract. Toxicol Environ Chem 

12:267-284. 

*Van den Berg M, Olie K, Hutzinger O. 1983. Uptake and selection in rats of orally administered 

chlorinated dioxins and dibenzofurans from fly-ash and fly-ash extract. Chemosphere 12:537-544. 

*Van den Berg M, Sinke M, Wever H. 1987c. Vehicle dependent bioavailability of polychlorinated 

dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs) in the rat. Chemosphere 16:1193-1203. 

*Van den Berg M, Van der Wielen FWM, Olie K, et al. 1986b. Presence of PCDDs and PCDFs in human 

breast milk from the Netherlands. Chemosphere 15:693-706. 

Van den Berg M, Van Greevenbroek M, Olie K, et al. 1986c. Bioavailability of polychlorinated 

dibenzo-p-dioxins and polychlorinated dibenzofurans on fly ash after semi-chronic oral ingestion by the rat. 

Chemosphere 15:509-518. 

*Van den Berg M, Van Wijnen J, Wever H, et al. 1989. Selective retention of toxic polychlorinated 

dibenzo-p-dioxins and dibenzofurans in the liver of the rat after intravenous administration of a mixture. 

Toxicology 55:173-182. 

*Van den Heuvel JP, Clark GC, Kohn MC, et al. 1994. Dioxin-responsive genes: Examination of 

dose-response relationships using quantitative reverse transcriptase-polymerase chain reaction. Cancer Res 

54(1):62-68. 

*Van den Heuvel JP, Clark GC, Thompson CL, et al. 1993. CYP1A1 mRNA levels as a human exposure 

biomarker: use of quantitative polymerase chain reaction to measure CYP1A1 expression in human 

peripheral blood lymphocytes. Carcinogenesis 14:2003-2006. 

Van den Heuvel JP, Lucier G. 1993. Environmental toxicology of polychlorinated dibenzo-p-dioxins and 

polychlorinated dibenzofurans. Environ Health Perspect 100:189-200. 

Van Der Kolk J, Van Birgelen A P JM, Poiger H, et al. 1992. Interactions of 2,2,4,4,5,5-hexachlorobiphenyl and 2,3,7,8-tetrachloeodibenzo-p-dioxin in a subchronic feeding study in the rat. Chemosphere 

25:2023-2027. 

Van der Welden MEJ, Craane LHJ, Evers EHG, et al. 1989. Bioavailability of PCDDs and PCDFs from 

bottom sediments and some associated biological effects in the carp (Cyprinus carpio). Chemosphere 

19:1009-1016. 
CDDs 663 

8. REFERENCES 

Van Logten MJ, Gupta BN, McConnell EE, et al. 1980. Role of the endocrine system in the action of 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the thymus. Toxicology 15:135-144. 

*Van Loveren H, Schuurman H-J, Kampinga J, et al. 1991. Reversibility of thymic atrophy induced by 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and bis(tri-n-butyltin)oxide (TBTO). Int J Immunopharmacol 

13(4):369-378. 

*Van Miller JP, Lalich JJ, Allen JR. 1977. Increased incidence of neoplasms in rats exposed to low levels 

of 2,3,7,8-tetrachlorodibenzo-p-dioxin. Chemosphere 6:537-544. 

*Van Miller JR, Marlar RJ, Allen JR. 1976. Tissue distribution and excretion of tritiated 

tetrachlorodibenzo-p-dioxin in non-human primates and rats. Food Cosmet Toxicol 14:31-34. 

VANR. 1988. Vermont Agency of Natural Resources. Department of Environmental Conservation, 

Groundwater Protection Regulations, effective 9/29/88. Washington, DC: Bureau of National Affairs, Inc., 

State of Vermont Chapter 12. 

*Vartiainen T, Jaakkola JJK, Saarikoski S, et al. 1998. Birth weight and sex of children and the correlation 

to the body burden of PCDDs/PCDFs and PCBs of the mother. Environ Health Perspect 106:61-66. 

*Vecchi A, Mantovani A, Sironi M, et al. 1980a. The effect of acute administration of 

2,3,7,8-tetrachlorodibenzo-p-dioxin on humoral antibody production and cell-mediated activities in mice. 

Arch Toxicol Suppl 4:163-165. 

*Vecchi A, Mantovani A, Sironi M, et al. 1980b. Effect of acute exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin on humoral antibody production in mice. Chem-Biol Interact 30:337-342. 

*Vecchi A, Sironi M, Canegrati M, et al. 1983b. Immunosuppressive effects of 2,3,7,8-tetrachlorodibenzop-dioxin in the strains of mice with different susceptibility to induction of aryl hydrocarbon hydroxylase. 

Toxicol Appl Pharmacol 68:434-441. 

*Vecchi A, Sironi M, Canegrati MA, et al. 1983a. Comparison of the immunosuppressive effects in mice of 

2,3,7,8-tetrachlorodibenzo-p-dioxin and 2,3,7,8-tetrachlorodibenzofuran. In: Choudhary G, Keith LH, 

Rappe C, eds. Chlorinated dioxins and dibenzofurans in the total environment. Boston, MA: Butterworth, 

397-402. 

Velzy CO. 1986. ASME standard sampling and analysis methods for dioxins/furans. Chemosphere 

15:1179-1185. 

*Vena J, Boffetta P, Becher H, et al. 1998. Exposure to dioxin and nonneoplastic mortality in the expanded 

IARC international cohort study of phenoxy herbicide and chlorophenol production workers and sprayers. 

Environ Health Perspect 106(Suppl 2):645-653. 

*Vieira I, Sonnier M, Cresteil T. 1996. Developmental expression of CYP2E1 in the human liver: 

hypermethylation control of gene expression during the neonatal period. European Journal of Biochemistry 

238:476-483. 
CDDs 664 

8. REFERENCES 

*Vikelsoe J, Madsen H, Hansen K. 1994. Emission of dioxins from Danish woodstoves. Chemosphere 

29(9-11):2019-2027. 

*Viluksela M, Stahl BU, Birnbaum LS, et al. 1998a. Subchronic/chronic toxicity of a mixture of four 

chlorinated dibenzo-p-dioxins in rats. I. Design, general observations, hematology, and liver concentrations. 

Toxicol Appl Pharmacol 151:57-69. 

*Viluksela M, Stahl BU, Birnbaum LS, et al. 1998b. Subchronic/chronic toxicity of a mixture of four 

chlorinated dibenzo-p-dioxins in rats. II. Biochemical effects. Toxicol Appl Pharmacol 151:70-78. 

*Viluksela M, Stahl BU, Rozman KK. 1994. Subchronic (13-week) toxicity of heptachlorodibenzop-dioxin in male Sprague-Dawley rats. Chemosphere 29:2381-2393. 

Vogg H, Metzger M, Stieglitz L. 1987. Recent findings on the formation and decomposition of 

PCDD/PCDF in municipal solid waste incineration. Waste Manage Res 5:285-294. 

*Vos JG, Kreeftenberg JG, Engel HWB, et al. 1978. Studies on 2,3,7,8-tetrachlorodibenzo-p-dioxininduced immune suppression and decreased resistance to infection: endotoxin hypersensitivity, serum zinc 

concentrations and effect of thymosin treatment. Toxicology 9:75-86. 

*Vos JG, Moore JA. 1974. Suppression of cellular immunity in rats and mice by maternal treatment with 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Int Arch Allergy 47:777-794. 

*Vos JG, Moore JA, Zinkl JG. 1973. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin on the immune system 

of laboratory animals. Environ Health Perspect 5:149-162. 

WAC. 1988. Control of hazardous pollutants. Wisconsin Administrative Code, Chapter NR 445. 

Department of Natural Resources. 

*Wacker R, Poiger H, Schlatter C. 1986. Pharmacokinetics and metabolism of 1,2,3,7,8-pentachlorodibenzo-p-dioxin in the rat. Chemosphere 15:1473-1476. 

*Wagner SL, Durand LR, Inman RD, et al. 1991. Residues of pentachlorophenol and other chlorinated 

contaminants in human tissues: Analysis by electron capture gas chromatography and electron capture 

negative ion mass spectrometry. Arch Environ Contam Toxicol 21(4):596-606. 

*Wahba ZZ, Lawson TW, Murray WJ, et al. 1989. Factors influencing the induction of DNA single strand 

breaks in rats by 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Toxicology 58:57-69. 

Wakimoto T, Tatsukawa R. 1985. Polychlorinated dibenzo-p-dioxins and dibenzofurans in fly ash and 

cinders collected from several municipal incinerators in Japan. Environ Health Perspect 59:159-162. 

