United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/600/BP-92/0016
June 1994
External Review Draft
oEPA Health Assessment
Document for 2,3,7,8-
Tetrachlorodibenzo-p-
Dioxin (TCDD) and
Related Compounds
Volume II of
Review
Draft
(Do Not
Cite or
Quote)
Notice
This document is a preliminary draft. It has not been formally
released by EPA and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its
technical accuracy and policy implications.
U S. Environmental Protection Agency
Region 5, Library -L-12J) .
77 West Jackson Boulevard, l2Ui
Chicago, IL 60604-3590
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Dioxin
External Review Drafts
Page 9 Of 9
Next Stages in tfteJteaaseagfnent Process,
As described previously, public briefings will be held during the first week of the
public comment period to be followed by formal public hearings in December 1994.
After the dose of the public comment period, the Agency's Science Advisory Board
(SAB) will review the draft documents in public session (early 1995). Following SAB
review, the draft documents will be revised, comments and revisions will be
Incorporated, and final documents will be Issued.
Acting Assistant Administrator
for Research and Development
Billing Code: 6560-50-P
9
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09-08-1994 12:25 202 260 8061 T1S P.09
ecosystems from exposure to dloxlns. Research efforts are focused on the study of
organisms in aquatic food webs to identify the effects of dloxin exposure that are likely
to result in significant population impacts. A report titled, Interim Report on Data and
Methods for the Assessment of 2,3,7,8-Tetrachlorodibenzo-p-Dloxln (TCDD) Risks to
Aquatic Organisms and Associated Wildlife (EPA/600/R-93/055), was published in
April 1993. This report will serve as a background document for assessing diox'n-
related ecological risks. Ultimately, these data will support the development of aquatic
life criteria which will aid In the Implementation of the Clean Water Act.
As mentioned previously, completion of the health assessment and exposure
documents Involves three phases: Phase 1 Involved drafting state-oMhe-science
chapters and a dose-response model for the health assessment document, expanding
the exposure document to address dloxin related compounds, and conducting peer-
review workshops by panels of experts. This phase has been completed.
Phase 2, preparation of the risk characterization, began during the September
1992 workshops with discussions by the peer-review panels and formulation of points
to be carried forward into the risk characterization. Following the September 1993
workshop, this work was completed and was incorporated as Chapter 9 (Volume III)
of the draft hearth assessment document. This phase has been completed.
Phase 3 Is currently underway. It includes making External Review Drafts of
both the health assessment document and the exposure document available for public
review and comment.
8
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09-08-1994 12:24 202 260 6061 T!S P.OS
scientific experts from outside the Agency reviewed the draft documents and provided
valuable comments. It also should be noted that outside scientists have been heavily
involved throughout the developmental process of writing and reviewing these draft
documents. With this notice, the External Review Drafts of both draft documents are
being released for a 120-day public review and comment period.
Stage of the Scientific Reassessment of Dioxip
The scientific reassessment of dloxin consists of five activities:
1. Update and revision of the health assessment document for dloxin.
2. Laboratory research In support of the dose-response model.
3. Development of a biologically based dose-response model for dloxin.
4. Update and revision of the dloxin exposure assessment document.
5. Research to characterize ecological risks in aquatic ecosystems.
The first four activities have resulted in' two draft documents (the health
assessment document and exposure document) for 2,3,7,8-tetrachlorodibenzo-p-dioxin
(TCDD) and related compounds. These companion documents, which form the basis
for the Agency's reassessment of dloxin, have been used In the development of the
risk characterization chapter that follows the health assessment (Chapter 9, Volume
III). The process for developing these documents consisted of three phases which are
outlined in later paragraphs.
The fifth activity, which la In progress at EPA's Environmental Research
Laboratory in Duluth, Minnesota, Involves characterizing ecological risks in aquatic
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09-08-1994 ;2:24 202 260 8O61 T15 P.07
to hear and receive public comments and reviews of the proposed plans, and to
receive any current, scientifically relevant Information.
In the Fall of 1992, the Agency convened two peer-review workshops to review
draft documents related to EPA'B scientific reassessment of the health effects of dloxln.
The first workshop was held September 10 and 11,1992, to review a draft exposure
assessment titled, Estimating Exposures to Dioxln-Uke Compounds. The second
workshop was held September 22-25,1992, to review eight chapters of a future draft
Health Assessment Document for 2,3,7,8'TetrachlorQdlbenzQ-p-dloxin (TCDD) and
Related Compounds. Peer-reviewers were also asked to Identify Issues to be
incorporated Into the risk characterization, which was under development.
In the Fall of 1993, a third peer-review workshop was held on September 7 and
8, to review a draft of the revised and expanded Epidemiology and Human Data
Chapter, which also would be part of the future health assessment document. The
revised chapter provided an evaluation of the scientific quality and strength of the
epidemiology data In the evaluation of toxic health effects, both cancer and noncancer,
from exposure to dloxln, with an emphasis on the specific congener, 2,3,7,8-TCDD.
Prior to each workshop, the draft documents or chapters were made available
In keeping with the Agency's continuing commitment to conduct the reassessment of
dloxin In an open and participatory manner, to keep the public Informed of its
progress, and to encourage public participation in the document development
process. The public also was invited to attend the workshops, to present oral
comments, and/or to submit written comments. At each workshop, a panel of
8
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adverse health effects of dioxin in people, of the pathways to human exposure, and of
the toxic effects of dloxln to the environment. The reassessment Is part of the
Agency's goals to Improve the research and science base and to incorporate
Improved research and science into EPA decisions.
History
In 1885 and 1988, the Agency prepared assessments of the human health risks
from environmental exposures to dioxin. Also, In 1988, a draft exposure document
was prepared that presented procedures for conducting site-specific exposure
assessments to dloxln-llke compounds. These assessments were reviewed by the
Agency's Science Advisory Board (SAB). At the time of the 1988 assessments, there
waa general agreement within the scientific community that there could be a
substantial improvement over the existing approach to analyzing dose response, but
there was no consensus as to a more biologically defensible methodology. The
Agency was asked to explore the development of such a method. The Agency's
reassessment activities are In response to this request.
Stages In the Reassessment Process That Have Been Completed
The EPA had endeavored to make each phase of the reassessment of dioxin
an open and participatory effort. On November 15, 1991, and April 28, 1992, public
meetings were held to discuss the Agency's plans and activities for the reassessment,
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For the exposure assessment document, send comments to: Dloxln Exposure
Assessment Comments, Technical Information Staff (8801), Office of Health and
Environmental Assessment, U.S. Environmental Protection Agency, 401 M Street,
S.W., Washington, DC 20460.
FOR FURTHER INFORMATION, CONTACT:
For questions on the overall reassessment of dioxin or technical questions on
the health assessment document: William Farland, Office of Health and Environmental
Assessment (8601), Office of Research and Development, U.S. Environmental
Protection Agency, 401 M Street, S.W., Washington, DC 20460; telephone (202) 260-
7315; fax (202) 260-0393.
For technical questions on the exposure assessment: John Schaum, Exposure
Assessment Group (8603), Office of Health and Environmental Assessment, U.S.
Environmental Protection Agency, 401 M Street, S.W., Washington, DC 20460;
telephone (202) 260-8909; fax (202) 260-1722.
SUPPLEMENTARY INFORMATION:
The Scientific Reassessment of Dioxin
In April 1991, EPA announced that It would conduct a scientific reassessment of
the health risks of exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and
chemically similar compounds collectively known as dioxin. The EPA has undertaken
this task In response to emerging scientific knowledge of the biological, human health,
and environmental effects of dioxin. Significant advances have occurred In the
scientific understanding of mechanisms of dioxin toxiclty, of the carcinogenic and other
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09-06-1994 12:22 202 260 6061 T15 P.04
Agency, 26 W. Martin Luther King Drive, Cincinnati, OH 45268; telephone (513) 569-
7562; fax (513) 569-7566. Please provide your name, mailing address, document title,
and EPA number.
Please note that the two summary volumes also will be made available as
WordPerfect 5.1 files on 3H" PC-DOS formatted disks. Please request by document
title and EPA number:
Rlak Characterization Chapter (Vol. Ill-Health), EPA/600/BP-92/OOlca (disk)
Executive Summary Chapter (Vol. l-Exposure), EPA/600/6-88/Q05Caa (disk)
The draft documents will be provided for Inspection at the ORD Public
Information Shelf, EPA Headquarters Library, 401 M Street, S.W., Washington, DC
20460, between the hours of 10:00 a.m. and 2:00 p.m., Monday through Friday,
except for Federal holidays, and at all of the EPA Regional and Laboratory libraries.
Submitting Comments
AH comments must be in writing. Commenters should submit three copies of
each comment, and If commenting on both documents—the health assessment
document and the exposure assessment-submit separate comments rather than
combined submissions.
For the hearth assessment document, send comments to: Dioxin Health
Assessment Comments, Technical Information Staff (8601), Office of Health and
Environmental Assessment, U.S. Environmental Protection Agency, 401 M Street,
S.W., Washington, DC 20460.
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09-08-1994 12:22 2C2 260 8061 T1S P. 03
ADDRESSES:
Requesting Documents
Due to the large size of both draft documents (each Is over 1,000 pages In
length), the documents will be available as follows:
Health Assessment Document for 2,3,7,8~Tetrachlorodlb9nzo-p'dlQxln (TCDD)
and Related Compound, EPA/800/BP-92/001a, 001 b, 001 c. (Note: The full
document is 3 volumes and approx. 1,100 pages.)
OR
Risk Characterization Chapter, EPA/SOO/BP-92/OOlc. (Note: This third volume
of the 3-volume set Integrates health and exposure information on dioxin and
related compounds; approx. 100 pages.)
AND/OR
Estimating Exposure to Qloxin-Uke Compounds, EPA/600/6-88/005Ca, Cb, Cc.
(Note: The full document is 3 volumes and approx. 1,300 pages.)
OR
Executive Summary Chapter of the Exposure Document, EPA/600/6-88/005Ca.
(Note: This first volume of the 3-volume set summarizes the exposure
information; approx. 100 pages.)
To obtain a paper copy of these draft documents, interested parties should
contact the ORD Publications Center, CERI-FRN, U.S. Environmental Protection
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09-08-1S94 12:22 202 260 8061 TIS P.02
Reassessment of 2,3,7,8-Tetrtchlorodlberuo-p-dloxln
(2,3,7,8-TCDD, dloxln)
AGENCY: U.S. Environmental Protection Agency (EPA)
ACTION: As part of the Agency's reassessment of 2,3,7,8-tetrachlorodibenzo-p-dloxln
(2,3,7,8-TCDD; hereinafter referred to as simply dloxln), two External Review Draft
documents are being made available for a 120-day public review and comment period.
SUMMARY: This notice announces the availability of two External Review Draft
documents for public review and comment:
1. Health Assessment Document for 2,3,7tB-Tetrachlorodibenzo-p-dloxln
(TCDD) and Related Compounds (EPA/600/BP-92/OOia-c)
2. estimating Exposure to DloxJn-Uke Compounds (EPA/600/6-88/005Ca-c)
During the public comment period, public comment meetings will be convened to take
formal comments on the draft documents. These meetings are being planned for the
first two weeks of December at five locations: Washington, DC; New York, NY/New
Jersey; Chicago, IL; Dallas, TX; and San Francisco, CA, Detailed information will be
provided in a future Federal Register notice.
The draft documents also will be reviewed at a Science Advisory Board meeting
to be held after the public comment period has ended, eariy next year. Information
about this meeting will be published In a future Federal Register notice.
DATES: The draft documents will be made available on September 13,1994.
Comments must be postmarked by January 13, 1994.
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DRAFT EPA/600/BP-92/001b
DO NOT QUOTE OR CITE June 1994
External Review Draft
Health Assessment Document for
2,3,7,8-Tetrachlorodibenzo-/?-dioxin (TCDD)
and Related Compounds
Volume II of III
U.S. Environmental Protection Agency
Region 5, Library (PL-12J)
77 West Jackson Boutevaitf, 22th
Oucagp, it 60604-3590
NOTICE
THIS DOCUMENT IS A PRELIMINARY DRAFT. It has not been formally released by
the U.S. Environmental Protection Agency and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its technical accuracy and
policy implications.
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, D.C.
Printed on Recycled Paper
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DISCLAIMER
This document is an external draft for review purposes only and does not constitute
Agency policy. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
Il-ii 06/30/94
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Health Assessment Document for
2,3,7,8-Tetrachlorodibenzo-p-dioxin(TCDD)
and Related Compounds
TABLE OF CONTENTS - OVERVIEW
Volume I
1. DISPOSITION AND PHARMACOKINETICS
2. MECHANISM(S) OF ACTION
3. ACUTE, SUBCHRONIC, AND CHRONIC TOXICITY
4. IMMUNOTOXICITY
5. DEVELOPMENTAL AND REPRODUCTIVE TOXICITY
6. CARCINOGENICITY OF TCDD IN ANIMALS
Volume II
7. EPIDEMIOLOGY/HUMAN DATA
PART A. CANCER EFFECTS
PART B. EFFECTS OTHER THAN CANCER
8. DOSE-RESPONSE MODELING
Volume III
9. RISK CHARACTERIZATION OF 2,3,7,8-TETRACHLORODIBENZO-p-DIOXIN
(TCDD) AND RELATED COMPOUNDS
II-iii 06/30/94
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Health Assessment Document for
2,3,7,8-Tetrachlorodibenzo-/?-dioxin(TCDD)
and Related Compounds
CONTENTS - VOLUME II
List of Tables II-x
List of Figures II-xvi
List of Abbreviations and Acronyms II-xvii
Preface II-xxv
Authors, Contributors, and Reviewers II-xxix
7. EPIDEMIOLOGY/HUMAN DATA
PART A: CANCER EFFECTS
7.1. INTRODUCTION 7-1
7.2. SCOPE 7-1
7.3. PREVIOUS EPA REVIEWS 7-4
7.4. REVIEW METHODS 7-4
7.5. FOLLOW-UP STUDIES OF CHEMICAL MANUFACTURING AND
PROCESSING WORKERS 7-8
7.5.1. United States 7-8
7.5.2. Germany 7-17
7.5.3. Ten-Country Study by International Agency for Research on Cancer 7-22
7.5.4. Other Studies 7-27
7.5.5. Summary 7-33
7.6. CASE-CONTROL STUDIES IN GENERAL POPULATIONS 7-35
7.6.1. Sweden 7-36
7.6.2. United States 7-45
7.6.3. New Zealand 7-51
7.6.4. Italy 7-58
7.6.5. Summary 7-60
7.7. STUDIES OF PULP AND PAPER MILL WORKERS 7-62
Il-iv 06/30/94
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CONTENTS (continued)
7.8. OTHER STUDIES 7-64
7.8.1. Vietnam Veterans 7-64
7.8.2. Residents of Seveso, Italy 7-66
7.8.3. Rice Oil Poisonings in Taiwan and Japan Involving Compounds
Structurally Related to Dioxin 7-71
7.9. CONCLUSIONS 7-73
REFERENCES FOR CHAPTER 7, PART A 7-78
PART B: EFFECTS OTHER THAN CANCER
7.10. INTRODUCTION 7-87
7.11. CROSS-SECTIONAL STUDIES: USES AND LIMITATIONS 7-88
7.12. DESCRIPTION OF PRINCIPAL STUDIES 7-90
7.12.1. Occupational Studies 7-91
7.12.1.1. U.S. Chemical Workers: West Virginia 7-91
7.12.1.2. U.S. Chemical Workers: The NIOSH Study 7-92
7.12.1.3. BASF Accident Cohort 7-93
7.12.2. Studies of Community Residents 7-94
7.12.2.1. The Missouri Experience 7-94
7.12.2.2. Seveso, Italy 7-96
7.12.3. Studies of Vietnam Veterans 7-97
7.12.3.1. The Vietnam Experience Study 7-97
7.12.3.2. U.S. Air Force Ranch Hand Study 7-98
7.13. REVIEW OF EFFECTS ASSOCIATED WITH EXPOSURE TO
2,3,7,8-TCDD 7-100
7.13.1. Dermal Effects 7-100
7.13.1.1. Chloracne 7-100
7.13.1.2. Dermatologic Disorders Other Than Chloracne .... 7-104
7.13.1.3. Comments 7-106
7.13.2. Gastrointestinal Effects 7-107
7.13.2.1. Hepatic Effects 7-107
7.13.2.2. Liver Size 7-107
7.13.2.3. Enzyme Levels 7-108
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CONTENTS (continued)
7.13.2.4. GOT 7-109
7.13.2.5. AST and ALT 7-112
7.13.2.6. D-Glucaric Acid 7-116
7.13.2.7. Comment 7-118
7.13.2.8. Porphyrin Metabolism 7-119
7.13.2.9. Comment 7-121
7.13.2.10. Lipid Levels 7-122
7.13.2.11. Total Cholesterol 7-122
7.13.2.12. Triglycerides 7-128
7.13.2.13. Comment 7-129
7.13.3. Other Gastrointestinal Disorders 7-130
7.13.4. Thyroid Function 7-130
7.13.4.1. Comment 7-134
7.13.5. Diabetes 7-137
7.13.5.1. Comment 7-139
7.13.6. Immunologic Effects 7-143
7.13.6.1. Comment 7-158
7.13.7. Neurologic Effects 7-159
7.13.7.1. Neurobehavioral Assessments 7-167
7.13.7.2. Neurologic Status 7-175
7.13.7.3. Comment 7-178
7.13.8. Circulatory System 7-178
7.13.8.1. Comment 7-185
7.13.9. Pulmonary Effects 7-187
7.13.9.1. Comment 7-188
7.13.10. Renal Effects 7-188
7.13.11. Reproductive Effects 7-189
7.13.12. Review of the Literature Prior to 1984 7-191
7.13.12.1. Occupational Studies 7-191
7.13.12.2. Environmental Studies 7-202
7.13.12.3. The Seveso, Italy, Dioxin Accident of 1976 7-202
7.13.12.4. Studies of Exposure to Agent Orange by Military
Veterans and Vietnamese Civilians 7-204
7.13.13. Review of the Literature From 1984 to 1992 7-207
7.13.14. Environmental Studies 7-208
7.13.14.1. The Times Beach, Missouri, 2,3,7,8-TCDD Episode 7-208
7.13.14.2. Studies of Vietnam Experience in Ground Troops
and Ranch Hands Published 1984-1992 7-211
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CONTENTS (continued)
7.13.15. Study Results 7-214
7.13.15.1. Atlanta Congenital Defects Program Study 7-214
7.13.15.2. CDC Vietnam Experience Study 7-215
7.13.15.3. American Legion Study 7-221
7.13.15.4. Boston Hospital Study 7-221
7.13.15.5. The Ranch Hand Study 7-222
7.13.15.6. Reproductive Hormones 7-229
7.13.15.7. Comment 7-230
7.14. NONCANCER EFFECTS OF INGESTION OF RICE OIL CONTAMINATED
WITH POLYCHLORINATED DIBENZOFURANS, QUATERPHENYLS,
AND BIPHENYLS IN JAPAN (YUSHO) AND TAIWAN (YU-CHENG) . 7-230
7.14.1. Acute Effects in Adults and Children Directly Exposed to
Contaminated Rice Oil 7-231
7.14.1.1. Yusho 7-231
7.14.1.2. Yu-Cheng 7-233
7.14.1.3. Effects Observed in Offspring of Yu-Cheng Cases . . 7-235
7.14.1.4. Comment 7-237
7.15. SUMMARY 7-238
7.15.1. Effects Having a Positive Relationship With Exposure to
2,3,7,8-TCDD 7-243
7.15.1.1. Chloracne 7-243
7.15.1.2. Gamma Glutamyl Transferase (GOT) Levels 7-245
7.15.1.3. Diabetes and Fasting Serum Glucose Levels 7-247
7.15.1.4. Reproductive Hormones 7-249
7.15.2. Possible Acute Effects of Exposure to 2,3,7,8-TCDD 7-250
7.15.2.1. Dermatologic Conditions Other Than Chloracne ... 7-250
7.15.2.2. Liver Enzymes Other Than GGT and Hepatomegaly . 7-251
7.15.2.3. Pulmonary Disorders 7-252
7.15.2.4. Neurologic Disorders 7-253
7.15.2.5. Porphyrias 7-253
7.15.3. Effects For Which Further Research Is Needed 7-254
7.15.3.1. Diseases of the Circulatory System 7-254
7.15.3.2. Reproductive Effects 7-256
7.15.3.3. Specific Reproductive End Points 7-258
7.15.3.4. Immunologic Effects 7-261
7.15.3.5. Lipids 7-261
7.15.3.6. Thyroid Function 7-262
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CONTENTS (continued)
REFERENCES FOR CHAPTER 7, PART B 7-263
8. DOSE-RESPONSE MODELING 8-1
8.1. INTRODUCTION 8-1
8.1.1. Introduction to Modeling for TCDD 8-13
8.1.2. Dose Delivery, Tissue Modeling, and Biochemical Modeling .... 8-24
8.2. TOXIC EFFECTS 8-38
8.2.1. Modeling Liver Tumor Response for TCDD 8-38
8.2.2. Tumor Incidence 8-41
8.2.2.1. A Potential Alternative Model for Promotion of
Carcinogenesis by Dioxin 8-56
8.2.3. Other Effects: Mammary/Uterine/Anticancer End Points 8-59
8.2.4. Noncancer End Points 8-61
8.2.5. Neurological and Behavioral Toxicity 8-62
8.2.6. Teratological and Developmental 8-63
8.2.6.1. Cleft Palate 8-63
8.2.6.2. Hydronephrosis 8-65
8.2.6.3. Thymic Atrophy 8-66
8.2.7. Immunotoxicity 8-66
8.2.8. Reproductive Toxicity 8-67
8.2.8.1. Female Reproductive Toxicity 8-67
8.2.8.2. Male Reproductive Toxicity 8-69
8.3. COMPARATIVE END POINTS/QUALITATIVE COMPARISONS 8-70
8.4. RELEVANCE OF ANIMAL DATA FOR ESTIMATING HUMAN RISKS . 8-73
8.5. HUMAN MODELS 8-77
8.5.1. Introduction 8-77
8.5.2. Modeling Toxic Effects in the Liver 8-78
8.5.3. Lung Cancer and All Cancers Combined 8-80
8.5.3.1. Dose-Response Models 8-83
8.5.3.2. Exposure and Dose Estimates 8-86
8.5.3.3. Calculation of Risk Estimates 8-93
8.5.3.4. Low-Dose Deviation From Linearity 8-98
8.5.3.5. Uncertainties in Estimates From Human Epidemiology 8-100
8.5.3.6. Conclusions 8-101
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CONTENTS (continued)
8.6. KNOWLEDGE GAPS 8-102
REFERENCES FOR CHAPTER 8 AND APPENDICES C AND D 8-107
APPENDIX A: A Mechanistic Model of Effects of Dioxin on Gene Expression
in the Rat Liver A-l
APPENDIX B: Modeling Receptor-Mediated Processes with Dioxin: Implications
for Pharmacokinetics and Risk Assessment B-l
APPENDIX C: Parameters for Analyzing Preneoplastic Lesions and Tumor
Incidence in Rat Hepatocytes C-l
APPENDIX D: Considerations for Using Dose Surrogates in Estimating Tumor
Incidence D-l
APPENDIX E: Tetrachlorodibenzo-p-Dioxin Empirical Relationships for
Non-Cancer End Points E-l
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LIST OF TABLES
7-1 Relative Risks of Selected Cancers in Study of Chemical Manufacturing
Workers Exposed to TCDD in United States, by Exposure Duration and Latency 7-10
7-2 Relative Risks of Lung Cancer in Subcohort of Chemical Manufacturing
Workers Exposed to TCDD in United States for at Least 1 Year and With at
Least 20 Years Latency, Adjusted for Alternative Hypotheses About Its
Smoking Distribution 7-14
7-3 Relative Risks of All Cancers Combined in Study of Chemical Manufacturing
Workers Exposed to TCDD in Germany, by Duration and Category of Exposure 7-19
7-4 Relative Risks of Selected Cancers in Study of Chemical Manufacturing Workers
Exposed to TCDD in Germany, by Median Blood TCDD Level and Latency . . . 7-23
7-5 Summary of Results for Selected Cancers From Follow-up Studies of Chemical
Manufacturing and Processing Workers Exposed to TCDD 7-34
7-6 Relative Risks of Soft Tissue Sarcomas and Malignant Lymphomas in Relation
to Phenoxy Acid and Chlorophenol Exposures in Five Case-Control Studies in
Sweden 7-38
7-7 Relative Risks of Malignant Lymphoma and Soft Tissue Sarcomas in Relation to
Phenoxy Acid and Chlorophenol Exposures in Three Case-Control Studies in
Sweden 7-41
7-8 Mantel-Haenszel Odds Ratios for Soft Tissue Sarcoma Among Persons Exposed
to All Dioxins, TCDD, and Dioxins Other Than TCDD in Four Case-Control
Studies Involving 434 Cases and 948 Controls 7-44
7-9 Relative Risks of Soft Tissue Sarcomas, Non-Hodgkin's Lymphomas, and
Hodgkin's Disease in Relation to Phenoxy Acid and Chlorophenol Exposures
in Two Case-Control Studies in Southern Sweden 7-46
7-10 Relative Risks of Non-Hodgkin's Lymphomas in Relation to Farm Use of
2,4,5-T in Case-Control Studies in Kansas, Eastern Nebraska, Iowa, and
Minnesota 7-47
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LIST OF TABLES (continued)
7-11 Relative Risks of Soft Tissue Sarcomas and Non-Hodgkin's Lymphomas in
Relation to Phenoxyacetic Acid and Chlorophenol Exposure in a Case-Control
Study in Western Washington State, 1981-1984 7-50
7-12 Relative Risks of Soft Tissue Sarcomas and Non-Hodgkin's Lymphomas in
Relation to Potential Exposure to Phenoxy Acids and Chlorophenols in
Case-Control Studies in New Zealand 7-54
7-13 Relative Risks of Soft Tissue Sarcomas in Relation to Phenoxy Acid Exposure
in Case-Control Study in Northern Italy, 1981-1983 7-59
7-14 Relative Risks for Selected Cancers from Follow-up Studies of Paper and
Pulp Mill Workers 7-63
7-15 Relative Risks for Selected Cancers Among Adults Exposed to TCDD in
Seveso, Italy 7-68
7-16 Relative Risks for Selected Cancers Among Adults Exposed to TCDD in
Seveso, Italy, in Contaminated Areas B and R 7-70
7-17 Serum 2,3,7,8-Tetrachlorodibenzo-/?-dioxin (TCDD) Levels in Seveso Residents
with Chloracne and Adipose Tissue Levels of 2,3,7,8-TCDD and
Hexachlorinated (HxCDD) Dioxins in German Chemical Workers 7-103
7-18 Mean Serum Levels of Gamma Glutamyl Transferase (GGT) Among Seveso and
Missouri Residents, TCP Production Workers, BASF Accident Cohort, and
Vietnam Veterans 7-110
7-19 Logistic Regression Model for an Out-of-Range Serum Gamma-
Glutamyltransferase (GGT) Level Using the Categorical TCDD Exposure
Measure .7-113
7-20 Serum Alanine Aminotransferase (ALT) Among Seveso Children and Missouri
Residents, TCP Production Workers, BASF Accident Cohort, and Vietnam
Veterans .7-115
7-21 Mean D-Glucaric Acid Levels Among Seveso Residents, TCP Production
Workers, and Vietnam Veterans 7-117
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LIST OF TABLES (continued)
7-22 Mean Total Cholesterol Levels Among Seveso and Missouri Residents, TCP
Production Workers, BASF Accident Cohort, and Vietnam Veterans 7-123
7-23 Mean Triglyceride Levels Among Seveso and Missouri Residents, TCP
Production Workers, BASF Accident Cohort, and Vietnam Veterans 7-125
7-24 Levels of Triiodothyronine Percent (T3%) Uptake or Free Thyroxine Index in
Vietnam Veterans 7-132
7-25 Levels of Thyroid-Stimulating Hormone (TSH) in Vietnam Veterans, Nursing
Infants, and BASF Accident Cohort 7-133
7-26 Levels of Thyroxine-Binding Globulin (TBG), Thyroxine (T4), or T4/TGB in
Nursing Infants and BASF Accident Cohort 7-135
7-27 Multiple Logistic Regression Model for Cases of Diabetes for TCP Production
Workers 7-138
7-28 Multiple Linear Regression Model for Log Fasting Serum Glucose for TCP
Production Workers and Unexposed Referents 7-140
7-29 Adjusted Relative Risk (RR) for Fasting Serum Glucose Levels, Cases of
Diabetes, and Mean 2-Hour Postprandial Glucose Levels by Category of Lip id
Adjusted Serum 2,3,7,8-TCDD in Ranch Hands 7-141
7-30 Effects of Exposure to 2,3,7,8-TCDD on Serum Glucose Levels in Nonhuman
Mammalian Species 7-142
7-31 CD4/CD8 Ratios in Missouri Residents, Vietnam Veterans, and BASF Accident
Cohort .7-145
7-32 Total Lymphocytes in 2,4,5-T Production Workers, Missouri Residents, Vietnam
Veterans, and BASF Accident Cohort 7-146
7-33 Bl Levels in Production Workers, 2,4,5-T Missouri Residents, Vietnam
Veterans, BASF Accident Cohort, and Extruder Personnel 7-148
7-34 CD4 Levels in Production Workers, Missouri Residents, Vietnam Veterans, and
BASF Chemical Workers 7-150
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LIST OF TABLES (continued)
7-35 CDS Levels in 2,4,5-T Production Workers, Missouri Residents, Vietnam
Veterans, and BASF Accident Cohort 7-152
7-36 IgG Levels in Missouri Residents, Vietnam Veterans, and BASF Accident 7-154
Cohort
7-37 IgM Levels in Missouri Residents, Vietnam Veterans, BASF Accident Cohort,
and Extruder Personnel 7-155
7-38 Levels of Natural Killer Cells in Missouri Residents, Vietnam Veterans, and
Extruder Personnel 7-157
7-39 Case Reports of Psychological and Neurologic Effects Among Individuals
Exposed to 2,3,7,8-TCDD-ContaminatedMaterials 7-160
7-40 Cross-Sectional Studies of Psychological and Neurologic Effects Among
Residents of Missouri and Seveso Exposed to 2,3,7,8-TCDD-Contaminated
Materials .7-165
7-41 Cross-Sectional Studies of Psychological and Neurologic Effects Among TCP
Production Workers and Vietnam Veterans Exposed to 2,3,7,8-TCDD-
Contaminated Materials 7-169
7-42 Mortality from Diseases of the Circulatory System in Populations Exposed to
2,3,7,8-TCDD 7-180
7-43 Results of Studies Examining the Effect of Dioxin on Reproductive Outcomes in
Humans, 1984-1992 7-193
7-44 2,3,7,8-TCDD Levels (pg/g of Lipid) for Selected Populations 7-200
7-45 Odds Ratios for Selected Categories of Birth Defects for the Telephone Interview
and Hospital Records Study in the Vietnam Experience Study, 1989 7-217
7-46 Results of the Misclassification Analyses for Birth Defects in the Hospital
Records Substudy, Vietnam Experience Study, 1989 7-218
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LIST OF TABLES (continued)
7-47 Rates of Miscarriage (per 1,000) by Pre- and Post-Vietnam Tour Status and Time
Since Tour of Duty, Among 1,475 Ranch Hands with > 10 pg/g Serum Dioxin,
Ranch Hand Study, 1992 7-226
7-48 Summary of Effects Observed in Humans 7-239
8-1 Examples of Levels of Information Available for Estimating Parameters
in Dose-Response Modeling 8-18
8-2 Design Considerations for Risk Assessment Purposes 8-22
8-3 Administered Dose, Tumor Response, and Number at Risk of Hepatocellular
Neoplasms in Male Sprague-Dawley Rats From Carcinogenicity Experiments
of Kociba et al. (1978) Using the Pathology Review of Sauer (1990) 8-40
8-4 Toxic End Points Database 8-72
8-5 Similarities Between Laboratory Animals and Humans in Biological Effects
of TCDD 8-75
8-6 Rat and Human Comparison of Daily TCDD Intakes and Body and Liver
Concentration for Equitoxic Response 8-81
8-7 Measured Serum TCDD Levels and Estimated Levels at Time of Last
Occupational Exposure to TCDD, Based on First-Order Elimination Kinetics
and a Half-Life for Elimination of 7.1 Years 8-88
8-8 Estimates of Lifetime Average Daily Dose for Oral Intake Equivalence
for TCDD Based on Total Concentration x Time Equivalence. Estimates
of TCDD Concentration Adjusted for Background at Time of Sampling and
Back-Calculated Using First Order Elimination Kinetics, by Cohort 8-92
8-9 Estimated Lifetime Average Daily Doses and Relative Risks by Individual
Study Cohort 8-94
8-10 Calculation of Incremental Unit Cancer Risk Estimates and 95% Lower
Limits for Both the Additive and Relative Risk Models Based on the Lung
Cancer Deaths Response in the Fingerhut, Zober, and Manz Studies 8-96
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LIST OF TABLES (continued)
8-11 Calculation of Incremental Unit Risk Estimates and 95% Lower Limits
for Both the Additive and Relative Risk Models Based on the Total Cancer
Deaths Response in the Fingerhut, Zober, and Manz Studies 8-97
8-12 Estimates of U.S. EPA Unit Cancer Risk for TCDD Oral Intake, Based
on Animal and Human Studies and U.S. EPA Current and Proposed
Estimates 8-99
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LIST OF FIGURES
8-1 Dose/response graph showing proportional relationship between receptor
occupancy and biological response (semilog scale) .................... 8-6
8-2 Dose/response graph showing proportional relationship between receptor
occupancy and biological response (arithmetic scale) ................... 8-8
8-3 Multistage carcinogenesis .................................... 8-9
8-4 The Armitage-Doll model of carcinogenesis ....................... 8-11
8-5 Developing a mechanistically based mathematical model ................ 8-16
8-6 A two-stage model of carcinogenesis ............................ 8-42
8-7 Fit of the two-stage model to the number of focal lesions from the data of
Maronpot et al. (1993) .................................... 8-47
8-8 Fit of the two-stage model to the size distribution for focal lesions from
the data of Maronpot et al. (1993) where the smooth line is derived
from the model and the step function results from the data .............. 8-48
8-9 Fit of the TS model to the data of Kociba et al. (1978) using parameters
for Mn-i(d), ft(d), and 6j(d) estimated from the focal lesion data ............ 8-49
8-10 Dose-response for tumor incidence using a two-stage model of
carcinogenesis ......................................... 8-52
8-11 Relative risks of lung cancer and all cancer mortality in three recent
cohort studies of workers exposed to TCDD by estimated LADD
equivalence ........................................... 8-95
8-12 Biologically based risk assessment approaches for dioxin: filling the gaps .... 8-104
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LIST OF ABBREVIATIONS AND ACRONYMS
ACTH Adrenocorticotrophic hormone
Ah receptor Aryl hydrocarbon receptor
AHH Aryl hydrocarbon hydroxylase
ALA Aminolevulinic acid
ALT L-alanine aminotransferase
AOR Adjusted odds ratio
APC Antigen-presenting cells
AST L-aspartate aminotransferase
ATPase Adenosine triphosphatase
BDD Brominated dibenzo-/?-dioxin
BDF Brominated dibenzofuran
BCF Bioconcentration factor
BGG Bovine gamma globulin
bHLH Basic helix-loop-helix
bw Body weight
cAMP Cyclic 3,5-adenosine monophosphate
CDC Centers for Disease Control and Prevention
CDD Chlorinated dibenzo-/?-dioxin
CDF Chlorinated dibenzofuran
cDNA Complementary DNA
cl Confidence level
CMI Cornell Medical Index
CNS Central nervous system
CSM Cerebrospinal malformation
CTL Cytotoxic T lymphocyte
DCDD 2,7-Dichlorodibenzo-p-dioxin
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
DEN Diethylnitrosamine
DHT 5a-Dihydrotestosterone
DIS Diagnostic Interview Schedule
DMBA Dimethylbenzanthracene
DMSO Dimethyl sulfoxide
DNA Deoxyribonucleic acid
DRE Dioxin-responsive enhancers
DTK Delayed-type hypersensitivity
EC50 Concentration effective for 50% of organisms tested
EC100 Concentration effective for 100% of organisms tested
ED50 Dose effective for 50% of recipients
ECOD 7-Emoxycoumarin-O-deethylase
EEG Electroencephalogram
EOF Epidermal growth factor
EGFR Epidermal growth factor receptor
ER Estrogen receptor
EROD 7-Ethoxyresorufin-O-deethylase
EOF Enzyme altered foci
EOI Exposure opportunity index
FEV Forced expiratory volume
FIQ Full-scale IQ
FSH Follicle-stimulating hormone
FTI Free thyroxine index
FVC Forced vital capacity
GC-ECD Gas chromatograph-electron capture detection
GC/MS Gas chromatograph/mass spectrometer
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GOT
GnRH
GST
GVH
HAH
HCB
HCDD
HDL
HLH
HP AH
HpCDD
HpCDF
HPLC
HRB
HRGC/HRMS
HTL
HxBB
HxCB
HxCDD
HxCDF
ICD-9
ID50
I-TEF
KVK
LADD
LD50
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
Gamma glutamyl transpeptidase
Gonadotropin-releasing hormone
Glutathione-S-transferase
Graft versus host
Halogenated aromatic hydrocarbons
Hexachlorobenzene
Hexachlorodibenzo-/?-dioxin
High density lipoprotein
Helix-loop-helix
Halogenated polycyclic aromatic hydrocarbon
Heptachlorinated dibenzo-p-dioxin
Heptachlorinated dibenzofuran
High performance liquid chromatography
Halstead-Reitan Battery
High resolution gas chromatography/high resolution mass spectrometry
Human tonsillar lymphocytes
Hexabrom-biphenyl
Hexachlorobiphenyl
Hexachlorinated dibenzo-/?-dioxin
Hexachlorinated dibenzofuran
International Classification of Diseases 9
Dose infective to 50% of recipients
International TCDD-toxic-equivalency
Kemisk Vaerk K0ge
Lifetime average daily dose
Dose lethal to 50% of recipients (and all other subscripter dose levels)
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
LDH
LH
LDL
LMS
LPL
LOAEL
LOEL
LPS
MACDP
3-MC
MCDF
MCF-7
MCMI
MCPA
MCPB
MCPP
MFO
MMPI
MLE
mRNA
MNNG
NADP
NADPH
NaTCP
NHL
NIEHS
L-lactate dehydrogenase
Luteinizing hormone
Low density liproprotein
Linearized multistage
Lipoprotein lipase activity
Lowest-observable-adverse-effect level
Lowest-observed-effect level
Lipopolysaccharide
Metropolitan Atlanta Congenital Defects Program
3-Methylcholanthrene
6-Methy 1-1,3,8-trichlorodibenzofuran
(breast cancer cell)
Millon Clinical Multiaxial Inventory
(4-Chloro-2-methylphenoxy)aceticacid
2-Methyl-4-chlorophenoxybutyric acid
2-(4-Chloro-2-methylphenoxy)-propanoicacid
Mixed function oxidase
Minnesota Multiphase Personality Inventory
Maximum likelihood estimate
Messenger RNA
W-methyl-./V-nitrosoguanidine
Nicotinamide adenine dinucleotide phosphate
Nicotinamide adenine dinucleotide phosphate (reduced form)
Sodium 2,4,5-trichlorophenate
Non-Hodgkin's lymphoma
National Institute of Environmental Health Sciences
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
NIOSH
NK
NOAEL
NOEL
NTP
OCDD
OCDF
OR
OVX
PAA
PAH
PBA
PBB
PBF
PEL
PB-PK
PCB
PCBA
PCDD
PCDF
PCP
PCPA
PCQ
PCT
PeCDD
PeCDF
National Institute for Occupational Safety and Health
Natural killer
No-observable-adverse-effect level
No-observed-effect level
National Toxicology Program
Octachlorodibenzo-p-dioxin
Octachlorodibenzofuran
Odds ratio
Ovariectomized
Phenoxyacetic acid
Polyaromatic hydrocarbon
Phenoxybutyric acid
Polybrominated biphenyl
Percent body fat
Peripheral blood lymphocytes
Physiologically based pharmacokinetic
Polychlorinated biphenyl
Phenoxybutyric acid
Polychlorinated dibenzodioxin
Polychlorinated dibenzofuran
Pentachlorophenol
Parachlorophenoxy acetic acid
Quaterphenyl
Porphyria cutanea tarda
Pentachlorinated dibenzo-/?-dioxin
Pentachlorinated dibenzo-/?-dioxin
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
PEPCK Phosphoenol pyruvate carboxykinase
PFC Plaque-forming cell
PGEj Prostaglandin E2
PGF2a Prostaglandin F2a
POST Placental glutathione-S-transferase
PGT Placental glutathione transferase
PHA Phytohemagglutinin
PIQ Performance IQ
PKC Protein kinase C
PNS Peripheral nervous system
POMS Profile of Mood States
ppb Parts per billion
ppm Parts per million
ppt Parts per trillion
PRR Prevalence risk ratio
PWM Pokeweed mitogen
RNA Ribonucleic acid
RR Relative risk
SAR Structure-activity relationships
SB-IQ Standford Binet IQ
SCL-90-R Self-Report Symptom Checklist-90-Revised
SD Standard deviation
SE Standard error
SEA Southeast Asia
SCOT Serum glutamic oxaloacetic transaminase
SGPT Serum glutamic pyruvic transaminase
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
SIR Standard incidence ratio
SMR Standardized mortality ratio
SRBC Sheep erythrocytes (red blood cells)
STS Soft tissue sarcoma
tiA Half-time
TBB Tetrabromobiphenyl
TBDD Tetrabrominated dibenzo-/?-dioxin
TBDF Tetrabrominated dibenzo-p-furan
TBG Thyroxine-binding globulin
TBP Thyroxine-binding protein
TCAOB Tetrachloroazoxybenzene
TCB Tetrachlorobiphenyl
TCDD Tetrachlorodibenzo-/?-dioxin
TCDF Tetrachlorodibenzofuran
TCP Trichlorophenol
TEF Toxic equivalency factors
TEQ Toxic equivalents
TGF Thyroid growth factor
TI T helper cell independent
TNF Tumor necrosis factor
tPA Tissue plasminogen activator
TPA Tetradecanoyl phorbol acetate
TSH Thyroid-stimulating hormone
TT Tetanus toxoid
TTR Transthyretrin
TxB2 Thromboxane B2
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
UDP Uridine diphosphate
UDPGT UDP-glucuronosyltransferase
URO-D Uroporphyrinogen decarboxylase
VIQ Verbal IQ
VLDL Very low density lipoprotein
v/v Volume per volume
w/w Weight by weight
WAIS Wechsler Adult Intelligence Scale
WISC-R Wechsler Intelligence Scale for Children, Revised
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PREFACE
In April 1991, the U.S. Environmental Protection Agency (EPA) announced that it
would conduct a scientific reassessment of the health risks of exposure to 2,3,7,8-
tetrachlorodibenzo-/?-dioxin (TCDD) and chemically similar compounds collectively known as
dioxin. The EPA has undertaken this task in response to emerging scientific knowledge of
the biological, human health, and environmental effects of dioxin. Significant advances have
occurred in the scientific understanding of mechanisms of dioxin toxicity, of the carcinogenic
and other adverse health effects of dioxin in people, of the pathways to human exposure, and
of the toxic effects of dioxin to the environment.
In 1985 and 1988, the Agency prepared assessments of the human health risks from
environmental exposures to dioxin. Also, in 1988, a draft exposure document was prepared
that presented procedures for conducting site-specific exposure assessments to dioxin-like
compounds. These assessments were reviewed by the Agency's Science Advisory Board
(SAB). At the time of the 1988 assessments, there was general agreement within the
scientific community that there could be a substantial improvement over the existing
approach to analyzing dose response, but there was no consensus as to a more biologically
defensible methodology. The Agency was asked to explore the development of such a
method. The current reassessment activities are in response to this request.
The scientific reassessment of dioxin consists of five activities:
1. Update and revision of the health assessment document for dioxin.
2. Laboratory research in support of the dose-response model.
3. Development of a biologically based dose-response model for dioxin.
4. Update and revision of the dioxin exposure assessment document.
5. Research to characterize ecological risks in aquatic ecosystems.
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PREFACE (continued)
The first four activities have resulted in two draft documents (the health assessment
document and exposure document) for 2,3,7,8-tetrachlorodibenzo-p-dioxin(TCDD) and
related compounds. These companion documents, which form the basis for the Agency's
reassessment of dioxin, have been used in the development of the risk characterization
chapter that follows the health assessment. The process for developing these documents
consisted of three phases which are outlined in later paragraphs.
The fifth activity, which is in progress at EPA's Environmental Research Laboratory
in Duluth, Minnesota, involves characterizing ecological risks in aquatic ecosystems from
exposure to dioxins. Research efforts are focused on the study of organisms in aquatic food
webs to identify the effects of dioxin exposure that are likely to result in significant
population impacts. A report titled, Interim Report on Data and Methods for the Assessment
of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) Risks to Aquatic Organisms and Associated
Wildlife (EPA/600/R-93/055), was published in April 1993. This report will serve as a
background document for assessing dioxin-related ecological risks. Ultimately, these data
will support the development of aquatic life criteria which will aid in the implementation of
the Clean Water Act.
The EPA had endeavored to make each phase of the current reassessment of dioxin an
open and participatory effort. On November 15, 1991, and April 28, 1992, public meetings
were held to inform the public of the Agency's plans and activities for the reassessment, to
hear and receive public comments and reviews of the proposed plans, and to receive any
current, scientifically relevant information.
In the Fall of 1992, the Agency convened two peer-review workshops to review draft
documents related to EPA's scientific reassessment of the health effects of dioxin. The first
workshop was held September 10 and 11, 1992, to review a draft exposure assessment titled,
Estimating Exposures to Dioxin-Like Compounds. The second workshop was held September
22-25, 1992, to review eight chapters of a future draft Health Assessment Document for
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PREFACE (continued)
2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related Compounds. Peer-reviewers were
also asked to identify issues to be incorporated into the risk characterization, which was
under development.
In the Fall of 1993, a third peer-review workshop was held on September 7 and 8,
1993, to review a draft of the revised and expanded Epidemiology and Human Data Chapter,
which also would be part of the future health assessment document. The revised chapter
provided an evaluation of the scientific quality and strength of the epidemiology data in the
evaluation of toxic health effects, both cancer and noncancer, from exposure to dioxin, with
an emphasis on the specific congener, 2,3,7,8-TCDD.
As mentioned previously, completion of the health assessment and exposure
documents involves three phases: Phase 1 involved drafting state-of-the-science chapters and
a dose-response model for the health assessment document, expanding the exposure document
to address dioxin related compounds, and conducting peer review workshops by panels of
experts. This phase has been completed.
Phase 2, preparation of the risk characterization, began during the September 1992
workshops with discussions by the peer-review panels and formulation of points to be carried
forward into the risk characterization. Following the September 1993 workshop, this work
was completed and was incorporated as Chapter 9 of the draft health assessment document.
This phase has been completed.
Phase 3 is currently underway. It includes making External Review Drafts of both
the health assessment document and the exposure document available for public review and
comment.
Following the public comment period, the Agency's Science Advisory Board (SAB)
will review the draft documents in public session. Assuming that public and SAB comments
are positive, the draft documents will be revised, and final documents will be issued.
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PREFACE (continued)
The Health Assessment Document for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
and Related Compounds has been prepared under the direction of the Office of Health and
Environmental Assessment, Office of Research and Development, which is responsible for
the report's scientific accuracy and conclusions. A comprehensive search of the scientific
literature for this document varies somewhat by chapter but is, in general, complete through
January 1994.
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
This draft Health Assessment Document was prepared under the leadership and
direction of the Office of Health and Environmental Assessment (OHEA) within EPA's
Office of Research and Development (ORD). The overall coordination and leadership of the
activities associated with EPA's reassessment of dioxin, which includes the development of
this draft document, is Dr. William H. Farland, Director of OHEA.
Authors and chapter managers for the Health Assessment Document are listed below.
Early drafts of some chapters were prepared by Syracuse Research Corporation under EPA
Contract No. 68-CO-0043. Other chapters were authored totally or in part by scientists
within EPA and other agencies within the federal government. The ORD chapter managers
were responsible for providing oversight, review, and technical editing of successive drafts,
and incorporating comments from reviewers to develop a comprehensive and consistent
document. In some cases, the chapter managers also authored sections or parts of the
chapter.
AUTHORS AND CHAPTER MANAGERS
Chapter
EPA Chapter Manager/Author
Outside Author
1. Disposition and
Pharmacokinetics
Jerry Blancato
U.S. EPA
Environmental Monitoring Systems
Laboratory
Las Vegas, NV
James Olson
Department of Pharmacology
and Therapeutics
State University of New York
Buffalo, NY
2. Mechanism(s) of Action
William H. Farland
U.S. EPA
Office of Health and Environmental
Assessment (OHEA)
Washington, DC
James Whitlock, Jr.
Department of Molecular
Pharmacology
Stanford University School of Medicine
Stanford, CA
3. Acute, Subchronic, and
Chronic Toxicity
Debdas Mukerjee
Environmental Criteria and
Assessment Office/OHEA
Cincinnati, OH
UlfG. Ahlborg
Karolinska Institute
Stockholm, SWEDEN
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Chapter
EPA Chapter Manager/Author
Outside Author
4. Immunotoxicity
Ralph Smialowicz
U.S. EPA
Health Effects Research Laboratory
Research Triangle Park, NC
Gary R. Burleson*
U.S. EPA
Health Effects Research Laboratory
Research Triangle Park, NC
Nancy Kerkvliet
Agricultural Chemistry
Oregon State University
Corvallis, OR
5. Developmental and
Reproductive Toxicity
Gary Kimmel
U.S. EPA
Human Health Assessment Group
OHEA
Washington, DC
Richard Peterson
School of Pharmacy
University of Wisconsin
Madison, WI
6. Carcinogenicity of
TCDD in Animals
Charalingayya B. Hiremath
U.S. EPA
Human Health Assessment Group
OHEA
Washington, DC
George Lucier
National Institute of Environmental
Health Sciences
Research Triangle Park, NC
7. Epidemiology/Human Data
Part A. Cancer Effects
Part B. Effects Other
Than Cancer
David L. Bayliss
U.S. EPA
Human Health Assessment Group
OHEA
Washington, DC
Charles Poole*
Epidemiology Research Institute
Cambridge, MA
Marie Haring-Sweeney
National Institute for Occupational
Safety and Health
Cincinnati, OH
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Chapter
8. Dose-Response
Modeling
9. Risk Characterization of
2,3,7,8-TCDD and Related
Compounds
EPA Chapter Manager/ Author
Steven P. Bayard
U.S. EPA
Human Health Assessment Group
OHEA
Washington, DC
William H. Farland
(See Chapter 2)
Outside Author
Dioxin Dose-Response Modeling Workgroup
Michael Gallo (Co-chair), Keith Cooper,
Panos Georgopolous, and Lynne McGrath
UMDNJ-Robert Wood Johnson Medical School
Environmental and Occupational Health
Sciences Institute (EOHSI)
Piscataway, NJ
George Lucier (Co-chair) and Christopher
Portier
National Institute of Environmental
Health Sciences
Research Triangle Park, NC
Melvin Andersen and Michael DeVito
U.S. EPA
Health Effects Research Laboratory
Research Triangle Park, NC
Steven Bayard and Paul White
U.S. EPA
Office of Health and Environmental
Assessment
Washington, DC
Lorrene Kedderis
University of North Carolina
Chapel Hill, NC
Jeremy Mills
Chemical Industry Institute of Toxicology
Research Triangle Park, NC
Ellen Silbergeld
University of Maryland
Baltimore, MD
*involved with an early draft, but no longer working on the reassessment project.
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
CONTRIBUTORS
Linda Birnbaum Director, Environmental Toxicology Division, Health Effects Research
Laboratory, U.S. Environmental Protection Agency, Research Triangle
Park, NC
Marilyn Fingerhut Chief, Industry-wide Studies Branch, National Institute for
Occupational Safety and Health, Cincinnati, OH
Dorothy Patton Executive Director, Risk Assessment Forum, Office of Research and
Development, U.S. Environmental Protection Agency, Washington, DC
Peter W. Preuss Director, Office of Science, Planning, and Regulatory Evaluation, U.S.
Environmental Protection Agency, Washington, DC
Dwain Winters Office of Prevention, Pesticides, and Toxic Substances, U.S.
Environmental Protection Agency, Washington, DC
REVIEWERS
Early drafts of Chapters 1 through 8 of this health assessment were reviewed by a
panel of experts at a peer-review workshop held September 22-25, 1992. Members of the
Peer Review Panel for this workshop were as follows:
Edward Bresnick Department of Pharmacology and Toxicology, Dartmouth
Medical School, Hanover, NH
M. Judith Charles Department of Environmental Sciences and Engineering,
University of North Carolina, Chapel Hill, NC
Michael Denison Department of Biochemistry, Michigan State University, East
Lansing, MI
Phillip Enterline Emeritus Professor of Biostatistics, University of Pittsburgh,
Pittsburgh, PA
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Mark Feeley Toxicity Evaluation Division, Bureau of Chemical Safety,
Health, and Welfare, Ottawa, Ontario, Canada
Thomas A. Gasiewicz
James Gillette
Claude Hughes
Curtis D. Klaassen
Daniel Krewski
Suresh Moolgavkar
Jay Silkworth
Thomas Webster
Department of Biophysics, University of Rochester,
Rochester, NY
Laboratory of Chemical Pharmacology, National Heart, Lung,
and Blood Institute, National Institutes of Health,
Bethesda, MD
Duke University Medical Center, Durham, NC
Department of Pharmacology, Toxicology and Therapeutics, The
University of Kansas Medical Center, Kansas City, KS
Biostatistics and Computer Applications, Environmental Health
Centre, Ottawa, Ontario, Canada
Professor of Epidemiology and Biostatistics, The Fred
Hutchinson Cancer Research Center, Seattle, WA
Wadsworth Center for Laboratories and Research, New York
State Department of Health, Albany, NY
Center for the Biology of Natural Systems, Queens College,
City University of New York, Flushing, NY
On September 7 and 8, 1993, a peer-review workshop was held to review a greatly
revised and expanded draft Chapter 7 (Epidemiology/ Human Data). Members of the Peer
Review Panel for this workshop are as follows:
John Andrews
Germaine Buck
Associate Administrator for Science, Agency for Toxic
Substances and Disease Registry, Atlanta, GA
Clinical Assistant Professor, Department of Social and
Preventive Medicine, State University of New York,
Buffalo, NY
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Harvey Checkoway Professor, Department of Environmental Health, University of
Washington, Seattle, WA
Phillip Enterline Emeritus Professor of Biostatistics, University of Pittsburgh,
Pittsburgh, PA
M. Gerald Ott Director of Epidemiology, BASF Corporation, Parsippany, NJ
Allan H. Smith Professor of Epidemiology, University of California,
Berkeley, CA
Anne Sweeney Assistant Professor of Epidemiology, School of Public Health,
University of Texas, Houston, TX
Karen Webb Medical Director, HealthLine Corporation Health,
St. Louis, MO
In addition, during the development of this draft Health Assessment Document,
selected sections, chapters, or volumes were peer reviewed by scientists and experts within
EPA and other federal agencies, as well as by experts in academia and the private sector.
A draft of Chapter 9, the risk characterization, was reviewed by an interagency
workgroup comprising scientists from the following agencies of the federal government:
Department of Agriculture
Department of Defense
Department of Health and Human Services*
Department of Labor (Occupational Safety and Health Administration)
Department of Veterans Affairs
*Drafts of Chapters 7 and 9 have been reviewed by the Subcommittee on Risk
Assessment of the Committee to Coordinate Health and Environmental Related Programs
(CCEHRP) under the direction of Bryan D. Hardin of the National Institute for Occupational
Safety and Health, Centers for Disease Control, Department of Health and Human Services,
and Ron Coene, Executive Secretary of CCEHRP.
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Executive Office of the President
Office of Science and Technology Policy
Council of Economic Advisors
Domestic Policy Council
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CHAPTER 7. EPIDEMIOLOGY/HUMAN DATA
PART A: CANCER EFFECTS
7.1. INTRODUCTION
Animal bioassay data provide substantial presumptive evidence of the human
carcinogenicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (see Chapter 6), but
confirmation must come from well-designed human studies. TCDD is a multiorgan
carcinogen in animals. Target organs include the liver, thyroid, lung, skin, and soft tissues.
There is no assumption of target tissue concordance from animals to humans, although site
concordance would add support to a causal interpretation. This chapter reports on the cancer
epidemiology evidence of TCDD and its congeners.
This review and analysis of the epidemiologic literature on dioxins and cancer begins
by defining the scope of chemical exposures, cancers, and research reports to be considered.
Then, following a brief summary of previous EPA assessments of epidemiologic literature, a
description is given of the methods used in the present review. The original research reports
are then discussed in four groups: 1) follow-up studies of chemical manufacturing and
processing workers, 2) case-control studies in general populations, 3) studies of pulp and
paper mill workers, and 4) other studies (including studies of pesticide applicators, Vietnam
veterans with potential exposure to Agent Orange, residents of Seveso, Italy, exposed to
TCDD during an accidental explosion, and victims of contaminated rice oil poisonings).
Because the discussions of the first two groups of studies are relatively lengthy, brief
summaries are given at the end of each of those sections. Conclusions are drawn following
an overall discussion of all the studies.
7.2. SCOPE
Epidemiologic studies of cancer among persons exposed to TCDD and other
polychlorinated dibenzodioxins (PCDDs) and dibenzofurans (PCDFs) are included in this
review. Primary emphasis is placed on studies with exposures to TCDD itself, occurring
primarily in the manufacture and use of 2,4,5-trichlorophenol, hexachlorophene, and the
herbicide 2,4,5-trichlorophenoxyacetic acid (2,4,5-T). Because exposures to 2,4,5-T and
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2,4-dichloroacetic acid (2,4-D) often occur among the same groups who manufacture and use
these herbicides, some studies of groups exposed only to 2,4-D are also included. Exposure
to lower-chlorinated PCDDs (dichlorinated and trichlorinated isomers) may occur in the
manufacture and use of 2,4-D. Also included are studies of groups exposed to higher-
chlorinated PCDDs (i.e., the hexachlorinated, heptachlorinated, and octachlorinated isomers),
occurring primarily in the manufacture and use of pentachlorophenol (PCP) and in the paper
and pulp industries.
A major weakness in nearly all of these studies is the lack of good exposure
information. Most studies rely solely on interviews and questionnaires of work history to
ascertain exposure surrogates. There is little, if any, verification of actual internal dose to
these compounds. Some studies use chloracne as a surrogate for exposure to TCDD. Three
of the recent cohort studies of production workers (Fingerhut et al., 1991; Manz et al., 1991;
Zober et al., 1990) do provide estimates of TCDD exposure in cohort samples via serum
blood levels taken decades after cessation of exposure. These can be used to determine
possible dose-response trends and estimate the risk of cancer to populations with low-level
exposure to TCDD (see Chapter 8). Measures of exposure by individual study will be
discussed.
At the time of EPA's last review in 1988, evidence of human carcinogenicity of
TCDD and the phenoxy herbicides focused on soft tissue sarcomas (STSs) and malignant
lymphomas. Consequently, this report will update and analyze the evidence pertaining to
these cancers. But also included are evaluations of the evidence of cancer at other sites.
The case-control studies reviewed for EPA's last analysis generally considered herbicide
applicators with potential exposures to both 2,4-D and 2,4,5-T. Recent case-control studies
of U.S. farmer groups in which exposure to 2,4-D and 2,4,5-T can be separated (Hoar et al.,
1986; Zahm et al., 1990; Cantor et al., 1992) provide a validation mechanism to separate
potential effects of these herbicides and, possibly, their different PCDD contaminants. Thus,
these and other recent studies (Hardell and Eriksson, 1988; Eriksson et al., 1990; Woods et
al., 1987) will be reviewed and compared with those discussed in EPA's earlier reports.
Four recent cohort mortality studies (Fingerhut et al., 1991; Manz et al., 1991; Zober
et al., 1990; Saracci et al., 1991) totaling over 23,000 workers potentially exposed to TCDD
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and/or phenoxy herbicides/chlorophenols provide a new important database for analyzing
cancer effects, and these will be reviewed first. The first three of these, and especially the
large U.S. study of Fingerhut et al. (1991), are considered to be the most important new
studies in the field of TCDD cancer epidemiology because of their attention to cohort
selection and to TCDD exposures or exposure surrogates (chloracne). The fourth study
(Saracci et al., 1991), while having the largest cohort, has less information on the TCDD-
exposed subcohort and no information that would allow a quantitative estimate of exposure.
Two other studies of phenoxy acid manufacturers are included (Lynge, 1985, 1993; Coggon
•
et al., 1986), but their usefulness is limited due to the low unsubstantiated exposure to
TCDD. Also included is a study of occupationally exposed women (Kogevinas et al., 1993)
who had probable exposure to TCDD.
Three recent cohort studies of workers in the pulp and paper mill industry are
included (Robinson et al., 1986; Jappinen et al., 1987; Henneberger et al., 1989) because of
potential for worker exposure to higher-chlorinated PCDDs, but not hexa-, hepta-, or
octaphenoxy herbicides. However, none of these studies provide any additional information
about which PCDD exposures were likely, and these studies are not given much weight.
Studies of Vietnam veterans potentially exposed to TCDD in Agent Orange are
reviewed briefly, with only one (Michalek et al., 1990) judged to have sufficient information
on potential TCDD exposure to be useful for analysis. Also, the recent studies of the
Seveso, Italy, residents are discussed (Bertazzi et al., 1992, 1993, 1989a, b; Mocarelli et al.,
1991; Pesatori et al., 1992); these studies provide some exposure data, but the cancer
response analysis is limited due to inadequate follow-up time.
Finally, the studies of the rice oil poisonings of residents in Taiwan and Japan with
polychlorinated biphenyl (PCB) and PCDF contaminants are reviewed (Kuratsune et al.,
1988, 1975; Chen et al., 1980; Koda and Masuda, 1975; Rogan et al., 1988). Even though
these poisoned oils did not contain TCDD, they did contain many TCDD-like congeners
currently considered by EPA to have carcinogenic potential, which can be compared to
TCDD. Also, certain dioxin-like PCBs are suspected carcinogens based on their receptor-
binding characteristics and on animal studies.
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Only follow-up and case-control studies are considered in this review. Case reports,
other clinical observations, and prevalence surveys are excluded. The review is restricted to
studies that have been published in full and that are available in the open scientific literature.
Prepublication reports and studies published only in abbreviated form (as well as abstracts or
letters to editors) are included only where they supplement the published articles. These
restrictions limit the review to studies that have received at least a minimum of peer review
and that have been described fully enough to permit a thorough assessment of materials,
methods, and results.
*
7.3. PREVIOUS EPA REVIEWS
In the Health Assessment Document for Polychlorinated Dibenzo-/?-Dioxins, dated
September 1985 (U.S. EPA, 1985), the majority of the epidemiology studies pertained to
groups of herbicide applicators with potential exposure to phenoxy acids and/or
chlorophenols. In that report, the analysis emphasized case-control studies of soft tissue
sarcomas and non-Hodgkin's lymphoma (NHL). That report concluded that the
epidemiologic research available at that time provided "limited evidence for the
carcinogenicity of phenoxy acids and/or chlorophenols in humans. However, with respect to
the dioxin impurities contained therein, the evidence for the human carcinogenicity for
2,3,7,8-TCDD based on the epidemiologic studies is only suggestive because of the difficulty
of evaluating the risk of 2,3,7,8-TCDD exposure in the presence of the confounding effects
of phenoxy acids and/or chlorophenols." In its next report, the review draft dated June 1988
of A Cancer Risk-Specific Dose Estimate for 2,3,7,8-TCDD (U.S. EPA, 1988), the focus
was essentially the same, and EPA concluded that "the human evidence supporting an
association between exposure to 2,3,7,8-TCDD and cancer is considered inadequate."
7.4. REVIEW METHODS
This review will follow the spirit of the EPA Risk Assessment Guidelines of 1986
(U.S. EPA, 1987) by considering alternative explanations for results observed in
epidemiologic studies. These explanations fall into the general categories of causality,
chance, bias, and confounding. The basic approach is akin to a process of elimination, by
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which one attempts to determine the direction and to quantify the magnitude of the influence
that chance, bias, and confounding may have had on the results of each study. Wherever
possible, the results of all studies will be reported in units of relative risk estimates and 95 %
confidence limits.
Most biases in epidemiologic studies can be placed into one of two categories: biases
of classification and biases of selection. Classification biases can result from inaccurate
ascertainment of exposure, disease, or confounders. Selection biases can result from
nonrepresentative sampling of populations, as in the selection of controls in case-control
studies, or from incomplete participation by study subjects. Any bias gains tenability as an
explanation for an observed result if empirical evidence can be adduced to buttress the mere
suggestion that the bias might have occurred. Only those biases considered to be potentially
important will be addressed explicitly in this review.
When imprecise exposure estimates are available, such as with much of the
epidemiologic data on dioxin, estimates of risk can be potentially biased toward the null.
Misclassification, if random, could potentially lead to a masking of a true effect.
Classification biases are of two kinds: differential versus nondifferential. Differential
misclassification will lead to either an exaggeration or an underestimation of an effect.
Nondifferential misclassification occurs when the exposure or disease classification is
incorrect in a portion of the subjects (cases or controls). This type of bias is generally
toward the null, and the risk estimate may reflect this. This can happen when some subjects
are classified as having exposure to dioxin when they really were not exposed. Similarly,
some actually may have had exposure but were classified as not having had it. In studies
where few or no effects were seen, researchers must seriously consider the problem of
nondifferential misclassification. This can be the reason why nonpositive risk estimates or
even disparate risk estimates are seen from different studies of the same effect. On the other
hand, in studies with significant results, nondifferential misclassification is not likely to be a
cause of a significant finding.
Recall bias may produce the opposite effect. Persons with a disease may tend to
remember exposure to a substance better when it is known to them that such exposure might
be associated with the disease. This could potentially lead to inflated risk estimates. In fact,
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it has been suggested that such biases are present in many of the case-control studies on
dioxin. The Swedish studies by Hardell and colleagues have been particularly singled out for
criticism in this respect. That some recall bias may be present is confirmed in a later case-
control study (Hardell and Eriksson, 1988) in which some reduction of risk estimates was
produced by the use of cancer controls.
Confounding bias is a tenable explanation for an association between an exposure and
a disease if the hypothetical confounder can be named and if a good case can be made that it
is a cause of the disease, that it was associated with the exposure in the study population, and
that it was not adequately controlled in the study design or data analysis. This review will
explicitly mention only those potential confounders that meet all of these criteria.
As stressed in previous EPA reviews (1985, 1988), concomitant exposures present a
special problem of potential confounding in the literature on TCDD and related chemicals.
As a noteworthy example, an association between 2,4,5-T exposure and a given cancer, if
causal, could be due to 2,4,5-T itself, to TCDD, or to some other contaminant. The
problem multiplies when it is recognized that, historically, many phenoxy acid herbicide
preparations were mixtures of 2,4,5-T and 2,4-D and that many persons who manufactured,
processed, and used these preparations were exposed to other chemicals as well.
Nevertheless, it may be possible by examining studies of persons exposed to different
combinations of chemicals to identify "threads" of commonality and differences in the
results, especially when specific cancers are considered separately.
Publication bias, sometimes considered a form of selection bias, is the tendency for
the results of a study to influence a judgment as to whether or not it will be published. The
direction and magnitude of publication bias is difficult, if not impossible, to quantify. It is
expected to be a much greater problem in literature reviews and in studies relying on existing
records than in original research in which substantial resources are devoted to collection of
data of relatively high quality. The level of effort required for such studies creates a strong
incentive to publish the results. There is a tendency to publish studies with positive results
as opposed to studies with nonpositive findings.
Strength of association, as measured by the magnitude of the estimated relative risk,
is an important feature of a study's results. The stronger the association, the stronger a bias
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or confounding factor would have to be to explain it. Because questions of bias and
confounding are study-specific, no defensible criteria can be set up in advance to place
relative risk values into categories of strength of association.
Trends in increased risk by degree of exposure and by time since first exposure
(latency) are also important. Different hypothetical causal mechanisms might predict
different exposure-response and latency patterns. Hypotheses of steadily increasing effect
with increasing exposure (i.e., monotonic exposure-response functions) and hypotheses of
effects early in the carcinogenic process (e.g., for factors that operate at the initiation stage)
predict that increases in risk will be greatest among persons with relatively high degrees of
exposure and after relatively long latency periods. Hypotheses of tumor promotion and/or
initiation will be discussed in the appropriate sections.
Replication of results is important in all scientific research. When several studies
have shown a positive association of effect with the same exposure but were conducted under
different circumstances, the possibility that an unknown confounder or chance produced the
observed elevated effect is minimized. When different investigators working with different
populations using different methods confirm an original finding, the results are more
believable.
The statistical aggregation of results from different studies (meta-analysis) has become
a popular feature of epidemiologic literature reviews. In this review, results from separate
studies are aggregated only when all key methodologic features and results are reasonably
similar. The method of aggregation used here is to take the ratio of the sum of the cause-
specific observed deaths to the sum of the cause-specific expected deaths for the individual
studies. Because investigators recognize the value of varying their methods to test
methodologic hypotheses, and because results often differ appreciably, aggregation of results
is not often indicated and is done here with caution.
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7.5. FOLLOW-UP STUDIES OF CHEMICAL MANUFACTURING AND
PROCESSING WORKERS
7.5.1. United States
Fingerhut and colleagues (1990, 1991) reported a study of 5,172 males who had
worked at 12 plants in the United States in the production of chemicals contaminated with
TCDD. Five thousand of the cohort members (97%) were identified in company records as
having been "assigned to a production or maintenance job in a process involving TCDD
contamination" (Fingerhut et al., 1991). The remaining 172 cohort members were
"identified in a previously published study on the basis of exposure to TCDD" (Fingerhut et
al., 1991). This cohort subsumed, and thereby supplanted, company-specific cohorts from
Dow Chemical USA (Ott et al., 1987; Cook, 1981) and the Monsanto Company (Zack and
Gaffey, 1983; Zack and Suskind, 1980) that had been the subject of previous reports. This
study was initiated in 1978 to determine whether health effects were apparent in humans who
were exposed to 2,4,5-T. In 1978, toxicological, teratogenic, and carcinogenic effects data
were released that indicated a cancer effect in animals. There was a concern about the
potential effects of exposure to Agent Orange on Vietnam veterans and workers who
produced products that were contaminated with dioxin (Fingerhut et al., 1992). Follow-up
began in 1940 or on the date of the "first systematically documented assignment to a process
involving TCDD contamination" (Fingerhut et al., 1991), whichever was later, and closed at
the end of 1987. Comparisons were made with the United States population.
The authors stated that approximately 13% of the cohort of 5,172 workers had
records of chloracne. The presence of chloracne in a group of people is an indicator of
relatively intense exposure to TCDD. It can be caused by higher-chlorinated PCDDs,
PCDFs, and PCBs as well (O'Malley et al., 1990). It is a highly specific indicator of
exposure because it virtually never occurs among unexposed persons. It is a nonsensitive
indicator, however, because many highly exposed persons do not develop it (Manz et al.,
1991; Caramaschi et al., 1981; Mocarelli et al., 1991). Exposure to other dioxin-like
chemicals that can be found in the workplace may produce a form of chloracne that could be
indistinguishable from that produced by dioxin (Ott et al., 1993). These chemicals can be
dioxin-like in their effects and act through the aryl hydroxylase (Ah) receptor.
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Although all members of the cohort had specific assignments to TCDD exposure areas
in common, exposures to multiple chemicals are the rule rather than the exception in the
chemical industry. At one plant, for instance, considerable overlap existed among persons
involved in the production of chlorophenols, 2,4,5-T, and 2,4-D (Ott et al., 1987; Bond
et al., 1988, 1989a), and who were thus exposed to TCDD and higher-chlorinated and lower-
chlorinated PCDDs. Presumably, many persons throughout the cohort had contact with
substantial numbers of other chemicals. Although comprehensive surveys of chemical
exposures were conducted in plant-specific cohorts, the results are not available.
Special attention was paid to results for the 3,036 workers who were followed for at
least 20 years after first exposure. This group was again divided into those with less than 1
year (N=l,516) and those with more than 1 year (1,520) of exposure (referred to below as
the long duration/latency subcohort). One year was chosen as the criterion for duration of
exposure because an analysis of 253 workers from 2 plants showed that every worker with 1
or more years of exposure had a lipid-adjusted serum TCDD level greater than the mean
value (7 ppt) in a comparison group of unexposed workers (Fingerhut et al., 1990).