*Walden R, Schiller CM. 1985. Comparative toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in 

four (sub)strains of adult male rats. Toxicol Appl Pharmacol 77:490-495. 

*Walker NJ, Miller BD, Kohn MC, et al. 1998. Differences in kinetics of induction and reversibility of 

TCDD-induced changes in cell proliferation and CYP1A1 expression in female Sprague-Dawley rat liver. 

Carcinogenesis 19:1427-1435. 
CDDs 665 

8. REFERENCES 

*Waller CL, McKinney JD. 1995. Three-dimensional quantitative structure-activity relationships of dioxins 

and dioxin-like compounds: Model validation and Ah receptor characterization. Chem Res Toxicol 

8:847-858. 

Walters RW, Guiseppi-Elie A. 1988. Sorption of 2,3,7,8-tetrachlorodibenzo-p-dioxin to soils from 

water/methanol mixtures. Environ Sci Technol 22:819-825. 

*Wan MT, Van Oostdam J. 1995. Utility and railroad rights-of-way contaminants: Dioxins and furans. J 

Environ Quality 24(2):61-69. 

*Wang DKW, Chiu DH, Leung PK, et al. 1983. Sampling and analytical methodologies for PCDDs and 

PCDFs in incinerators and wood burning facilities. Environ Sci Res 26:113-126. 

*Wang GL, Jiang B-H, Rue EA, et al. 1995. Hypoxia-inducible factor 1 is a basic-helix-loop-helix-PAS 

heterodimer regulated by cellular O2 tension. Proc Natl Acad Sci USA 92: 5510-5514. 

*Wang GL, Semanza GL. 1995. Purification and characterization of hypoxia-inducible factor 1. J Biol 

Chem 270: 1230-1237. 

*Wang X, Narasimhan TR, Morrison V, et al. 1991. In situ and in vitro photoaffinity labeling of the nuclear 

aryl hydrocarbon receptor from transformed rodent and human cell lines. Arch Biochem Biophys 

287:186-194. 

*Wang X, Santostefano MJ, Evans MV, et al. 1997. Determination of parameters responsible for 

pharmacokinetic behavior of TCDD in female Sprague-Dawley rats. Toxicol Appl Pharmacol 147:151-168. 

*Ward CT, Matsumura F. 1978. Fate of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in a model aquatic 

environment. Arch Environ Contam Toxicol 7:349-357. 

Wassom JS, Huff JE, Loprieno N. 1978. A review of the genetic toxicology of chlorinated 

dibenzo-p-dioxins. Mutat Res 47:141-160. 

WDNR. 1987. Surface water quality criteria for toxic substances. Order of the State of Wisconsin Natural 

Resources Board repealing, renumbering, renumbering and amending, and amending, repealing and 

recreating, and creating rules. Section 27, Chapter NR 105. Wisconsin Department of Natural Resources. 

*Webb K, Ayres S, Slavin R, et al. 1984. Results of a pilot study of health effects due to 

2,3,7,8-tetrachlorodibenzodioxin contamination-Missouri. MMWR 33:54-61. 

*Webb KB, Evans RG, Knutsen AP, et al. 1989. Medical evaluation of subjects with known body levels of 

2,3,7,8-tetrachlorodibenzo-p-dioxin. J Toxicol Environ Health 28:183-193. 

*Weber H, Birnbaum LS. 1985. 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and 

2,3,7,8-tetrachlorodibenzofuran (TCDF) in pregnant C57BL/6N mice: Distribution to the embryo and 

excretion. Arch Toxicol 57:159-162. 

*Weber H, Harris MW, Haseman JK, et al. 1985. Teratogenic potency of TCDD, TCDF, and TCDD-TCDF 

combinations in C57BL/6N mice. Toxicol Lett 26:159-167. 
CDDs 666 

8. REFERENCES 

*Weber H, Poiger H, Schlatter C. 1982. Acute oral toxicity of TCDD-metabolites in male guinea pigs. 

Toxicol Lett 14:117-122. 

*Weber LW, Lebofsky M, Stahl BU, et al. 1992a. Comparative toxicity of four chlorinated 

dibenzo-p-dioxins (CDDs) and their mixture. Part II: Structure-activity relationships with inhibition of 

hepatic phospholpyruvate carboxykinase, pyruvate carboxylase, and gamma-glutamyl transpeptidase 

activities. Arch Toxicol 66(7):478-483. 

*Weber LW, Lebofsky M, Stahl BU, et al. 1992b. Comparative toxicity of four chlorinated 

dibenzo-p-dioxins (CDDs) and their mixture. Part III: Structure-activity relationship with increased plasma 

tryptophan levels, but no relationship to hepatic ethoxyresorufin o-deethylase activity. Arch Toxicol 

66(7):484-488. 

*Weber LWD, Lebofsky M, Greim H, et al. 1991a. Key enzymes of gluconeogenesis are dose-dependently 

reduced in 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)-treated rats. Arch Toxicol 65:119-123. 

*Weber LWD, Lebofsky M, Stahl BU, et al. 1991b. Reduced activities of key enzymes of gluconeogenesis 

as possible cause of acute toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in rats. Toxicology 

66:133-144. 

*Weber LWD, Lebofsky M, Stahl BU, et al. 1995. Correlation between toxicity and effects on intermediate 

metabolism in 2,3,7,8-tetrachlorodibenzo-p-dioxin-treated male C57BL/6J and DBA/2J mice. Toxicol Appl 

Pharmacol 131:155-162. 

*Weber LWD, Palmer CD, Rozman K. 1994. Reduced activity of trytophan 2,3,-dioxygenase in the liver of 

rats treated with chlorinated dibenzo-p-dioxins (CDDs): Dose-responses and structure-activity relationship. 

Toxicol 86:63-69. 

*Weber LWD, Stahl BU, Rozman K. 1992c. Are serotonergic mechanisms involved in the acute toxicity of 

chlorinated dibenzo-p-dioxins (CDDs)? Chemosphere 25:161-164. 

*Weber LWD, Zesch A, Rozman K. 1991d. Penetration, distribution and kinetics of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in human skin in vitro. Arch Toxicol 65:421-428. 

Weber LWD, Zesch A, Rozman K. 1992d. Decontamination of human skin exposed to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in vitro. Arch Environ Health 47:302-308. 

*Weber TJ, Ou X, Hart C, et al. 1991c. Modulation of protein phosphorylation in rat aortic smooth muscle 

cells by 2,3,7,8-tetrachlorodibnezo-p-dioxin (TCDD) in vivo [Abstract 1335]. Toxicologist 11: 340. 

*Webster GRB, Friesen KL, Sarna LP, et al. 1985. Environmental fate modelling of chlorodioxins: 

Determination of physical constants. Chemosphere 14:609-622. 

Webster GRB, Sarna LP, Friesen K, et al. 1984. HPLC techniques for the evaluation of the interactions of 

organic contaminants with individual environmental compartments. Am Chem Soc Div Environ Chem 

24:42-45. 
CDDs 667 

8. REFERENCES 

*Weerasinghe NCA, Schecter AJ, Pan JC, et al. 1986. Levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(2,3,7,8-TCDD) in adipose tissue of Vietnam veterans seeking medical assistance. Chemosphere 

15:1787-1794. 

Weeren RD, Asshauer J. 1985. Problem and results of trace analysis of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

in 2,4,5-trichlorophenoxyacetic acid and its esters. J Assoc Off Anal Chem 68:917-921. 

*Weib C, Kolluri SK, Kiefer F, et al. 1996. Complementation of Ah receptor deficiency in hepatoma cells: 

negative feedback regulation and cell cycle control by the Ah receptor. Exp Cell Res 226:154-163. 

*Weisglas-Kuperus N, Sas TCJ, Koopman-Esseboom C, et al. 1995. Immunologic effects of background 

prenatal and postnatal exposure to dioxins and polychlorinated biphenyls in Dutch infants. Pediatr Res 

38:404-410. 

*Weisiger R, Gollan J, Ockner R. 1981. Receptor for albumin on the liver cell surface may mediate uptake 

of fatty acids and other albumin-bound substances. Science 211:1048-1050. 

*Weissberg JB, Zinkl JG. 1973. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin upon hemostasis and 

hematologic function in the rat. Environ Perspect 5:119-124. 

Welch KJ, Hoffman NE. 1993. Analysis of polychlorinated biphenyls, dibenxodioxins and dibenzofurans 

by online coupled micro-column HPLC capillary GC-MS. Journal of Liquid Chromatography 16:307-313. 