Although the average level for all 253 workers was 233 ppt, the average increased to 418 ppt
in those 119 who were exposed for 1 or more years (Fingerhut et al., 1991). The
researchers described a plan to replace this duration-based exposure scale by using "a dioxin
exposure matrix constructed from historic process descriptions, analytic measurements of
TCDD and industrial hygiene data ... to develop the relative ranking of workers exposed to
TCDD" (Fingerhut etal., 1990).
The cohort as a whole experienced an estimated 15% (95% CI=1.0-1.3) elevation of
mortality from all cancers combined, with a 46% elevation (95% CI=1.2-1.8) among those
in the long duration/latency subcohort (Table 7-1). An excess of deaths from cancers of
connective and soft tissues (STSs) was apparent in the total cohort (RR=3.4, CI=0.9-8.6)
and in the long duration/latency subcohort (RR=9.2, CI= 1.9-27.0), but these results were
based on only four deaths and three deaths, respectively, from two different plants. A 40%
overall elevation in deaths (CI =0.7-2.5) from non-Hodgkin's lymphoma was confined to
workers in the total cohort and was not seen in the long duration/latency subcohort. Results
for Hodgkin's disease were highly imprecise, based on only three deaths (vs. 2.5 expected)
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Table 7-1. Relative Risks of Selected Cancers in Study of Chemical Manufacturing Workers Exposed to TCDD in United States,
by Exposure Duration and Latency
Latency ^20 years
Cancer
Connective and
soft tissues
Hodgkin's
disease
Non-Hodgkin's
lymphomas
Lung cancer
Stomach cancer
All combined
Measure* Latency <20 years
Observed deaths
Relative risk
95 % confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95 % confidence interval
Observed deaths
Relative risk
95 % confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95% confidence interval
1
1.4
0.1 -7.0
2
1.1
0.2 - 3.5
6
1.6
0.7 - 3.4
32
0.9
0.7 - 1.3
3
0.6
1.5- 1.6
103
1.05
0.8 - 1.2
Exposure < 1 year
0
0.02
0.0 - 15.0
0
0.0
0.0 - 15.0
2
1.5
0.2 - 4.9
17
1.0
0.6 - 1.5
3
1.8
0.4 - 5.2
48
1.02
0.76 - 1.4
Exposure ^1 year
3
9.22
1.90 - 27.0
1
2.8
0.1 - 15.3
2
0.9
0.1 -3.30
40
1.4
1.0- 1.9
4
1.4
0.4 - 3.5
114
1.46
1.2- 1.8
Total cohort
4
3.4
0.9 - 8.6
3
1.2
0.3 - 3.3
10
1.4
0.7 - 2.5
89
1.1
0.9 - 1.4
10
1.0
0.5 - 1.9
265
1.15
1.0- 1.3
I
O
90
n
H
m
"Relative risks and confidence intervals are based on rounded values and may differ slightly from those in the original reports (Fingerhut
etal.,
1990, 1991).
Source: Fingerhut et al., 1991.
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in the total cohort. Lung cancer was elevated by 10% overall but by 40% (CI=1.0-1.9) in
the long duration/latency subcohort. A similar 40% excess of stomach cancer (CI=0.4-3.5)
in this subcohort was based on only four deaths; no excess was seen in the total cohort.
The investigators conducted a special study of connective and soft tissue cancers. A
review of all available hospital records and tissue specimens failed to confirm the indications
of soft tissue sarcomas on two of the four death certificates that had been assigned to this
cause-of-death category (Fingerhut et al., 1990). The review also provided evidence that two
persons in plant 8 whose deaths had been assigned to other causes of death had actually had
soft tissue sarcomas. Because only the exposed cohort's death certificates were subjected to
detailed review, the analytic comparisons with the United States population were required to
be based strictly on death certificate information. The basic and well-known rule in such
situations is that, absent evidence to the contrary, erroneous information on death certificates
must be considered to have been equally frequent in the two groups being compared. Suruda
et al. (1993), in a study of STS diagnoses in cohorts exposed to dioxins and chlorinated
naphthalenes, found that death certificates are "relatively insensitive" for detecting STS and
that the power of life table analysis to detect excess risks of STS may be reduced compared
with its utility in correctly estimating the risk of other cancers, such as colon cancer or rectal
cancer. However, the correct identification of STS as the underlying cause of death on death
certificates appears to be much better (82%), based on medical records, than first thought
according to the results of this study. Medical records on the remaining 18% could not be
found. If a death certificate gives as an underlying cause of death a soft tissue sarcoma and
it is coded as 171x, it is very likely correctly coded according to the authors. In an earlier
study, this figure was estimated by Percy et al. (1981) to be 55%.
Four of the soft tissue sarcomas that are discussed in Fingerhut's study (two are
included in Fingerhut's life table analysis while two others are discussed but did not qualify
for inclusion in the life table analysis) are actually from the Nitro, West Virginia, plant. The
remaining two cases that were included in Fingerhut's life table analysis are from a different
plant in the study. One of these cases also suffered chloracne. The four cases from the
Nitro, West Virginia, plant are also the subject of a later study by Collins et al. (1993). The
two cases included in Fingerhut's life table analysis had previously suffered chloracne from
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the 1949 accident (Zack and Gaffey, 1983) in which 121 workers developed chloracne as a
result of a trichlorophenol process accident on March 8, 1949. Collins included one more
STS not previously discussed by Fingerhut from the same plant. The Collins study is
reviewed later.
If the properly diagnosed STSs were correctly assigned to their appropriate plants and
the two incorrectly diagnosed STSs were removed, there would be one in plant 9 and three in
plant 8. No STSs were found at the other plants. This unusually skewed distribution might
possibly be related to the accident that happened in plant 8 some 44 years earlier. Or
perhaps physicians in Nitro, West Virginia, are more likely to diagnose STS than are
physicians in other areas of the country. Because of the rarity of STS, there was inadequate
statistical power to expect to see STS in the other individual plants. If this cancer has a long
latent period and an excess risk is associated with exposure to 2,3,7,8-TCDD, then it might
be several more years before this cohort will produce STSs at the other facilities that are part
of the study.
In any case, this nondifferential misdiagnosis can potentially bias risk estimates
downward.
Cases of soft tissue sarcoma include a diverse group of histological entities. All soft
tissue sarcomas arise only from mesenchymal tissue and all share common features that make
them alike in their basic intercellular and intracellular composition rather than different in
their morphology and location. Characterizing them as fundamentally different because they
are found in different sites of the body is perhaps inadequate for the determination of the risk
of cancer (Enzinger and Weiss, 1988).
The histological classification of STS is centered on a dozen distinctly different
classes of mesenchymal cells that form six relatively well-defined but widely distributed
organ systems. By considering the growth pattern and cell morphology with an evaluation of
intracellular and extracellular products of the tumor cells, fairly precise histogenetic
classification of soft tissue sarcomas is possible (Hajdu, 1981). For human cancer risk
assessment, all connective tissues developing from the same mesodermal tissue, expressing
the same set of "proto-oncogenes" and surrounded by the same chemical milieu of the
extracellular matrix, are expected to develop cancer following exposure to certain
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carcinogens. The grouping of these end-target organs together is therefore a necessary
measure to evaluate 2,3,7,8-TCDD human carcinogenicity.
Confounding by cigarette smoking must be considered in interpreting the approximate
40% excess of lung cancer deaths in the long duration/latency subcohort (Table 7-2). For
the United States as a whole, the authors (Fingerhut et al., 1990) computed age-adjusted
proportions of 24% never smokers, 19% former smokers, and 57% current smokers in 1965
(roughly midway through the follow-up period). The corresponding proportions were 28%
never smokers, 14% former smokers, and 59% current smokers among the 87 workers from
the study of serum TCDD levels who were members of the long duration/latency subcohort
as well. Assuming relative risks of lung cancer of 4.7 for former smokers and 10.9 for
current smokers, the authors used a standard technique (Axelson and Steenland, 1988) to
adjust the number of expected lung cancer deaths and found essentially no change in the
results. It should be kept in mind that the sample of smoking histories was taken from only
2 plants, numbers 1 and 2, but the excess in lung cancer risk was chiefly in plants 8 and 10.
The generalization of smoking habits of employees in 2 of the 12 participating plants to that
of the entire cohort may not be representative of the true smoking impact of the risk of lung
cancer to the entire cohort. Most of the lung cancers came from 3 facilities (56 of 89
observed lung cancer deaths), plant numbers 8, 9, and 10. The remaining 7 plants
contributed the remaining 33 lung cancers to the total due to the small sizes of the respective
subcohorts and insufficient latency. It then would be possible to evaluate the effect of
smoking on the risk of lung cancer at each individual plant. It should be remembered that
national U.S. rates were used to derive expected deaths in each of the 12 plants. It is
possible that local or regional rates at the locality of each of the plants may be a more
appropriate comparison population for the cancer sites examined by the authors, although
local rates may be unstable. In addition, if biomonitoring could be extended to the remaining
10 plants, a better idea could be derived concerning the dose levels that are associated with
those plants experiencing higher lung cancer rates. The authors point out that deaths from
other diseases associated with smoking such as diseases of the heart and circulatory system
were either not increased or significantly decreased in this cohort (Fingerhut et al., 1990).
This led the authors to conclude that "cancers of the respiratory tract . . . may result from
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Table 7-2. Relative Risks of Lung Cancer in Subcohort of Chemical Manufacturing Workers Exposed to TCDD in United States
for at Least 1 Year and With at Least 20 Years Latency, Adjusted for Alternative Hypotheses About Its Smoking Distribution
Never
24
28
25
20
15
10
15
10
Proportion of subcohort
Former
smokers smokers
19
14
10
15
15
20
10
15
(percent)3
Current
smokers
57
59
65
65
70
70
75
75
Expected lung
cancer deaths
(40 observed)
28.8
29.2
30.5
31.2
33.2
33.9
34.4
35.1
Relative risk
1.4
1.4
1.3
1.3
1.2
1.2
1.2
1.1
95 % confidence interval
1.0-
1.0-
0.9 -
0.9 -
0.9 -
0.9-
0.8 -
0.8 -
1.9
1.8
1.8
1.7
1.6
1.6
1.6
1.5
o
o
o
m
o
?o
o
"The first set of proportions assumes no difference between the subcohort and the United States population. The second set is based
on 87 surviving members of the subcohort (Fingerhut et al., 1990). The remaining sets are hypothetical values used to test the
sensitivity of the results. Relative risks of lung cancer are assumed to be 4.7 for former smokers and 10.9 for current smokers
(Fingerhut et al., 1990).
Source: Fingerhut et al., 1991.
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exposure to TCDD, although we cannot exclude the possible contribution of factors such as
smoking and occupational exposure to other chemicals" (Fingerhut et al., 1991).
One possible explanation for the increase follows from the fact that the comparison of
smoking habits between the United States population and the 87 surviving members of the
long duration/latency subcohort did not show the well-known tendency for smoking to be
more common among blue-collar workers than in the general population. A possible
explanation for this result is that because smoking appreciably elevates the overall death rate,
fewer and fewer smokers will remain in a fixed group of persons as time goes by. Thus, the
use of the 87 surviving members of the long duration/latency subcohort may have
underestimated the proportions of former and current smokers in the subcohort as a whole
over the course of mortality follow-up. However, this same phenomenon is also present in
the population from which expected deaths were generated, so the effect is probably
nullified. As shown in Table 7-2, the actual proportions would not have to be inordinately
high for smoking to have exerted appreciable confounding on the estimated relative risks of
lung cancer. The lack of increased mortality from cardiovascular diseases as well as cancer
of the buccal cavity and pharynx in this cohort, however, makes this explanation less likely.
Furthermore, as the authors point out, mortality from nonmalignant respiratory disease
(standard mortality ratio [SMR] = 96), which is often associated with smoking, was less
than expected.
On the other hand, the authors report a correlation coefficient of 0.72 between length
of exposure and serum TCDD tissue level. This negative bias is potentially much greater
than the positive bias that could possibly be produced by smoking. And because
nondifferential bias is the only type of bias that could occur in the Fingerhut study, risk
estimates are more than likely to be lower than the true risk. The only way that dioxin
exposure could have a positive bias is if it prevents cancer and the exposure classification is
100% wrong.
The authors of this study also report that two of these cancer deaths were of
mesothelioma, a finding that indicates more than likely exposure to asbestos. Although these
two mesotheliomas occurred at plants 9 and 12, it is not known whether these persons were
exposed in the current job or perhaps in some previous job. If exposure occurred throughout
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this cohort, it is possible that asbestos also could be a confounding factor for lung cancer.
The combination of smoking and exposure to asbestos fiber might have a synergistic
confounding effect on the risk of lung cancer.
The positive association of lung cancer in male workers observed in this study is also
consistent with an excess of pulmonary tumors found in male mice and rats exposed to
TCDD (see Chapter 6). These same animal data suggest the possibility of a protective
hormonal effect from TCDD and the risk of pulmonary cancer in female rats. Because this
study dealt with only male workers, this hypothesis could not be verified in female workers.
On the other hand, no elevated risk of liver cancer is evident in male workers even in the
long duration/latency subcohort. This is also consistent with rat data where the tumors were
only observed in female rats. If there is a promoting effect on liver cancer in females due to
hormonal effects as suggested by the rat studies, it could not be verified.
In a recently published paper (Collins et al., 1993), it is suggested that there is an
association of STS with exposure to 4-aminobiphenyl. The authors reported that workers
who developed chloracne from an accident in which a chemical mix containing TCDD was
scattered throughout a 2,4,5-trichlorophenol plant had increased mortality from soft tissue
sarcoma, bladder cancer, and respiratory cancer. All individuals who were identified in this
study of 754 chemical employees as having STSs or lung cancers, and who were employed at
the time of the 1949 accident, were potentially exposed to 4-aminobiphenyl as well as to
2,3,7,8-TCDD. However, it is not known how many were actually working inside the plant
when the accident occurred. 4-Aminobiphenyl is thought to be a bladder carcinogen from
previous studies. However, this chemical has not been shown to be associated with STS or
lung cancer in humans. Unfortunately, no tissue measurements are available to substantiate
exposure to 2,3,7,8-TCDD or exposure to 4-aminobiphenyl. It is also of interest that no
significant increase in STSs was noted in the "chloracne-free" subgroup exposed to 4-
aminobiphenyl. However, the authors report that an additional 106 persons also had
indications of chloracne-type conditions noted in their medical files, presumably as a result of
exposure to 2,3,7,8-TCDD and not as a result of the accident. Although these individuals
probably were heavily exposed to dioxin as well, they were included in the "no chloracne"
subgroup for the purpose of analysis. It would be of some interest to see how much of an
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effect would occur to the risk estimates if they were included in the "chloracne" subgroup.
Furthermore, plant employees who left work before March 8, 1949, or began work after
November 22, 1949, were not included even if they had received exposure to dioxin or even
developed chloracne as a result of exposure. The major interest according to the authors was
in the 122 workers who developed chloracne from the 1949 accident. The authors point out
that the numbers are small and that confounding factors, such as misclassification of
exposure, cannot be ruled out. This study presents an interesting theory that needs to be
investigated further. However, at this time, it has not been substantiated anywhere.
It was also noted by Collins that toxicological evidence is available that supports the
idea that STSs (i.e., angiosarcomas) have resulted from exposure to 4-aminobiphenyl
(Schieferstein et al., 1985). Angiosarcomas and bladder cancer in females were found to be
dose-related with oral consumption of 4-aminobiphenyl in drinking water. However,
angiosarcomas of the liver arise mainly from the endothelial lining of blood vessels. This
type of tissue is more likely to be susceptible to a hydrophilic carcinogen such as 4-
aminobiphenyl during its passage through the blood vessel. Hydrophobic carcinogens such
as 2,3,7,8-TCDD might be expected to exert an influence on mesenchymal tissue from which
most STSs arise, i.e., fibrosarcoma, histiocytoma, liposarcoma, leiomyosarcoma,
rhabdomyosarcoma, synovial sarcoma, schwannoma, myxoid neurogenic sarcoma, and
others. Therefore, the author's assumption that 4-aminobiphenyl can cause other types of
STSs remains unproven and highly unlikely.
7.5.2. Germany
Manz and colleagues (1991) reported a study of 1,583 persons (1,184 men and 399
women) employed at a German chemical manufacturing facility that produced 2,4,5-T and its
precursor, 2,4,5-trichlorophenol. In 1954, a chloracne outbreak had occurred in the working
population of the plant, and after that, production of the TCDD contaminant was reduced.
Cohort members worked at least 3 months from 1952 through 1984. The start of follow-up
was not stated in the report, but presumably began on the date of accumulation of 3 months
of employment. The follow-up period closed at the end of 1989. The cohort's mortality
experience was compared with that of the West German population and with that of a cohort
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of workers at a gas supply company. Because the results did not differ materially between
the analyses, and because limited data on the gas workers forced that comparison to be based
on a subset of the TCDD-exposed cohort, only the results of the comparisons with West
Germany are reported here.
The cohort was postdivided by duration of employment and by a three-category
exposure scale based on TCDD measurements "in nonsystematic samples of precursor
materials, products, waste, and soil from the grounds of the plant, mainly after the plant had
closed in 1984" (Manz et al., 1991). This scale was validated to some extent by adipose
tissue TCDD levels in 48 volunteers (mean = 296 ng/kg in 37 persons from the highest
group, 83 ng/kg in 11 persons from the other two groups). Based on these results, the low
and intermediate groups are combined for the present analysis.
For the males, this study, with 75 total cancer deaths expected and 24 expected in the
high-exposure subcohort, was considerably smaller than the study by Fingerhut et al. (1991),
which had 230 total cancer deaths expected and 78 in its long duration/latency subcohort
(Table 7-3). Manz et al. presented detailed analyses only for all cancers combined. The
high-exposure subcohort, and especially those with longer employment duration, experienced
an excess of total cancer deaths (RR = 1.4, CI = 1.0-2.0 for the high-exposure group and
RR=2.6, CI=1.2-4.9 for the high exposed/long duration subcohort) (Table 7-3). The
authors concluded that "the increase in (total) cancer risk of 1.24-1.39 . . . cannot be
explained completely by confounding factors, and ... is associated with exposure to TCDD"
(Manz et al., 1991). In a later abstract of an update of this same paper, Dwyer (1992)
reported that using a Cox regression analysis of nine major areas of employment within the
plant, the area of work with the strongest relative risk for cancer mortality was found to be
in 2,4,5-T production (RR=2.7, CI = 1.7-4.2). These findings are similar to those of
Fingerhut et al. (1991).
For the cohort as a whole, the estimated relative risk of lung cancer was 1.4
(CI = 1.0-2.0, 30 observed deaths). Smoking as an explanation for the observed increase in
lung cancer mortality is less likely because a comparison using the gas worker reference
actually leads to an increased RR of 1.7 (CI = l.l-2.4). Although smoking histories were not
available for the entire Boehringer cohort, of the 361 men, 73% reported that they smoked.
7_18 06/30/94
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Table 7-3. Relative Risks of All Cancers Combined in Study of Chemical Manufacturing Workers Exposed to TCDD in Germany,
by Duration and Category of Exposure
Exposure duration
<20 years
>20 years
Total
Measure
Observed deaths
Relative risk
95 % confidence interval
Observed deaths
Relative risk
95 % confidence interval
Observed deaths
Relative risk
95 % confidence interval
Exposure category ('median
Low and medium (60 ng/kg)
49
1.1
0.8 - 1.4
10
1.5
0.8 -2.7
59
1.2
0.9 - 1.5
adipose TCDD level)
High (137 ng/kg)
26
1.2
0.8- 1.8
8
2.6
1.2-4.9
34
1.4
1.0-2.0
Total
75
1.1
0.9 - 1.4
18
1.9
1.1 -2.9
93
1.2
1.0- 1.5
6
o
O
g
n
t-H
H
w
Source: Manz et al., 1991.
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Similarly, 76% of 2,860 gas workers smoked. Substantial confounding based on smoking
does not appear to be present because smoking seems to be similar in both plants. The
estimate for stomach cancer was 1.2 (CI=0.7-2.1, 12 observed deaths). Three deaths from
non-Hodgkin's lymphomas and no connective and soft tissue cancer deaths were observed.
(The authors described an additional three deaths from chronic lymphocytic leukemia as non-
Hodgkin's lymphoma deaths, but these deaths would not have been classified as non-
Hodgkin's lymphomas in the other studies in this review.) Expected numbers of deaths from
these cancers were not given. Based on the proportions of expected cancer deaths in the
Fingerhut study, one might estimate that approximately 2.4 non-Hodgkin's lymphoma deaths
and 0.4 connective and soft tissue cancer deaths would have been expected in this cohort as a
whole and about 0.1 connective and soft tissue cancer deaths in the high-exposure subcohort.
(The numbers of expected deaths from lung cancer, stomach cancer, or non-Hodgkin's
lymphoma in the high-exposure subcohort were not estimated because information is lacking
on how many of the observed deaths from these cancers were in that subcohort.)
The authors reported exposure to other industrial chemicals, such as benzene and
dimethylsulphite. In addition, manual laborers were exposed to asbestos "probably" to some
extent. However, the authors maintain that this exposure explains neither the increased
mortality from all cancers nor the patterns of associations with TCDD exposure groups.
Other possible sources of bias include a potential lack of comparability between cause
of death ascertainment based on medical records in some, perhaps many, cohort members
versus use of death certificates only for cause of death certification in the derivation of
German national death rates. This is somewhat alleviated by the use of gas workers as a
second comparison group. In these workers, the same methods were used for medical
certification, making comparison of cause of death somewhat more comparable. However,
this is offset by the fact that the gas workers may have somewhat better mortality experience
because they had to work a minimum of 10 years in order to obtain entrance into the gas
workers cohort, whereas the dioxin cohort had to work only a minimum of 3 months. This
could have introduced survivorship bias in this group and consequently lower mortality and
higher risk estimates when compared with dioxin workers.
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Furthermore, the lack of an analysis of the mortality data by time since first exposure
for individual causes such as lung cancer makes it impossible to assess latent effects.
Of the 399 female cohort members, only 7% worked in high-exposure departments.
In total, there were 54 deaths and an overall RR=0.8 (CI=0.6-1.0). The RR for all cancers
was 0.9 (CI=0.6-1.4), but the RR=2.2 (CI=1.0-4.1) for breast cancer was significantly
increased based on 9 deaths. This is an interesting result in view of a suggestion of reduced
mammary cancer based on mechanistic studies and animal bioassays. However, at this point
the data do not provide a sufficient basis for any conclusions. Of all the worker follow-up
studies with TCDD, this is the only one that reports on a cohort of females.
In another investigation in Germany, Zober and colleagues (1990) studied persons
employed at a German chemical manufacturing facility where 2,4,5-trichlorophenol was
produced. An uncontrolled decomposition reaction in 1953 and subsequent cleanup activities
resulted in substantial TCDD exposures. The cohort contained 247 persons who had worked
at the plant from 1953 through 1987, 51% of whom had developed chloracne or erythema (a
skin condition suggestive of chloracne), with mortality follow-up covering the same calendar
period. Seventy-eight persons had died, RR=0.95; 23 had died of cancer, RR=1.2
(CI=0.8-1.7). Expected deaths were based on national mortality rates in the Federal
Republic of Germany. When workers with chloracne were looked at separately, the risk of
cancer as expressed by the SMR rose to 1.4 (CI=0.9-2.1). Again, within this highly
exposed subgroup, if the analysis is restricted to only those workers who were observed 20
or more years after first employment, the SMR was significant at 2.0 (CI=1.2-3.2). For
lung cancer, the SMR is of borderline significance at 2.5 (CI= 1.0-5.3). The authors report
that the results ". . .do not support a strong association between cancer mortality and
TCDD, but they do suggest that some hazard may have been produced."
Three subcohorts were defined on the basis of procedures by which cohort members
were identified as having potential for varying degrees of exposure. Subcohort Cl contained
69 persons known to be exposed to TCDD during the accident period. Cohorts C2 (84
persons) and C3 (94 persons) contained workers thought to be exposed to lesser amounts of
TCDD. Recent TCDD levels in blood samples from small numbers of persons in each group
suggested that exposures had been higher in Cl (median 24.5 ppt, 11 samples) than in C2
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(median 9.5 ppt, 7 samples) or C3 (median 8.4 ppt, 10 samples). Thus, C2 and C3 are
grouped together in this review. A separate analysis by the authors divides the group into
the 127 persons with chloracne (N=114) and erythema (N=13) versus those 120 persons
with neither. The average serum TCDD levels in the two subcohorts are 15 ppt and 5.8 ppt,
based on 16 samples (with chloracne) and 12 samples (without chloracne), respectively. The
two stratifications provide similar results. Only the first is presented here. Section 8.5
provides an additional analysis.
This study, with only 20 expected cancer deaths in the total cohort and 4 expected
cancer deaths in the members of subcohort Cl with 20 or more years of latency, is much
smaller than the studies by Fingerhut et al. (1991) and Manz et al. (1991). The authors,
however, did provide detailed analyses of data on specific cancers (Table 7-4). Elevated
mortality rates from lung cancer, stomach cancer, and all cancers combined were confined
largely to the members of subcohort Cl with long latency. The confidence intervals for the
relative risk estimates are extremely wide, however. No deaths from cancers of connective
and soft tissues or from non-Hodgkin's lymphomas were observed, and expected numbers of
deaths from these cancers were not reported. However, one mesothelioma was reported in a
plant supervisor with known asbestos exposure. Based on the proportions of all expected
cancer deaths due to these cancers in the study by Fingerhut et al. (1991), one might estimate
that approximately 0.6 non-Hodgkin's lymphoma deaths and 0.1 connective and soft tissue
cancer deaths would have been expected in this cohort as a whole, and about 0.2 non-
Hodgkin's lymphoma deaths and less than 0.1 connective and soft tissue cancer deaths among
the members of subcohort Cl with long latency. This study lacks power to detect a
significant site-specific cancer risk at most sites due to its small size.
7.5.3. Ten-Country Study by International Agency for Research on Cancer
A historical cohort study of cancer mortality in 18,390 production workers or
sprayers exposed to chlorophenoxy herbicides and/or chlorophenols was reported on by
Saracci et al. (1991). Exposure was reconstructed through questionnaires, factory or
spraying records, and job histories. Workers were classified as exposed (N = 13,482),
probably exposed (N=416), exposure unknown (N=541) and nonexposed (N=3,951). The
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to
o
o\
u3
o
Table 7-4. Relative Risks of Selected Cancers in Study of Chemical Manufacturing Workers Exposed to TCDD in Germany, by
Median Blood TCDD Level and Latency
Subcohort (median
blood TCDD
level) Cancer
Cl (24.5 ppt) Lung
Stomach
All combined
C2 (9.5 ppt) and Lung
C3 (8.4 ppt)
Stomach
All combined
Total cohort Lung
Stomach
All combined
Measure
Observed deaths
Relative risk
95 % confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95 % confidence interval
Observed deaths
Relative risk
95% confidence interval
Observed deaths
Relative risk
95% confidence interval
Time since
<20 years
1
1.2
0.1 -6.2
1
2.0
0.1 -9.7
2
0.7
0.1 -2.4
0
0.0
0.0- 1.8
0
0.0
0.0 - 3.2
5
0.8
0.3 - 1.9
1
0.4
0.0 - 2.0
1
0.7
0.0-3.4
7
0.8
0.4 - 1.6
first exposure
^20 years
3
2.5
0.6 - 6.9
2
4.0
0.7- 13.2
7
1.7
0.7-3.3
2
1.0
0.2- 3.4
0
0.0
0.0-3.9
9
1.3
0.6-2.4
5
1.6
0.6-3.5
2
1.6
0.3 -5.2
16
1.5
0.9 - 2.3
Total
4
2.0
0.6 - 4.8
3
3.0
0.8-8.1
9
1.3
0.6 - 2.4
2
0.5
0.1 - 1.8
0
0.0
0.0- 1.7
14
1.1
0.6- 1.8
6
1.1
0.4 - 2.2
3
1.1
0.3 - 3.0
23
1.2
0.8 - 1.7
Source: Zober et al., 1990.
O
O
2
3
o
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exposed group contains everyone known to have sprayed chlorophenoxy herbicides and
everyone who had worked in any of certain specified departments at factories producing
chlorophenoxy herbicides. The criteria for duration or level of exposure required for
selection was reported for only 3 of the 10 countries and only 4 of the 20 cohorts; these
ranged from at least 1 month to 1 year. For all the other cohorts, the criterion for inclusion
was to have ever been employed in production or spraying of these herbicides. The cohort
contained 1,537 female workers, but results were not presented separately, except for female
breast and genital organ cancers by phenoxy herbicide exposure. Average follow-up for the
cohort was 17 years; 5% of eligible workers were lost to follow-up. Three of the cohorts
comprising over 10,000 workers have also been reported in separate publications but for
different follow-up periods (Lynge, 1985, 1987, 1993; Coggon et al., 1986; Kogevinas et
al., 1993); these are discussed briefly in this report following this discussion.
Also included in the analysis was a division of the cohort (probable vs. unlikely) by
whether or not exposure to TCDD occurred. No definition is given for that which is
considered "probable" exposure to TCDD. Exposure to phenoxy herbicides does not
necessarily imply exposure to TCDD in this study. The "probably exposed" category
includes production workers at two plants producing PCP, 2-(2,4-dichlorophenoxy)propanoic
acid (2,4-DP; dichlorprop), 4-(2,4-dichlorophenoxy)butanoic acid (2,4-DB), (4-chloro-2-
methylphenoxy)acetic acid (MCPA), or 2-(4-chloro-2-methylphenoxy)-propanoic acid
(MCPP; mecoprop). Those "unlikely exposed" appear to be so classified because they
appeared to work in different factories. There were 181 cases of chloracne among workers
in the cohort.
The results are presented below by each of the two divisions of the total cohort: a)
phenoxy herbicide (ph) and/or chlorophenols and b) probable TCDD exposure. For the
cohort division by ph and chlorophenols, no excess was observed for all-cause mortality, for
all malignant neoplasms, for most common epithelial cancers, or for lymphomas. The four
STS deaths were all in the "exposed to phenoxy herbicides and chlorophenols" subcohort
(RR=2.0, Cl=0.5-5.2) and all appeared 10 to 19 years after first exposure (RR=6.1,
CI = 1.6-15.5), with the excess risk limited to exposed sprayers (RR = 8.8, CI = 1.8-25.8)
based on three observed deaths. None were observed in the 20 years or more category.
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Increases were also noted in the exposed group for mortality from thyroid cancer (RR=3.7,
CI= 1.0-9.4) based on four deaths, cancer of the testis (RR=2.2, CI=0.9-4.6) based on
seven deaths, other endocrine glands (RR=4.6, CI=0.9-13.5) based on three deaths, and
nose and nasal cavities (RR=2.9, CI=0.6-8.5) based on three deaths. An increase in lung
cancer mortality was limited to the "probably exposed" (to phenoxy herbicide and
chlorophenols) group (RR=2.2, CI=1.1-4.0) based on 11 observed deaths.
The authors provided an additional analysis of STS, including five additional cases
who were either alive at the end of follow-up or who had died from another cause. They
concluded that the results suggest that STS in these workers "is compatible with a causal role
for chlorophenoxy herbicides, though not specifically for those probably contaminated with
TCDD."
The authors present only a limited analysis based on 215 and 294 expected total
cancer deaths in the "probable" vs. "unlikely" exposed groups, respectively. There was a
slight increase in mortality from all cancers for the probably versus the unlikely exposed
groups (RR=1.1, CI = 1.0-1.2 versus RR=0.9, CI=0.8-1.1), but no increase in either STS
or NHL based on 4 and 11 total cases, respectively. There was also an increased mortality
for testicular cancer in the group probably exposed to TCDD versus those probably not
exposed (RR=3.0 vs. 1.6) based on seven total deaths and for thyroid cancer (RR=4.3 vs.
3.1) based on four total deaths. These latter two differences are not significant and, while
interesting because of TCDD's known effects on these organs, add little to the information
base.
While the Saracci et al. cohort is significantly larger than the other three worker
cohorts (Fingerhut et al., 1991; Manz et al., 1991; Zober et al., 1990), the lack of both a
clear definition of exposure and uniformity of exposure classification between and within
plants makes the results difficult to interpret and lessens the confidence in these results.
When several studies have shown a positive association of effect with the same exposure but
were conducted under different circumstances, the possibility that an unknown confounder or
chance produced the observed elevated effect is minimized. When different investigators
working with different populations using different methods confirm an original finding, the
results are more believable. TCDD tissue levels were available only from a sample of 9 of
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the 181 workers with chloracne (median = 340 ng/kg, range 98 to 659 ng/kg). For 17
external controls, the median was 16 ng/kg while the range was 0 to 23.3 ng/kg. This
suggests that some of the controls were exposed to TCDD. Unfortunately, no further
analysis was presented on these workers.
There are several problems with this study. A portion of the Saracci et al. cohort
consists of Danish workers from the Lynge (1985) study. None of them are reported by
Saracci et al. (1991) as having had any exposure to 2,4,5-T. Lynge indicates that 2,4,5-T
was produced at the Kemisk Vaerk K0ge (KVK) facility in Denmark from 1951 until the end
of 1980. However, in a later update (Lynge, 1993), she maintains that the excess is due to
exposure to phenoxy herbicides other than 2,4,5-T because only 5.3 tons of 2,4,5-T were
produced in 1951-1952. This statement is somewhat contradicted in her methods paper
(Lynge, 1987), where she discloses that 350 tons of 2,4,5-T esters were produced during the
period 1951-1981 based on purchased 2,4,5-T acid. Perhaps there occurred more exposure
to 2,4,5-T than was asserted by the author. This suggests the possibility that exposure
misclassification may be present in the Saracci et al. study. There may be potentially as
many as 3,844 workers who had exposure to 2,4,5-T and consequently 2,3,7,8-TCDD.
Many or all of them were considered as unexposed in the Saracci et al. study.
Lynge in her studies reported on five histologically confirmed cases of soft tissue
sarcoma. These are listed as cases 2, 3, 4, 5, and 9 in Table IV of the Saracci et al. study.
Two are considered alive for the purposes of the Saracci et al. study even though in the later
1993 Lynge study these same two are listed as deceased. The remaining three are reported
by Saracci et al. as deceased. Two of these three are coded to cancer sites other than STS.
Only one is correctly coded to STS. This suggests that underreporting of STS as the
underlying cause of death is a problem in this study, which is consistent with the findings of
Suruda (1993). that STS is underreported generally on death certificates. Added evidence of
underreporting of STS is provided by the death certificate's cause of death for the two who
were deceased after 1984. Both were coded to a cancer site other than STS. Altogether,
four out of five of the confirmed STSs in the Lynge study were coded to causes other than
STS.
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7.5.4. Other Studies
Four studies containing portions of the same cohort reported above were reported
elsewhere (Lynge, 1985, 1987, 1993; Coggon et al., 1986; Kogevinas et al., 1993; Bueno de
Mesquita et al., 1993). Lynge (1985) reported a study of cancer incidence among persons
employed in the manufacture of phenoxy herbicides in Denmark. The cohort consisted of
4,459 persons from two factories. One factory contributed 615 cohort members who had
worked in the years 1951-1981. The only phenoxy acids manufactured and packaged at this
plant were MCPA and mecoprop unlikely to contain TCDD. The other factory, Kemisk
Vaerk K0ge, contributed 3,844 cohort members who had worked in the years 1933-1981. At
this plant, MCPA, 2,4-D, and lesser amounts of mecoprop, dichlorprop, and 2,4,5-T were
manufactured and packaged. The investigators were unable to classify cohort members by
the specific types of phenoxy herbicides to which they were exposed. However, in this
plant, where exposure to TCDD-contaminated 2,4,5-T probably did occur, a significant
excess risk of STS (4 observed vs. 1.00 expected, CI=1.09-10.24) was noted by the author
in those workers who had achieved a minimum 10 years of latency. Unfortunately,
individual tissue measurements of 2,3,7,8-TCDD were not included in this study.
In an update of the earlier study, Lynge continues to report an increase in the risk of
STS with four cases reported to be in persons exposed to phenoxy herbicides (standardized
incidence ratio [SIR] =2.3, CI=0.6-5.8). Just as before, this excess occurred in workers
employed for more than 1 year in the Kemisk Vaerk K0ge factory (SIR=6.4, CI=1.3-18.7).
The author concluded that her study continues to provide evidence that exposure to phenoxy
herbicides increases the risk of STS.
However, Lynge maintains that only small amounts of 2,4-D and "negligible"
amounts of 2,4,5-T were produced at the KVK factory. That this amount of 2,4,5-T was
negligible is somewhat at odds with data from an earlier paper in which she discussed the
design of her ongoing cohort study (Lynge, 1987). In the 1987 paper, she reported that
although 5.3 tons of 2,4,5-trichlorophenol were produced in 1951 and 1952, 350 tons of
2,4,5-T esters were produced from 1951-1981 based on purchased 2,4,5-T acid. This varied
from zero to as much as 63 tons in any one year. Very likely, the term "negligible" is used
in a relative sense—relative to the amounts of the other herbicides produced in the KVK
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factory, which were produced in much larger amounts. Actual exposure to 2,4,5-T may
have been greater than the impression given in Lynge's 1993 study. Most of the potential
exposure was to MCPA, MCPP, 2,4-DP, and various dyes and pigments. MCPA, MCPP,
2,4-DP, and the nondioxin-containing phenoxy herbicides have not heretofore been seriously
thought of as possible causes of cancer in humans.
Lynge also found that the risk of non-Hodgkin's lymphoma was not elevated in
persons potentially exposed to phenoxy herbicides. She did find what she calls a "puzzling"
3.5-fold excess risk in employees of KVK employed in other manufacturing departments.
No detailed information on production in these areas was included in the study.
Little additional information is provided concerning any increased risks of other forms
of cancer, except for a statement that multiple myeloma and cervical cancer in women and
malignant melanoma in men were significantly increased. No numbers are given for these
statements. A significant excess risk of lung cancer seen in the earlier study is now
borderline significant in this study (obs. = 13, SIR=1.6, C.I. =0.9-2.8). The author is now
arranging to have serum tissues in some of her subjects analyzed for their dioxin content.
Coggon and colleagues (1986) conducted a study of 5,754 workers at a British plant
that manufactured and formulated MCPA from 1947 until 1982 and operated its own aerial
and tractor-mounted spraying service from 1947 until 1972. The authors stated that other
phenoxy acids were handled "at times" and that, "in comparison with MCPA, 2,4,5-T was
handled only on a small scale."
Coggon et al. (1991) conducted a study of four British cohorts of manufacturers of
phenoxy herbicides, including 2,4,5-T, comprising 2,239 men employed sometime during the
period 1963 to 1985. All four of these cohorts were included in the Saracci et al. study
previously discussed. Follow-up was to the end of 1987 through the National Health Service
Central Register and the National Insurance Index. Comparisons were with the national
population. Factory A produced 2,4,5-T only beginning in 1968, while the remaining three
factories formulated it only beginning in 1959, 1960, and 1970, respectively. No tissue
measurements were conducted on any members of the cohort. A slight excess of lung cancer
was noted (19 observed, 14.2 expected). Two non-Hodgkin's lymphomas also were
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observed (0.87 expected). No STSs were observed (0.18 expected). Total cancer also was
not increased (37 observed, 36.85 expected).
This cohort has not been followed for a sufficiently long enough time to expect latent
effects to manifest themselves. The authors assumed that the slight increase in lung cancer
was probably due to cigarette smoking or a chance occurrence based on the observation that
most of the lung cancer deaths occurred less than 10 years after first exposure to phenoxy
compounds. Phenoxy herbicides produced or formulated at these factories include 2,4-D,
MCPA, 2,4-DP, 2-methyl-4 chlorophenoxy butyric acid (MCPB), MCPP, phenoxybutyric
acid (PBA), parachlorophenoxyacetic acid (PCPA), and phenoxyacetic acid (PAA). In
addition, other herbicides were also made here. The author says they were exposed to a
multiplicity of chemicals. Except for the slight increase in lung cancer, which is in the same
direction as the findings from the earlier cohort studies, this study contributes little to the
elucidation of the risk of cancer from exposure to 2,3,7,8-TCDD.
Kogevinas and her colleagues (1993) studied a group of 701 occupationally exposed
women who were enrolled in lARC's International Registry of Persons Exposed to Phenoxy
Herbicides and Their Contaminants. These workers were also included in the earlier Saracci
et al. (1991) study. The likelihood of exposure to TCDD was based on individual job
histories, company records, and company exposure questionnaires. Actual measurements of
TCDD serum levels in women were not available according to the authors, so that
confirmation of exposure could not be performed. Both national cancer incidence rates and
national death rates were used to generate expected cases and deaths utilizing the methods of
the Saracci I ARC study.
The overall cancer risk did not exceed expected (SIR=96, CI=0.6-1.4) based on 29
cases. However, the group with the greatest potential for exposure to TCDD-contaminated
chlorophenoxy herbicides produced a significant excess risk of cancer of all sites (SIR=222,
CI = 1.0-4.2) based on nine cases. The risk was observed within the first 10 years of
exposure, with no elevated risk appearing after the 10th year of observation. For those
women who had probable exposure to TCDD, the risk of dying from cancer was slightly
elevated as well (SMR=165, CI=0.4-4.8) based on three deaths.
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This study suffers from many of the same problems as the Saracci et al. (1991) study.
In addition, it is a study of a small population and as such cannot be considered sensitive to
the detection of small risks. These same workers were also subject to exposure to other
toxic chemicals in the workplace, which may also have an effect on the risk of cancer.
However, the elevated cancer risk in women exposed to TCDD-contaminated phenoxy
herbicides is consistent with a hypothesis of overall increased cancer risk seen in other
studies from exposure to TCDD or TCDD-like contaminants.
Recently, another study was published (Bueno de Mesquita et al., 1993) of a cohort
of 2,310 workers in two plants involved in the manufacture and preparation of phenoxy
herbicides (not necessarily 2,4,5-T) in the Netherlands. These workers were also included in
the IARC International Registry of Persons Exposed to Phenoxy Herbicides and Their
Contaminants and hence part of the Sarraci et al. (1991) study. Some 963 were considered
by the author to be exposed to phenoxy herbicides, while 1,111 were considered not
exposed. The follow-up periods were somewhat skewed between the two subcohorts as well.
The workers of one plant were followed from 1955 to 1985, and those at the other were
followed from 1965 to 1986.
Only a slight increase occurred in total cancer mortality based on 31 deaths
(SMR=107, 95% CI=73-152) utilizing The Netherlands' national rates. A slightly higher
risk of total cancer was seen based on 10 deaths (SMR=137, CI=66-252) in 139 workers
probably exposed to dioxins during or immediately after a 1963 industrial accident in which
dioxin was released into the atmosphere.
When compared with nonexposed workers, mortality due to all cancers was
insignificantly elevated (RR = 1.7, 95% CI=0.9-3.4) while that due to respiratory cancer was
also insignificantly elevated (RR=1.7, 95% CI=0.5-6.3). This group was too small to
provide enough power to detect significant site-specific cancers.
Although the size of the cohort seems large, actually only about 549 workers in
Factory A had a potential for exposure to TCDD-contaminated 2,4,5-T and the higher
chlorinated dioxins. No one at Factory B was exposed to 2,4,5-T because it was not
produced there. At Factory A, the SMR for lung cancer was elevated insignificantly to 165
based on 6 deaths in the 20-year latent category. No serum dioxin measurements are
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available to substantiate exposure to dioxin. No soft tissue sarcomas were reported although
only 0.28 were expected. However, the authors did conclude that the SMR of 73 for
Factory A, the SMR of 118 for Factory B, and the SMR of 137 for the cohort exposed to the
accident are not inconsistent with the possibility of a carcinogenic effect of TCDD in
humans.
Wiklund and Holm (1986) studied a massive cohort of 354,620 Swedish men who
were recorded as having an agriculture or forestry job according to the census of 1960 versus
1,725,845 Swedish men in all other industries. The primary exposure in those jobs was
postulated to be MCPA; 2,4-D and 2,4,5-T were also used to a lesser extent. The authors
found that the relative risk of STS was only 0.9. This study has several deficiencies that
reduce its usefulness in determining the risk of STS due to exposure to 2,3,7,8-TCDD: (1) a
lack of individualized exposure data, (2) only 15% of Swedish agricultural and forestry
workers were estimated to be exposed to phenoxyacetic acids and 2% to chlorophenols, (3)
Swedish agricultural workers have a decreased cancer risk and tend to use health services
less frequently, (4) classifying workers according to a 1-week employment status in October
of 1960 as reported in a census invites the possibility of misclassification, and (5) the crude
rate of STS in agricultural and forestry workers based on data in the study is 5.45 per
100,000 person-years, and in the remaining workers it is 5.00 per 100,000 person-years.
Both rates are high compared with rates from other nations (1 to 3 per 100,000 person-
years).
A few years later, Wiklund et al. (1988, 1989) produced two new cohort studies that
superficially appear to contradict the earlier findings of Hardell and Eriksson. Wiklund et al.
followed some 20,245 licensed pesticide applicators in Sweden from date of license in 1965
or after until December 31, 1984. Some 72% were estimated to have been exposed to
phenoxy herbicides 1 day or longer (based on questionnaire data sent to a random sample of
273 persons in the cohort). The relative risk for soft tissue sarcoma reported in the first
study was found to be 0.9, with a mean follow-up time of 13.9 years. Even after a 10-year
latency, the risk for soft tissue sarcoma was only 1.0 based on four deaths. With respect to
the second study of all other cancer sites, major significant deficits were found in several
sites followed for an average of 12.2 years until December 31, 1982. No report is given
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concerning loss to follow-up or vital status. The authors also report that "the cohort may
have been observed for too short a time to reveal any excess risk . . . ."
The authors describe a major disadvantage of these studies to be a "lack of individual
exposure data" and that information is available "for only a sample of the cohort." Even the
length of exposure of individual applicators to phenoxy herbicides is not available. What is
presented is information that herbicide use in the 1950s was only 19%, and in the 1960s it
increased to just 49%. By the 1970s, however, it was up to 67%. On the other hand,
pesticide use is reported to be 92% during the same period. This seems to indicate that
perhaps less than half had any exposure to the phenoxy herbicides at the time of licensing
and perhaps for many years afterwards. Furthermore, information is presented about the
presence or absence of chloracne, another marker of exposure to 2,3,7,8-TCDD.
In addition to missing important information regarding the vital status of this cohort
by the end of the follow-up, no information is available concerning the distribution of
person-years "at risk" generated by the "lost to follow-up" group. Among the cancer sites
reported to have significantly reduced risks are total cancer, liver, pancreas, lung, and
kidney.
Furthermore, the extent of exposure to the agent of concern, 2,3,7,8-TCDD, may not
be extensive among licensed applicators in Sweden. The entire discussion is centered on
exposure to "phenoxy herbicides." The authors state that the most widely used phenoxy
herbicides in Sweden are MCPA, mecoprop, and dichlorprop. None of these contain
2,3,7,8-TCDD as a contaminant. 2,4-D and 2,4,5-T have also been used to a "lesser extent"
according to the authors. Sweden prohibited the use of 2,4,5-T in 1977. If, as the authors
state, only 19% used herbicides in the 1950s, increasing to 49% in the 1960s, it suggests that
far fewer applicators were exposed to small quantities of 2,3,7,8-TCDD for a long enough
period of time to produce any effects. Furthermore, only 68.2% of the cohort could have
attained the age of 59 by the close of the study in 1984. This hints at the likelihood that the
full impact of exposure on mortality has not yet been achieved. In fact, the authors report
that the "latency time may anyhow be too short to detect increasing risks of cancer . . . ."
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7.5.5. Summary
The cohorts assembled by Fingerhut et al. (1991), Manz et al. (1991), and Zober
et al. (1990) are important because they contain sizable proportions of persons with
substantial TCDD exposures. These exposures were documented, at least in subsets of the
cohorts, by blood and/or adipose tissue measurements, workplace measurements, and the
occurrence of chloracne. The Saracci et al. cohort, while significantly larger, is assembled
with nonuniform exposure criteria for TCDD exposure, leading to less confidence in the
results.
The exposures and methods in the three former studies in males were similar enough
to warrant aggregating the results. Within each study, relative risks were estimated by
summing the observed and expected numbers of deaths across categories of age, race, and
calendar time, and then dividing the totals to produce relative risk estimates in the form of
standardized mortality ratios. Thus, aggregate relative risks can be obtained simply by
summing the observed and expected numbers of deaths across the studies. Alternatively, the
aggregate relative risk could have been derived by weighting the individual relative risks
from each study by the inverse of the variance. A separate analysis of the three studies using
estimates of lifetime dose intake is presented in Section 8.5.
As shown in Table 7-5, the studies by Manz et al. (1991) and Zober et al. (1990) add
little to the information provided by the study by Fingerhut et al. (1991), except to increase
the precision of the relative risk estimates (as indicated by a narrowing of the confidence
intervals). Even when aggregated, however, the estimates for cancers of connective and soft
tissues, non-Hodgkin's lymphoma, and stomach cancer are highly imprecise. With this
important limitation, the results suggest little or no increase in the risk of non-Hodgkin's
lymphomas. They also suggest increased risk—especially among persons with relatively high
exposure and relatively long latency—for connective and soft tissue cancers, for lung cancer,
and for all cancers combined. Some confounding from exposure to asbestos cannot be ruled
out in the Fingerhut study.
The estimates of increased risk for lung cancer and for all cancers combined are
considerably more precise (Table 7-5). The elevations for these cancers also appear to be
more pronounced in the subcohorts of relatively high exposure and relatively long latency
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Table 7-5. Summary of Results for Selected Cancers From Follow-up Studies of Chemical Manufacturing and Processing Workers
Exposed to TCDD
Total cohorts
Cancer
Connective and
soft tissue
cancers
Non-Hodgkin's
lymphomas
Lung cancer
Stomach cancer
All cancers
combined
Study
Fingerhut (1991)
Manz (1991)
Zober (1990)
Total
Fingerhut (1991)
Manz (1991)
Zober (1990)
Total
Fingerhut (1991)
Manz (1991)
Zober (1990)
Total
Fingerhut (1991)
Manz (1991)
Zober (1990)
Total
Fingerhut (1991)
Manz (1991)
Zober (1990)
Total
Observed
deaths
4
0
0
4
10
3
0
13
89
30
6
125
10
12
3
25
265
93
23
381
Expected
deaths
1.2
0.4"
0.1"
1.7
7.3
2.4"
0.6"
10.3
80.1
21.3
5.6
107.0
9.7
9.9
2.7
22.3
229.9
75.2
19.7
324.8
Relative
risk
3.3
0.0
0.0
2.4
1.4
1.2
0.0
1.3
1.1
1.4
1.1
1.2
1.0
1.2
1.1
1.1
1.2
1.2
1.2
1.2
95%
Confidence
interval
1.1
0.0
0.0
0.7
0.7
0.3
0.0
0.7
0.9
1.0
0.4
1.0
0.5
0.7
0.3
0.7
1.0
1.0
0.8
1.1
-8.0
-7.5
-30.0
-5.7
-2.4
-3.4
-5.0
-2.1
- 1.4
-2.0
- 2.2
- 1.4
- 1.8
-2.1
-3.0
- 1.6
- 1.3
- 1.5
- 1.7
- 1.3
Subcohorts
Observed
deaths
3
0
0
3
2
NA
0
2
40
NA
3
43
4
NA
2
6
114
34
7
155
with high exposure, lone latency, or
Expected
deaths
0.3
0.1'
0.0"
0.4
2.1
NA
0.2"
2.3
28.8
NA
1.2
30.0
2.9
NA
0.5
3.4
78.0
23.9
4.2
106.1
Relative
risk
10.0
0.0
00
7.5
1.0
NA
0.0
0.9
1.4
NA
2.5
1.4
1.4
NA
4.0
1.8
1.5
1.4
LI
1.5
both
95%
Confidence
interval
2.5-
0.0-
0.0-
1.9-
0.2-
NA
0.0-
0.1 -
1.0-
NA
0.6-
1.1 -
0.4-
NA
0.7-
0.7-
1.2-
1.0-
0.7-
1.2-
27.3
30.0
99.9
20.4
3.1
15.0
2.9
1.9
6.8
1.9
3.3
13.2
3.7
1.8
2.0
3.3
1.7
o
o
o
O
c
en
O
»
n
HH
H
W
"Estimated as a proportion of expected deaths from all cancers combined (see text).
NA, not available.
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than in the total cohorts. Because they come from comparisons between blue-collar workers
and national populations, it is reasonable to suspect that these estimates—especially for lung
cancer—are inflated to some degree by confounding due to cigarette smoking, but the limited
analyses presented suggest that the association is not a chance occurrence (Table 7-2). On
the other hand, the counterinfluence of the healthy worker effect, exposure misclassification,
and diagnostic error would tend to force risk estimates downward. The risk estimates may
be expected to be too small in this event.
The results in males are consistent with results from animal studies to some degree.
In Chapter 6, it was shown that in a lifetime TCDD bioassay male rats developed lung
cancer, and in an initiation-promotion study ovariectomized rats exposed to TCDD developed
lung tumors, while intact rats similarly exposed did not. Furthermore, the mice in the NCI
study developed fibromas and fibrosarcomas. Also, TCDD affects the immune system and
has been shown to be a tumor promotor in animal liver and skin assays. Either or both of
these actions could lead to increased total cancer. With respect to health effects in females,
the study by Saracci et al. (1991) suggests a possible increase in breast cancer, but the results
are considered preliminary in view of the small numbers and less certain exposure. On the
other hand, Kogevinas et al. (1993) reported an increase in cancer incidence, all causes,
among women who were exposed to chlorophenoxy herbicides contaminated with TCDD.
However, no excess was observed for breast cancer. Still, in another study by Bertazzi et al.
(1993) to be discussed later in the section on Seveso, Italy, a deficit of breast cancer and
endometrial cancer was seen in women living in geographical areas contaminated by dioxin.
TCDD exposure might be expected to result in decreased breast cancer in females, based on
similar observations in rats and on TCDD's action on downregulation of the estrogen
receptor. However, this is very species-, tissue-, and age-specific.
7.6. CASE-CONTROL STUDIES IN GENERAL POPULATIONS
In this section on case-control studies, the discussion will focus chiefly on the cancer
sites, i.e., STS and non-Hodgkin's lymphoma, that have been suggested by the earlier
Swedish studies as being associated with exposure to the phenoxy herbicides. This is a
reflection of the interest shown in these two cancer sites over the last decade. Few, if any,
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case-control studies have been completed on other cancer sites. And, of course, if other
cancer sites are not studied, the risk of cancer cannot be evaluated. It is meaningless to do
case-control studies on total cancer.
7.6.1. Sweden
Hardell, Eriksson, and colleagues conducted four studies of soft tissue sarcomas
(Hardell and Sandstrom, 1979; Eriksson et al., 1981, 1990; Hardell and Eriksson, 1988) and
one study of malignant lymphomas (Hardell et al., 1981) among men living in different parts
of Sweden. In all studies, cases and their matched controls were considered exposed if they
reported phenoxy acid or chlorophenol exposures lasting at least 1 day and occurring at least
5 years before the case's date of diagnosis.
The nature of the phenoxy acid exposures differed across the Swedish study locales.
In northern Sweden, most exposures occurred in the use of 2,4,5-T and 2,4-D in combination
in forestry applications, often by knapsack spraying (Hardell and Sandstrom, 1979; Hardell
and Eriksson, 1988; Hardell et al., 1981; Hardell, 1981a, b). Phenoxy acid exposures not
involving 2,4,5-T became progressively more common, on a proportional basis, in the
central and southern regions in which agricultural herbicide uses predominated (Eriksson
et al., 1981, 1990). Whereas exposures not involving 2,4,5-T made up only 22 percent of
all phenoxy acid exposures in the first northern sarcoma study (Hardell and Sandstrom, 1979;
Hardell, 1981a, b), they accounted for 27 percent in the study in central Sweden (Eriksson et
al., 1990) and 58 percent in the study in southern Sweden (Eriksson et al., 1981).
Exposures defined only as phenoxy acid exposures are therefore less useful as indicators of
exposure to TCDD and related compounds in southern Sweden than in the central and
northern parts of the country. Furthermore, none of these studies provide information
dealing with how much exposure each subject may have had.
The reports contain little information on the specific chlorophenol preparations to
which the cases and controls were exposed. Occasional statements in some of the
manuscripts suggest that most chlorophenol exposures occurred in the sawmill and pulp
industries, that they primarily involved pentachlorophenol, and that they seldom involved
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trichlorophenols. Thus, most of the reported chlorophenol exposures entailed exposures to
the higher-chlorinated PCDDs and PCDFs but not to TCDD.
Because exposure prevalences were generally low and because phenoxy acids and
chlorophenols tend to be used in different occupations, very few persons reported joint
exposures. Thus, it is efficient to control potential confounding in the analysis of data from
these studies by comparing each exposure category (phenoxy acids vs. chlorophenols) with
the category composed of all persons who reported no exposure to phenoxy acids or
chlorophenols. Based on this method of analysis, relative risk estimates from all five studies
are presented in Table 7-6, with the sarcoma studies arranged in order of publication. The
measure of effect here is the "odds ratio" that is an estimate of the relative risk when the
cancer is relatively rare and henceforth will be called the relative risk.
The results for 2,4,5-T are the only results pertinent to TCDD exposures only.
Because of the small number of cases and controls reporting 2,4,5-T use in southern Sweden,
the confidence interval for the relative risk estimate from the study in that part of the country
(Eriksson et al., 1981) is extremely wide. A separate relative risk estimate for 2,4,5-T could
not be computed from the data in the second northern sarcoma study (Hardell and Eriksson,
1988). The published report, however, did state that all of the cases and most of the controls
exposed to phenoxy acids were exposed to preparations including 2,4,5-T (Hardell and
Eriksson, 1988). The report also gave a relative risk of 3.5 for TCDD exposure, but no
confidence interval or counts of cases and controls was provided.
The studies were conducted in two phases. The lymphoma study (Hardell et al.,
1981), the first northern sarcoma study (Hardell and Sandstrom, 1979), and the southern
sarcoma study (Eriksson et al., 1981) were published between 1979 and 1981. The
remaining two sarcoma studies (Hardell and Eriksson, 1988; Eriksson et al., 1990) appeared
about a decade later. The relative risk estimates from these more recent studies are
consistently lower than those from the earlier studies (Table 7-6). Thus, systematic
differences between the two sets of studies may be an explanation for the heterogeneity of
results.
The first set of studies received a considerable amount of criticism concerning the
methods by which the exposure information had been obtained (Hardell, 1981b; Cole, 1980).
7-37 06/30/94
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DRAFT-DO NOT QUOTE OR CITE
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06/30/94
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DRAFT-DO NOT QUOTE OR CITE
The basic concern was the possibility of bias from differential exposure misclassification
between cases and controls (sometimes called "observational bias" or "interviewer and recall
bias"), with false-negative reports of exposure suspected as being more common among the
controls and false-positive reports more common among the cases. Much of the discussion
focused on telephone interviews that were conducted by research staff who were aware of the
purpose of the study and of the case or control status of the respondents. These interviews
were conducted with selected participants to confirm reported exposures and to resolve
uncertainties on postal questionnaires, which were the primary sources of exposure
information. For living cases, controls were selected from the Swedish National Population
Registry. For deceased cases, controls were selected from the Swedish National Registry for
Causes of Death. As the criticisms of these procedures have echoed through the years (Bond
et al., 1989b; Colton, 1986), no quantitative analysis has been made of the degree of bias
that would have been required to produce the very strong associations reported in the first
three studies (Table 7-6). Of greater importance, analyses by Hardell based solely on the
questionnaire information (Hardell, 1981b) have been largely overlooked. These analyses
produced relative risk estimates very similar to those obtained when the information from the
supplemental interviews was used.
Hardell also went so far as to enroll a series of colon cancer patients (Hardell, 1981)
as a sort of "positive control" group. In contrast to soft tissue sarcomas and malignant
lymphomas, colon cancer turned out not to be associated strongly with phenoxy acid or
chlorophenol exposures. Hardell and Eriksson made a similar finding in one of the newer
studies (Hardell and Eriksson, 1988) when they included a control group consisting of a
variety of cancers along with a set of general population controls. The cancer controls were
drawn at random from the Swedish Regional Cancer Registry.
Despite Hardell's conclusion that "the previously reported associations . . . cannot to
any essential degree be explained by observational bias in the studies" (Hardell, 1981b), he
and his colleagues imposed procedures designed to reduce the potential for such bias in their
subsequent studies (Eriksson et al., 1990; Hardell et al., 1981). When lower relative risk
estimates were produced (Table 7-6), the researchers suggested that one explanation might
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DRAFT-DO NOT QUOTE OR CITE
have been the improved methods of exposure assessment. Again, controls were from
national population registries.
The investigators suggested that another explanation for the reduced relative risk
estimate for phenoxy acids and soft tissue sarcomas in the study in central Sweden (Table 7-
6) "could be the decade in which exposure occurred" (Eriksson et al., 1990), with the
implication that exposures were higher in earlier decades. They supported this suggestion
with an analysis in which only those phenoxy acid exposures occurring in the 1950s were
considered. This analysis yielded a higher relative risk estimate of 2.3 (95% CI=1.0-5.4).
Unfortunately, no basis of comparison exists because analyses by calendar time of exposure
were not conducted in any of the other studies.
As an alternative explanation for the lack of an elevated relative risk estimate in
connection with chlorophenols in the second northern sarcoma study (Table 7-6), the authors
offered "random variation" due to a low number of exposed subjects (Hardell and Eriksson,
1988). The prevalence of chlorophenol exposures among the controls in that study (10.9%)
was several times higher than in the earlier northern sarcoma study (2.9%) (Hardell and
Sandstrom, 1979) and virtually identical to the prevalence in the lymphoma study (10.4%)
(Hardell et al., 1981). The phenoxy acid exposure prevalences were highly uniform in all
three northern studies: 7.2% in the lymphoma study (Hardell et al., 1981), 6.8% in the first
sarcoma study (Hardell and Sandstrom, 1979), and 7.1% in the second sarcoma study
(Hardell and Eriksson, 1988). The chlorophenol exposure prevalence among the controls in
the first northern sarcoma study (Hardell and Sandstrom, 1979) seems to have been low.
However, the overall consistency of some excess risk is perhaps more important than actual
levels of the excess.
For three of the studies by Hardell and colleagues, relative risk estimates can be
computed restricting the data to persons who had worked in agriculture and the other
occupational categories in which the exposures of interest tend predominantly to occur
(Table 7-7). For the lymphoma study (Hardell et al., 1981) and the sarcoma study in
southern Sweden (Eriksson et al., 1981), the results for exposure to phenoxy acids,
chlorophenols, or both in the restricted analyses are virtually identical to those obtained with
the data for all subjects (Table 7-6). For the sarcoma study in central Sweden (Eriksson et
7-40 06/30/94
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Table 7-7. Relative Risks of Malignant Lymphoma and Soft Tissue Sarcomas in Relation to Phenoxy Acid and Chlorophenol
Exposures in Three Case-Control Studies in Sweden"
Cancer and study locale
Malignant lymphoma,
northern Sweden (Hardell
etal., 1981; Hardell,
1981)b
Soft tissue sarcoma,
southern Sweden (Eriksson
etal., 1981)°
Soft tissue sarcoma,
central Sweden (Eriksson
etal., 1990)d
Exposure category
Exposed to phenoxy acids,
chlorophenols, or both
Unexposed
Exposed to phenoxy acids,
chlorophenols, or both
Unexposed
Exposed to phenoxy acids
Unexposed
Cases
51
49
14
17
22
33
Controls
28
145
8
39
15
51
Relative risk (95% confidence interval)
1.0
5.4(3.1 -9.5)
1.0
4.0(1.4- 11.3)
1.0
2.3 (1.0-5.0)
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"Restricted to persons who worked in the occupational categories in which these exposures predominantly occur.
bAnalysis restricted to persons employed in agriculture, forestry, or the wood products industry.
cAnalysis restricted to persons employed in agriculture or forestry.
dAnalysis restricted to persons employed in agriculture, horticulture, or forestry.
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DRAFT-DO NOT QUOTE OR CITE
al., 1990), however, the relative risk for phenoxy acids was higher (2.3) within the special
occupational categories than among all subjects.
Interestingly, Eriksson and his colleagues stated that the association with the risk of
STS seemed to strengthen with exposure to the higher-chlorinated dioxin isomers. His
conclusion was that not only may 2,3,7,8-TCDD be a risk factor for STS but also that other
higher-chlorinated dioxins may be risk factors.
The risk calculated for exposure to 2,4,5-T in the 1950s was somewhat higher at 2.94
(95% CI= 1.1-8.0). However, exposure to 2,4,5-T during the span of the study was
nonsignificant at 1.8 (95% CI=0.9-3.9) excluding the chlorophenols. Exposure to dioxin-
containing phenoxyacetic acids or chlorophenols, excluding nondioxin-containing herbicides,
produced a significant risk estimate of 2.4 (95% CI=1.3-4.5). Exposure to high-grade
pentachlorophenols produced a risk ratio of 3.9 (95% CI = 1.2-12.9).
These analyses are important because many of the mechanisms by which biases might
occur would be related to occupation. For instance, biases in case identification, control
selection, or nonparticipation that might be related to occupational status (e.g., by its link to
socioeconomic status) would not be expected to be as great in analyses conducted within the
occupational categories as in analyses of the overall data. The potential for confounding by
occupational exposures encountered in the same lines of work would also be reduced in the
occupationally restricted analyses. Several researchers and reviewers (Johnson, 1990; Pearce
et al., 1985; Blair et al., 1985) have noted reports of farmers being at increased risk of
malignant lymphomas and other cancers and have mentioned a wide range of potentially
responsible exposures, "including pesticides, solvents, oils and fuels, dusts, paints, welding
fumes, zoonotic viruses, microbes, and fungi" (Blair et al., 1985). (Studies of farmers and
other agricultural workers are not included in this review because mere membership in these
occupational categories is insufficient as an indicator of exposure to such substances as 2,4,5-
T or chlorophenols.)
In the original reports, occasional attempts to assess exposure-response trends
produced mixed results. In general, the reported exposure periods were short in all of the
studies. In the first northern sarcoma study, for instance, 93% of all reported phenoxy acid
exposures lasted 1 year or less, 74% lasted 6 months or less, and 33% lasted 30 days or less
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DRAFT-DO NOT QUOTE OR CITE
(Hardell and Sandstrom, 1979; Hardell, 1981b). Reported exposures in southern Sweden
were even briefer, with 53% lasting 30 days or less (Eriksson et al., 1981).
Hardell et al. recently aggregated the four soft tissue sarcoma studies in a re-analysis
examining exposures to herbicides contaminated with TCDD and other dioxins (Hardell
et al., 1991). Increasing trends in risk with duration of exposure (< 1 year and > 1 year)
and "latency" (5-19 years and >20 years since first exposure) were numerically impressive,
being based on the totals of 434 cases and 948 controls from all the studies (Table 7-8). The
problem of concomitant exposures was not solved in these analyses, however, and an
analysis of the aggregated data obscured the pronounced heterogeneity of results among the
individual studies (Table 7-6).
Regardless of the exposure definition, considerable heterogeneity exists among the
relative risk estimates from the four soft tissue sarcoma studies (Table 7-6). (Tests of
homogeneity yield two-tailed p-values of 0.002 for phenoxy acids, chlorophenols, or both;
0.02 for phenoxy acids; 0.03 for 2,4,5-T; and 0.01 for chlorophenols.) In this circumstance,
aggregation of results across studies is not indicated and, instead, a search should be made
for explanations for the heterogeneity.
Two additional studies of malignant lymphomas and one study of soft tissue sarcomas
were conducted by independent research teams in southern Sweden. Olsson and Brandt's
(1988) study consisted of 167 men diagnosed with non-Hodgkin's lymphoma in the years
1978-1981 and 140 controls from the Swedish National Population Registry. Men who
reported handling phenoxy acids or chlorophenols for at least 1 day were considered
exposed. However, the main focus of the study was to evaluate the contribution of organic
solvent exposure to the risk of non-Hodgkin's lymphoma. Persson et al. (1989) studied 54
cases of Hodgkin's disease, 106 cases of non-Hodgkin's lymphoma, and 275 controls of both
genders from the population registry of Sweden. The cases were diagnosed in the years
1964-1986, but only those who were still alive in 1986 were included. The authors did not
ask specific questions about phenoxy acid use. Wingren et al. (1990) studied 96 men with
soft tissue sarcomas diagnosed in the years 1975-1982, 450 general population controls, and
200 cancer controls from the regional cancer registry. Because the results did not differ
substantially between the two control groups, only those obtained from analyses with the
7-43 06/30/94
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Table 7-8. Mantel-Haenszel Odds Ratios for Soft Tissue Sarcoma Among Persons Exposed to All Dioxins, TCDD, and Dioxins
Other Than TCDD in Four Case-Control Studies Involving 434 Cases and 948 Controls*
Substance and
variable
All dioxins
No. of cases
No. of controls
OR
90% CI
TCDD
No. of cases
No. of controls
OR
90% CI
Other dioxins
No. of cases
No. of controls
OR
90% CI
No exposure Exposure < 1 Yrb Exposure 2: 1 Yr
Latency Latency Latency Latency
5 - 19 Yr S 20 Yr 5 - 19 Yr ^ 20 Yr
352 24 34 3 21
865 22 52 0 9
1.0 2.4 6.4
1.7 - 3.4 3.5 - 12
352 18 22 1 5
865 14 25 0 2
1.0 3.0 7.2
2.0 - 4.5 2.6 - 20
352 6 12 2 16
865 8 27 0 7
1.0 1.7 6.2
0.98 - 2.9 2.9 - 13
<—) aOR denotes odds ratio and CI confidence interval.
5j bAll subjects were exposed for at least 1 day. Data for latency periods were combined to determine the odds ratios.
g Source: Hardell et al. (1991).
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DRAFT-DO NOT QUOTE OR CITE
general population controls are reported here. The authors had to resort to job-associated
uses, of which one was called "unspecified chemical work, potential exposure to phenoxy
herbicides and chlorophenols," because only limited information could be obtained about
specific chemical exposures from postal questionnaires and selected, supplemental telephone
interviews.
Results from these three studies are summarized in Table 7-9. Persson et al. (1989)
found strong associations, Wingren et al. (1990) found an association of intermediate
strength, and Olsson and Brandt (1988) found very little association. These studies are
limited by the lack of specificity in their exposure information.
7.6.2. United States
Zahm, Cantor, and colleagues from the National Cancer Institute have reported results
from three case-control studies of exposure to 2,4,5-T or 2,4-T as well as other pesticides
and herbicides in four Great Plains states (Hoar et al., 1986; Zahm et al., 1990; Cantor et
al., 1992). The first study was conducted in Kansas (Hoar et al., 1986). It included soft
tissue sarcomas, Hodgkin's disease, and non-Hodgkin's lymphomas, but detailed analyses
were confined to the non-Hodgkin's lymphomas. The two subsequent studies, one conducted
in eastern Nebraska (Zahm et al., 1990) and the other in Iowa and Minnesota (Cantor et al.,
1992), evaluated non-Hodgkin's lymphomas. Cancer risks at other sites from exposure to
2,4-D calculated from these groups are the subject of later studies. These studies did not
consider chlorophenol exposures, and only those persons who ever lived or worked on a
farm were asked questions about pesticide exposures. Farmers and nonfarmers were asked
about home and garden use of pesticides. Thus, all nonfarmers were considered unexposed.
As in southern Sweden (Table 7-6), the vast majority of phenoxy acid exposures did not
involve 2,4,5-T and those that did virtually always involved 2,4-D as well. The relevance of
these studies to the focus on TCDD and related compounds in this review is therefore
somewhat limited.
Results for 2,4,5-T from the three studies are summarized in Table 7-10. Among all
subjects and among farmers, only the study in eastern Nebraska (Zahm et al., 1990) suggests
an increase in risk. All three studies were conducted with virtually identical methods and no
7-45 06/30/94
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Table 7-9. Relative Risks of Soft Tissue Sarcomas, Non-Hodgkin's Lymphomas, and Hodgkin's Disease in Relation to Phenoxy
Acid and Chlorophenol Exposures in Two Case-Control Studies in Southern Sweden
ON
Authors, study period
Cancer
Exposure
Relative risk (confidence interval)8
Olsson and Brandt, 1978-1981
(Olsson and Brandt, 1988)
Perssonetal., 1964-1986
(Perssonet al., 1989)
Wingrenetal., 1975-1982
(Wingrenet al., 1990)
Non-Hodgkin's lymphomas Phenoxy acids
Chlorophenols
Non-Hodgkin's lymphomas
Hodgkin's disease
Soft tissue sarcomas
Herbicides
Herbicides
Unspecified chemical work,
potential exposure to phenoxy
herbicides or chlorophenols
aConfidence intervals are 95% in the Olsson and Brandt study and 90% in the other two studies.