*Welschpausch K, McLachlan MS, Umlauf G. 1995. Determination of the principal pathways of 

polychlorinated dibenzo-p-dioxins and dibenzofurans to Lolium multiflorum ,Welsh ray grass. Environ Sci 

Technol 24(4):1090-1098. 

*Wendling JM, Orth RG, Poiger H. 1990. Determination of [3H]-2,3,7,8-tetrachlorodibenzo-p-dioxin in 

human feces to ascertain its relative metabolism in man. Anal Chem 62:796-800. 

*Wenning RJ, Erickson G. 1994. Interpretation and analysis of complex environmental data using 

chemometric methods. Trends in Analytical Chemistry 13(10):446-456. 

Wenning RJ, Harris MA, Finley B, et al. 1993a. Application of pattern recognition techniques to evaluate 

polychlorinated dibenzo-p-dioxin and dibenzofuran distributions in surficial sediments from the lower 

Passaic River and Newark Bay. Ecotoxicol Environ Saf 25(1):103-125. 

Wenning RJ, Harris MA, Paustenbach DJ, et al. 1992. Potential sources of 1,2,8,9-tetrachlorodibenzo-p-dioxin in the aquatic environment. Ecotoxicol Environ Saf 23(2):133-146. 

*Wenning RJ, Paustenbach D, Johnson G, et al. 1993b. Chemometric analysis of potential sources of 

polychlorinated dibenzo-p-dioxins and dibenzofurans in surficial sediments from Newark Bay, New Jersey. 

Chemosphere 27(1-3):55-64. 

*West JR, Smith HW, Chasis H. 1948. Glomerular filtration rate, effective renal blood flow, and maximal 

tubular excretory capacity in infancy. J. of Pediatrics 32a:10-18. 
 

CDDs 668 

8. REFERENCES 

*White DH, Hoffman DJ. 1995. Effects of polychlorinated dibenzo p-dioxins and dibenzofurans on nesting 

wood ducks (aix sponsa) at Bayou Meto, Arkansas. Environmental Health Perspectives 

103(suppl.4):37-39. 

*White KL Jr, Lysy HH, McCay JA, et al. 1986. Modulation of serum complement levels following 

exposure to polychlorinated dibenzo-p-dioxins. Toxicol Appl Pharmacol 84:209-219. 

*Whitelaw M, Pongratz I, Gustafsson J-A, et al. 1993. Ligand-dependent recruitment of the Arnt 

coregulator determines DNA recognition by the dioxin receptor. Mol Cell Biol 13:2504-2514. 

Whitlock JB Jr. 1990. Genetic and molecular aspects of 2,3,7,8-tetrachlorodibenzo-p-dioxin action. Ann 

Rev Pharmacol Toxicol 30:251-277. 

*Whitlock JP Jr. 1987. The regulation of gene expression by 2,3,7,8-tetrachlorodibenzo-p-dioxin. 

Pharmacol Rev 39:147-161. 

*Whitlock JP Jr. 1993. Mechanistic aspects of dioxin action. Chem Res Toxicol 6:754-763. 

*Whittle DM, Mageau C, Duncan RK, et al. 1993. Canadian national dioxin sampling program: Dioxins 

and furans in biota near 46 puop and paper mills using the chlorine bleaching process. Chemosphere 

27(1-3):279-86. 

WHO. 1988. Environmental Health 29. No. 29. PCBs, PCDDs, and PCDFs in breast milk. Assessments 

of Health Risks. World Health Organization, Copenhagen Denmark. 

*WHO. 1991. Consultation on tolerable daily intake from food of PCDDs and PCDFs. Summary report. 

Bilthoven, Netherlands: World Health Organization, Copenhagen Denmark. 

*WHO. 1998. WHO toxic equivalent factors (TEFs) for dioxin-like compounds for humans and wildlife. 

Environ Health Perspect 106(12):775-792.. 

*WHO. 1998a. WHO revises the tolerable daily intake (TDI) for dioxins. World Health Organization. 

18th Symposium on Halogenated Environmental Organic Pollutants. Stockholm, Sweden. August 17-21, 

1998. Organohalogens Compounds. Volume 38, 295-298. 

*WHO. 1998b. World Health Organization. WHO Information. 1998 Press Releases. Press Release 

WHO/45. June 3, 1998. http://www.who.int/inf-pr-1998/en/pr98-45.html. 

*Wiberg K, Lundstr÷m K, Glas B, et al. 1989. PCDDs and PCDFs in consumers' paper products. 

Chemosphere 19:735-740. 

*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. 

*Wiklund K, Dich J, Holm LE. 1987. Risk of malignant lymphoma in Swedish pesticide appliers. Br J 

Cancer 56:505-508. 
CDDs 669 

8. REFERENCES 

*Wiklund K, Holm LE. 1986. Soft tissue sarcoma risk in Swedish agriculture and forestry workers. J Natl 

Cancer Inst 76:229-234. 

*Wikstrom E, Lofvenius G, Rappe C, Marklund S. 1996. Influence of level and form of chlorine on the 

formation of chlorinated dioxins, dibenzofurans, and benzenes during combustion of an artificial fuel in a 

laboratory reactor. Environmental Science & Technology 30(5):1637-1644. 

*Wild SR, Harrad SJ, Jones KC. 1993. The influence of sewage sludge applications to agricultural land on 

human exposure to polychlorinated dibenzo-p-dioxins (PCDDs) and -furans (PCDFs). Environ Pollution 

357-369. 

*Wild SR, Jones KC. 1992. Organic chemicals entering agricultural soils in sewage sludges: Screening for 

their potential to transfer to crop plants and livestock. Sci Total Environ 119:85-119. 

*Wilker C, Johnson L, Safe S. 1996. Effects of developmental exposure to indole-3-carbinol or 

2,3,7,8-tetrachlorodibenzo-p-dioxin on reproductive potential of male rat offspring. Toxicol Appl Pharmacol 

141:68-75. 

*Williams DT, Cunningham HM, Blanchfield BJ. 1972. Distribution and excretion studies of 

octachlorodibenzo-p-dioxin in the rat. Bull Environ Contam Toxicol 7:57-62. 

Williams DT, LeBel GL, Benoit FM. 1992. Polychlorodibenzodioxins and polychlorodibenzofurans in 

dioxazine dyes and pigments. Chemosphere 24(2):169-180. 

Williams PL. 1983. Commercial PCP: Toxic impurities. Occup Health Saf 52:14-16. 

Wilson JD. 1986. Airborne dioxins and dibenzofurans: Sources and fates. Comments. Environ Sci 

Technol 20:1185-1186. 

*Winters D,Cleverly D, meier K, et al. 1996. A statistical survey of dioxin-like compounds in United States 

beef: a progress report. Chemosphere 32(3):469-478. 

Wipf HK, Schmid J. 1983. Seveso--an environmental assessment. In: Tucker RE, ed. Human and 

environmental risk of chlorinated dioxins and related compounds. New York, London: Plenum Publishing 

Corp, 255-274. 

*Wolfe RJ, Walker RJ. 1987. Subsistence economies in Alaska: Productivity, geography, and development 

impacts. Arctic Anthropology 24(2):56-81. 

*Wolfe WH, Lathrop GD, Albanese RA, et al. 1985. An epidemiologic investigation of health effects in Air 

Force personnel following exposure to herbicides and associated dioxins. Chemosphere 14:707-716. 

*Wolfe WH, Michalek JE, Miner JC, et al. 1994. Determinants of TCDD half-life in veterans of Operation 

Ranch Hand. J Toxicol Environ Health 41:481-488. 

*Wolfe WH, Michalek JE, Miner JC, et al. 1995. Paternal serum dioxin and reproductive outcomes among 

veterans of Operation Ranch Hand. Epidemiol 6:17-22. 
CDDs 670 

8. REFERENCES 

*Wood SC, Jeong HG, Morris DL, et al. 1993. Direct effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) on human tonsillar lymphocytes. Toxicology 81(2):131-43. 

Wood SC, Karras JG, Holsapple MP. 1992. Integration of the human lymphocyte into immunotoxicological 

investigations. Fundam Appl Toxicol 18(3):450-9. 

Woods JS, Polissar L, Severson RK, et al. 1989. Soft tissue sarcoma and non-Hodgkin's lymphoma in 

relation to phenoxyherbicide and chlorinated phenol exposure in western Washington. J Natl Cancer Inst 

78:889-910. 