1.3 (0.8 -2.1)
1.2(0.7-2.0)
4.9(1.3 - 18)
3.8 (0.7 - 21)
1.6 (0.8 -3.3)
O
o
1
o
a
3
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n
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a
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o
-------
Table 7-10. Relative Risks of Non-Hodgkin's Lymphomas in Relation to Farm Use of 2,4,5-T in Case-Control Studies in Kansas,
Eastern Nebraska, Iowa, and Minnesota
Occupational
category
All subjects
Farmers
Exposed farmers
and unexposed
nonfarmers
Exposure category and
measure
Exposed
Unexposed
Relative risk
95 % confidence interval
Exposed
Unexposed
Relative risk
95 % confidence interval
Exposed
Unexposed
Relative risk
95 % confidence interval
Kansas (Hoar et al,. 1986}
Cases
3
167
3
130
3
37
Controls
18
930
0.9
0.3 - 3.2
18
644
0.8
0.2 - 2.8
18
286
1.3
0.4 - 4.6
Eastern Nebraska
(Zahmetal., 19901
Cases
13
188
13
134
13
54
Controls
27
696
1.8
0.9 - 3.5
27
512
1.8
0.9 - 3.7
27
184
1.6
0.8 - 3.4
Iowa and Minnesota
(Cantor et al.. 1992)
Cases
25
597
1.0
0.6- 1
25
331
1.0
0.6- 1
25
266
1.1
0.6- 1
Controls
48
1,197
.7
48
650
.7
48
547
.8
o
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g
n
Source: Hoar et al., 1986; Zahm et al., 1990; Cantor et al., 1992.
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DRAFT-DO NOT QUOTE OR CITE
information on herbicide application methods in any of the reports indicate any exposure
conditions peculiar to eastern Nebraska.
The third set of relative risk estimates in Table 7-10 were computed using the
investigators' procedure of including only the exposed farmers and the unexposed
nonfarmers, with the unexposed farmers excluded. Comparing exposed farmers with
unexposed nonfarmers shows that it is possible that risk estimates could be influenced by
potential confounding effects that are germane to farmers, i.e., exposure to other pesticides
or herbicides in large quantities. Furthermore, if the results of follow-up efforts are
markedly different in farmers versus follow-up efforts in nonfarmers, then this difference
also might add some uncertainty to the accuracy of risk estimates. In these analyses, the
relative risk estimates from the studies in Kansas, Iowa, and Minnesota are somewhat higher
and the estimate from the eastern Nebraska study is somewhat lower than in the two more
conventional analyses.
Formal homogeneity tests across the three studies yield two-tailed p-values of 0.4 in
the analysis of all subjects, 0.3 in the analysis restricted to farmers, and 0.6 in the third
analysis. Ordinarily, especially considering the virtually identical methods used in the three
studies, these results would be considered sufficient justification to compute summary
estimates. Summary (maximum likelihood) estimates of relative risk are virtually identical in
all three groups of subjects, with point estimates of 1.2, lower 95% confidence limits of 0.8,
and upper 95% confidence limits of 1.7 to 1.8.
Woods and colleagues (1987) conducted a study of soft tissue sarcomas and non-
Hodgkin's lymphomas in western Washington State. In this study, the principal method of
phenoxy acid and chlorophenol exposure assessment was to place job titles, activities, and
chemical preparations reported during interviews into categories of potential exposure. The
categories were created "in consultation with local industrial and university representatives
who had long-term experience with forestry, wood products, and agricultural industries in the
Pacific Northwest" (Woods et al., 1987). No statement is given about the relative prevalence
of 2,4-D and 2,4,5-T among the phenoxy acids used in this region. The authors did not
present any information regarding tissue levels of TCDD in either cases or controls.
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DRAFT-DO NOT QUOTE OR CITE
The results (Table 7-11) show no association between soft tissue sarcomas or non-
Hodgkin's lymphomas and estimated potential for exposure to phenoxy acids or
chlorophenols. The authors report, however, that the relative risk of non-Hodgkin's
lymphomas associated with more than 15 years of potential exposure to phenoxy acids
increased with time since the accumulation of that exposure. The relative estimates were 1.3
(95% CI=0.9-2.2) for exposures more than 5 years before diagnosis, 1.7 (95% CI=1.0-2.8)
for exposures more than 15 years before, and 2.5 (95% CI=0.5-13.0) for exposures more
than 25 years before. The authors stated that similar trends were not seen in any of the
analyses of soft tissue sarcomas and phenoxy acids or of either cancer in connection with
chlorophenol exposures. It is not possible with the available data from this study to conduct
analyses restricted to persons who worked in forestry, agriculture, and the wood products
industry, and in which the exposed persons are those who reported specific exposures to
phenoxy acids or chlorophenols.
The western Washington State study reported two unique results. One consisted of
elevated relative risks in connection with a history of chloracne based on personal interviews:
3.3 (95% CI=0.8-14.0) for soft tissue sarcomas and 2.1 (95% CI=0.6-7.0) for non-
Hodgkin's lymphomas. However, the diagnoses were not medically confirmed and because
only 1 % of all cases and controls reported chloracne histories, the confidence intervals were
extremely wide. The other intriguing result consisted of elevated relative risks of soft tissue
sarcomas among persons with Scandinavian surnames (12% of the cases and controls). The
estimates from this analysis were 2.8 (95% CI=0.5-15.6) for "high" estimated potential for
phenoxy acid exposure and 7.2 (95% CI =2.1-24.7) for "high" estimated potential for
exposure to chlorophenols. The authors noted that similarly elevated relative risks were not
found for non-Hodgkin's lymphomas.
In a later study of the same study population, Woods and Polissar (1989) found a
significant excess risk (OR = 1.33, 95% CI = 1.03-1.7) of non-Hodgkin's lymphoma among
farmers compared with nonfarmers. When further examined to determine if 2,4-D or 2,4,5-
T was responsible, both tended to be nonsignificant^ decreased. However, frequency of use
of these herbicides was not considered by the authors.
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DRAFT-DO NOT QUOTE OR CITE
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Brown et al. (1990) conducted a population-based, case-control interview study of 578
white males with leukemia in Iowa and Minnesota matched to 1,245 controls living in those
same states. The purpose of the study was to investigate potential agricultural hazards that
may be related to a diagnosis of leukemia. The cases were derived from the Iowa Tumor
Registry and a network of hospitals and pathology laboratories in Nebraska between March
1981 and October 1983. Areas with little farm activity were excluded from the study.
There was a slight but marginally significant elevation of leukemia risk (OR = 1.2,
CI = 1.0-1.5) in farmers versus nonfarmers. But for those who mixed, handled, or applied
2,4,5-T, the risk was slightly but nonsignificantly elevated (OR = 1.3, CI=0.7-2.2).
In another case-control study of Iowa agricultural influences on multiple myeloma,
Brown et al. (1993), using similar methodology as in her earlier study, matched 173 white
males with multiple myeloma to 650 controls from Iowa. Although a slight nonsignificant
elevated risk (OR = 1.2, CI=0.8-1.7) was seen in farmers, the risk of multiple myeloma
from exposure to 2,4,5-T was found to be nonsignificant (OR=0.9, CI=0.4-2.1). The same
was true for numerous other herbicides, pesticides, and insecticides.
The authors concluded that there was little evidence to suggest any association of
multiple myeloma with farming or pesticides. Neither of these studies on leukemia or
multiple myeloma has shown that exposure to dioxin occurred. This is only presumed. No
actual measurements were taken. Both of these studies could be considered hypothesis-
generating studies because they involved multiple exposures to many different chemicals used
in farming.
The major problem with U.S. case-control studies is that specific exposure to TCDD
and related compounds is not identified or quantified, although information on the use of
2,4,5-T and 2,4-D is available in some studies. In some, only potential exposure to phenoxy
herbicides is the exposure surrogate. This limits the usefulness of these studies.
7.6.3. New Zealand
Smith, Pearce, and colleagues conducted two studies of soft tissue sarcomas (Smith et
al., 1982a, 1983, 1984; Smith and Pearce, 1986) and one study of non-Hodgkin's
lymphomas (Pearce et al., 1986, 1987) among men in New Zealand. In these studies,
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DRAFT-DO NOT QUOTE OR CITE
persons were first asked whether or not they "had worked in particular occupations in which
there was potential for exposure to phenoxyherbicides or chlorophenols" (Smith et al., 1984).
If the response was affirmative, "a series of subsidiary questions were asked to clarify the
work done and the actual potential for exposure, firstly in general terms, and then in specific
terms, seeking the identity of the chemicals used" (Pearce et al., 1986). The authors
indicated that 2,4,5-T was widely used as a phenoxy acid herbicide in New Zealand over the
years pertinent to these studies (i.e., prior to the early 1980s) (Smith et al., 1984). Thus, in
these studies, the phenoxy acid exposure designation was considered a suitable indicator of
exposure to 2,4,5-T and, thus, to TCDD. Typical uses of 2,4,5-T were in the spraying of
gorse, blackberry, pasture, cereal, and peas. No actual measurements of 2,3,7,8-TCDD
were made in these studies.
In the analyses of phenoxy acids, the authors distinguished between "potential" and
"probable or definite" exposure. The latter category was created by deleting persons with
only "possible" exposures from those with "potential" exposures. It is not clear whether the
"probable or definite" designation included inferences from job titles, activities, and the like,
or whether it was based solely on affirmative responses to specific questions about phenoxy
acid exposures. For chlorophenols, only the "potential" designation was employed.
In these studies, the controls were patients diagnosed with other cancers. As opposed
to a control group selected from the entire study population, a cancer control group offers
less certainty about the degree to which its exposure distribution represents that of the study
population, but greater certainty that bias from differential exposure misclassification through
elimination of interviewer bias and recall bias is negligible. However, inclusion of cancer
sites in the controls that may be associated with the exposure could potentially bias the risk
estimate toward the null. In these particular studies, the cancer controls had an additional
advantage in minimizing any bias that might have resulted from the inability of the
researchers to include patients diagnosed at private hospitals. Private hospitals in New
Zealand have only recently been contributing to the National Cancer Registry. In an interim
report of the non-Hodgkin's lymphoma study, a second control group was drawn from the
New Zealand electoral roll. The authors concluded that this control group "gave very similar
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DRAFT-DO NOT QUOTE OR CITE
findings to those obtained with the main control group of other cancer patients" (Pearce
etal., 1986).
Another unique feature of the New Zealand soft tissue sarcoma studies is that, like the
mortality follow-up studies of chemical manufacturing and processing workers previously
reviewed, they included only those cases classified to the International Classification of
Diseases (World Health Organization, 1977) category 171, malignant neoplasms of the soft
and connective tissues. This category, which does not include soft tissue sarcomas occurring
in parenchymatous organs such as the stomach or uterus, accounted for about 60% of the soft
tissue sarcoma cases in the studies in Sweden (Fingerhut et al., 1984). There is no
indication from the Swedish studies, however, that the associations with phenoxy acids or
chlorophenols differed between soft tissue sarcomas that would be classified in category 171
and those that would be classified in the categories for the involved organs (Hardell and
Sandstrom, 1979; Eriksson et al., 1981; Hardell and Eriksson, 1988; Eriksson et al., 1990;
Hardell, 1981).
In the New Zealand Study, investigators divided their soft tissue sarcoma research
into two studies with very similar, but not identical, methods. The first study (Smith et al.,
1982a, 1983, 1984) consisted of patients and controls with cancer registrations in the years
1976-1980. The second study (Smith and Pearce, 1986) extended case-finding through 1982
and was the subject of an extremely abbreviated report. The controls in the second study
consisted of 315 of the 338 cancer controls from the non-Hodgkin's lymphoma study (Pearce
et al., 1985) whose cancer registrations were during the period 1977-1981. (The results for
the additional 23 controls, who were interviewed near the end of the non-Hodgkin's
lymphoma study, evidently were unavailable at the time the analyses for the second soft
tissue sarcoma study were conducted.)
The first sarcoma study (Smith et al., 1982a, 1983, 1984) reported very similar
results for phenoxy acids and chlorophenols when all subjects were included in the analyses,
with relative risk estimates of 1.3 for any "potential" exposure and 1.6 for exposures
("definite or probable" for phenoxy acids, "potential" for chlorophenols) lasting more than
1 day and occurring more than 5 years prior to diagnosis (Table 7-12). For phenoxy acid
exposures classified by the latter definition, sufficient data were presented to permit an
7-53 06/30/94
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Table 7-12. Relative Risks of Soft Tissue Sarcomas and Non-Hodgkin's Lymphomas in Relation to Potential Exposure to Phenoxy
Acids and Chlorophenols in Case-Control Studies in New Zealand
Measure
Cases
Controls
Phenoxv acids
Any potential exposure
Cases
Controls
Relative risk (95 % CI)
Probable or definite exposure >1 day, >5
years before cancer registration
Cases
Controls
Relative risk (95% CI)
Chlorophenols
Any potential exposure
Cases
Controls
Relative risk (95% CI)
Potential exposure >1 day, >5 years
before cancer registration
Cases
Controls
Relative risk (95% CI)
Soft tissue
First study (1976-1980) (Smith
et al., 1983, 1984)
82
92
21
19
1.3 (0.7 -2.7)
17
13
1.6(0.7-3.5)"
8
7
1.3 (0.5 -3.8)
8
6
1.6(0.5-4.7)
sarcomas
Second study (1981-1982)
(Smith and Pearce, 1986)
51
315
NR
NR
NR
6
46
0.8 (0.3 - 1.9)
NR
NR
NR
NR
NR
NR
Non-Hodgkin's lymphomas
(1978-1981)
(Pearce etal., 1987)
183
338
44
72
1.2(0.8 - 1.8)
29
50
1.1 (0.7- 1.8)
21
27
1.5 (0.8-2.7)
20
27
1.4 (0.8 - 2.6)
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"Among farmers, 3.0 (1.1-8.3); controlling for fanning by standardization, 1.9 (0.8-4.5).
CI, confidence interval; NR, not reported.
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DRAFT-DO NOT QUOTE OK Li i,
analysis restricted to farmers. Thirty of the 82 cases, 13 of the 17 exposed cases, 44 of the
controls, and 9 of the 13 exposed controls were farmers. Thus, the estimated relative risk is
3.0 (95% CI= 1.1-8.3) among farmers. Controlling for farming by ("indirect")
standardization yields an estimated relative risk of 1.9 (95% CI=0.8-4.5). Thus, as in some
studies previously reviewed, accounting for the farmer/nonfarmer distinction has a material
impact on the results from this study.
Very few details were presented for the second sarcoma study (Smith and Pearce,
1986). In comparison with a relative risk of 1.6 in the first study, the second study reported
a relative risk of 0.8 for the principal measure of phenoxy acid exposure (Table 7-12,
homogeneity-test p-value = 0.2 contrasting the two studies), The exposme prevalences in
the two control groups were virtually identical (14.1% in the first study and 14.6% in the
second), but the prevalences in the two case groups differed (exposure odds ratio = 2.0,
95% CI=0.7-5.4). Because of this difference, and because a relative risk estimate restricted
to farmers cannot be computed with the data available from the second study, aggregation of
the results would not be warranted.
The non-Hodgkin's lymphoma study (Pearce et al., 1986, 1987) reported little or no
association with phenoxy acids and a somewhat stronger association with chlorophenols
(Table 7-12). The latter association did not increase when the more restrictive measure of
exposure was used. The various activities involving exposure to chlorophenols include the
treatment of fence posts as well as treating pelts in meat works tanneries. Data that would
permit an analysis restricted to farmers were not reported.
The authors continue to maintain that herbicide spraying is a full-time occupation in
New Zealand and that none of the soft tissue sarcoma or malignant lymphoma cases had been
commercial sprayers. Smith et al. (1984) estimated the prevalence of current and former
commercial sprayers at approximately 1,500, which would be 0.17% of the male population
of New Zealand in the early 1970s (Waterhouse et al., 1982). On the null hypothesis,
therefore, only about 0.1 commercial sprayers would be expected among the cases in each of
the two soft tissue sarcoma studies and about 0.3 commercial sprayers would be expected
among the non-Hodgkin's lymphoma cases. Thus, soft tissue sarcoma risk could have been
increased manyfold and non-Hodgkin's lymphoma risk could have been increased about 3-
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fold before even one commercial sprayer would be expected in any of the case groups. As a
consequence, the absence of commercial sprayers in any of the case groups is not strong
evidence against an effect.
In a letter to the editor, Pearce (1989) produced tabular data from his earlier case-
control study by duration of use and by frequency of use. Although he maintained that his
data exhibited little evidence of an association with non-Hodgkin's lymphoma, a
nonsignificant increase in the risk was seen in the category 10-19 days of use per year
(OR=2.2, 95% CI=0.4-12.6) before dropping back to 1.1 in the category greater than 19
years.
In a study of nine selected applicators in New Zealand who had sprayed herbicides for
a minimum of 180 months (and hence 2,4,5-T), Smith et al. (1992) found a high correlation
between tissue levels of TCDD and months sprayed. This is analogous to Fingerhut's
finding that tissue levels of TCDD correlate well with duration of employment in the
herbicide manufacturing industry. TCDD serum levels ranged from 131.0 ppt in a sprayer
with 31 years of spraying to a low of 3.0 ppt in a sprayer who sprayed for only 7 years.
The average was 53 ppt for the nine sprayers who sprayed an average of 16 years. Actually,
the mean average TCDD serum level in Fingerhut's lowest exposure group who worked less
than 1 year was higher (69 ppt) than the mean average of sprayers in the Smith et al. study.
Smith's conclusions were based on his analysis that brief exposures to TCDD probably do
not contribute to the increased cancer risks seen in studies in other countries. Although it is
an interesting inference, this conclusion may be somewhat overstated without some
information regarding what the tissue levels of TCDD were in the individual cases and
controls of those other studies, information that only recently is becoming available and not
for all studies. Unfortunately, the hypothesis suggested by several occupational accidents,
such as in Seveso, Italy, and Nitro, West Virginia, that one-time large doses of exposure to
2,3,7,8-TCDD could, in fact, lead to residual high tissue levels of serum 2,3,7,8-TCDD
years later, could not be tested in the studies of Swedes. Hardell and colleagues have
provided little information regarding tissue levels of TCDD, past or present, in individuals
who were participants in their studies. One study by Nygren et al. (1986) is cited frequently
as evidence that Swedish subjects who were involved in spraying phenoxy herbicides have
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DRAFT-DO NOT QUOTE OR CITE
low levels of TCDD in adipose tissue samples. Thirty-one patients from the Regional
Hospital in Umea were each relieved of a sample of adipose tissue for analysis of dioxin
content. After "careful interviewing," it was determined that 13 of these patients had
"sprayed" herbicide at some time during their past. Adipose tissue measurements indicated a
mean of 2 ppt of 2,3,7,8-TCDD. The remaining 18 nonsprayers revealed a mean of 3 ppt.
However, in correspondence with C. Rappe (1987) regarding this study, it was
determined that of the only three that were soft tissue sarcomas, adipose tissue measurements
of 2,3,7,8-TCDD are reported to be 2, 2, and 9 ppt. The one STS with the highest 2,3,7,8-
TCDD level (9 ppt) of any of the 31 subjects is stated by the authors to have had only 10
days of "knapsack spraying" some 25 to 29 years earlier. What is striking about these 13
"cases" is that the total levels of all chlorinated dioxins are considerably greater, i.e., from
168 ppt to as much as 936 ppt per patient, and that TCDD levels are a mere fraction of the
total. The vast majority consist of the higher-chlorinated PCDDs. It is not clear that
spraying herbicides is or ever was a major occupational endeavor of this group of 13 or that
any of these patients was exposed to 2,3,7,8-TCDD in large quantities. Based on recent
correspondence with Hardell (1993), none of the patients reported by Nygren were members
of any of his case-control studies. The total PCDD levels in the three STSs ranged from 674
ppt to 792 ppt (Rappe, 1987). These were nearly all highly chlorinated. Although the cases
with tissue samples came from Hardell's clinic, they did not come from any case-control
studies.
This analysis provides little information on tissue levels in Swedish applicators with
STS. Furthermore, there may be some differences in applicator practices between New
Zealand and other countries such as Sweden. Professional applicators in New Zealand are
registered with the New Zealand Agricultural Chemicals Board (Smith et al., 1982b, 1992).
And although it might appear that they could be expected to receive a great deal of exposure
to TCDD, more than half of the applicators show serum TCDD levels below 50 ppt (Smith
et al., 1992). Considering that they were spraying 2,4,5-T for 7 to 31 years until just
recently, it seems remarkable that the distribution of serum TCDD levels is as low as it is.
The authors report that professional pesticide applicators in New Zealand are "perhaps the
group most heavily exposed to agricultural use of 2,4,5-T in the world." In other studies,
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shorter exposures to large quantities of TCDD-containing herbicides have occurred to a few
personnel such as in the Ranch Hands Cohort, Seveso, Italy, and the Nitro, West Virginia,
accident. In Fingerhut's study, employees with less than 1 year of exposure to
phenoxyacetic acids had mean serum levels averaging 69 ppt 2,3,7,8-TCDD.
7.6.4. Italy
Vineis and colleagues conducted a case-control study of soft tissue sarcomas in three
provinces in northern Italy (Vineis et al., 1986). Phenoxy acid exposure classifications were
based on job information provided on interviews or questionnaires. The assessments were
made by "[t]wo experts with experience in chemical aspects of agriculture." Cases and
controls were classified into three categories: "certainly unexposed," "exposure could not be
ruled out" (abbreviated below as "possibly exposed"), and "certainly exposed." The authors
implied that phenoxy acid herbicides of all types (2,4-D, 2,4,5-T, and MCPA) were used in
the area during the periods of interest, but were able to document only the use of 2,4-D and
MCPA (MCPA and MCPP like 2,4-D do not contain dioxin and dibenzofuran impurities).
Thus, this study may have limited relevance to an interest in TCDD exposures.
The study indicated an inverse association between possible or certain phenoxy acid
exposure and soft tissue sarcoma risk among men and a positive association among women
(Table 7-13). This latter association was restricted to women who were alive at the time the
exposure information was collected. (In the other studies in this review, in which the results
were stratified by vital status at the time of the interview, no appreciable differences were
found.) As shown in Table 7-13, when the analysis is restricted to persons who had ever
worked in agriculture ("farmers"), the relative risk among all women is reduced from 1.9 to
1.1. Sufficient data are not available for an analysis that is both restricted to farming women
and stratified by vital status.
The authors offered overmatching by location of residence as an explanation for the
lack of association among deceased subjects. It would be extraordinary for overmatching or
nondifferential misclassification (the latter being the usual explanation when reduced relative
risks are obtained with exposure information from proxy respondents) to be so strong as to
bias a relative risk of 2.4 all the way down to 0.8.
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Table 7-13. Relative Risks of Soft Tissue Sarcomas in Relation to Phenoxy Acid Exposure in Case-Control Study in Northern
Italy, 1981-1983
Men
Category
Living
Deceased
Total
Total, farmers
only
Measure
Cases
Controls
Relative risk
(95 % confidence interval)
Cases
Controls
Relative risk
(95 % confidence interval)
Cases
Controls
Relative risk
(95 % confidence interval)
Cases
Controls
Relative risk
(95 % confidence interval)
Unexposed
21
54
1.0
NA
13
17
1.0
NA
34
71
1.0
NA
12
12
1.0
NA
Possibly or
certainly exposed
2
8
0.6
(0.1 - 3.3)
1
6
0.2
(0.0 - 2.0)
3
14
0.4
(0.1 - 1.7)
3
14
0.2
(0.0 - 0.9)
Women
Unexposed
16
53
1.0
NA
6
7
1.0
NA
22
60
1.0
NA
5
8
1.0
NA
Possibly or
certainly exposed
5
7
2.4
(0.7 - 8.5)
4
6
0.8
(0.1 -4.1)
9
13
1.9
(0.7 - 5.0)
9
13
1.1
(0.3 - 4.5)
NA, not applicable.
Source: Vineis et al., 1986.
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Rice is the principal agricultural crop in the study area and rice weeding was
historically a predominantly female occupation. (Of 29 rice weeders in the study, all but two
were women.) Rice weeding during the period 1950-1955 was manual and contact with the
phenoxy herbicides was mainly through the skin.
Among all women in this study, rice weeding during the early 1950s is associated
with a relative risk of 2.3 (95% CI=0.7-7.7). When the analysis is restricted to women who
were farmers, however, the relative risk drops to 1.4 (95% CI=0.3-6.5).
7.6.5. Summary
From the standpoint of exposures to TCDD, the most important results from general-
population case-control studies come from those studies conducted in northern Sweden
(Hardell and Sandstrom, 1979; Hardell and Eriksson, 1988; Hardell et al., 1981), central
Sweden (Eriksson et al., 1990), and New Zealand (Smith et al., 1982a, 1983, 1984; Smith
and Pearce, 1986; Pearce et al., 1986, 1987). These studies were conducted in areas in
which high proportions of phenoxy acid exposures involved 2,4,5-T. The exposure-
assessment methods in these studies included the posing of specific questions about particular
chemicals and herbicide preparations. Moreover, for all but the non-Hodgkin's lymphoma
study in New Zealand (Pearce et al., 1986, 1987), available data permit analyses restricted to
farmers and the other occupational categories within which the relevant exposures
predominantly occur.
For soft tissue sarcomas, the Swedish studies are perhaps best represented by a
relative risk of 2.3 (95% CI = 1.0-5.4) for phenoxy acids among workers in agriculture,
horticulture, and forestry in the study in central Sweden (Table 7-7) (Eriksson et al., 1990).
This is justified by the following factors: the proportion of exposures to 2,4,5-T was high in
this study, the methods of assessment were better, and there were analyses within relevant
occupational categories. The authors in their current studies have redesigned their methods
to accommodate readers' criticisms of their earlier studies and have made an effort to present
risk estimates that are adjusted to reflect these criticisms.
The relative risk estimate of 3.0 (95% CI = l.l-8.3) for phenoxy acid exposure among
farmers in the first soft tissue sarcoma study in New Zealand (Smith et al., 1982a, 1983,
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DRAFT-DO NOT QUOTE OR CITE
1984) seems to indicate that fanning may be a confounder in this study. Indirect
standardization for farming produces a relative risk of 1.9 (95% CI=0.8-4.5).
For malignant lymphomas, the case-control studies provide less evidence of a positive
association. The relative risk estimates from the study by Hardell and colleagues (Hardell
et al., 1981) were very high, even among persons employed in the special occupational
groups, but this study was conducted before the researchers had improved their data
collection methods. The studies in New Zealand (Pearce et al., 1986, 1987), Kansas (Hoar
et al., 1986), eastern Nebraska (Zahm et al., 1990), and Iowa and Minnesota (Cantor et al.,
1992) are more consistent with a much smaller increase in risk, or no increase at all, from
exposures to TCDD.
The remaining case-control studies (Eriksson et al., 1981; Olsson and Brandt, 1988;
Persson et al., 1989; Wingren et al., 1990; Woods et al., 1987) offer mixed results, some
suggesting increases in the risk of soft tissue sarcoma or malignant lymphoma and others
suggesting little or no increase. The informativeness of each of these studies, however, is
limited by one or more of the following important drawbacks: study areas in which most
phenoxy acid exposures did not involve 2,4,5-T, a lack of information on specific chemicals
and preparations to which cases and controls were exposed, and an inability with available
data to conduct analyses restricted to farmers and the other occupational groups in which the
exposures of interest primarily occur. Apparently, farming as an occupation appears to
affect risk estimates based on the findings from several studies where occupation is
considered and should be considered as a potential confounder.
Vineis et al. (1992) presents the hypothesis that the excess risk of non-Hodgkin's
lymphoma seen among farmers exposed to phenoxy herbicides may be caused by viruses.
Such viruses induce proliferation and immortalization of B-cells, followed by T-cell
impairment leading to cell-mediated immunity. Increased risks of NHL have been observed
in immunologically deficient individuals. Hypothetically, the same effect could be the result
of exposure to TCDD as suggested in some mouse studies (see Chapter 4, Immunotoxicity).
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7.7. STUDIES OF PULP AND PAPER MILL WORKERS
Table 7-14 summarizes results for cancers of interest from three follow-up studies of
pulp and paper mill workers. These studies are important because of the potential for
exposure to PCDDs in this line of work. The study by Robinson et al. (1986) was of 3,572
persons who had worked for at least 1 year between 1945 and 1955 at any of five mills in
the states of California, Oregon, or Washington. The study by Jappinen et al. (1987) was of
3,454 workers in the Finnish pulp and paper industry who had worked continuously for at
least 1 year between 1945 and 1961. The study by Henneberger et al. (1989) was of 883
persons who had worked for at least 1 year at a mill in New Hampshire. Jappinen et al.
(1987) studied cancer incidence. The other two studies were mortality studies.
Individually and in the aggregate, these studies give little indication of appreciable
increases in the risk of non-Hodgkin's lymphomas, lung cancer, or stomach cancer among
pulp and paper mill workers. Overall, the rate of all cancers combined was somewhat lower
than expected. None of the studies examined connective and soft tissue cancers specifically.
Analyses of specific cancers by work location, duration of employment, and latency were
only occasionally conducted in these studies. No consistent results were found that would
alter substantially the impression given by the results for the total cohorts. These studies do
not specifically mention exposure to the PCDDs and are not designed to evaluate the risk of
cancer to PCDDs.
Other studies of cancer among paper and pulp mill workers have been restricted to
information on deaths, using either proportional mortality ratios (Milham, 1976; Milham and
Demers, 1984; Schwartz, 1988; Solet et al., 1989) or mortality odds ratios (Wingren et al.,
1991) as measures of relative risk. These studies are not highly informative because they
usually rely on minimal information in death records and because they are subject to an
upward bias due to the "healthy-worker effect" (i.e., a tendency for employed groups to have
favorable total mortality experience and causes of death other than cancers, when compared
with the general population). The degree of bias in such studies varies, but it can be
appreciable. For instance, in the cohort studied by Robinson et al. (1986), 915 deaths from
all causes were observed and 1,150.3 were expected. If the relative risk estimates for
stomach cancer and non-Hodgkin's lymphomas had been computed as proportional mortality
7-62 06/30/94
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Table 7-14. Relative Risks for Selected Cancers from Follow-up Studies of Paper and Pulp Mill Workers
Cancer
Non-Hodgkin's
lymphomas
Lung cancer
Stomach cancer
All cancers
combined
Study
Robinson (1986)
Jappinen (1987)
Henneberger (1989)
Total
Robinson (1986)
Jappinen (1987)
Henneberger (1989)
Total
Robinson (1986)
Jappinen (1987)
Henneberger (1989)
Total
Robinson (1986)
Jappinen (1987)
Henneberger (1989)
Total
Observed deaths
12
2
4
18
50
78
25
153
17
24
5
46
160
196
97
453
Expected deaths
8.9
3.5
3.8
16.2
62.1
62.6
28.0
152.7
13.8
28.8
4.2
46.8
211.5
203.8
87.9
503.2
Relative risk
1.3
0.6
U.
1.1
0.8
1.2
0.9
1.0
1.2
0.8
1.2
1.0
0.8
1.0
1.1
0.9
95% confidence
interval
0.7
0.1
0.3
0.7
0.6
1.0
0.6
0.8
0.7
0.5
0.4
0.7
0.6
0.8
0.9
0.8
-2.3
- 1.9
-2.5
- 1.7
- 1.1
- 1.5
- 1.3
- 1.2
- 1.9
- 1.2
-2.6
- 1.3
-0.9
- 1.1
- 1.3
- 1.0
0\
o
VO
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DRAFT-DO NOT QUOTE OR CITE
ratios or mortality odds ratios, they would have been 1.5 and 1.7, respectively, instead of the
values of 1.2 and 1.3 that were obtained from the authors' more valid comparisons of
mortality rates (Table 7-14).
7.8. OTHER STUDIES
Studies of pesticide applicators are not informative because they contain little
information on specific compounds and preparations to which individual persons were
exposed and as such there is no evidence of exposure to TCDD. Studies with no information
of this type include studies of licensed pesticide applicators by Wang and MacMahon (1979),
Barthel (1981), Blair et al. (1983), Wiklund et al. (1987), Corrao et al. (1989), and a study
of gardeners by Hansen et al. (1992). These studies contribute little or nothing to the
discussion of TCDD or compounds like TCDD.
Axelson and Sundell assembled a cohort of 348 Swedish railroad workers who had
applied amitrol, 2,4-D, and 2,4,5-T (Axelson and Sundell, 1974). In the most recent report
(Axelson et al., 1980), 17 deaths from tumors were observed (11.85 expected, relative risk
1.43, p = .09). The relative risk estimate for lung cancer was 1.4 (three observed deaths,
p=.37) and the estimate for stomach cancer was 2.2 (three observed deaths, p = .15). Again
as in most studies, no actual measurements of TCDD are available from this paper. Only
potential exposure to the herbicides 2,4-D and 2,4,5-T are mentioned without any effort to
quantify the exposure.
Riihimaki et al. (1982, 1983) followed a cohort of 1,971 Finnish men who had
applied 2,4-D and 2,4,5-T. With allowance for a 10-year latency period, 20 cancer deaths
were observed (24.3 expected, RR=0.8, 95% CI=0.5-1.2). The relative risk for lung
cancer was 1.1 (12 deaths observed, 95% CI=0.6-1.8). The author points out that because
of limitations in the study materials, only powerful carcinogenic effects are likely to be seen.
7.8.1. Vietnam Veterans
Distributions of TCDD levels in serum and adipose tissue are typically
indistinguishable between Vietnam veterans and comparison populations unless the Vietnam
veterans group has been carefully defined on the basis of military records to have engaged in
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activities known to have involved herbicide exposure (Centers for Disease Control Veterans
Health Studies, 1988; Devine et al., 1990; Gross et al., 1984; Kahn et al., 1988; Kang et
al., 1991; Pirkle et al., 1989; Schecter et al., 1989). Thus, the mere designation, "Vietnam
veteran," is insufficient as an indicator of exposure to 2,4,5-T or TCDD exposure. This
conclusion is also supported by Stellman and Stellman's (1986) review of military records for
the purpose of developing an Agent Orange exposure index. Stellman and Stellman drew the
further conclusion that "it is impossible to give any credence to any health effects study in
which assignment of herbicide exposure levels to individual veterans is based solely on self-
reports" (Stellman and Stellman, 1986). It is also insufficient to base an exposure index
among Vietnam veterans on such crude information as military branch (Army, Marine, etc.),
corps, or region of duty within Vietnam. Thus, a large number of studies of cancer
experience among Vietnam veterans are uninformative from the standpoint of hypothetical
effects of TCDD. These include studies by Breslin et al. (1988), the Centers for Disease
Control (1987), Dalager et al. (1991), Fett et al. (1987), Greenwald et al. (1984), Kang et
al. (1986), Kogan and Clapp (1988), Lawrence et al. (1985), and O'Brien et al. (1991).
One Vietnam veteran study by Kang et al. (1987) that further examined mortality in a
subgroup of veterans who had ventured into the areas at the time when Agent Orange was
being sprayed reported a nonsignificant odds ratio of 8.64 (CI=0.77-111.84) for soft tissue
sarcoma. The number of cases was not provided. Presumably, these would be ground
troops with a high likelihood of exposure.
The only study of cancer among Vietnam veterans at present with information on
activities involving TCDD exposure is a small mortality study by Michalek et al. (1990) of
1,261 Air Force veterans of Operation Ranch Hand. These persons were responsible for the
aerial herbicide spraying missions in Vietnam. The researchers compared the Ranch Hand
group with a group of 19,101 other Air Force veterans who were mainly involved in cargo
missions in Southeast Asia and who did not have herbicide exposure. A total of 12 cancer
deaths were observed in the Ranch Hand cohort (17.0 expected, RR=0.7, 95% CI=0.3-1.1).
Calculated death rates of all specific cancers of interest in this review were equal to or less
than the rates in the comparison group, with the exception of bone, connective tissue, skin,
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breast, and genitourinary organs. These numbers are too small for any meaningful
comparisons.
Serum TCDD measurements were taken on 888 Ranch Hands (total). A few of the
Ranch Hands, who were enlisted ground crew, exhibited tissue levels above 200 ppt of
TCDD, but the median serum level was 12.4 ppt (range 0 to 618 ppm) for the entire group.