*Worobec SM, DiBeneditto JP. 1984. Perspectives on occupational dermatoses. In: Drill VA, Lazar P, eds. 

Cutaneous toxicity. New York, NY: Raven Press, 253-268. 

*Wroblewski VJ, Olson JR. 1985. Hepatic metabolism of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in 

the rat and guinea pig. Toxicol Appl Pharmacol 81:231-240. 

*Wu L, Whitlock J P JR. 1992. Mechanism of dioxin action: Ah receptor-mediated increase in promotor 

accessibility in vivo. Proc Natl Acad Sci USA 89(11):4811-4815. 

*Wu L, Whitlock JP Jr. 1993. Mechanism of dioxin action: Receptor-enhancer interaction in intact cells. 

Nucleic Acids Res 21:119-125. 

Wurrey CJ, Bourne S, Kloepfer RD. 1986. Application of gas chromatography/ matrix isolation/Fourier 

transformer infrared spectrometry to dioxin determinations. Anal Chem 58:482-483. 

*Wuthe J, Link B, Walther J, et al. 1993. Dioxin and furan (PCDD/PCDF) levels in human blood from 

persons living in a contaminated area. Chemosphere 27(103):287-293. 

Yamagishi T, Miyazaki T, Akiyama K, et al. 1981. Polychlorinated dibenzo-p-dioxins and dibenzofurans in 

commercial diphenyl ether herbicides and in fresh water fish collected from the application area. 

Chemosphere 10:1137-1144. 

*Yamamoto T, Fukushima M. 1993. Modeling study on contribution of combustion source complex to 

PCDD/PCDF levels in urban air. Chemosphere 27(1-3):295-300. 

*Yamashita N, Tanabe S, Ludwig JP, et al. 1992. Embryonic abnormalities and organochlorine 

contamination in double-crested cormorants (phalacrocorax auritus) and Caspian Terns (hydroprogne caspia) 

from the upper Great Lakes in 1988. Environ Pollution 19:163-173. 

*Yanders AF, Orazio CE, Puri RK, et al. 1989. On translocation of 2,3,7,8-tetrachlorodibenzo-p-dioxin: 

Time dependent analysis at the Times Beach experimental site. Chemosphere 19:429-432. 

*Yang KH, Croft WA, Peterson RE. 1977. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on plasma 

disappearance and biliary excretion of foreign compounds in rats. Toxicol Appl Pharmacol 40:485-496. 

*Yang XP, Masten SA, Clark GC, et al. 1997. Markers of exposure and susceptibility to dioxin in german 

chemical plant workers (meeting abstract). Proceedings of the American Association for Cancer Research 

38:352-353. 
CDDs 671 

8. REFERENCES 

*Yang YG, Lebrec H, Burleson GR. 1994. Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on 

pulmonary influenza virus titer and natural killer (NK) activity in rats. Fund Appl Toxicol 23:125-131. 

Yasuda M, Nariat M. 1989. Modification of teratogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in mice 

with various factors [Abstract]. Teratology (Japanese Teratology Society Abstracts) 40:678. 

Yasuhara A, Morita M. 1989. Formation of chlorodibenzofurans by thermal decomposition of 

vinylidene-chloride polymer. Chemosphere 18:1737-1740. 

*Yockim RS, Isensee Ar, Jones GE. 1978. Distribution and toxicity of TCDD and 2,4,5,-T in an aquatic 

model ecosystem. Chemosphere 3:215-220. 

Young AL. 1984a. Analysis of dioxins and furans in human adipose tissue. Public Health Risk Dioxins 

Proc Symp 63-75. 

Young AL. 1984b. Determination and measurement of human exposure to the dibenzo-p-dioxins. Bull 

Environ Contam Toxicol 33:702-709. 

Young AL, Cockerham LG, Thalken CE. 1987. A long-term study of ecosystem contamination with 

2,3,7,8-tetrachlorodibenzo-p-dioxin. Chemosphere 16:1791-1815. 

Young AL, Kang HK. 1985. Status and results of federal epidemiologic studies of populations exposed to 

TCDD. Chemosphere 14:779-790. 

Young AL, Kang HK, Shepard BM. 1983. Chlorinated dioxins as herbicide contaminants. Environ Sci 

Technol 17:530A-540A. 

Young AL, Thalken CE, Harrison DD. 1981. Persistence, bioaccumulation, and toxicology of TCDD in an 

ecosystem treated with massive quantities of 2,4,5-T herbicide. Proc West Soc Weed Sci 34:70-77. 

*Yousefi Z, Walters RW. 1987. Use of soil columns to measure sorption of dioxins to soils. Toxic Hazard 

Wastes 19:181-193. 

Zabik ME, Zabik MJ. 1980. Dioxin levels in raw and cooked liver, loin steaks, round and patties from beef 

fed technical grade pentachlorophenol. Bull Environ Contam Toxicol 24:344-349. 

*Zacharewski T, Harris M, Biegel L, et al. 1992. 6-Methyl-1,3,8-trichlorodibenzofuran (MCDF) as an 

antiestrogen in human and rodent cancer cell lines: Evidence for the role of the Ah receptor. Toxicol Appl 

Pharmacol 113:311-318. 

*Zacharewski T, Harris M, Safe S. 1991. Evidence for the mechanism of action of the 

2,3,7,8-tetrachlorodibenzo-p-dioxin-mediated decrease of nuclear estrogen receptor levels in wild-type and 

mutant Hepa 1c1c7 cells. Biochem Pharmacol 41:1931-1939. 

*Zacharewski T, Safe L, Safe S, et al. 1989. Comparative analysis of polychlorinated dibenzo-p-dioxin and 

dibenzofuran congeners in Great Lakes fish extracts by gas chromatography and in vitro enzyme induction 

activities. Environ Sci Technol 23:730-735. 
CDDs 672 

8. REFERENCES 

*Zack JA, Suskind RR. 1980. The mortality experience of workers exposed to tetrachlorodibenzodioxin in 

a trichlorophenol process accident. J Occup Med 22:11-14. 

Zeiger E. 1983. Memorandum from Dr. Zeiger to Dr. E.E. McConnell on the results of test performed for 

the Environmental Mutagenesis Development Program. NTP, NIEHS. 

*Zhong Y, Overcash MR, McPeters AL. 1993. Near sunlight zone model for photodegradation of TCDD in 

soils containing organic solvents. Chemosphere 26(7):1263-1272. 

*Ziegler EE, Edwards BB, Jensen RL et al. 1978. Absorption and retention of lead by infants. Pediatr Res 

12:29-34. 

*Zimmering S, Mason JM, Valencia R, et al. 1985. Chemical mutagenesis testing in Drosophila: II. 

Results of 20 coded compounds tested for the National Toxicology Program. Environ Mutagen 7:87-100. 

*Zinkl JG, Vos JG, Moore JA, et al. 1973. Hematologic and chemical clinical chemistry of 

2,3,7,8-tetrachlorodibenzo-p-dioxin in laboratory animals. Environ Health Perspect 5:111-118. 

Zitko V. 1989a. A graphical presentation of the composition and properties of chlorinated dibenzodioxins 

and dibenzofurans. Sci Total Environ 83:191-194. 

Zitko V. 1989b. Composition of chlorinated dibenzodioxins and dibenzofurans in various samples. Sci 

Total Environ 80:127-137. 

Zitko V, Hutzinger O, Choi PMK. 1972. Contamination of the Bay of Fundy - Gulf of Maine area with 

polychlorinated biphenyls, polychlorinated terphenyls, chlorinated dibenzodioxins, and dibenzofurans. 

Environ Health Perspect 1:47-50. 

*Zober A, Messerer P, Huber P. 1990. Thirty-four-year mortality follow-up of BASF employees exposed to 

2,3,7,8-TCDD after the 1953 accident. Int Arch Occup Environ Health 62:139-157. 

*Zober A, Ott M, Fleig I, et al. 1993. Cytogenic studies in lymphocyte of workers exposed to 

2,3,7,8-TCDD. Int Arch Occup Environ Health 65:157-161. 

*Zober A, Ott MG, Messerer P. 1994. Morbidity follow up study of BASF employees exposed to 

2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) after a 1953 chemical reactor incident. Occup Environ Med 

51:479-486. 

*Zober A, Papke O. 1993. Concentrations of PCDDs and PCDFs in human tissue 36 years after accidental 

dioxin exposure. Chemosphere 27:413-418. 