The median serum level in the controls averaged 4.2 ppt (Wolfe et al., 1990). The majority
of the Ranch Hands probably received little exposure. The subgroups of Ranch Hands that
appear to have had the greatest exposure are nonflying enlisted personnel. The median
serum 2,3,7,8-TCDD levels in 407 of them was 23.6 ppt. The next highest levels were in
flying enlisted personnel with a median of 17.2 ppt. The remaining Ranch Hands exhibited
levels that were not much elevated (under 10 ppt) from background (flying officers, both
pilots and navigators, as well as nonflying officers). Not a great deal of cancer mortality is
to be expected to date in this relatively youthful group, which has not quite reached the 20-
year latency milestone. It would be of greater interest to reevaluate mortality and continue
follow-up in the enlisted Ranch Hands only. They appear to be the subgroup with the
greatest exposure. Further follow-up of the officers will probably not reveal any useful
information that could be attributed to exposure to 2,3,7,8-TCDD. A more appropriate
group in which to observe effects are members of the South Vietnamese Army who did the
mainstay of the spraying around the perimeters of the military bases in Vietnam.
7.8.2. Residents of Seveso, Italy
Residents of Seveso, Italy, were exposed to 2,3,7,8-TCDD in a chemical accident in
1976. This group is important because of the high exposures, with approximately 200 cases
of chloracne reported (Caramaschi et al., 1981). Children (Bertazzi et al., 1992) and adults
(Bertazzi et al., 1989a, b) at the time of the accident are being studied separately. The group
residing in the zone of highest estimated exposure consists of 556 adults and 306 children.
The group residing in a zone of intermediate estimated exposure is larger, with 3,920 adults
and 2,727 children. The group with lowest estimated exposure is larger still, with 26,227
adults and 16,604 children. The accuracy of the three estimated exposure zones has been
questioned (Caramaschi et al., 1981; Merlo et al., 1986; Ratti et al., 1987; Merlo and
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Puntoni, 1986), especially because the ranking does not seem to correspond to the occurrence
of chloracne in the area.
There is no question that at least some of the residents of the most heavily exposed
area (zone A) received massive exposure to 2,3,7,8-T (Mocarelli et al., 1991). The 1990
analysis, based on tissue specimens taken in 1976, found that the highest detected levels were
recorded just after the accident. Six children at the time who subsequently developed severe
chloracne had serum 2,3,7,8-T levels ranging from 12,100 ppt to 56,000 ppt. Four other
persons with slightly less severe chloracne exhibited levels ranging from 828 ppt to 17,300
ppt. These were similar to those of nine other residents of zone A who did not develop
chloracne whose serum 2,3,5,8-TCDD levels ranged from 1,770 to 10,400 ppt (Mocarelli et
al., 1991). None of the latter group of nine were reported to be ill at the time of sampling.
Thus far, the population in zones A, B, and R around Seveso has been followed for
10 years (Bertazzi et al., 1992, 1989a, b). Ten cancer deaths, too few to support a
meaningful analysis of specific cancers, have been observed among children (Bertazzi et al.,
1992). Results for the cancers of interest among adults are summarized in Table 7-15
(Bertazzi et al., 1989b). No excesses of mortality from lung cancer, stomach cancer, or all
cancers combined are apparent. A moderate and statistically imprecise elevation in the death
rate from a subset of the cancers that make up the non-Hodgkin's lymphomas is evident in
the second 5-year period of follow-up. An excess of greater relative magnitude, but even
more imprecisely estimated, in mortality from cancers of connective and soft tissues appears
to have occurred in the same time period. Because the exposed Seveso residents have been
followed for only 10 years since the accident (Bertazzi et al., 1989a, b), highly informative
results will not come until additional time has elapsed.
In a preliminary study of cancer incidence in the same Seveso population (Pesatori et
al., 1992), the relative risk estimate of connective, subcutaneous soft tissue sarcoma of males
living in zone R is reported to be significantly elevated at 2.81 based on 6 cases
(CI = l.l-7.4). In zones A and B, none were observed but 0.4 were expected to occur in
males and 0.2 in females. For females, the risk in zone R of STS is 1.43 based on two
cases. Other cancer sites that are also elevated are certain hematologic neoplasms in males
(lymphoreticulosarcoma) and hepatobiliary tract cancers in both males and females.
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Table 7-15. Relative Risks for Selected Cancers Among Adults Exposed to TCDD in Seveso, Italy
Cancer
Connective and soft
tissues
"Other" lymphatic
tissue3
Lung cancer
Stomach cancer
All combined
Gender
Male
Female
Male
Female
Male
Female
Male
Female
Male
Female
1976-1981
RR(95%
(NA)
(NA)
(NA)
1.0(0.1 -
0.7 (0.5 -
0.7 (0.3 -
0.8 (0.5 -
0.9 (0.5 -
0.8 (0.7 -
0.8 (0.7 -
CI)
8.6)
1.0)
1.7)
1.3)
1.7)
1.0)
1.0)
Calendar period
1982-1986
RR (95% CI)
2.8 (0.3
3.0 (0.3
1.3 (0.4
1.6(0.4
1.0 (0.8
0.7 (0.3
0.9 (0.6
1.0 (0.5
1.0 (0.8
0.8 (0.7
-31.1)
- 33.0)
-4.5)
-5.7)
-1.4)
-1.9)
-1.4)
-2.0)
-1.1)
-1.1)
No. deaths
2
1
3
4
99
10
40
22
325
176
1976-1986
RR(95% CI)
5.4 (0.8
2.0 (0.2
0.9 (0.3
1.4 (0.5
0.9 (0.7
0.7 (0.4
0.8 (0.6
1.0 (0.6
0.9 (0.8
0.8 (0.7
aA subset of non-Hodgkin's lymphomas.
NA, not available.
1 /^Ort1_
- 38.6)
- 19.0)
-2.9)
-4.1)
-1.1)
-1.3)
-1.2)
-1.5)
-1.0)
-1.0)
fe
i
D
0
2
H
O
0
w
JO
H
~J
I
00
tn
I
u>
o
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Bertazzi et al. (1993) refined this study to include a more complete vital status
ascertainment without adding additional years of follow-up to the cancer incidence data in the
contaminated areas surrounding the factory where the accident took place. Cancer
occurrence ascertainment was confined chiefly to the Lombardy region of Italy (with a
population of 9 million persons) because only there can be found an efficient hospital and
discharge registration system, according to the authors. Lombardy hospitals routinely
provide hospitalization data to the regional health department. These results can be found in
Table 7-16.
Of note in this update is the information that only 14 cases of cancer were reported to
have occurred in zone A. But this is not unexpected. Only 724 residents were reported to
have lived there. This number was based on the assignment of addresses by the municipal
vital statistics offices. But it is far too small to produce any meaningful results. However,
in the more populated region B, with 4,824 residents, and region R, with 31,647 residents,
several findings were noted (Table 7-16). In zone B, hepatobiliary cancer in females
(RR=3.3, 95% CI = 1.3-8.1), lymphoreticulosarcoma in men (RR=5.7, 95% CI=1.7-19.0),
and multiple myeloma in women (RR=5.3, 95% CI= 1.2-22.6) were significantly elevated.
In zone R, soft tissue sarcomas in men (RR=2.8, 95% CI=1.0-7.3) were the only
site-specific cancers that were significantly elevated. Cancer of the genitourinary system was
significantly depressed (RR=0.8, 95% CI=0.6-1.0). This is consistent with the animal data
that suggest estrogen-induced protective effects in female rats.
The authors explain that the absence of cancer among chloracne victims is not
unexpected at this time because of the relatively young age of the group and the small
number of individuals affected.
Much has been made of the knowledge that these cancers have appeared within a
relatively short period of time in the Seveso population for latent effects to have played any
major role in the development of cancer; that is, they occurred too soon after the accident.
And, indeed, it would appear that such is the case. But it should be remembered that a
production plant making 2,4,5-T existed in that area many years prior to the date of the
accident. It is possible that earlier exposure to dioxin from the preexisting plant could have
been the initiating event for the cancers including the STSs. That would ensure enough time
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Table 7-16. Relative Risks for Selected Cancers Among Adults Exposed to TCDD in Seveso, Italy, in Contaminated Areas B
andR
Cancer
Region B
All Malignancies
Trachea, Bronchus, Lung
Hepatobiliary
Liver
Hematopoietic System
Non-Hodgkin's Lymphoma
Lymphoreticulosarcoma
Hodgkin's Disease
Multiple Myeloma
Leukemia
Myeloid Leukemia
Region R
All Malignancies
Trachea, Bronchus, Lung
Hepatobiliary
Liver
Connective & Soft
Non-Hodgkin's Lymphoma
Lymphoreticulosarcoma
Multiple Myeloma
Myeloid Leukemia
Genitourinary Organs
Breast
OBS
76
18
5
4
8
3
3
1
2
2
1
447
96
11
3
6
12
4
1
5
75
1
Males
RR
1
1
1
2
2
2
5
1
3
1
2
0
0
0
0
2
1
1
0
1
1
1
.1
.1
.8
.1
.1
.3
.7
.7
.2
.6
.0
.9
.8
.5
.2
.8
.3
.1
.2
.4
.0
.2
95%
0
0
0
0
1
0
1.
0.
0.
0
0.
0
0
0
0
1
0
0
0
0
0
0.
.9 -
.7-
.7 -
.8-
.0-
.7-
7 -
2-
8 -
.4-
2-
.9 -
.7-
.3 -
.1 -
.0-
.7-
.4-
.0-
.5-
.8-
1 -
CI
1.4
1.8
4.4
5.8
4.3
7.4
19.0
12.8
13.3
6.5
14.6
1.0
1.0
1.0
0.7
7.3
2.5
3.2
1.6
3.8
1.3
10.2
OBS
36
0
5
0
6
1
1
1
2
2
2
318
16
12
2
2
10
6
2
2
106
Females
RR
0
-
3
-
1
0
2
2
5
1
3
0
1
0
0
1
1
1
0
0
1
.8
~
.3
-
.9
.9
.3
.1
.3
.8
.7
.9
.5
.9
.5
.6
.2
.7
.6
.5
.1
95%
0.6-
-
1.3-
-
0.8-
0.1 -
0.3 -
0.3 -
1.2-
0.4-
0.9-
0.8-
0.8-
0.5-
0.1 -
0.3 -
0.6-
0.7-
0.2-
0.1 -
0.9-
CI
1.1
-
8.1
-
4.4
6.4
16.9
15.7
22.6
7.3
15.7
1.1
2.5
1.7
2.1
7.4
2.3
4.2
2.8
2.1
1.3
o
o
3
o
H
O
W
O
?d
n
H
W
Source: Bertazzi et al., 1993.
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DRAFT-DO NOT QUOTE OR CITE
for latent effects to manifest themselves. The relatively intense exposures received by these
victims may also have contributed to a shortening of the latent period. Furthermore,
hematopoietic tumors have a shorter latency than most carcinomas.
7.8.3. Rice Oil Poisonings in Taiwan and Japan Involving Compounds Structurally
Related to Dioxin
This section discusses two similar incidents involving ingestion of rice cooked with
oils accidently poisoned with PCBs and PCDFs. PCBs and PCDFs are structurally similar to
the polychlorinated dioxins, and some of these are considered to be dioxin-like in their
activity. The dioxin-like effects of these compounds are felt to be mediated through a
cytosolic receptor (Poland, 1984; Poland et al., 1987; Goldstein and Safe, 1989). The
dioxin-like polychlorinated biphenyl congeners, chlorinated dibenzofurans, and dioxins that
have a high affinity to bind the Ah receptor induce similar effects in both animals and
humans but appear to differ quantitatively in toxicity (McConnell, 1989; Ahlborg, Chapter 3,
this document; Schecter, 1991). They appear to harm growth and reproduction, they may
damage the immune system, and they also appear to cause cancer. These same effects have
been observed in a number of different species including humans.
Two accidents involving ingestion of food contaminated with PCBs and
dibenzofurans, in the Yusho (Japan) and Yu-Cheng (Taiwan) incidents, have been reported.
The Yusho incident involved 1,900 people who in 1968 accidentally consumed up to 2 grams
each of PCBs that had leaked into the rice oil at the facility where the rice oil was canned.
The PCBs were primarily Kanechlor 400 that had been used as a heat exchange medium
thousands of times. Commercial preparation Kanechlor 400 had a concentration that was
49% chlorinated. The use of this medium for exchange of heat resulted in an increase in the
dibenzofuran contamination approximately 250 times. The final mixture that was actually
present in the rice oil had a ratio of one molecule of dibenzofuran to every 200 molecules of
PCBs.
These victims suffered many ill effects from their massive exposure that lasted only a
few months. Tissue studies by the Japanese of the victims indicated that some of the PCBs
and PCDFs were retained for many years after the initial exposure. The PCDFs are
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eliminated at a slower rate than PCBs. Concentrations measured several years following the
Yusho accident indicated that the ratio of PCDFs to PCBs remaining in the adipose tissue of
the victims was about 1 to 4 (Kuratsune et al., 1975). Japanese researchers have attributed
most of the noncancer toxic effects to the presence of the PCDFs although these effects are
consistent with PCB exposure. These toxic effects include comedo formation, acneform
eruptions, hyperpigmentation, and hyperkeratosis. In addition, ocular lesions such as swollen
meibomian glands filled with yellow infarct-like material and pigmentation of the conjunctiva
were seen, similar to those effects of TCDD. For further discussion of these and other
effects, see Part B of Chapter 7, which addresses noncancer health effects in humans.
Kuratsune et al. (1988) reported a significantly increased risk of liver cancer in male
victims (9 observed vs. 1.6 expected; SMR=559, p<.01) and a nonsignificant^ increased
risk in female victims (2 observed vs. 0.66 expected; SMR=304), as well as a significantly
increased risk of lung cancer in male victims (8 observed vs. 2.45 expected; SMR=326,
p<.01). Some 1,761 patients (887 males and 874 females) were followed from date of
registration to the end of 1983, 15 years after the accident in October of 1968. Thirty-three
male cancer deaths had occurred by this time versus 15.51 expected. In female victims, 8
cancer deaths had occurred while 10.55 were expected. Comparisons were with the age-,
sex-, and cause-specific death rates of Japan and, separately, of Nagasaki and Fukuoka
prefectures in 1970, 1975, and 1980. The author reports that the risk of liver cancer
remained elevated even after the influence of latency, alcohol consumption, and liver disease
had been evaluated. Kuratsune said that because there was an uneven distribution of deaths
in the provinces where most of the victims lived, it was too early to draw any conclusions.
Apparently, most of the liver cancers occurred in Fukuoka prefecture. A statistically
significant excess mortality was still present in males of Fukuoka even when liver cancer
deaths occurring less than 9 years after the accident were eliminated. The author stated,
"Such a markedly uneven geographical distribution of deaths can hardly be explained by
exposure to the toxic rice oil alone." However, he cautioned that his findings suggest that
the poisoning might have caused liver cancer at least in male patients. He concludes, "Our
findings should not be disregarded, however, because the hepatocarcinogenicity of PCBs in
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animals has been well documented." Deaths from chronic liver diseases and cirrhosis are
also elevated but not significantly.
An outbreak of illness similar to Yusho was reported among some 2,000 persons in
the Taichung and Changhwa provinces of Taiwan in March 1979. The illness consisted of
chloracne, hyperpigmentation, and meibomian gland dilatation. In October 1979, the illness
was found to be the result of the ingestion of cooking oil contaminated with PCBs and
PCDFs. Chen et al. (1980) reported on the blood PCB levels of 66 victims for which gas
chromatograms had been prepared. Basically, blood concentration residues ranged from 11
ppb to 720 ppb in these patients. The mean value was 49 ppb; most values were under 100
ppb. In only two instances were the concentrations greater, at 120 ppb and 720 ppb. The
authors reported that the higher value of 720 ppb occurred in a patient who had difficulty
metabolizing and excreting PCB components. They also maintain that blood PCB levels of
these patients are "much higher" than those of 72 Japanese Yusho patients (Koda and
Masuda, 1975). Koda and Masuda reported the mean PCB value in Yusho patients was 5.9
ppb with a standard deviation of 4.5 ppb in 1973 and 1974. Chen et al. (1980) maintained
that this difference is due to a lengthy time lapse from the exposure to PCB in Yusho patients
before measurements were taken compared with a much shorter time lapse in Yu-Cheng
patients. Furthermore, the patients of Yu-Cheng consumed a larger proportion of higher-
chlorinated PCBs compared with those of Yusho and, as a result, the substance will be
retained longer in the body, according to the authors. Ten years have now passed.
Researchers should take a close look at this cohort now that the latent period has almost been
achieved for liver cancer in order to confirm or deny that an excess liver cancer risk is
present in these patients.
7.9. CONCLUSIONS
Of all the cancers examined in both the case-control and follow-up studies, STS
provides the greater evidence of an association with TCDD relative to other sites. The
original reports by Hardell, Sandstrom, and Eriksson of an association between STS and
exposures involving TCDD-contaminated phenoxy herbicides have stood up to extensive
criticism and a great deal of subsequent research. The degree of risk, as estimated in later
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studies by Hardell's research group, does not appear to be as great as originally suggested,
but an association with TCDD exposure is strengthened by tissue measurements of 2,3,7,8-
TCDD in potentially exposed groups.
It is also possible that the herbicides, phenoxyacetic acid, and/or chlorophenols may
exert a confounding effect. The results of the recent Lynge (1993) study suggest that
TCDD-free phenoxyherbicides and/or chlorophenols may by themselves increase the risk of
STS. Eriksson et al. (1990) further suggested that the higher-chlorinated dioxin isomers may
also be carcinogenic.
Not every study that has looked for an association between TCDD exposure and soft
tissue sarcoma risk has found one, but several studies of sound design and adequate size have
done so to a greater or lesser extent. The results from the important cohort study by
Fingerhut et al. (1991) of 5,000 chemical production and processing workers exposed to
TCDD are corroborative, as are those from the second 5 years of follow-up of the persons
exposed to TCDD in Seveso (Bertazzi et al., 1989a, b). The large IARC Registry cohort
study also suggested an association between STS and phenoxy herbicide exposure, but the
TCDD exposure component was less certain. The first New Zealand sarcoma study (Smith
et al., 1983, 1984) also appeared to produce positive results when the analysis, presented
above, was restricted to farmers to minimize bias. These results were produced by
independent investigators using substantially different research methods and studying
populations exposed under conditions much different from those in the studies by Hardell and
colleagues, with TCDD exposure being the common link.
Moreover, no persuasive case has been made that the entirety of the association in
these studies is real and not due to selection bias, differential exposure misclassification,
confounding, or chance.
The evidence on malignant lymphomas in connection with TCDD exposure has not
been substantiated, but recent evidence suggests an association between NHL and exposure to
the herbicide 2,4-D (Zahm and Blair, 1992), which may contain dioxins other than 2,3,7,8-
TCDD. The evidence from two large industrial cohort studies (Fingerhut et al., 1991;
Saracci et al., 1991), and from the Seveso population suggest little if any evidence of
increased risk. The limited evidence on TCDD exposure that can be extracted from the
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extensive case-control studies on NHL by the National Cancer Institute (Hoar et al., 1986;
Zahm et al., 1990; Cantor et al., 1992) also does not indicate a consistent and pronounced
increase in risk. It would be interesting to see results restricted to farmers from the New
Zealand study of NHL. At the present time, however, the existing studies do not present
even a minimally consistent picture of increased risks of malignant lymphoma among persons
most probably exposed to TCDD.
The evidence for lung cancer and TCDD exposure comes from the three recent cohort
follow-up studies (Fingerhut et al., 1991; Manz et al., 1991; Zober et al., 1990), all of
which provided good TCDD exposure surrogates and some actual TCDD serum level
samples. All three studies showed increased risks of borderline statistical significance of
about 40% to 100% in their highly exposed groups and low risks in their less exposed
groups. In addition to the above studies, the report of a significantly increased lung cancer
risk in male victims of the Japanese rice oil poisoning accident (Kuratsune, 1988) is also
suggestive of a TCDD-like effect (see discussion of animal carcinogenicity in Chapter 6).
While confounding or synergism by tobacco smoke cannot be excluded, the limited analyses
conducted suggest that smoking cannot explain the entire increase and that the association is
real.
It must be remembered that the confounding influence of other occupational chemicals
and presence of asbestos in the workplace could have added to the observed lung cancer
cases in these studies and helped to increase the lung cancer risk calculated for these studies.
The evidence for an association with stomach cancer is less than that for lung cancer.
While the high exposure/long duration cohorts in both the Fingerhut and Manz studies
suggest increased risks of at least 40%, the estimates are based on too few deaths for any
conclusions to be made. There are no case-control studies relating to TCDD exposure and
either lung or stomach cancer because the cancers are too common with too many potential
causes. Further research is needed.
While males comprise all the case-control studies and the bulk of the cohort study
analyses, animal and mechanism studies suggest that males and females might respond
differently to TCDD, which reduces estrogen levels in reproductive tissues and reduces
estrogen-receptor binding in rat and mouse liver. These antiestrogenic effects are thought to
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be responsible for decreased tumor incidences seen in the mammary gland, uterus, and
pituitary of TCDD-treated female rats and may also be responsible for increased liver cancer
seen in female but not male rats (see Sections 6.4.1 and 6.5.4) although in the Japanese rice
oil poisonings, the reverse was the case. These female rat liver tumors may be ovary-
dependent, while at the same time the ovaries appear to protect against TCDD-mediated
tumor promotion in the rat lung (see Section 6.4.2). Thus, these complex mechanisms might
very well affect human carcinogenicity in males and females differently. The only reported
female cohort with good TCDD exposure surrogate information was that of Manz et al.
(1991), which had a borderline statistically significant increase in breast cancer. While
Saracci et al. (1991) did report reduced female breast and genital organ cancer mortality, this
was based on few observed deaths and chlorophenoxy herbicide, rather than TCDD
exposures. Bertazzi et al. (1993) reported a deficit of breast cancer and endometrial cancer
in women living in geographical areas around Seveso contaminated by dioxin. Although
Kogevinas et al. (1993) saw an increase in cancer incidence among female workers most
likely exposed to TCDD, no increase in breast cancer was observed in her small cohort. In
sum, TCDD cancer experience for women may differ from that of men but currently there
are few results.
Other TCDD-related hormonal effects, including immune suppression, may result in
multiorgan sensitivity and may contribute to the overall increased mortality from all
malignancies combined seen in all four cohort production worker subcohorts with higher
estimated TCDD exposures (Fingerhut et al., 1991; Manz et al., 1991; Zober et al., 1990;
Saracci et al., 1991; Kogevinas et al., 1993). These increased relative risks, while not large
(10% to 70%) are consistent and are either statistically significant or of borderline
significance. While no one tissue site can account for this observed increase, lung cancer is
also increased in four of these. Although these data suggest a hypothesis of "general
carcinogenicity" consistent with a tumor promotor effect or immune suppression, it must be
kept in mind that no single agent has ever been determined to cause an overall cancer
response by the epidemiologic method. Although many carcinogens have been shown to
increase the risk at several sites, the magnitude of the risk has differed with respect to each
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site. One would not expect to see the same magnitude of increase in risk at every site
affected from exposure to a carcinogen.
In conclusion, although there are uncertainties associated with the epidemiologic
evidence that could have influenced risk estimates, the overall weight of evidence from the
epidemiologic studies suggests that the generally increased risk of cancer is more than likely
due to exposure to TCDD. The consistency of this finding in the four major cohort studies
is corroborated by animal studies that show TCDD to be a multisite, multisex, and
multispecies carcinogen.
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United States Environmental Protection Agency, Office of Health and Environmental Assessment. (1987) The
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PART B: EFFECTS OTHER THAN CANCER
7.10. INTRODUCTION
Human exposure to 2,3,7,8-TCDD has been associated with noncancer effects in most
systems. The majority of effects have been reported among occupationally exposed groups,
such as chemical production workers, pesticide users, and individuals who handled or were
exposed to materials treated with 2,3,7,8-TCDD-contaminated pesticides, and among
residents of communities contaminated with tainted waste oil (Missouri, USA) and industrial
effluent (Seveso, Italy).
These effects represent a complex network of responses ranging from changes in
hepatic enzyme levels, which, based on current evidence, do not appear to be related to
clinical disease, to observable alterations in the character and physiology of the sebaceous
gland, as in chloracne (Calvert et al., 1992; Taylor, 1979). This section of Chapter 7
describes by system the noncancer effects associated with exposure to 2,3,7,8-TCDD. The
characterization of the effects by system provides a context within which to compare the
results of the various studies. However, it is important to recognize that the observed effects
are not independent events but rather may be one outcome in a series of interrelated
outcomes, some of which we may be incapable of measuring with the present technology or
which we currently do not recognize as an outcome of exposure to 2,3,7,8-TCDD. A
summary of the effects observed in humans is outlined in Tables 7-17 to 7-43.
The information describing human effects attributed to exposure to 2,3,7,8-TCDD-
contaminated materials is derived from a wide variety of sources, including clinical
assessments (case reports) of exposed individuals and analytic epidemiologic studies using
case-control, cross-sectional, and cohort designs. The case reports describe the acute
outcomes of exposure to 2,3,7,8-TCDD and provide the basis for hypothesis generation for
controlled epidemiologic studies; however, they are not suitable for testing causal
relationships between exposure and related effects (Ashe and Suskind, 1950; Suskind et al.,
1953; Bauer etal., 1961; Goldman, 1972).
As described in the previous section, cohort and case-control studies have been used
to investigate hypothesized increases in malignancies among the various 2,3,7,8-TCDD-
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exposed populations (Fingerhut et al., 1991a, b; Manz et al., 1991; Eriksson et al., 1990).
Cross-sectional studies have been conducted to evaluate the prevalence or extent of disease in
living 2,3,7,8-TCDD-exposed groups (Suskind and Hertzberg, 1984; Moses et al., 1984;
Lathrop et al., 1984, 1987; Roegner et al., 1991; Sweeney et al., 1989; Centers for Disease
Control Vietnam Experience Study, 1988a; Webb et al., 1989). Many of the earliest studies
were unable to define exposure-outcome relationships due to a variety of shortcomings,
including small sample size, poor participation, short latency periods, selection of
inappropriate controls, and the inability to quantify exposure to 2,3,7,8-TCDD or to identify
confounding exposures. In more recent cross-sectional studies of U.S. chemical workers
(Sweeney et al., 1989), U.S. Air Force Ranch Hand personnel (Roegner et al., 1991), and
Missouri residents (Webb et al., 1989), serum or adipose tissue levels of 2,3,7,8-TCDD
were measured to evaluate 2,3,7,8-TCDD-associated effects in exposed populations. The
ability to measure tissue or serum levels of 2,3,7,8-TCDD for all or a large sample of the
subjects confirmed exposure to 2,3,7,8-TCDD and permitted the investigators to test
hypothesized dose-response relationships.
7.11. CROSS-SECTIONAL STUDIES: USES AND LIMITATIONS
Most of the studies that describe nonmalignant effects were designed as cross-
sectional medical studies. These types of studies are useful for assessing the current status of
the surviving study population; however, they are inherently limited by a number of factors,
including survivor and participation biases, exposure and disease misclassification, recall
bias, and interobserver variability. Survivor and participation biases may have occurred
because the studies included only those who were living at the time of the study and did not
or could not obtain similar information on those who died or were too ill to participate.
Studies of groups exposed to agents that contribute to early deaths or cause severe illnesses
may exclude the populations who were at highest risk. Exclusion of the sick and deceased
whose condition was associated with 2,3,7,8-TCDD may erroneously cause the risk estimate
to be closer to the null than the true risk.
Disease misclassification may be introduced in a variety of ways. Medical tests in
many reviewed studies were most often performed once without follow-up. For some
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disorders, multiple testing is preferred to obviate normal variations in some test parameters,
e.g., immunologic tests, hormone levels. In other situations, self-reported medical histories
were collected and, by design, were not or could not be confirmed by medical records.
When the exposed group incorrectly reports more disease than the unexposed group, recall
and reporting biases may falsely raise the risk estimate.
Exposure misclassification, particularly in the early studies, was a major limitation.
In the earlier studies of production workers or community residents, exposure to 2,3,7,8-
TCDD was determined only by an individual's presence (residing or working) in an area that
was contaminated with 2,3,7,8-TCDD (Suskind and Hertzberg, 1984; Moses et al., 1984;
Hoffman et al., 1986; Poland et al., 1971; May, 1973, 1982; Martin, 1984; Bond et al.,
1983, 1989; Filippini et al., 1981; Ideo et al., 1985; Mocarelli et al., 1986). The lack of a
measurement to quantify exposure hindered the ability to confirm exposure and to assess the
magnitude of an exposure-response relationship. If the misclassification is nondifferential, it
tends to bias the measure of effect toward the null.
As described above, in later studies (Roegner et al., 1991; Sweeney et al., 1989;
Centers for Disease Control Veterans Health Studies, 1988; Webb et al., 1989) researchers
were able to confirm and quantify exposure to 2,3,7,8-TCDD in serum or adipose tissue.
This breakthrough helped establish that certain previously exposed populations had 2,3,7,8-
TCDD levels well above the background level of less than 20 picograms per gram of lipid
(pg/g) (Patterson et al., 1989) and that the nonexposed comparison group was truly not
exposed. Yet, because the occupational populations were last exposed to 2,3,7,8-TCDD-
contaminated substances for as many as 40 years before being tested, levels attained at the
time of exposure can only be estimated. It appears that for some populations with higher
exposures, such as workers, the estimate reflects continuous exposure over an extended
period. For example, despite the intervening period between last exposure to 2,3,7,8-TCDD
and the determination of serum 2,3,7,8-TCDD levels in workers employed in the production
of 2,4,5-TCP and 2,4,5-T (15 to 37 years after last occupational exposure), the duration of
occupational exposure was highly correlated to serum levels of 2,3,7,8-TCDD obtained at the
time of the study (1987-1988) (Pearson product moment correlation coefficient [r]=0.7)
(Sweeney et al., 1989). These data suggest a strong relationship between length of
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occupational exposure and serum 2,3,7,8-TCDD regardless of the length of the intervening
period.
The deposition, metabolism, and excretion of high doses of 2,3,7,8-TCDD in the
human system have not been fully described. A study by Pirkle et al. (1989) suggests that
2,3,7,8-TCDD decays by one-half in approximately 7.1 years, based on a one-compartment
model and using a standard half-life equation. If this is true, exposures to trichlorophenol
production workers may have been as high as 30,000 pg/g (Fingerhut et al., 1991a) and in
excess of 50,000 pg/g in some residents of Seveso (Mocarelli et al., 1991). The data may be
limited by the lack of complete information on the manner in which human metabolism
handles 2,3,7,8-TCDD exposure.
This section of Chapter 7 is a selective review of studies that, to date, provide the
most information on the relationship between nonmalignant outcomes and exposure to
2,3,7,8-TCDD-contaminated materials. Animal studies have been reviewed in other chapters
of this document and will not be discussed in detail. Case reports will not be reviewed but
will be used to provide support for the analytic studies. In the assessment of mortality from
nonmalignant causes of death, only cohort studies in which standardized mortality ratios
(SMR) or equivalent population-based risk ratios were calculated will be discussed.
Many of the reviewed studies describe a variety of outcomes that will be discussed by
system throughout the text. Presented but not yet published works described by investigators
from the U.S. Air Force and the National Institute for Occupational Safety and Health
(NIOSH) will be included in the review. Before the results of any analysis are presented,
each institution requires extensive external peer review of the findings. Other criteria for
including studies in this review of noncancer end points are described in the beginning of
Chapter 7.
7.12. DESCRIPTION OF PRINCIPAL STUDIES
In the following section we have provided a summary of the population description
and methods of the studies that reported results for two or more systems or effects. Studies
that are referenced only once will be described when cited. Results of these studies will be
described in subsequent sections.
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7.12.1. Occupational Studies
7.12.1.1. U.S. Chemical Workers: West Virginia
In March 1949, an explosion of a TCP reaction kettle in a chemical plant in Nitro,
West Virginia, and the subsequent cleanup exposed approximately 450 workers to 2,3,7,8-
TCDD-contaminated substances. Examination of the workers in 1949 revealed a number of
acute symptoms "characterized by skin, eye and respiratory tract irritation, headache,
dizziness and nausea" (Suskind and Hertzberg, 1984). The acute symptoms subsided but
were followed within a week or two by "acneform eruption, severe muscle pain affecting the
extremities, thorax and shoulders, fatigue, nervousness and irritability, dyspnea, complaint of
decreased libido and intolerance to cold" (Suskind and Hertzberg, 1984). Thirty years later,
two independent, cross-sectional medical studies were conducted to evaluate the long-term
consequences of exposure to 2,3,7,8-TCDD-contaminated substances among the surviving
workers (Suskind and Hertzberg, 1984; Moses et al., 1984).
In a study by Suskind and Hertzberg (1984), a group of 204 (of a total of 419) active
and retired white male workers exposed between 1948 and 1969 to the 2,4,5-T production
process and to the reactor release were included in a clinical examination program. The
control group consisted of 163 (46% participation) current or former employees of the same
plant but who had no self-reported exposure to 2,4,5-T production or maintenance of the
facility. The study collected demographic and medical histories and performed clinical
chemistries, urinalysis, pulmonary function tests, dermatologic examination, and conduction
velocities of the sural sensory and peroneal motor. Multiple linear regression analysis was
used to compare the exposed and nonexposed groups.
For participation in a separate study of workers at the Nitro plant, Moses et al. (1984)
invited all workers documented in union records to have worked in 2,4,5-T production and a
systematic random sample of workers with no known 2,4,5-T production exposure. Fifty-
five percent (N=226) of the persons invited participated in the study. Lifetime occupational
and medical histories were collected and clinical chemistries, urinalysis, and dermatologic
examinations were conducted. Exposure to 2,3,7,8-TCDD could not be discerned due to
irreconcilable inconsistencies in self-reported work histories and the lack of good company
records to estimate and confirm the likelihood of exposure. Although the authors recognized
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that absence of chloracne did not preclude exposure, nevertheless, they compared the results
of the group with chloracne with those without chloracne. Thus, the study design was
revised to explore the differences in health status in individuals with and without chloracne.
Because exposed workers may have been included in the group diagnosed without chloracne,
the usefulness of the data to quantify exposure-disease relationships is limited.
Two studies of this cohort also examined cancer and noncancer mortality of subsets of
workers from this plant (Zack and Suskind, 1980; Collins et al., 1993).
7.12.1.2. U.S. Chemical Workers: The NIOSH Study
The study conducted by the National Institute for Occupational Safety and Health is a
cross-sectional medical study of living workers who were previously employed for at least
one day in one of two plants located in Newark, New Jersey, and Verona, Missouri. From
1951 to 1969, 490 workers employed at the New Jersey "plant produced sodium
2,4,5-trichlorophenate (NaTCP), 2,4,5-trichlorophenoxy acetic acid (2,4,5-T), and
2,4-dichlorophenoxy acetic acid (2,4-D). A high proportion of chloracne and other
dermatologic abnormalities and cases of porphyria and hypomania were reported among the
workers at the New Jersey facility (Poland et al., 1971; Bleiberg et al., 1964), which
produced some of the most heavily 2,3,7,8-TCDD-contaminated NaTCP and 2,4,5-T among
production facilities whose products were surveyed (Fee et al., 1975). At the Missouri plant,
NaTCP and 2,4,5-T were produced intermittently for 4 months in 1968, and NaTCP and
hexachlorophene were produced continuously for 22 months between April 1970 and January
1972. Prior to the NIOSH cross-sectional study, the health of the 96 Missouri production
workers had not been previously studied.
For comparison, unexposed neighborhood referents were recruited using a random
sampling procedure described by Sweeney et al. (1989). Referents were selected if they
reported no prior history of occupational exposure to 2,3,7,8-TCDD and matched the worker
by age (within 5 years), race, and gender. A total of 586 workers were eligible for inclusion
in the study, of which 400 (68.3%) were living, 142 (24.2%) were deceased, and 44 (7.5%)
could not be located. All 400 living workers were invited to participate in the study; 281
(70%) were examined. A description of the study population is included in the results.
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Worker and referent health and exposure status was assessed in 1987-1988 through an
interviewer-administered medical and occupational history and comprehensive physical and
psychological examinations (Sweeney et al., 1989). A lifetime medical history was obtained
from each examined participant by interviewers who were blind to the exposure status of the
respondent. Results of the pulmonary, hepatic, gastrointestinal, porphyria, mood
dysfunction, and neurologic examinations have been published or accepted for publication
(Calvert et al., 1991, 1992, 1993; Sweeney et al., 1993). Findings for diabetes and serum
glucose level, reproductive endocrine function, and p450 metabolism have been presented in
public meetings (Sweeney et al., 1992; Egeland et al., 1994; Halperin et al., 1992).