Zook DR, Rappe C. 1994. Environmental sources, distribution, and fate of polychlorinated dibenzodioxins, 

dibenzofurans, and related organochlorines. In: Schecter A, ed. Dioxins and health. New York: Plenum 

Press, 80-113. 
CDDs 673 

9. 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)—is 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—is 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—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. 

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. 
CDDs 674 

9. 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. 
CDDs 675 

9. GLOSSARY 

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. 

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. 

Immunological Effects—are functional changes in the immune response. 

Immunologic Toxicity—The occurrence of adverse effects on the immune system that may result from 

exposure to environmental agents such as chemicals. 

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. 
CDDs 676 

9. GLOSSARY 

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. 

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—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—is 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. 
CDDs 677 

9. GLOSSARY 

Pharmacokinetic Model—is 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 physiologicallybased model compartments represent real anatomic regions of the body. 

Physiologically Based Pharmacodynamic (PBPD) Model—is a type of physiologically-based doseresponse model which quantitatively describes the relationship between target tissue dose and toxic end 

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—is 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 (NOAELfrom 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. 
CDDs 678 

9. GLOSSARY 

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. 
CDDs A-1 

APPENDIX A 

ATSDR MINIMAL RISK LEVELS 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. MRLs are intended only to serve as a 
CDDs A-2 

APPENDIX 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. 
CDDs A-3 

APPENDIX A 

MRL WORKSHEET 

Chemical Name: 2,3,7,8-Tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) 

CAS Number: 1746-01-6 

Date: December 10, 1998 

Profile Status: Final Draft 

Route: [ ] Inhalation [X] Oral 

Duration: [X ] Acute [ ] Intermediate [] Chronic 

Key to Figure: 78m 

Species: Mice 

Minimal Risk Level: 0.0002 (2×10-4) [X] μg/kg/day [ ] ppm 

Reference: Burleson et al. 1996 

Experimental design: (human study details or strain, number of animals per exposure/control groups, sex, 

dose administration details): 

Groups of 20 female B6C3F1 mice were administered a single gavage dose of 0, 0.001, 0.005, 0.01, 0.05, 

or 0.1 μg/kg 2,3,7,8-TCDD in corn oil. Seven days after 2,3,7,8-TCDD exposure, the mice were infected 

intranasally with influenza A/Hong Kong/8/68 (H3N2) virus diluted at 10-48, 10-50, 10-52, or 10-54. In a 

separate experiment, groups of 18 female mice received a single gavage dose of 0, 0.001, 0.01, or 0.1 

μg/kg 2,3,7,8-TCDD and were infected 7 days later with influenza A virus at a dose not known to cause 

mortality (10-54 and 10-58) or were sham-infected. Body weight, thymus weight, and wet lung weights were 

measured 3, 9, or 12 days postinfection. Pulmonary virus titers were determined in groups of 72 mice 

exposed to 0, 0.001, 0.01, or 0.01 μg/kg 2,3,7,8-TCDD and infected with influenza A virus seven days 

later. For the virus titer study, groups mice were killed 2 hours, 1, 4, 6, 7, 8, 9, 10, and 11 days postinfection. 

Effects noted in study and corresponding doses: 

Statistically significant increases in mortality were observed in the influenza A infected mice exposed to 

0.01, 0.05, or 0.1 μg/kg 2,3,7,8-TCDD. However, no between group differences in mortality were 

observed at these 2,3,7,8-TCDD dosages. Mortality in mice receiving 0.001 or 0.005 μg/kg did not 

significantly differ from the mortality in the control group. Exposure to 2,3,7,8-TCDD did not enhance the 

increase in relative lung weight normally seen in mice infected with influenza A virus. As compared to 

controls, no significant alterations in thymus weights were observed in 2,3,7,8-TCDD-exposed mice 

sham-infected or those infected with influenza A virus. 2,3,7,8-TCDD exposure did not result in a 

significant increase in viral titers in the lung, as compared to titers from the control group. The authors 

noted that the lack of dose-response in mortality and the lack of effect on the relative lung weight, thymus 

weight, and viral titers suggest that 2,3,7,8-TCDD may be exerting an effect via an indirect mechanism 

such as through an effect on cytokines. 

Dose and end point used for MRL derivation: Impaired resistance to influenza A virus infection, as 

evidence by the significant increase in mortality, was observed in female B6C3F1 mice administered a 

single gavage dose of $0.01 μg/kg. No significant effects were observed at lower doses. Thus, 0.005 and 

0.01 μg/kg are the NOAEL and LOAEL, respectively, for impaired resistance. 
 

 

CDDs A-4 

APPENDIX A 

[X] NOAEL [ ] LOAEL 

Uncertainty Factors used in MRL derivation: 

[ ]	 10 for use of a LOAEL 

[X] 3 for extrapolation from animals to humans- A comparison of species sensitivity suggests that even 

though there are wide ranges of sensitivity for some 2,3,7,8-TCDD-induced health effects, for most 

health effects, the LOAELs for the majority of animal species cluster within an order of magnitude. 

Based on the weight of evidence of animal species comparisons and human and animal mechanistic 

data, it is reasonable to assume that human sensitivity would fall within the range of animal 

sensitivity. 

[X] 10 for human variability 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? 

No. A modifying factor of 0.7 was applied to adjust for the difference in higher bioavailability of 

2,3,7,8-TCDD from gavage with an oil vehicle than from food. Support for this modifying factor comes 

from toxicokinetic studies in Sprague Dawley rats. In rats fed 0.35 or 1 μg/kg/day 2,3,7,8-TCDD in the 

diet for 42 days, approximately 60% of the administered dose was absorbed (Fries and Marrow 1975). In 

contrast, 70-84% of a single or repeated gavage dose of 0.01-50 μg/kg 2,3,7,8-TCDD in corn oil was 

absorbed in rats (Piper et al. 1973; Rose et al. 1976). Thus, the ratio of 2,3,7,8-TCDD absorption from the 

diet to gavage with an oil vehicle is 0.71-0.85. 

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: 

2,3,7,8-TCDD is a known immunosuppressant in animals in acute-, intermediate, and chronic-duration 

studies (Kerkvliet 1995). Suppression of the antibody response to sheep erythrocytes was observed in 

B6C3F1 mice administered 14 daily doses of 0.1 μg 2,3,7,8-TCDD/kg/day (Holsapple et al. 1986), and a 

significant increase in mortality was observed in B6C3F1 mice administered 1.0 μg/kg/day 2,3,7,8-TCDD 

for 14 days and challenged with Streptococcus pneumoniae (White et al. 1986). Decreased survival after 

viral infection was also reported in female B6C3F1 mice after a single intraperitoneal dose of 0.1 μg 

2,3,7,8-TCDD/kg (House et al. 1990). A significant suppression of complement hemolytic activity was 

observed in mice administered 0.01 μg/kg/day via gavage for 14 days (White et al. 1986). Furthermore, 

2,3,7,8-TCDD alters the immune system of offspring when exposed through lactation and/or in utero. For 

example, a dose-related decrease in relative thymus weights were seen in offspring of rats dosed at levels 

of 0.005-0.35 μg 2,3,7,8-TCDD/kg on day 16 of pregnancy (Madsen and Larsen 1979). 

Agency Contact (Chemical Manager): Hana Pohl 
CDDs A-5 

APPENDIX A 

MRL WORKSHEET 

Chemical Name: 2,3,7,8-Tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) 

CAS Number: 1746-01-6 

Date: December 10, 1998 

Profile Status: Final Draft 

Route: [ ] Inhalation [X] Oral 

Duration: [ ] Acute [X ] Intermediate [] Chronic 

Key to Figure: 187g 

Species: Guinea pig 

Minimal Risk Level: 0.00002 (2×10-5) [X] μg/kg/day [ ] ppm 

Reference: DeCaprio et al. 1986 

Experimental design: (human study details or strain, number of animals per exposure/control groups, sex, 

dose administration details): 

Groups of weanling Hartley guinea pigs (10/sex) were administered a diet containing 2, 10, 76 or 430 ppt 

for 90 days. These diets provided an average of 0.0001, 0.0007, 0.005, or 0.028 μg 2,3,7,8-TCDD/kg/day. 

The average doses were estimated by the investigators. A group of control guinea pigs was fed a diet 

without added 2,3,7,8-TCDD. Body weights and food consumption were monitored throughout the 

experiment. At the end of the dosing period the animals were sacrificed and clinical chemistries, 

hematology, organ weights and histopathology examinations were performed. The recovery following 

treatment was studied in groups of 10 guinea pigs fed a diet containing 430 ppt 2,3,7,8-TCDD for 11, 21, 

or 35 days and allowed to recover for 79, 69, or 55 additional days, respectively. 

Effects noted in study and corresponding doses: 

The highest dietary level of 2,3,7,8-TCDD caused net body weight loss and mortality. Four males and 

four females died and additional animals had to be sacrificed due to poor health. Food consumption was 

significantly reduced in the highest dose group only. Body weight gain in the 0.0007 and 0.005 μg/kg/day 

male groups was reduced by 9% and 20%, respectively. In the corresponding female groups, body weight 

gain was reduced by 6% and 15%. Gross lesions were observed only in the highest dose group and 

included thymic atrophy, depletion of body fat, and liver enlargement. Significant changes in organ 

weights included a decrease in absolute kidney weight and in absolute and relative thymus weight in males 

dosed with 0.005 μg/kg/day, increase in relative liver weight in males and females at the 0.005 μg/kg/day 

level, and increase in relative brain weight in males at this same dose level. Organ weights from high dose 

animals were not monitored. Administration of 2,3,7,8-TCDD did not cause any significant hematological 

effect (blood was not collected from the highest dose group). In the 0.005 μg/kg/day groups, serum ALT 

was significantly reduced in females whereas triglycerides were elevated in males. No other significant 

changes in clinical chemistries were observed. Treatment-related histological alterations were observed 

only in the two higher dose groups and consisted of hepatocellular cytoplasmic inclusion bodies and 

atrophy of the thymic cortex. In the recovery study there was 10% mortality in the groups treated for 11 

and 21 days and 70% mortality in the group treated for 35 days. Surviving animals in all groups exhibited 

significantly reduced body weight gain. 
 

 

CDDs A-6 

APPENDIX A 

Dose and end point used for MRL derivation: The dose of 0.0007 μg/kg/day represents a NOAEL for 

decreased thymus weight, whereas the 0.005 μg/kg/day is a LOAEL. 

[X ] NOAEL [ ] LOAEL 

Uncertainty Factors used in MRL derivation: 

[ ] 10 for use of a LOAEL 

[X] 3 for extrapolation from animals to humans-A comparison of species sensitivity suggests that even 

though there are wide ranges of sensitivity for some 2,3,7,8-TCDD-induced health effects, for most 

health effects, the LOAELs for the majority of animal species cluster within an order of magnitude. 

Based on the weight of evidence of animal species comparisons and human and animal mechanistic 

data, it is reasonable to assume that human sensitivity would fall within the range of animal 

sensitivity. 

[X] 10 for human variability 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? 

No. The doses were estimated by the investigators. 

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: 

2,3,7,8-TCDD is a known immunosuppressant in animals in acute-, intermediate, and chronic-duration 

studies (Kerkvliet 1995). Reduction of thymus weight was also observed in intermediate-duration oral 

studies in rats (Van Birgelen et al. 1995; Viluksela et al. 1994). Another sensitive species for 

immunological effects of 2,3,7,8-TCDD is the marmoset monkey in which alterations in lymphocyte 

subsets have been reported after subcutaneous application of an average 0.0015 μg 2,3,7,8-TCDD/kg/day 

for 26 weeks (Neubert et al. 1992). Furthermore, 2,3,7,8-TCDD alters the immune system of offspring 

when exposed through lactation and/or in utero. For example, a dose-related decrease in relative thymus 

weights were seen in offspring of rats dosed at levels of 0.005-0.35 μg 2,3,7,8-TCDD/kg on day 16 of 

pregnancy (Madsen and Larsen 1979). 

Agency Contact (Chemical Manager): Hana Pohl 
CDDs A-7 

APPENDIX A 

MRL WORKSHEET 

Chemical Name: 2,3,7,8-Tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) 

CAS Number: 1746-01-6 

Date: December 10, 1998 

Profile Status: Final Draft 

Route: [ ] Inhalation [X] Oral 

Duration: [ ] Acute [ ] Intermediate [X] Chronic 

Key to Figure: 226k 

Species: Monkey 

Minimal Risk Level: 0.000001 (1×10-6) [X] μg/kg/day [ ] ppm 

Reference: Schantz et al. 1992 

Experimental design: (human study details or strain, number of animals per exposure/control groups, sex, 

dose administration details): 

Groups of 8 female rhesus monkeys were fed a diet containing 0, 5, or 25 ppt 2,3,7,8-TCDD for a total of 

16.2 ± 0.4 months. After 7 months of exposure, the monkeys were mated with unexposed males. (Only 1 

monkey in the 25 ppt group delivered a viable offspring; this offspring was not studied behaviorally). The 

monkeys were fed the 2,3,7,8-TCDD diet throughout the mating period, gestation, and lactation. The 

authors estimated that the total 2,3,7,8-TCDD intake over the course of the study was 59.6 ± 5.0 ng/kg for 

the 5 ppt group. The offspring were weaned at 4 months and individually housed. Mesenteric fat samples 

were collected from the offspring at age 5 months; the average 2,3,7,8-TCDD levels in the fat samples was 

377 ±141 ppt (range of 290-950) for the 5 ppt group and below the detection limit of 2-200 ppt for the 

controls. When the offspring were 8.6 months of age, they were placed in peer groups of 4 monkeys and 

allowed to play for 1.5 hours without interference. The peer groups consisted of two 2,3,7,8-TCDDexposed monkeys and two control monkeys. Behavioral patterns (social interactions and other behaviors 

such as vocalization, locomotion, self-directed behavior and environmental exploration) were monitored 

4 days/week for 9 weeks. 

Effects noted in study and corresponding doses: 

No overt signs of toxicity were observed in the mothers or offspring, and birth weights and growth were 

not adversely affected by 2,3,7,8-TCDD exposure. Significant alterations were observed in play behavior, 

displacement, and self-directed behavior in the 2,3,7,8-TCDD -exposed offspring. 2,3,7,8-TCDD-exposed 

monkeys tended to initiate more rough-tumble play bouts and retreated less from play bouts than controls, 

were less often displaced from preferred positions in the playroom than the controls, and engaged in more 

self-directed behavior than controls. No other significant alterations in behavior were observed. 

Dose and end point used for MRL derivation: 

Although the mothers were exposed to 5 or 25 ppt 2,3,7,8-TCDD, only the offspring from the 5 ppt group 

underwent behavioral testing. The 5 ppt dietary concentration is equivalent to a daily dose of 1.2 x 10-4 

μg/kg/day. This dose is a LOAEL for altered social behavior. 

[ ] NOAEL [X] LOAEL 
 

 

CDDs A-8 

APPENDIX A 

Uncertainty Factors used in MRL derivation: 

[X] 3 for use of a minimal LOAEL 

[X] 3 for extrapolation from animals to humans A comparison of species sensitivity suggests that even 

though there are wide ranges of sensitivity for some 2,3,7,8-TCDD-induced health effects, for most 

health effects, the LOAELs for the majority of animal species cluster within an order of magnitude. 

Based on the weight of evidence of animal species comparisons and human and animal mechanistic 

data, it is reasonable to assume that human sensitivity would fall within the range of animal 

sensitivity. 

[X] 10 for human variability. 

Was a conversion factor used from ppm in food or water to a mg/body weight dose? 

Monkeys were exposed to a dietary concentration of 5 ppt 2,3,7,8-TCDD; the authors estimated that the 

total maternal intake during the 16.2 months of exposure (492 days) was 59.6 ng/kg. 

Daily dose = (59.6 ng/kg) / (492 days) = 0.12 ng/kg/day (1.2 x 10-4 μg/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: 

A behavioral teratology test battery was performed in monkey infants exposed to 2,3,7,8-TCDD during 

gestation and lactation; the results of this test battery was published in a series of papers (Bowman et al. 