As a surrogate for cumulative exposure, serum 2,3,7,8-TCDD levels were measured
in 237 workers and a random sample of 79 referents. Procedures for sample collection,
preparation, adjustment for lipids, and statistical analysis were described in earlier reports
(Fingerhut et al., 1989; Patterson et al., 1986a; Sweeney et al., 1990). The mean lipid-
adjusted serum 2,3,7,8-TCDD level for workers was 220 pg/g, median 80 pg/g, ranging to
3,400 pg/g. The mean level was statistically significantly greater than that for referents (7
pg/g) (p< 0.001). Analyses of other congeners of dioxins and dibenzofurans were also
conducted; only the 2,3,7,8-TCDD levels were different in the two exposure groups
(Piacitellietal., 1992).
7.12.1.3. BASF Accident Cohort
"On 17 November 1953, an uncontrolled decomposition reaction occurred during the
production of 2,4,5-trichlorophenol at a BASF AG facility in Ludwigshafen, Germany" (Ott
et al., 1993a). The reactor contents, which contained 2,3,7,8-TCDD, contaminated the
building in which the TCP autoclave was housed. A series of studies have documented the
effects and mortality experience of the workers present at the time of the decomposition
reaction and exposed during the initial cleanup and equipment maintenance (May 1954
medical department list) (Cohort Cl, N=69), individuals present during subsequent clean-
demolition activities between 1954 and 1969 (Cohort C2, N=84), and a mixed group of
workers identified as of December 1987 through interviews that include individuals who
worked in the laboratory as safety inspectors and others who participated in the 1968-1969
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demolition activities (Cohort C3, N=101) (Zober et al., 1990; Ott et al., 1993b). Two
hundred forty-seven study subjects were included in a mortality study that found a
significantly elevated standardized mortality ratio (SMR) for all malignant neoplasms among
workers with chloracne and 20 or more years since first exposure to 2,3,7,8-TCDD-
contaminated chemicals (Zober et al., 1990).
Among 79% of the living subjects, lipid-adjusted serum 2,3,7,8-TCDD levels were
measured in 138 (54%) of 254 study subjects during the period 1988-1992 (Ott et al.,
1993a). The geometric mean of the 2,3,7,8-TCDD levels in the entire group was 15.4 ppt
(ranging from < 1 to 553.0 ppt) (Ott et al., 1993a) or 43 pg/g of lipid (M.G. Ott, personal
communication, 1993). Geometric means for the cohorts are as follows: Cl = 1,009.5 pg/g
lipid; C2=48.8 pg/g of lipid; and C3=83.7 pg/g of lipid. Background levels were
determined in separate analyses of 102 unexposed individuals from Germany (Papke et al.,
1992). The geometric mean for 2,3,7,8-TCDD of the external referent group was 3.0,
ranging from 0.6 to 9.1 pg/g of lipid. Based on regression analyses, serum 2,3,7,8-TCDD
levels were highly correlated to duration of exposure and location of exposure. Chloracne
severity was positively and significantly related to 2,3,7,8-TCDD concentrations.
Comprehensive batteries of clinical chemistry measurements were also measured for
the 138 subjects between 1988 and 1993 (Ott et al., 1993b). Referents were selected from
among BASF employees between the ages of 50 and 69 who participated in routine
occupational medical examinations from 1989 to 1991. For some tests, there were as many
as 6,000 referent values. For the immunologic parameters, the referent values were obtained
from a group of 42 unexposed BASF employees who participated in a separate study that
examined the immunologic function of 21 extruder personnel exposed to 2,3,7,8-
tetrabrominated dibenzo-dioxins (2,3,7,8-TBDD) and -furans (2,3,7,8-TBDF) (Zober et al.,
1992).
7.12.2. Studies of Community Residents
7.12.2.1. The Missouri Experience
During 1971, 2,3,7,8-TCDD-contaminated stillbottoms were removed from a
hexachlorophene production facility and mixed with waste oil. This mixture was deposited
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on 45 residential, recreational, and industrial sites in southeastern Missouri in 1971 and 1972
(Daryl Roberts, personal communication). Waste oil mists were commonly used in the
summer for dust control on roadways, horse arenas, truck depots, and other unpaved
surfaces. Estimated contamination of the areas ranges from 1 to 2,200 ppb (Hoffman and
Stehr-Green, 1989). A listing of potentially exposed persons was created (volunteers), and a
survey was conducted to obtain baseline information to identify persons at high risk of
exposure. Beginning in 1984, a series of studies were conducted to evaluate potential effects
(Hoffman et al., 1986; Evans et al., 1988; Webb et al., 1989), including reproductive
events, of residential exposure to 2,3,7,8-TCDD (Stockbauer et al., 1988).
The study by Hoffman et al. (1986) was conducted on 154 individuals (74% of total
eligible) who were residents of the Quail Run Mobile Home Park between 1971 and 1983
because soil concentrations around the site were 2,200 ppb 2,3,7,8-TCDD. The comparison
group of 155 (77% of total eligible) individuals was recruited from residents of a nearby
mobile home park. The examination included tests for delayed hypersensitivity (the multitest
Cornell Medical Index [CMI]; Merieux Institute, Miami, FL) and neurobehavioral effects,
blood chemistries, urinalysis, height, weight, vital signs, and examination of the skin,
peripheral pulses, lymph nodes, abdomen, and peripheral nervous system. The results of this
study were plagued by the exclusion of skin test results of 150 of 294 participants due to
high reader error. Furthermore, information on subject exposure to 2,3,7,8-TCDD was
limited because it was based on a minimum residence of 6 months in areas with contaminated
soil. Actual contact with contaminated soil was not assessed.
In the follow-up study by Evans et al. (1988), 50 persons from the initial study who
did not respond to the delayed hypersensitivity skin tests were retested. These subjects were
thought to have impaired immune function. The multitest CMI was reapplied to all test
subjects.
Webb and colleagues (1989) examined 41 of 51 persons with various histories of
exposure to 2,3,7,8-TCDD (residential, recreational, and occupational exposure) and for
whom adipose tissue levels of 2,3,7,8-TCDD were measured. Of the 41 participants, 16 had
adipose 2,3,7,8-TCDD levels less than 20 pg/g (within background range), 13 had levels
between >20 and 60 pg/g, and 12 subjects had levels above 60 pg/g. Standard medical
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examinations were conducted, and complete blood count with differential, a panel of
automated chemistry tests, serum immunoglobulins, tests for porphyrins, and the multitest
CMI were performed.
7.12.2.2. Seveso, Italy
In 1976, an explosion of a trichlorophenol reactor in a 2,4,5-T production facility in
Medina, Italy, caused the contamination by 2,3,7,8-TCDD of the neighboring city of Seveso,
Italy. The contaminated area was subdivided into three zones (A, B, and R) of decreasing
mean soil levels of 2,3,7,8-TCDD (Mocarelli et al., 1988). The mean 2,3,7,8-TCDD
concentration in Zone A was 230 /xg/m2; in Zone B, 3.0 /ug/m2; and in Zone R, 0.9 /xg/m2.
In 1979, Pocchiari et al. (1979) reported on initial efforts to screen residents in Zones
A, B, and R for 2,3,7,8-TCDD-related effects. Since then, a series of cross-sectional
medical studies have reported the final results of the screening (Caramaschi et al., 1981; Ideo
et al., 1985; Mocarelli et al., 1986; Assennato et al., 1989). Within 1 year of the reactor
release, 193 cases of chloracne were identified among residents of Zones A, B, and R, most
of which resolved with time (Assennato et al., 1989). For Seveso residents, a standard
diagnosis of chloracne was developed, in which all cases were stratified by severity: 0, no
lesions; 1, a few comedones (up to 10, minimum stage); 2, numerous comedones and cysts
(light stage); 3, comedones and cysts in specific regions (medium stage); and 4, comedones
and cysts spreading from the face to other regions of the body (serious stage) (Caramaschi et
al., 1981). Four studies investigated possible biochemical changes, particularly liver enzyme
induction and lipid levels, among the 170 children diagnosed with chloracne and control
groups (Caramaschi et al., 1981; Ideo et al., 1985; Mocarelli et al., 1986; Assennato et al.,
1989).
In addition to chloracne, several other studies also evaluated peripheral neuropathy.
In an early report by Pocchiari et al. (1979), tests for presence of peripheral nerve
dysfunction were conducted for Seveso residents and for workers at the Icmesa production
facility. Assennato et al. (1989) and Filippini et al. (1981) assessed the prevalence of
peripheral neuropathy, comparing residents with and without chloracne (Assennato et al.,
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1989) or comparing residents having chloracne or abnormal serum hepatic enzyme levels
with residents with no manifestations of 2,3,7,8-TCDD exposure (Filippini et al., 1981).
Two mortality studies, one of children ages 1-19 years and another of adults 20 years
and older, examined death rates in residents of Zones A, B, and R 10 years after the
explosion (1976-1986) (Bertazzi et al., 1989, 1992). The comparison population was
composed of approximately 100,000 inhabitants of uncontaminated areas surrounding Seveso.
Follow-up for both the young and older cohorts was 99 %.
To date, the general limitation of the studies conducted in Seveso residents is the
classification of exposure of subjects by residence in Zones A, B, or R, which is based on
mean soil concentration of 2,3,7,8-TCDD. Although serum levels of some individuals
exposed in Zone A at the time of the reactor release were in excess of 10,000 pg/g, levels of
residents living in the other areas were not measured (Mocarelli et al., 1991). The
weaknesses of using soil 2,3,7,8-TCDD levels to classify extent of exposure to 2,3,7,8-
TCDD were aptly described by Bertazzi et al. (1989): "This (use of soil contamination) is a
rather poor surrogate of exposure, and by no means an indicator of intake, since it does not
take into consideration all the possible sources and ignores interindividual variability." This
is the same problem encountered by some researchers investigating 2,3,7,8-TCDD-related
effects among Missouri residents. One might also expand this limitation to each of the
studies where environmental rather than personal levels of 2,3,7,8-TCDD contamination
were used for exposure classification.
7.12.3. Studies of Vietnam Veterans
7.12.3.1. The Vietnam Experience Study
The Vietnam Experience Study is described by its authors as a "multidimensional
assessment of the health of Vietnam veterans" (Centers for Disease Control Vietnam
Experience Study, 1988a, b, c, d). This study was designed to examine effects among men
who served in Vietnam.
The study population was composed of a random sample of men who enlisted in the
U.S. Army from 1965 through 1971, whose military occupational status was other than "duty
soldier," who enlisted for a single term with a minimum of 16 weeks active duty and who
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were discharged at pay grades of E-l to E-5. The controls were selected from among
veterans enlisting during the same period but whose duty station was the United States,
Germany, or Korea. Participation involved completion of a telephone survey of current and
past health status by 7,924 veterans who served in Vietnam and 7,364 veterans who served
outside of Vietnam. A random subsample of 2,940 Vietnam and 1,972 non-Vietnam veterans
participated in the health evaluation component.
In a separate study, serum 2,3,7,8-TCDD levels were measured in a subset of the
examined population: 646 Vietnam veterans who served in Vietnam during 1967 and 1968
and 97 non-Vietnam veterans (Centers for Disease Control Veterans Health Studies, 1988).
The mean serum 2,3,7,8-TCDD level was not different between Vietnam (mean = 4.1 pg/g
lipid [standard deviation (SD) ± 2.3]) and non-Vietnam (4.2 pg/g, SD+2.6) veterans. Two
Vietnam veterans had levels above the background level of 20 pg/g: 25 pg/g and 45 pg/g.
The overall strengths of this study are that it is a large study, with good power to
detect many common disorders; participation in the questionnaire part of the study was good
(87% for Vietnam veterans; 84% for non-Vietnam veterans); there was good comparability
between the two cohorts in demographic characteristics, although there were differential
participation rates in the examination; and validation of selected self-reported effects was
conducted. This study is limited primarily by the differential participation rates in the
examination (75% for Vietnam veterans; 63% for non-Vietnam veterans). The low level of
2,3,7,8-TCDD in the sample of veterans made it impossible to conduct dose-response
analyses.
7.12.3.2. U.S. Air Force Ranch Hand Study
One of the largest epidemiologic studies of U.S. military personnel stationed in
Vietnam is being conducted by the U.S. Air Force. The study population consists of Air
Force personnel who served in Operation Ranch Hand units in Vietnam from 1962 to 1971
and who were employed in the dissemination of Agent Orange through aerial spraying.
Comparisons included Air Force personnel who flew or maintained C-130 aircraft in
Southeast Asia during the same time period.
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The study design includes a series of cross-sectional medical studies conducted at 5-
year intervals beginning with the baseline study in 1982 (N= 1,045 exposed, 1,224
unexposed). Two follow-up evaluations were conducted in 1985 (N= 1,016 exposed, 1,293
unexposed) and 1987 (N=995 exposed, 1,299 unexposed). Each cross-sectional study
included comprehensive physical and psychological evaluations. In the 1982 baseline and
1985 and 1987 follow-up studies, exposure was based on the comparison of the Ranch Hand
group versus the comparison group. An additional analysis approximated exposure (low,
medium, high) for the Ranch Hand group by using historical military data and herbicide
procurement and usage records. The results of these analyses were prepared by Lathrop and
colleagues (1984 and 1987). In 1988, serum 2,3,7,8-TCDD levels were measured for a
sample of the 1987 Ranch Hand group (N=866) and the 1987 comparison group (N=804).
The 1987 examination data were then reanalyzed using lipid-adjusted serum 2,3,7,8-TCDD
levels as the relative measure of exposure. The median serum 2,3,7,8-TCDD level adjusted
for lipids for the Ranch Hand group was 12.8 pg/g, ranging to 618 pg/g. For the
comparison group, the median level was 4.2, ranging to 54.8 pg/g (Roegner et al., 1991).
The overall strengths of this study are that it is a large study, with good power to
detect many common disorders; follow-up was very good as is continued participation of
Ranch Hand and comparison populations. The physical and psychological examinations are
extensive, planned to evaluate most, if not all, outcomes hypothetically associated with
2,3,7,8-TCDD. Continued reevaluation of the subjects (every 5 years) permits investigators
to monitor the development of chronic diseases and to test for additional outcomes as new
biochemical and toxicologic data become available. Finally, the determination of serum
2,3,7,8-TCDD levels permitted validation of the exposure matrix based on historical records
and the subsequent development of disease-specific dose-response models. Repeated
measures of serum 2,3,7,8-TCDD over time will also provide valuable information on its
half-life in humans.
Noteworthy caveats in the study include the fact that the majority of the population
currently has serum levels under the background level of 20 pg/g (median = 12.8 pg/g,
range to 600 pg/g). These data suggest that, although there are some Ranch Hands who
were exposed to very high levels of 2,3,7,8-TCDD, most of the study group had lower
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exposures. In addition, the serum 2,3,7,8-TCDD levels indicated that the exposure matrix
used in the analysis of the baseline and 1984 studies did not appropriately describe the
potential for exposure. Therefore, data described in this chapter will refer only to the
baseline and 1984 results where Ranch Hands as a group were compared to the comparisons
and, most often, to the reanalysis of the 1987 data using serum levels.
Many analyses were performed for this study. For ease of discussion in this report,
only the adjusted odds ratios for the three categories of serum 2,3,7,8-TCDD selected by
Roegner and colleagues (1991) are presented and discussed. The categories are: < 10 pg/g
2,3,7,8-TCDD; 15-<33.3 pg/g 2,3,7,8-TCDD; and >33.3 pg/g 2,3,7,8-TCDD.
A consequence of conducting a comprehensive study in which a large number of
statistical tests are performed is an increased possibility of spurious findings. The reader
should be cognizant of this limitation, looking for consistencies with the results of other
studies and in the toxicologic literature rather than statistical significance.
7.13. REVIEW OF EFFECTS ASSOCIATED WITH EXPOSURE TO 2,3,7,8-TCDD
The following section contains a review of the case reports and epidemiologic studies
that describe effects associated with exposure to materials contaminated with 2,3,7,8-TCDD.
7.13.1. Dermal Effects
7.13.1.1. Chloracne
The most widely recognized dermal effect of exposure to 2,3,7,8-TCDD-
contaminated substances is chloracne. Chloracne is a persistent acneiform condition
characterized by comedones, keratin cysts, and inflamed papules with hyperpigmentation and
a unique anatomic distribution, occurring subsequent to acute and chronic exposure to a
variety of chlorinated aromatic compounds (Crow et ah, 1978; Moses and Prioleau, 1985).
This acne-like condition is reported to have occurred with and without other effects in at
least a few workers after all reported accidents at TCP production facilities (Ashe and
Suskind, 1950; Suskind et ah, 1953; Goldman, 1972; May, 1973; Zober at ah, 1990),
among individuals involved in daily production of 2,3,7,8-TCDD-contaminated products
(Bleiberg et ah, 1964; Poland et ah, 1971; Pazderova-Vejlupkova et ah, 1981; Moses et ah,
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1984; Moses and Prioleau, 1985; Suskind and Hertzberg, 1984; Bond et al., 1989), among
three laboratory workers exposed to pure 2,3,7,8-TCDD (Oliver, 1975), and among at least
193 (0.6%) Seveso residents, mostly children (Reggiani, 1978; Caramaschi et al., 1981; Ideo
et al., 1985; Mocarelli et al., 1986; Assennato et al., 1989). Chloracne was not found
among Missouri residents (Hoffman et al., 1986; Webb et al., 1989) examined 10 years after
exposure or among Ranch Hand personnel (Lathrop et al., 1984; Roegner et al., 1991). In
U.S. Army Vietnam veterans, chloracne-like skin lesions were rarely observed on
examination (0.9% in Vietnam veterans versus 0.8% in non-Vietnam veterans, OR=1.4,
95% CI=0.7-2.9) (Centers for Disease Control Vietnam Experience Study, 1988a).
Based on reports from Seveso and studies of chemical workers, chloracne appeared
shortly after exposure to 2,3,7,8-TCDD-contaminated chemicals (Caramaschi et al., 1981;
Zober et al., 1990). The eruption of blackheads, usually accompanied by cysts, was
observed between 2 weeks and 2 months after the reactor release (Reggiani, 1980). In
Seveso, within 6 months of the explosion, 34 cases of chloracne were identified among
children, whereupon a more intensive search was undertaken among school children
(Caramaschi et al., 1981). In chemical workers involved in the TCP reactor release at BASF
Ludwigshafen, Germany, most cases of chloracne (that were also diagnosed with cancer)
developed within 2 days after first exposure (Ott et al., 1993a; Zober et al., 1990). One
case of chloracne did not develop for 2 years, but the authors suggest that the etiology of this
case is unclear.
For many affected individuals, the condition disappeared after discontinuation of
exposure (Assennato et al., 1989) despite high serum 2,3,7,8-TCDD levels (Mocarelli et al.,
1991). But for a few, the chloracne remained for many years (Suskind and Hertzberg, 1984;
Moses and Prioleau, 1985). Of the 204 exposed workers in the study by Suskind and
Hertzberg (1984), 52% had persistent chloracne for at least 10 years after the TCP and
2,4,5-T processes ceased, 34% reported a history of chloracne, and 14% reported no history
of chloracne. Moses et al. (1984) reported that the mean duration for persistent chloracne
was 26.1+5.9 years.
There are very few human data from which to determine definitively the threshold
level of 2,3,7,8-TCDD at which chloracne occurs or who is at greatest risk to develop
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chloracne. Data from analyses of chloracne cases among chemical workers and its
relationship to serum and adipose tissue levels of 2,3,7,8-TCDD and hexachlorinated
(HxCDD) dioxins provide some basic yet useful information on the characteristics of
chloracne cases, particularly the interindividual susceptibility to chloracne (Bond et al., 1989;
Ott et al., 1987; Beck et al., 1989; Mocarelli et al., 1991). Bond et al. (1989) described 325
cases (15%) of chloracne among 2,192 workers exposed to 2,3,7,8-TCDD and HxCDDs or
octachlorinated (OCDD) dioxins between 1938 and 1982 during chemical production
activities (not as a result of a reactor accident). Cases were identified through company-
maintained medical records; age- and calendar year-specific incidence rates were estimated
based on age- and calendar year-specific person-years of employment contributed by the
cohort; and risk factors were adjusted for by logistic regression. The analysis found that risk
of chloracne was highest among workers who were exposed at younger ages, among those
who had the longest length of exposure to 2,4,5-trichlorophenol or pentachlorophenol
production operations, and among jobs rated at the highest intensity of exposure (Ott et al.,
1987). These characteristics of chloracne cases in Michigan workers are consistent with
those observed in chloracne cases from the BASF accident cohort (Ott et al., 1993a).
Although this study may underestimate the incidence rate of chloracne due to possible
underreporting or misdiagnosis of cases, and misclassification of exposure may have
occurred, this study was the first of its kind to explore analytically the risk factors associated
with occupationally acquired chloracne. It is unfortunate that these exposure estimates have
not been validated by serum or adipose tissue 2,3,7,8-TCDD, HxCDD, and OCDD levels.
Serum levels of 2,3,7,8-TCDD and HxCDD have been measured in chloracne cases
of Seveso residents (Mocarelli et al., 1991) and German chemical workers (Beck et al.,
1989; Ott et al., 1993a). Mocarelli et al. (1991) described chloracne in persons from Zone
A who had very high serum 2,3,7,8-TCDD levels ranging from 820 to 56,000 pg/g
measured within 1 year of the reactor release (Table 7-17). The study also included other
individuals from Zone A, but without chloracne, who had serum 2,3,7,8-TCDD levels that
ranged from 1,770 to 10,400 pg/g. With the exception of one person with chloracne who
was 16 years old at the time of the accident, all of the cases were in children under age 11.
Those without chloracne, for the most part, were over age 30. It is not clear whether the
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Table 7-17. Serum 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) Levels in Seveso Residents with
Chloracne and Adipose Tissue Levels of 2,3,7,8-TCDD and Hexachlorinated (HxCDD) Dioxins in German
Chemical Workers
Author
Beck et al.,
1989
Mocarelli et
al., 1991
Population
Chemical
workers0
Seveso
residents
2,3,7,8-TCDD
level (pg/g)*
174
99
147
61
50
16
1,280
49
50
2,252
158
6
56,000d
27,800
26,400
15,900
12,100
17,300
7,420
1,690
828
HxCDD
level
(Pg/g)'
247
166
5,101
172
517
58
1,019
3,442
9,613
3,087
1,191
283
—
—
—
—
—
—
—
—
—
Year of
chloracne
diagnosis
1955
1955
1963
1957(?)
1969
1955
1978
1974(?)
1972(?)
1984
1977(?)
1970
1976-1977
1976-1977
1976-1977
1976-1977
1976-1977
1976-1977
1976-1977
1976-1977
1976-1977
Half-life
extrapolated
2,3,7,8-
TCDD11
750
4,010
2,350
2,470
380
650
3,380
210
260
2,850
460
40
Half-life
extrapolated
HxCDD"
10,000
6,720
81,620
6,970
3,940
2,360
2,690
14,760
50,740
3,910
3,490
1,970
"pg/g of lipid (parts per trillion).
bHalf-life extrapolation calculated by authors (Beck et al., 1989) using the formula C0 = C, X 2"
where C0 = original concentration of 2,3,7,8-TCDD or HxCDD, C, = concentration at time t, n = number
of half-life periods, and t = half-life period of 5.8 years. Exposures occurred between 1949 and 1986.
°Measured in 1984.
•"Measured in 1976.
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children were more susceptible to the chloranegenic effects or whether they had greater
exposure to 2,3,7,8-TCDD-contaminated soil or airborne effluent.
As documented by others, adult TCP production workers also developed chloracne
(Beck et al., 1989; Suskind and Hertzberg, 1984; Bond et al., 1989). Adipose tissue levels
of 2,3,7,8-TCDD and HxCDD measured in adult chloracne cases of German chemical
production workers suggest that these cases may have been a function of the combined
exposure, making it difficult to isolate the contribution of the different chlorinated
compounds. All cases had estimated adipose levels of greater than 200 pg/g 2,3,7,8-TCDD
and in excess of 2,000 pg/g lipid HxCDD at the time of diagnosis. Estimated levels were
based on the half-life extrapolation of the adipose tissue level measured in 1986 to the date of
last occupational exposure, which may have occurred between 1949 and 1984. Similarly, Ott
(1993a) found that 80% of the severe chloracne cases had estimated (back-calculated) levels
of 250 pg/g. Yet, 26% of nonchloracne cases had estimated 2,3,7,8-TCDD concentrations
of 250 pg/g.
Data from studies of Seveso residents conducted from 1982 to 1985 indicate that,
despite high serum 2,3,7,8-TCDD levels, the chloracne resolved in all but one person by
1983 (Assennato et al., 1989). The fact that the cases of chloracne in Seveso residents
resolved within 10 years may explain why no chloracne was observed in the Ranch Hand
group although some serum 2,3,7,8-TCDD levels exceeded 600 pg/g in 1988 (Roegner et
al., 1991) and may have been as high as 2,400 pg/g at the time of last occupational
exposure, assuming 21 years since last exposure and a 7-year half-life. Nevertheless,
residual chloracne was observed 30 years after first exposure among workers from Nitro,
West Virginia, which may suggest that chronic high exposure to 2,3,7,8-TCDD or exposures
higher than experienced by the Ranch Hands may account for long-term persistence of
chloracne.
7.13.1.2. Dermatologic Disorders Other Than Chloracne
Dermal effects other than chloracne attributed to 2,3,7,8-TCDD exposure include a
variety of symptoms and conditions that occurred less frequently than chloracne but appeared
in several groups subsequent to acute and continuous exposure to 2,3,7,8-TCDD-
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contaminated TCP and 2,4,5-T. Two reports indicated that after acute episodes of exposure,
e.g., accidents, individuals complained of red and irritated eyes, conjunctivitis, and
blepharitis (inflammation of the eyelids) (Ashe and Suskind, 1950; Baader and Bauer, 1951).
Other investigators also found cases of eyelid cysts several months after acute exposure
(Suskind et al., 1953; Kimmig and Schulz, 1957a, b; Poland et al., 1971; Reggiani, 1980)
and up to 25 years after exposure (Moses et al., 1984; Suskind and Hertzberg, 1984).
Hyperpigmentation and hirsutism (also known as hypertrichosis or abnormal
distribution of hair) were diagnosed among chemical workers in the United States (West
Virginia and New Jersey) (Ashe and Suskind, 1950; Suskind et al., 1953; Bleiberg et al.,
1964; Poland et al., 1971), Germany (Bauer et al., 1961; Goldman, 1972), and
Czechoslovakia (Jirasek et al., 1974) who were exposed to 2,3,7,8-TCDD-contaminated TCP
during manufacturing processes or industrial accidents and among laboratory workers in
England exposed while synthesizing pure 2,3,7,8-TCDD (Oliver, 1975). Upon
reexamination 25 years later, hypertrichosis was observed in exposed workers (5.4% exposed
vs. 1.8% unexposed) from the West Virginia plant, particularly among workers with
persistent chloracne on clinical examination (10.3% with persistent chloracne vs. 0% with
history of chloracne only) (p< 0.001) in one of two independent studies (Suskind and
Hertzberg, 1984). A second study by Moses et al. (1984) found no evidence of
hypertrichosis, although 31% of the exposed workers had evidence of residual chloracne
(Moses et al., 1984). Studies of Vietnam veterans have reported no significant increase in
the prevalence of either hyperpigmentation or hypertrichosis (Roegner et al., 1991; Centers
for Disease Control Vietnam Experience Study, 1988a). Three cases of hypertrichosis but
not hyperpigmentation were observed among Missouri residents, one with serum levels less
than 20 pg/g and two with levels between 20 and 60 pg/g (Webb et al., 1989). Neither
disorder was noted on examination among residents of the Quail Run Mobile Home Park
(Hoffman etal., 1986).
Actinic or solar elastosis was found to be more prevalent among West Virginia
workers diagnosed with active chloracne at the time of their examinations in 1979 (Suskind
and Hertzberg, 1984) (exposed =59.1% vs. unexposed = 30.1%, p<0.01). No significant
difference was observed in the age-adjusted prevalence of actinic elastosis in workers with or
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without chloracne in the study by Moses et al. (1984). Actinic elastosis is known to be
directly related to sun exposure; however, the amount of sun exposure, skin type, or other
factors contributing to the sensitivity of the skin to sunlight were not assessed in the report.
No other studies of TCP production workers, the Ranch Hands, or U.S. Army Vietnam
veterans have found an increase in the prevalence of actinic elastosis.
Among the group of workers studied by Suskind and Hertzberg (1984), three cases of
Peyronie's disease were noted. Peyronie's disease is a rare condition characterized by
progressive scarring of the penile membrane. No explanation for this finding was expressed
nor has the condition been noted (or perhaps looked for) in other studies (Bond et al., 1989;
Moses et al., 1984; Roegner et al., 1991).
In 1984, a statistically significant excess of nonmelanotic skin cancer was reported
among Ranch Hand personnel involved in the aerial spraying of herbicides over Vietnam
compared with a matched comparison group (Lathrop et al., 1984). The comparison group
was composed of Air Force personnel assigned to cargo missions outside the sprayed areas of
Vietnam. A follow-up study of the same cohorts in 1987 confirmed the excess of basal cell
carcinoma and attributed the increase to sunlight exposure (Lathrop et al., 1987). However,
in the reanalysis of the 1987 examination data, skin neoplasms of any kind were not related
to serum 2,3,7,8-TCDD level (Roegner et al., 1991).
7.13.1.3. Comments
From an epidemiologic perspective, chloracne is a common consequence of exposure
to chemicals contaminated with 2,3,7,8-TCDD and some other polyhalogenated
hydrocarbons. Available data on serum or adipose tissue levels of 2,3,7,8-TCDD have not
determined the threshold at which a case of chloracne occurs. Evidence from Bond et al.
(1989) suggests that for chemical production workers the risk of chloracne may be related to
the part of the process in which the workers were engaged, the amount of time spent in the
contaminated region, and the intensity of the exposure while in the area. Chloracne is also
related to short-term high-intensity exposures as observed in Seveso residents (Reggiani,
1980) and occupational cohorts (Ott et al., 1993a). Few studies have been successful in
evaluating the relationship between history of chloracne and long-term nonmalignant effects.
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However, Zober et al. (1990) noted that the SMR for all malignant neoplasms for workers
with chloracne and who had at least 20 years of latency (time since first exposure) was
statistically significantly elevated (SMR 201; 90% CI = 122, 135). Future studies that are
able to evaluate the association between history of chloracne and effects would provide useful
information in this regard.
Other conditions, such as hyperpigmentation and hypertrichosis, may be more acute
effects of 2,3,7,8-TCDD exposure that resolve over time, because they were not observed in
studies where the cohorts were examined years after the cessation of exposure. Actinic
keratosis, Peyronie's disease, and basal cell carcinoma may not be due to 2,3,7,8-TCDD
because actinic keratosis and Peyronie's disease have been observed in a single cohort.
Likewise, the excess basal cell carcinoma was noted only in one study group, and the results
could not be replicated when a better indicator of personal exposure to 2,3,7,8-TCDD, serum
2,3,7,8-TCDD, was used as a surrogate for exposure in the statistical models.
7.13.2. Gastrointestinal Effects
7.13.2.1. Hepatic Effects
Changes in liver function and structure after exposure to 2,3,7,8-TCDD are
commonly observed in experimental animals (Greig et al., 1973; Vos et al., 1974; Jones and
Greig, 1975; McConnell et al., 1978a, b; Jones et al., 1981; Zinkl et al., 1973; Kociba et
al., 1976, 1978; Gasiewicz et al., 1980; DeCaprio et al., 1986). The changes are not always
consistent from one species to another, but they have prompted examination of hepatic effects
among exposed human populations. As with animals, there is wide variation in the type and
degree of hepatic effects reported in humans after exposure to 2,3,7,8-TCDD-contaminated
materials. This section describes selected hepatic effects associated with 2,3,7,8-TCDD
exposure observed in humans, including hepatomegaly and hepatic enzyme changes.
7.13.2.2. Liver Size
Increased liver size is consistently reported in treated animals after exposure to
2,3,7,8-TCDD (Vos et al., 1974; Allen et al., 1977; McConnell et al., 1978a, b; Kociba et
al., 1978; Gasiewicz et al., 1980). Among exposed human populations, four case reports in
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three populations, but not controlled epidemiologic studies, described evidence of enlarged
livers or hepatomegaly. Liver size was reportedly increased among two TCP production
workers in West Virginia within a few months after a TCP reactor explosion (Ashe and
Suskind, 1950; Suskind et al., 1953) and among "several" production workers in
Czechoslovakia exposed to TCP, the butyl ester of 2,4,5-T, and sodium pentachlorophenol
(Jirasek et al., 1974). Temporary liver enlargement was observed in 5 of 22 Seveso
residents who had severe chloracne (Reggiani, 1980). The hepatomegaly lasted "several"
months without concomitant elevation in hepatic enzymes. Fortunately, the effect appeared
to be transient. Cross-sectional medical studies of TCP production workers (Bond et al.,
1983; Suskind and Hertzberg, 1984; Moses et al., 1984; Calvert et al., 1992), Vietnam
veterans (Centers for Disease Control Vietnam Experience Study, 1988a; Roegner et al.,
1991), and Missouri residents (Webb, 1989; Hoffman et al., 1986) have found little evidence
of excess hepatomegaly in the exposed populations. Additionally, no dose-response
relationship was observed between serum levels of 2,3,7,8-TCDD and physical findings of an
enlarged liver for either the study of Ranch Hands (< 10 pg/g 2,3,7,8-TCDD, RR=0.39,
95% CI=0.11-1.33; 15-<33.3 pg/g 2,3,7,8-TCDD, RR = 1.47, 95% CI=0.57-3.79; >33.3
pg/g 2,3,7,8-TCDD, RR= 1.6.9, 95% CI=0.60-4.75) (Roegner et al., 1991) or the NIOSH
study of TCP productions workers (two workers; four referents; OR=0.46, 95% CI=0.09,
2.43) (Calvert etal., 1992).
One Missouri resident was found to have hepatomegaly, but he also was suffering
from diabetes mellitus. His adipose tissue 2,3,7,8-TCDD level was 430 pg/g (Webb et al.,
1989). The differences in findings between the case reports and the controlled epidemiologic
studies suggest that hepatomegaly may be a resolvable, acute effect as a result of exposure to
high levels of 2,3,7,8-TCDD.
7.13.2.3. Enzyme Levels
Laboratory studies have demonstrated changes in hepatic enzyme levels after 2,3,7,8-
TCDD exposure, although there is considerable interspecies variation in the observed effect
(Zinkl et al., 1973; Kociba et al., 1976, 1978; Gasiewicz et al., 1980; Olson et al., 1980).
Epidemiologic studies and case reports describe elevated liver enzymes among exposed TCP
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production workers and among Seveso residents (Mocarelli et al., 1986; May, 1982; Martin,
1984; Moses et al., 1984; Calvert et al., 1992; Ott et al., 19935).
7.13.2.4. GGT
Increased levels of gamma glutamyl transferase (GGT) may suggest activity such as
cholestasis, liver regeneration, or drug or xenobiotic metabolism (Table 7-18). The studies
of Seveso children demonstrate an increase in GGT levels occurring shortly after the
explosion and then a gradual decline to near normal levels within 5 years. In one of the
earliest studies of Seveso children with (N=141) and without chloracne (N=138), 2.8% of
the children with chloracne had out-of-range GGT levels, but none of the children without
chloracne had an out-of-range level (p<0.001) (Caramaschi et al., 1981). These results
were echoed in a study of children from Zones A, B, and R of Seveso, in which enzyme
levels were measured yearly between June 1977 and June 1982 (Mocarelli et al., 1986).
GGT levels were elevated in children of Zone A, particularly in boys, during the first 2
years after the explosion (1977: exposed = 9.73 U/L; unexposed = 7.28 U/L; p<0.01;
1978: exposed = 9.88 U/L; unexposed = 8.26 U/L; p<0.05) (Mocarelli et al., 1986).
Levels in girls during the same years were elevated but did not achieve statistical
significance. For the next 4 years of the study, GGT levels remained elevated in boys and
girls from Zone A compared to unexposed children, but the values declined with time.