1989a; Schantz and Bowman 1989; Schantz et al. 1992). No significant alterations in reflex development, 

visual exploration, locomotor activity, or fine motor control were found (Bowman et al. 1989a). In tests of 

cognitive function, object learning was significantly impaired, but no effect on spatial learning was 

observed (Schantz and Bowman 1989). When the monkey infants were placed in social groups, altered 

social behavior was observed (Bowman et al. 1989a; Schantz et al. 1992). Additional data on the 

neurodevelopmental toxicity of 2,3,7,8-TCDD are limited to a study in which prenatal exposure to 

2,3,7,8-TCDD resulted in masculinized behavior in female rats (Schantz et al. 1991). No chronic duration 

animal neurotoxicity studies were located, decreased motor activity was reported in rats acutely exposed to 

2 (Giavini et al. 1983) or 5 (Seefeld et al. 1984a) μg/kg/day. The Schantz and Bowman studies are the 

only available chronic developmental toxicity studies. Acute and intermediate duration studies provide 

evidence that 2,3,7,8-TCDD is a potent developmental toxicant. Other sensitive developmental effects that 

have been observed included cleft palate [lowest LOAEL- 0.1 μg/kg/day (Giavini et al. 1982)], 

hydronephrosis [lowest LOAEL- 1 μg/kg (Moore et al. 1973)], immunosuppression [lowest LOAEL0.005 μg/kg (Madsen and Larsen 1979)], impaired development of the reproductive system [lowest 

LOAEL- 0.064 μg/kg (Mably et al. 1992a, 1992b, 1992c)], and increased newborn mortality [lowest 

LOAEL-0.7 μg/kg (Bjerke et al. 1994a)]; NOAELs were not identified for these effects in the most 

sensitive species or strain. 
CDDs A-9 

APPENDIX A 

Some human studies have reported effects on the central and peripheral nervous systems shortly after 

exposure to high levels of 2,3,7,8-TCDD (Filippini et al. 1981; Moses et al. 1984; Pazderova-Vejlupkova 

et al. 1981; Pocchiari et al. 1979; Suskind 1977). However, follow-up studies did not find neurological 

effects years after exposure termination (Barbieri et al. 1988), suggesting that the effects may be transient. 

No human studies examined the effect of 2,3,7,8-TCDD on the developing neurologic system. 

It should be also noted that 10 years after termination of 2,3,7,8-TCDD exposure in the Schantz et al. 

(1992) study, Rier et al. (1993) reported a dose-related increase in the incidence and severity of 

endometriosis in these same rhesus monkeys. Rier et al. (1993) identified a less serious LOAEL of 5 ppt 

(0.00012 μg/kg/day) for moderate endometriosis. However, monkeys appear to be more susceptible to 

endometriosis, based on a background incidence of endometriosis in monkeys of 30% (Rier et al. 1993) 

compared to a background incidence of 10% in humans (Wheeler et al. 1992). Thus, derivation of a 

chronic oral MRL based on endometriosis would necessitate using an uncertainty factor of less than 1 (or 

at most, 1) to account for the increased sensitivity of monkeys to endometriosis as compared to humans. If 

the Rier et al. (1993) study was used to calculate an oral MRL, the LOAEL of 0.00012 μg/kg/day would 

be divided by an uncertainty factor of 100 (10 to extrapolate from a LOAEL, 10 for human variability and 

1 for interspecies differences). This would result in a computed MRL essentially the same as the chronic 

oral MRL of 1 pg/kg/day based on developmental toxicity as described in the preceding paragraph. 

Moreover, (1) the clinical history for these rhesus monkeys during the 10 year period between the Schantz 

et al. (1992) study and examination by Rier et al. (1993) is unknown (not reported); (2) Boyd et al. (1995) 

did not find an association between exposure to CDDs, CDFs, or PCBs and endometriosis in a clinical 

study in women; and (3) the EPA (1997) concluded that “the evidence for supporting the hypothesis that 

CDDs and PCBs are causally related to human endometriosis via an endocrine-disruption mechanism is 

very weak.” So, even though there is information to indicate that endometriosis may also be a sensitive 

toxicological end point for 2,3,7,8-TCDD exposure, the developmental end point (altered social behavior) 

reported in the Schantz et al. (1992) study was determined to be the most appropriate end point for 

derivation of an MRL for chronic oral 2,3,7,8-TCDD exposure. 

Agency Contact (Chemical Manager): Hana Pohl 
 B-1 CDDs 

APPENDIX B 

Update to the ATSDR Policy Guideline for 

Dioxins and Dioxin-Like Compounds in Residential Soil 

Purpose 

The Agency for Toxic Substances and Disease Registry (ATSDR) is updating its Policy 

Guideline for Dioxins and Dioxin-Like Compounds in Residential Soil. 

The objective of this update is to ensure that ATSDR health assessors evaluate dioxin levels that 

exceed the ATSDR established screening level of 0.05 ppb as described in the ATSDR Public 

Health Assessment Guidance Manual (PHAGM) (ATSDR 2005). The 0.05 ppb value should be 

used as the comparison value when following the PHAGM. The comparison value is not a 

threshold for toxicity and should not be used to predict adverse health effects (ATSDR 2005). 

This update replaces Appendix B in the Toxicological Profile for Chlorinated Dibenzo-p-dioxins 

(CDDs) (December, 1998). It does not reflect a change in ATSDR’s scientific assessment on 

dioxin toxicity or the ATSDR Minimal Risk Level (MRL). This update does not impact the EPA 

guidance which continues to identify 1 ppb as the preliminary remediation goal for residential 

exposure scenarios. (EPA 1998). 

History of the Dioxin Policy Guideline 

In 1998, ATSDR adopted a Policy Guideline for Dioxin and Dioxin-like Compounds (ATSDR, 

1998). The policy was developed to guide health assessors in evaluating the public health 

implications of dioxin and dioxin-like compounds (including 2,3,7,8-tetrachlorodibenzo-p-dioxin 

and other structurally related halogenated aromatic hydrocarbons) in residential soils near or on 

hazardous waste sites. The 1998 guideline established three levels as criteria for comparing 

dioxin levels in residential soil: 

• a screening level, 

• an evaluation level, and 

• an action level. 

The 1998 guideline also recommended specific considerations for public health actions within 

each of these levels. 

Since the release of the Policy Guideline in 1998, ATSDR issued the PHAGM. By issuing this 

update to the guideline, ATSDR is ensuring that health assessors will use the screening level as 

the appropriate comparison value for following the PHAGM, rather than the “action level” 

described in the earlier version of this policy guidance. This does not reflect a change in dioxin 

science; it is simply a reiteration to ensure that the appropriate value is used as a starting point 

when following the procedures described in the PHAGM. 

November 2008 
 

If health assessors follow the PHAGM, the evaluation and action levels values, as set in 1998, 

are no longer necessary. 

Changes Being Made to the ATSDR Policy Guideline for Dioxins and Dioxin-Like 

Compounds in Residential Soil 

The specific changes to the policy guideline, the reason for those changes, and the expected 

impact of those changes are summarized in the following table: 

Change Reason for Change Impact of Change 

Elimination of Confusion about interpretation of the This change brings the guidelines upthe “evaluation evaluation and action levels was a to-date with ATSDR’s PHAGM 

level” and the barrier to a more consistent evaluation which uses only screening levels

“action level” of exposure to dioxin in residential soils. 

The public health actions described in 

the 1998 policy guideline remain 

options that may be applied as 

appropriate rather than being triggered 

by a prescribed soil concentration. 

The minimal risk level (MRL) for 

dioxin exposure described in the 1998

Toxicological Profile remains the 

same. 

Ensure PHAGM was not referenced in the Consistency with 2005 PHAGM will

consistency previous policy. ensure more comprehensive 

with ATSDR evaluation, for instance assessing both

PHAGM direct and indirect exposure pathways

should result in a more comprehensive 

evaluation of exposure conditions at 

sites with dioxin contamination. 

Summary 

This policy update replaces Appendix B in the Toxicological Profile for Chlorinated Dibenzo-pdioxins (CDDs) (December, 1998). ATSDR will no longer refer to an Action Level for dioxin in 

these evaluations. The 0.05 ppb screening level is retained as an initial comparison value for 

health assessments. The update does not change the assessment of health hazards associated 

with dioxin exposure, as summarized in the 1998 ATSDR Toxicological Profile and in the 

derivation of the Minimal Risk Level (MRL). The policy update impacts site-specific health 

assessments evaluating exposure to dioxin directly from residential soils. The update ensures 

consistency in the methodology ATSDR uses for site-specific evaluations of health risks for all 

chemicals. 

B-2 

CDDs 

APPENDIX B 

November 2008 
EPA’s preliminary remediation goal for dioxin in soil has not changed and remains at 1 ppb. 

ATSDR does not establish clean-up goals or preliminary remediation goals, but ATSDR believes 

that health risks associated with levels of dioxins in soil below 1 ppb would be low under most 

scenarios where the primary exposure pathway is incidental ingestion through direct exposure to 

soil. In such instances, ATSDR public health recommendations may include community health 

education or limiting access to contaminated areas. Consistency with 2005 PHAGM also 

ensures that a comprehensive evaluation of dioxins from contaminated soils includes the 

consideration of scenarios where dioxins may enter the food chain pathway. 

References 

ATSDR. 1998. Toxicological profile for Chlorinated Dibenzo-p-Dioxins. US Department of Health and 

Human Services, Agency for Toxic Substances and Disease Registry, Atlanta, GA. 

ATSDR. 2005. Public health assessment guidance manual. US Department of Health and Human 

Services, Agency for Toxic Substances and Disease Registry, Atlanta, GA. 

EPA. 1998. Approach for Addressing Dioxin in Soil at CERCLA and RCRA Sites. Washington, DC: 

US Environmental Protection Agency. OSWER Directive 9200.4-26; April 13, 1998. 

B-3 

CDDs 

APPENDIX B 

November 2008 
CDDs	 C-1 

APPENDIX C 

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 

Tables and Figures for Levels of Significant Exposure (LSE) 

Tables (2-1, 2-2, and 2-3) and figures (2-1 and 2-2) 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). 

The legends presented below demonstrate the application of these tables and figures. Representative 

examples of LSE Table 2-1 and Figure 2-1 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 2-1 

(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 2-1, 2-2, and 2-3, respectively). LSE figures are limited to the inhalation (LSE Figure 2-1) 

and oral (LSE Figure 2-2) routes. Not all substances will have data on each route of exposure and 

will not therefore have all five of the tables and figures. 
CDDs	 C-2 

APPENDIX C 

(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 2-1). 

(5)	 Species The test species, whether animal or human, are identified in this column. Section 2.5, 

"Relevance to Public Health," covers the relevance of animal data to human toxicity and Section 2.3, 

"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. 

(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 8 of the profile. 
CDDs	 C-3 

APPENDIX C 

(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 2-1 

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, 18r NOAEL is the critical end point for which an intermediate inhalation 

exposure MRL is based. As you can see from the LSE figure key, the open-circle symbol indicates 

to a NOAEL for the test species-rat. 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. 

(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. 
SAMPLE

1 6 TABLE 2-1. Levels of Significant Exposure to [Chemical x] – Inhalation 

Key to figurea Species 

Exposure frequency/ duration System 

NOAEL (ppm) Less serious (ppm) 

LOAEL (effect) 

Serious (ppm) 

3 

4 

2 6 INTERMEDIATE EXPOSURE

5 6 7 8 9 

6 Systemic 9 9 9 9 9 

6 18 Rat 13 wk Resp 3b 10 (hyperplasia) 

5d/wk 6hr/d 

CHRONIC EXPOSURE 

11 

Cancer 938 Rat 18 mo 20 (CEL, multiple org

5d/wk7hr/d

39 Rat 89–104 wk 10 (CEL, lung tumors

5d/wk nasal tumors) 6hr/d 

40 Mouse 79–103 wk 10 (CEL, lung tumors

5d/wk hemangiosarcoma6hr/d 

a The number corresponds to entries in Figure 2-1. 

6 b Used to derive an intermediate inhalation Minimal Risk Level (MRL) of 5 x 10-3 ppm; dose adjusted for intermittent exposure 

12 uncertainty factor of 100 (10 for extrapolation from animal to humans, 10 for human variability). 

CEL = cancer effect level; d = days(s); hr = hour(s); LOAEL = lowest-observed-adverse-effect level; mo = month(s); NOAEL = effect level; Resp = respiratory; wk = week(s) 
CDDs	 C-6 

APPENDIX C 

Chapter 2 (Section 2.5) 

Relevance to Public Health 

The Relevance to Public Health section 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 section 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 section. 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 Data Needs section. 

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 

2.5, "Relevance to Public Health," contains basic information known about the substance. Other sections 

such as 2.8, "Interactions with Other Substances,” and 2.9, "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). 
CDDs C-7 

APPENDIX C 

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. 
CDDs D-1 

APPENDIX D 

ACRONYMS, ABBREVIATIONS, AND SYMBOLS 

2,4-D 2,4-dichlorophenoxyacetic acid 

2,4,5-T 2,4,5-trichlorophenoxyacetic acid 

2,4,5-TCP 2,4,5-trichlorophenol 

AAH arylhydrocarbon hydroxylase 

ACGIH American Conference of Governmental Industrial Hygienists 

ACOH acetanylide-4-hydroxylase 

ACTH adenocorticotropin 

ADI Acceptable Daily Intake 

ADME Absorption, Distribution, Metabolism, and Excretion 

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 

AST aspartate aminotransferase 

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 

CL ceiling limit value 

CLP Contract Laboratory Program 

cm centimeter 

CML chronic myeloid leukemia 

CNS central nervous system 

CPSC Consumer Products Safety Commission 

CTL cytotoxic T-lymphocyte 

CWA Clean Water Act 

d day 

Derm dermal 

DHEW Department of Health, Education, and Welfare 
 

CDDs D-2 

APPENDIX D 

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 

DOT/UN/ Department of Transportation/United Nations/

 NA/IMCO North America/International Maritime Dangerous Goods Code 

DTH delayed-type hypersensitivity 

DWEL Drinking Water Exposure Level 

ECD electron capture detection 

ECG/EKG electrocardiogram 

EEG electroencephalogram 

EEGL Emergency Exposure Guidance Level 

EGF epidermal growth factor 

EPA Environmental Protection Agency 

EROD ethoxyresorufin-O-deethylase 

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 

ft foot 

FR Federal Register 

FSH follicle-stimulating hormone 

g gram 

GC gas chromatography 

Gd gestational day 

gen generation 

GGT gamma-glutamyl transferase 

GLC gas liquid chromatography 

GPC gel permeation chromatography 

HDL high density lipoprotein 

HPLC high-performance liquid chromatography 

hr hour 

HRGC high resolution gas chromatography 

HSDB Hazardous Substance Data Bank 

IDLH Immediately Dangerous to Life and Health 

IARC International Agency for Research on Cancer 

ILO International Labor Organization 

in inch 

IRIS Integrated Risk Information System 

Kd adsorption ratio 

kg kilogram 

kkg metric ton 
CDDs D-3 

APPENDIX D 

Koc organic carbon partition coefficient Kow octanol-water partition coefficient 

L liter 

LC liquid chromatography 

LCLo lethal concentration, low 

LC50 lethal concentration, 50% kill 

LDL low density lipoprotein 

LDLo lethal dose, low 

LD50 lethal dose, 50% kill 

LH lteinizing 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 

MFO mxed-function oxidase 

mg milligram 

min minute 

mL milliliter 

mm millimeter 

mm Hg millimeters of mercury 

mmol millimole 

mo month 

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 

NCI National Cancer Institute 

NIEHS National Institute of Environmental Health Sciences 

NIOSH National Institute for Occupational Safety and Health 

NIOSHTIC NIOSH's Computerized Information Retrieval System 

NFPA National Fire Protection Association 

ng nanogram 

NK cells natural killer cells 

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 
CDDs D-4 

APPENDIX D 

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 

OSHA Occupational Safety and Health Administration 

OSW Office of Solid Waste, EPA 

OTS Office of Toxic Substances 

OW Office of Water 

OWRS Office of Water Regulations and Standards, EPA 

PAH Polycyclic Aromatic Hydrocarbon 

PBPD Physiologically Based Pharmacodynamic 

PBPK Physiologically Based Pharmacokinetic 

PCE polychromatic erythrocytes 

PEL permissible exposure limit 

PEPCK phosphoenolpyruvate carboxykinase 

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 

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 

sec second 

SIC Standard Industrial Classification 

SIM selected ion monitoring 

SMCL Secondary Maximum Contaminant Level 

SMR standard mortality ratio 
CDDs D-5 

APPENDIX D 

SNARL Suggested No Adverse Response Level 

SPEGL Short-Term Public Emergency Guidance Level 

SRBC sheep red blood cell 

STEL short-term exposure limit 

STORET Storage and Retrieval T3 triidothyronine T4 thyroxine 

TdO 2,3-dioxygenase 

TD50 toxic dose, 50% specific toxic effect 

TLV threshold limit value 

TOC Total Organic Compound 

TPQ Threshold Planning Quantity 

TRI Toxics Release Inventory 

TSCA Toxic Substances Control Act 

TSH thyroid-stimulating hormone 

TRI Toxics Release Inventory 

TTR transthyretin 

TWA time-weighted average 

UDPGT UDP-glucuronosyltransferase 

U.S. United States 

UF uncertainty factor 

VLDL very low density lipoprotein 

VOC Volatile Organic Compound 

yr year 

WHO World Health Organization 

wk week 

> greater than 

> greater than or equal to 

= equal to 

< less than 

< 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