GGT was found to be elevated among TCP production workers from one plant in
Great Britain and in workers from West Virginia, Missouri, and New Jersey up to 30 years
after last occupational exposure to 2,3,7,8-TCDD-contaminated chemicals (May, 1982;
Martin, 1984; Moses et al., 1984; Calvert et al., 1992). The findings of two studies of
British workers were similar. Mean GGT levels were increased, but not statistically
elevated, in workers with chloracne compared to unexposed controls tested 10 years after
exposure to 2,3,7,8-TCDD-contaminated chemicals as a result of a TCP reactor explosion
(chloracne, GGT=39 U/L; controls, 27.7 U/L [May, 1982]) (chloracne, GGT=32 U/L;
controls, 32 U/L [Martin, 1984]). Similarly, a statistically significant excess in the
proportion of individuals with abnormally high GGT levels was found among West Virginia
workers with chloracne who were examined as many as 30 years after exposure (Moses et
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Table 7-18. Mean Serum Levels of Gamma Glutamyl Transferase (GGT) Among Seveso and Missouri
Residents, TCP Production Workers, BASF Accident Cohort, and Vietnam Veterans
Author
Mocarelli et al. ,
1986
Assennato et
al., 1989
Webbetal.,
1989
Hoffman et al.,
1986
May, 1982
Martin, 1984
Moses et al. ,
1984
Calvertetal.,
1992
Population
Seveso residents
(1977) (Boys)
(Girls)
(1982) (Boys)
(Girls)
1976
1982
1984
1985
Missouri residents
<20*
20-60
>60
Missouri residents in Quail
Run Mobile Home Park
TCP production workers in
Great Britain
TCP production workers in
Great Britain
TCP and 2,4,5-T production
workers in West Virginia
TCP and 2,4,5-T production
workers in Missouri and
New Jersey
Exposed
Mean
N level" (SD)
52b 9.73 (4.72-20.04)
32b 9.10 (3.62-22.85)
106b 9.70 (4.29-21.93)
117b 9.04 (4.25-19.23)
193d 11.94 (16.80)
152d 11.67 (7.94)
142d 10.53 (7.07)
141d 10.57 (7.38)
16 24.0 (15.5)
13 17.7 (7.23)
12 32.8 (23.90)
140 3Q.& (88.3)
41 39.0
41 40.0 (14-91)
22d 26. 3C (27.0)
280 5S.5c-i (73.7)
Unexposed
Mean
N level' (SD)
42b 7.28 (3.71-14.26C)
43b 8.05 (2.99-21.68)
138b 8.99 (4.45-18.20)
140b 8.59 (3.85-19.14)
e
123f 11.23 (7.41)
196f 11.19 (7.05)
167f 10.94 (4.68)
—
141 20.1 (23.4)
31 27.7° —
120 32.0 (11-90)
9f 17.4 (11.0)
259 47.4 (41.1)
•Units = U/L.
bNumber of samples.
cp<0.05.
"thloracne.
eNo data for controls in 1976.
fNo chloracne.
8Adipose tissue level of 2,3,7,8-TCDD in pg/g of lipid.
b% Abnormal: exposed = 3.6; unexposed = 3.6.
!% Abnormal: exposed = 10.7; unexposed = 5.0; OR=227 (95% CI=1.17, 4.39).
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Table 7-18. (continued)
Author
Ott et al., 1993b
Centers for
Disease Control
Vietnam
Experience
Study, 1988a
Roegner et al.,
1991
Population
BASF accident cohort
U.S. Army ground troops
U.S. Air Force Ranch Hand
personnel
Unknown ^10'-ra
Low 15-<33.3'-m
High >33.3'-m
Exposed
Mean
N level" (SD)
133 30.5 (58.4)
2,490 43. 2j
338 31.49
191 38.28
182 40.82
Unexposed
Mean
N level' (SD)
6,708 29.9 (43.5)
1,972 41. lk —
777 34.64
•Units = U/L.
••Geometric mean.
k% Abnormal (Vietnam veterans, 5.5%; non-Vietnam veterans, 4.4; OR=1.3, 95% CI=1.0-1.8).
1 Serum 2,3,7,8-TCDD level in pg/g of lipid.
"Contrasted to unexposed comparison population.
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al., 1984) (chloracne, mean GGT=26.3 U/L, 23% abnormal; no chloracne, mean
GGT=17.4 U/L, 9% abnormal; p<0.003). Yet, compared to controls, GGT was not
elevated in another study of West Virginia workers (Suskind and Hertzberg, 1984).
In a study by Calvert et al. (1992), the mean GGT level and the proportion of
workers with out-of-range levels were statistically significantly elevated among TCP workers
from New Jersey and Missouri (workers, mean GGT=58.5 U/L; unexposed referents, mean
GGT=47.4 U/L, p<0.03; workers, 11% abnormal; referents, 5% abnormal, OR=2.27,
95% CI=1.17-4.39). Based on the logistic regression model in Table 7-19, the increases in
GGT were limited to workers with high serum 2,3,7,8-TCDD levels (> 100 pg/g) and
extremely high lifetime alcohol consumption (> 30 alcohol years) (alcohol year = 1 alcoholic
beverage/day for 1 year). The contribution of other potentially confounding exposures that
may have affected GGT levels was not explored in this study. Other studies of TCP
production workers in Michigan and West Virginia or the BASF accident cohort did not
report elevations in GGT levels (Bond et al., 1983; Suskind and Hertzberg, 1984; Ott et al.,
1993b).
Both the Vietnam Experience Study and the U.S. Air Force Ranch Hand Study found
statistically significant elevations in GGT levels (Centers for Disease Control Vietnam
Experience Study, 1988a; Roegner et al., 1991). In Army Vietnam veterans, mean GGT
levels were 43.2 U/L compared with 41.1 U/L in non-Vietnam veterans (OR for out-of-range
value = 1.3, 95% CI=1.0-1.8) (Centers for Disease Control Vietnam Experience Study,
1988a). In the 1987 follow-up study, the comparison of the adjusted mean GGT level in the
comparison group and in each of the three Ranch Hand groups defined by 2,3,7,8-TCDD
level found statistically significant increases in the Ranch Hand population (< 10 pg/g
2,3,7,8-TCDD, p<0.017; 15-<33.3 pg/g 2,3,7,8-TCDD, p<0.043; >33.3 pg/g 2,3,7,8-
TCDD, p<0.001) (Roegner et al., 1991).
7.13.2.5. AST and ALT
Abnormal levels of aspartate aminotransferase (AST) and alanine aminotransferase
(ALT) may indicate liver cell damage from a number of causes, including hepatic necrosis,
metastatic carcinoma, or obstructive jaundice (AST and ALT) or infectious or toxic hepatitis
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Table 7-19. Logistic Regression Model for an Out-of-Range Serum Gamma-Glutamyltransferase" (GGT)
Level Using the Categorical TCDDb Exposure Measure0
Variable
Intercept
Exposure
(worker = 1, referent = 0)
Per alcohol-year
Current alcohol drinker
(yes = 1, no = 0)d
Alcohol-years/exposure interaction'-'
Triglyceride level
Beta
-4.12
0.37
-2.4 x 10-6
0.90
7.2 x ID'3
3.8 x 10'3
Standard error
of the estimate
0.50
0.43
2.6 x ID'3
0.42
3.5 x 10-3
1.1 x 10'3
X2
67.01
0.74
0.00
4.70
4.24
11.94
P
<0.001
0.195
0.999
0.030
0.039
<0.001
"Reference value: 96 IU/L; a level was considered out of range if it exceeded the reference value. Reference
values were defined as the 95th percentile for the referent cohort.
bTCDD = 2,3,7,8-tetrachlorodibenzo-para-dioxin.
°N=536 observations.
dThe results from this logistic regression analysis change little when this term is dropped from the model.
'Interaction between alcohol-years and exposure.
'Exposure odds ratios (ORs) for an abnormal y-glutamyltransferase level among workers by selected
alcohol-year levels, adjusting for all variables in the model, are as follows: OR=2.96 (95% confidence
interval [CI]= 1.34-6.54) for 100 alcohol-years; OR=1.79 (95% CI=0.81-3.97) for 30 alcohol-years;
OR=1.45 (95% CI=0.63-3.36) for 1 alcohol-year; and OR=1.44 (95% CI=0.62-3.34) for 0 alcohol-years.
Source: Calvert et al., 1992b. Used with permission from the author.
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and cirrhosis (AST). Elevated levels of these enzymes may also be due to nonhepatic
origins, such as myocardial infarction, acute pancreatitis (AST and ALT), or skeletal,
cerebral, or renal necrosis (AST).
With respect to exposure to 2,3,7,8-TCDD, elevations in serum ALT and AST appear
to be transient effects of acute exposure (Table 7-20). Case reports note that some
populations have increased serum ALT levels shortly after exposure (Seveso children, British
and Czechoslovakian TCP production workers) (May, 1973; Jirasek et al., 1974; Caramaschi
et al., 1981; Mocarelli et al., 1986). Whereas epidemiologic studies conducted 10 to 30
years after last exposure reported no effects in exposed workers, Vietnam veterans and
Missouri residents compared to unexposed control groups (Table 7-20) (Suskind and
Hertzberg, 1984; May, 1982; Martin, 1984; Bond et al., 1983; Calvert et al., 1992; Centers
for Disease Control Vietnam Experience Study, 1988a; Roegner et al., 1991; Hoffman et al.,
1986; Webb et al., 1989; Ott et al., 1993b), or workers with and without chloracne (Moses
et al., 1984). Furthermore, it appears from a single study that current exposure to 2,3,7,8-
TCDD-contaminated substances must exceed some threshold to produce enzyme elevations
(Poland et al., 1971). Normal levels of AST (AST=10.6 U/L) were found in workers who
volunteered to participate in a medical study conducted concurrently with their employment
in a New Jersey chemical facility producing TCP and 2,4,5,-T (Poland et al., 1971). This
was the same plant included in a later cross-sectional medical study of workers that found, in
1988, a high mean serum 2,3,7,8-TCDD level for the group (220 pg/g) but no elevations in
AST or ALT (Calvert et al., 1992; Fingerhut et al., 199la). Similarly, no increases in AST
were noted in the BASF accident cohort (Ott et al., 1993b).
Several reports illustrate the probable transiency of ALT and AST elevations after
heavy 2,3,7,8-TCDD exposure. Both enzymes were reported to be elevated among TCP
production workers employed in Czechoslovakia who were described as exhibiting symptoms
of "chemical intoxication" (Jirasek et al., 1974). The authors reported that AST and ALT
levels were "different" in 11 (20%) of the 55 examined workers. These workers were
exposed to 2,3,7,8-TCDD-contaminated chemicals, pentachlorophenol, and its production by-
products, which include hexa-, hepta-, and octachlorinated dioxins and dibenzofurans. The
exposures occurred between 1965 and 1968, with the ensuing effects beginning shortly
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Table 7-20. Serum Alanine Aminotransferase (ALT) Among Seveso Children and Missouri Residents,
TCP Production Workers, BASF Accident Cohort, and Vietnam Veterans
Author
Mocarelli et ah,
1986
Caramaschi et ah,
1981
Moses et ah, 1984
Calvert et ah, 1992
Ott et ah, 1993a
Hoffman et ah,
1986
Webb et ah, 1989
Centers for Disease
Control Vietnam
Experience Study,
1988a
Roegner et ah,
1991
Population
Seveso children
1977 Boys
Girls
1982 Boys
Girls
Seveso children
TCP production workers;
10-30 years postexposure
TCP production workers;
15-37 years postexposure
BASF accident cohort
Missouri residents in Quail
Run Mobile Home Park
Missouri residents
<20'
20-60
>60
U.S. Army ground troops
U.S. Air Force Ranch
Hand personnel
Unknown <10'
Low 15- =£33. 3'
High >33.3'
Exposed
Mean
N leveP (SD)
46b 12.25C (7.14-20.99)
100 12.97 (6.68-25.18)
33 10.74 (5.09-22.65)
119 12.27 (6.5-23.17)
141d 3.5e-c
105d 15.9 (13.0)
2808 33.8 (22.6)
133 14.8 (8.4)
134 — h
16 22.7 (2.76)
12 22.5 (2.39)
12 23.3 (3.06)
2,490 26.4i'k
339 19.16
191 20.83
182 20.09° —
Unexposed
Mean
N leveP (SD)
45b 9.33" (3.73-23.33)
141 11.99 (5.51-26.12)
39 10.74 (5.09-22.65)
136 12.19 (6.46-23.01)
Of 0"
101f 15.7 (13.0)
259 33.0 (21.2)
6,721 15.1 (10.0)
135 — h
1,972 25.8
777 20.34
•AhT; units U/L.
""Number of samples.
cp<0.05.
dChloracne.
e% Abnormal.
fNo chloracne.
g% Abnormal: workers 4.3; unexposed 5.0; OR=0.85 (95% CI=0.38, 1.89).
h% Abnormal: exposed 6.0; unexposed 3.0.
JAdipose 2,3,7,8-TCDD level in pg/g of lipid.
•"Geometric mean.
k% Abnormal; Vietnam veterans, 5.3; non-Vietnam veterans, 4.4; OR=1.2, 95% CI=0.9-1.5.
'Serum 2,3,7,8-TCDD in pg/g of lipid.
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thereafter. The observation period lasted from 1967 to 1973. In a follow-up report of the
same population conducted in approximately 1977, Pazderova-Vejlupkova et al. (1981) did
not report liver enzyme levels but suggested that the levels were not abnormal.
Similarly, ALT was increased in 5 of 14 TCP workers from Great Britain who were
in the manufacturing building at the time of a TCP reactor explosion in 1968 (May, 1973).
Levels for AST were not reported. In 1977, workers from the same facility were
reevaluated. No elevations in AST or ALT were found in production and laboratory workers
with chloracne (May, 1982).
Finally, during the first year after the TCP reactor explosion, Caramaschi et al.
(1981) evaluated AST and ALT among Seveso children with and without chloracne. Only
ALT was statistically significantly elevated in children with chloracne. In a larger study,
Mocarelli et al. (1986) tested liver enzyme levels yearly from 1977 to 1982 in male and
female children from Seveso and from the unexposed surrounding area. In the 1977, 1979,
1980, and 1981 test series ALT, but not AST, was statistically significantly (p<0.05)
elevated among male children in Seveso compared with unexposed comparisons. Female
children had normal levels compared with controls for all years. In 1982, ALT levels in the
exposed boys returned to normal.
None of the studies reporting elevations in ALT or AST identified clinical evidence of
liver disease in the study populations. Therefore, in the absence of reports of hepatic or
nonhepatic diseases related to changes in ALT or AST levels among exposed individuals, it
is possible that the increases in ALT and AST are related to high-level, acute exposure to
2,3,7,8-TCDD-contaminated chemicals and that, barring additional exposure, the enzyme
levels decrease with time.
7.13.2.6. D-Glucaric Acid
In five studies of Seveso residents, TCP production workers, and Vietnam veterans,
urinary excretion of D-glucaric acid was measured to determine if exposure to 2,3,7,8-
TCDD induced hepatic microsomal activity (Table 7-21). D-glucaric acid excretion is an
indirect but valid indicator of enzyme induction.
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Table 7-21. Mean D-GIucaric Acid Levels Among Seveso Residents, TCP Production Workers, and
Vietnam Veterans
Author
Ideo et
al., 1985
May,
1982
Martin,
1984
Calvert
etal.,
1992
Roegner
etal.,
1991
Population
Seveso adults
Levels measured in
1978 (Zone B)
Seveso children
Zone A, 1976
Zone B
1979
1980
TCP production
workers in Great
Britain
TCP production
workers in Great
Britain
TCP and 2,4,5-T
production workers
in Missouri and
New Jersey
U.S. Air Force
Ranch Hand
personnel
Unknown ^10"
Low 15-<33.3
High >33.3
Exposed
N
117
14e
26.8
17.0
41
39
273
317
180
173
Mean (SD)
27.1a-b-c
39a,b.c
— e
—
2.07h
2.09C (0.7-7.9)
14.11 (11.1)
14.11k''
14.43'
15.22'
Unexposed
N
127
17f
—
—
31
126
256
727
—
—
Mean (SD)
19 ga,b,d
20.5b
—
—
1.52h
1.59 (0.8-8.3)
13.2! (7.9)
14.11k
— —
—
"Units: /tmol/g of creatinine.
bMedian level.
cp<0.05.
dResidents of unexposed community.
eSkin lesions.
fNo skin lesions.
sd-Glucaric acid levels measured in 1979 were significantly higher than levels measured in 1980 (p<0.05), no
data presented.
hd-Glucaric acid/creatinine ratio.
'Units = jtg/g of creatinine.
jSerum 2,3,7,8-TCDD levels in pg/g of lipid.
kUnits = /xM.
'Contrasted to unexposed comparison population.
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Ideo and colleagues (1985) measured urinary D-glucaric acid in adults and children
from all zones of Seveso and from nearby uncontaminated towns. Of adults tested in 1978,
D-glucaric acid excretion was significantly elevated in adults residing in Seveso, Italy, at the
time of the reactor explosion compared with residents of unexposed communities (Seveso =
27.1 /xmol/g of creatinine vs. unexposed = 19.8 /xmol/g of creatinine, p<0.05). No further
studies of adults have been published. A series of studies evaluated D-glucaric acid excretion
in Seveso children (Ideo et al., 1985) (Table 7-21). In 1976, the levels in children from
Zone A with chloracne (39 /xmol/g of creatinine) were significantly greater than in children
without chloracne (20.5 ^mol/g of creatinine). Additional studies, conducted until 1981,
found significant yearly decreases in urinary D-glucaric acid excretion. By 1981, levels were
within normal range.
D-glucaric acid-creatinine ratios were used to assess urinary excretion in TCP
production workers tested within a year of suspension of TCP production and 10 years after
a TCP reactor explosion. For exposed workers with or without chloracne at the time of the
study, the D-glucaric acid-creatinine ratio was significantly higher than that of the unexposed
controls (exposed = 2.09; unexposed controls = 1.59; p<0.05) (Martin, 1984).
Other studies of Air Force Ranch Hands or TCP production workers that examined
D-glucaric acid excretion did not find increases in exposed populations 10 to 37 years after
last exposure to 2,3,7,8-TCDD-contaminated chemicals (Roegner et al., 1991; Calvert et al.,
1992).
7.13.2.7. Comment
The evidence presented by the large number of case reports and epidemiologic studies
of groups exposed to 2,3,7,8-TCDD-contaminated chemicals suggests that hepatic enzyme
induction occurred in some populations within a short time after high-level 2,3,7,8-TCDD
exposure. In most cases, enzyme levels decreased as the time from exposure increased.
However, even after more than 15 years since last exposure, levels of GGT continue to be
significantly elevated in relation to serum 2,3,7,8-TCDD in TCP production workers with
above-average alcohol consumption (Calvert et al., 1992) and in Air Force Ranch Hands
(Roegner et al., 1991). Other Vietnam veterans, U.S. Army ground troops, also have
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significantly increased GGT levels compared with non-Vietnam veterans, but this increase in
the Army veterans is probably not due to exposure to high levels of 2,3,7,8-TCDD. In the
population of Army Vietnam veterans studied, the mean serum 2,3,7,8-TCDD was
approximately 4 pg/g (Centers for Disease Control Veterans Health Studies, 1988), compared
to a mean of 220 pg/g (range to 3,400 pg/g) in the production workers and a median of 12
pg/g (range to 600 pg/g) in the Ranch Hands.
The finding of continued elevation of GGT may be a spurious result or it may reflect
activity related to the continued presence of above-background levels of 2,3,7,8-TCDD in
exposed individuals.
7.13.2.8. Porphyrin Metabolism
In rats and mice, exposure to 2,3,7,8-TCDD has been clearly shown to produce
alterations in porphyrin metabolism (Goldstein et al., 1973, 1982; Smith et al., 1982; Jones
et al., 1981; DeVerneuil et al., 1983; Cantoni et al., 1981). Whether 2,3,7,8-TCDD is
associated with porphyrin changes in humans, particularly porphyria cutanea tarda (PCT), is
a subject of some debate. PCT is a form of acquired or inherited porphyria caused by a
deficiency of the enzyme uroporphyrinogen decarboxylase and the resulting overproduction
and excretion of uroporphyrin (Sweeney, 1986). The predominant characteristics of PCT
include skin fragility, blistering upon sun exposure, dark pigmentation, excess hair growth,
hepatomegaly, reddish-colored urine, and urinary excretion of uro- and heptacarboxyl-
icoporphyrins (Strik, 1979). PCT has been associated with excessive alcohol intake, oral
estrogens, iron overload, hepatomas, and exposure to polyhalogenated hydrocarbons (Strik,
1979). A particularly large outbreak of PCT occurred after consumption of grain treated
with hexachlorobenzene (HCB) (Cam and Nigogosyan, 1963). Cases of PCT were described
in two populations of TCP production workers (Bleiberg et al., 1964; Jirasek et al., 1974)
and among members of a family with inherited uroporphyrin decarboxylase deficiency who
were living in Seveso at the time of the reactor explosion (Strik, 1979).
In 1964, Bleiberg reported that based on the Watson-Schwartz test, 11 of 29 New
Jersey TCP production workers with chloracne had porphyria cutanea tarda as a result of
increased urinary uroporphyrins, coproporphyrins, and urobilinogen. In a later study of 73
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workers from the same plant in New Jersey, including four of the individuals that Bleiberg et
al. (1964) found to have elevated urinary porphyrins, Poland et al. (1971) identified one
individual with uroporphyrinuria. The report did not explain if this individual was one of the
four described by Bleiberg et al. (1964). In the NIOSH study that examined workers from
the same plant in New Jersey, the pattern of urinary porphyrin excretion for each participant
was assessed to determine the presence of PCT (Calvert et al., 1993). No difference was
found between workers and an unexposed control group in the prevalence of PCT
(OR=0.93, 95% CI=0.19, 4.54). Furthermore, there were no differences in the risk
between workers and the control group for an out-of-range uroporphyrin concentration or an
out-of-range coproporphyrin concentration. Because this study was conducted at least 15
years after last occupational exposure to TCDD, it was not possible to determine whether
porphyrinuria occurred during the years more proximal to occupational 2,3,7,8-TCDD
exposure.
There is some question as to the appropriateness of the clinical test Bleiberg et al.
(1964) used to measure porphyrin levels. The Watson-Schwartz test was capable of
measuring only the presence of porphobilinogen. The test was rarely positive in cases of
exposure to hepatotoxins. Bleiberg's findings suggest that either other unspecified tests were
used to measure uro- and coproporphyrin levels or the authors misinterpreted the function of
the Watson-Schwartz test.
Jirasek et al. (1974) found 11 of 55 Czechoslovakian TCP production workers to have
elevated urinary uroporphyrins that decreased during the observation period; the authors did
not describe the test used to measure urinary uroporphyrins or coproporphyrins. Ten years
later in a follow-up study, Pazderova-Vejlupkova et al. (1981) found no evidence of
increased excretion of uroporphyrins or dermatological indications of PCT in the same group
of workers.
There is some question that the porphyria noted in the New Jersey (Bleiberg et al.,
1964) and Czechoslovakian workers was due to 2,3,7,8-TCDD exposure. Jones and Chelsky
suggest that the observed cases of PCT in both plants may be due to exposure to HCB or a
combination of both 2,3,7,8-TCDD and HCB (Jones and Chelsky, 1986). HCB was
manufactured at the New Jersey plant from 1951 until 1960 and was produced at the facility
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in Czechoslovakia during the production of pentachlorophenol and TCP (Jirasek et al.,
1973). Evidence of porphyria in other studies of individuals exposed to 2,3,7,8-TCDD-
contaminated substances is minimal. Although Suskind and Hertzberg (1984) sampled urine
for porphyrins in the examination of West Virginia TCP workers, the authors report that the
data are not valid. However, there was no dermatologic evidence of porphyria identified
among the exposed workers. Moses et al. (1984) found no difference in porphyrin levels
when comparing TCP workers with and without chloracne. Finally, no PCT was reported
among three laboratory workers exposed to pure 2,3,7,8-TCDD (Oliver, 1975), the only
humans known to have documented unintentional exposure to uncontaminated
2,3,7,8-TCDD.
In 1977, 60 Seveso residents were tested for elevated porphyrins, exclusive of the
family with inherited deficiency of uroporphyrinogen decarboxylase. None of the 60
residents developed PCT; however, 13 (22%) exhibited secondary coproporphyrinuria, 5 of
whom showed a slight increase of urocarboxyporphyrins, heptacarboxyporphyrins, and
coproporphyrins classified as a "transition constellation to CHP type A" (Doss et al., 1984).
Porphyrin levels were retested in 1980. Porphyrin levels returned to normal in 12
individuals. In three of those with transition CHP, porphyrin levels were higher than those
in 1977 and were attributed to liver damage and alcohol consumption. Doss et al. (1984)
suggested that in the Seveso family with uroporphyrinogen dicarboxylase deficiency, the
exposure to 2,3,7,8-TCDD-contaminated effluent caused an exacerbation of a preexisting
enzyme deficiency.
7.13.2.9. Comment
It is possible that the PCT and elevated urinary porphyrins observed in the New
Jersey and Czechoslovakian workers were a direct result of exposure to hexachlorobenzene.
In the follow-up studies, urinary porphyrin levels in workers were not elevated (Pazderova-
Vejlupkova et al., 1981; Poland et al., 1971) or did not differ from levels in the control
group (Calvert et al., 1992). The transient elevations in coproporphyrins among 22 Seveso
residents described by Doss et al. (1984) may be a direct result of acute exposure to 2,3,7,8-
TCDD.
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The association is not clear. However, 2,3,7,8-TCDD is a potent porphyrigen in rats
and mice and, therefore, high acute exposures may have contributed to the observed changes
in porphyrin levels in these populations.
7.13.2.10. Lipid Levels
Animal studies provide conflicting evidence on the relationship between exposure to
2,3,7,8-TCDD and serum lipid levels. Some studies suggest that short-term high exposure to
2,3,7,8-TCDD increases serum cholesterol (Greig et al., 1973; Zinkl et al., 1973; Poli et al.,
1980; Gasiewicz et al., 1980; Schiller et al., 1986; Gasiewicz and Neal, 1979; Olson et al.,
1980) and triglyceride fractions (Schiller et al., 1986; McConnell et al., 1978a, b; Gasiewicz
and Neal, 1979), whereas other studies suggest a decrease (Gasiewicz et al., 1980; Olson et
al., 1980) or no change in triglyceride levels (Poli et al., 1980). The human data appear to
be similarly confusing. A number of case reports and epidemiologic studies have described
increases in the levels of serum lipid fractions, particularly total cholesterol and triglycerides,
in TCP production workers, laboratory workers, Seveso and Missouri residents, and Vietnam
veterans. Others report no differences between subject and reference levels. A summary of
the reported levels is included in Tables 7-22 and 7-23.
7.13.2.11. Total Cholesterol
Two case reports of workers with presumably high exposures to 2,3,7,8-TCDD-
contaminated chemicals described elevations in total cholesterol. In 50% of 55
Czechoslovakian TCP production workers examined between 1968 and 1969 who exhibited
signs of "chemical intoxication," total cholesterol was noted as elevated (Jirasek et al.,
1974). In a follow-up study 10 years later, lipid levels among workers removed from
exposure were not significantly different from referent levels, but total cholesterol levels
remained significantly increased (Pazderova-Vejlupkova et al., 1981). In a separate report,
three laboratory workers who were exposed during the synthesis of pure 2,3,7,8-TCDD
developed serum cholesterol levels in excess of 7.7 mmol/L (Oliver, 1975). No information
on pre-exposure cholesterol levels was provided in either report.
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Table 7-22. Mean Total Cholesterol Levels Among Seveso and Missouri Residents, TCP Production
Workers, BASF Accident Cohort, and Vietnam Veterans
Author
Mocarelli et
al., 1986
Caramaschi
etal., 1981
Assennato et
al., 1989
May, 1982
Martin, 1984
Poland et
al., 1971
Moses et al.,
1984
Population
Seveso children
1977
1982
Seveso children
Seveso residents
1976
1982
1984
1985
TCP production
workers in Great
Britain
TCP production
workers in Great
Britain
TCP production
workers in New
Jersey
TCP production
workers in West
Virginia
Exposed
N
16"
182b
138
193d
152d
142d
141d
41
39
71
105d
Mean*
level (SD)
4.62 3.26-5.98
4.48 2.47-5.99
15.2c-d
4.78 (0.99)
4.06 (0.80)
4.09 (0.88)
4.14 (0.91)
5.97
6.02g
6.12 (0.82)
5.38 (0.88)
Unexposed
N
28b
250b
120
f
123e
196e
167e
31
126
—
101e
Mean"
level (SD)
4.45 (3.12-5.77)
4.41 (2.99-5.83)
12.5c-e
—
4.14 (0.77)
4.12 (0.86)
4.13 (0.78)
6.6
5.6
— —
5.37 (0.85)
"Units = mmol/L.
bNumber of samples.
c% abnormal.
dChloracne.
"No chloracne.
fNo data for controls in 1976.
gp<0.05.
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Table 7-22. (continued)
Author
Suskind and
Hertzberg,
1984
Ottetal.,
1993b
Calvert et
al., 1993
Hoffman et
al., 1986
Webbetal.,
1989
Centers for
Disease
Control
Vietnam
Experience
Study, 1988a
Roegner et
al., 1991
Population
TCP production
workers in West
Virginia
TCP production
workers:
Chloracne vs.
never chloracne
BASF accident
cohort
TCP and 2,4,5-T
production workers
in Missouri and
New Jersey
Missouri residents
in Quail Run
Mobile Home Park
Missouri residents
<20h
20-6011
>60h
U.S. Army
Vietnam veterans
Air Force Ranch
Hand personnel
Unknown <1&
Low 15-<33.3k
High >33.3k
Exposed
N
200
105d
135
278
142
16
12
12
2,490
338'
191
182
Mean*
level (SD)
5.46 (0.07)
5.44 (0.08)
6.14m (1.01)
5.7"
4.97g (0.96)
5.88 (1.10)
6.60 (0.93)
6.76 (0.97)
5.43;J
5.53
5.55
5.68g
Unexposed
N
163
28"
6,581
259
148
—
1,972
777
Mean"
level (SD)
5.28 (0.08)
5.30 (0.18)
6.37m (1.17)
5.6°
5.2 (1.09)
—
5.36
5.51
hAdipose tissue levels of 2,3,7,8-TCDD in pg/g of lipid.
'Geometric mean.
J% abnormal: Vietnam veterans, 5.1; non-Vietnam
veterans, 4.7; OR=1.1, 95% CI = 0.8-1.5.
kSerum 2,3,7,8-TCDD levels in pg/g of lipid.
'Contrasted to unexposed comparisons.
"Adjusted for age, body mass index,
smoking history.
"Adjusted for age, body mass index,
smoking, gender.
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Table 7-23. Mean Triglyceride Levels Among Seveso and Missouri Residents, TCP Production Workers,
BASF Accident Cohort, and Vietnam Veterans
Author
Mocarelli
etal., 1986
Assennato
etal., 1989
May, 1982
Martin,
1984
Moses et
al., 1984
Suskind
and
Hertzberg,
1984
Ott et al.,
1993b
Calvert et
al., 1993
Population
Seveso children
1977
1982
Seveso residents
1976
1982
1984
1985
TCP production
workers in Great
Britain
TCP production
workers in Great
Britain
TCP production
workers in West
Virginia
TCP production
workers in West
Virginia
BASF accident
cohort
TCP and 2,4,5-T
production workers
in Missouri and
New Jersey
Exposed
N
38b
207b
193
152
142
141
41"
39d
93d
200
135
273
Mean" (SD)
0.97 (0.6-1.50)
0.91 (0.52-1.60)
0.99 (0.43)
0.87 (0.40)
0.94 (0.59)
0.84 (0.44)
2.03
1.97f (0.4-4.0)
1.69g (1.26)
1.65 (0.08)
1.91h (1.19)
1.20 (-)
Unexposed
N
36b
257b
123
196
167
31°
126e
93e
163
4,471
259
Mean" (SD)
0.95 (0.63-1.51)
0.86 (0.47-1.56)
0.85 (0.37)
0.88 (0.46)
0.87 (0.55)
1.83
1.41 (0.3-3.2)
1.46 (0.73)
1.76 (0.08)
1.97h (1.65)
1.151 (-)
"Units = mmol/L.
""Number of samples.
cAdipose tissue 2,3,7,8-TCDD levels in pg/g of lipid.
dChloracne.
eNo chloracne.
fp<0.01.
sp = 0.056.
hAdjusted for age, body mass index, smoking history.
'Adjusted for age, body mass index, smoking, gender.
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Table 7-23. (continued)
Author
Hoffman et
al., 1986
Webb et
al., 1989
Centers for
Disease
Control
Vietnam
Experience
Study,
1988a
Roegner et
al., 1991
Population
Missouri residents
in Quail Run Mobile
Home Park
Missouri residents
<20C
20-60C
>60C
U.S. Army Vietnam
veterans
U.S. Air Force
Ranch Hand
personnel
Unknown <10'
Low 15-<33.3'
High >33.3'
Exposed
N
141
16
12
12
2,490
338
191
182
Mean' (SD)
1.07 (0.73)
2.17 (2.08)
1.81 (1.19)
2.69 (1.06)
1.06i-k
1.02f'm
1.37f-m
1.35f-m
Unexposed
N
146
—
—
—
1972
777
Mean" (SD)
1.19 (1.07)
—
—
—
1.05
1.16
J% Abnormal: Vietnam veterans, 4.7; non-Vietnam veterans, 5.3; OR=0.9, 95% CI=0.7-1.2.
kGeometric mean.
'Serum 2,3,7,8-TCDD levels in pg/g of lipid.
"Contrasted to unexposed comparison population.
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The results of epidemiologic studies conflict. Among British TCP production workers
whose last exposure to 2,3,7,8-TCDD-contaminated chemicals was less than 1 year at the
time of the study, total cholesterol levels in exposed workers with (6.02 mmol/L) and
without (6.14 mmol/L) chloracne were significantly elevated compared to unexposed controls
(5.6 mmol/L) (Martin, 1984) (Table 7-22), whereas May (1982) found unexposed workers
(6.6 mmol/L) to have cholesterol levels higher than those of exposed workers with chloracne
(5.97 mmol/L). Martin (1984) also found reduced, but not significantly, HDL cholesterol
among exposed workers with chloracne (1.19 mmol/L) compared to unexposed controls (1.25
mmol/L). The differences in the results may be due to differences in the control groups and
inclusion of different workers in the exposed groups.
Cholesterol levels in West Virginia TCP production workers were compared with
unexposed workers from the same plant (Suskind and Hertzberg, 1984). No difference was
identified in mean cholesterol levels between workers and controls. However, when lipid
fractions were examined, there was a larger but nonsignificant percentage of exposed
workers with elevated LDL cholesterol (7.7%) compared to unexposed controls (6.3%). A
comparison of workers with persistent chloracne, no chloracne, or a history of chloracne
found a significant association (p<0.05) between the proportion of out-of-range LDL
cholesterol values and persistent chloracne. An out-of-range LDL was defined as above the
90th percentile of the total range of values. In a second study that compared West Virginia
workers with and without chloracne, no difference was found in mean cholesterol levels
(Moses et al., 1984). In the NIOSH study, there was little difference between the adjusted
mean total cholesterol levels for workers (5.7 mmol/L) and referents (5.6 mmol/L) and no
relation to increasing serum 2,3,7,8-TCDD levels (Calvert et al., 1993). The mean levels
were adjusted for age, body mass index, age, and gender.
Mean cholesterol levels were no different between workers in the BASF accident
cohort (6.14 mmol/L) and the referent population (6.37 mmol/L) and were not related to
current or log TCDD back-calculated levels (Ott et al., 1993b). In addition, no significant
differences were noted between the exposed and unexposed populations for HDL and LDL
levels.
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In general, cholesterol levels among exposed community residents were not increased.
Despite their high exposure to 2,3,7,8-TCDD-contaminated TCP, neither children nor adults
from Seveso were found to have elevated serum cholesterol levels compared to controls
(Mocarelli et al., 1986; Assennato et al., 1989). Evaluated from 1976 through 1985,
cholesterol levels in this population remained constant throughout the study period (Table
7-22). Similarly, among Missouri residents, serum cholesterol was not related to residence
in the Quail Run Mobile Home Park (Hoffman et al., 1986) or to adipose tissue 2,3,7,8-
TCDD levels (Webb et al., 1989).
Among U.S. Army veterans, there was no difference in total cholesterol levels
between groups serving in Vietnam or other arenas (Centers for Disease Control Vietnam
Experience Study, 1988a). In contrast, there was a statistically significant positive
relationship between Ranch Hands with serum 2,3,7,8-TCDD levels above 33.3 pg/g and
total cholesterol in Air Force Ranch Hands (Roegner et al., 1991). The total cholesterol-
HDL ratio was also highest in this serum 2,3,7,8-TCDD category.
7.13.2.12. Triglycerides
Elevated triglyceride levels were reported only in three of the studied populations.
Among British TCP workers, triglycerides were significantly higher in exposed workers with
(1.97 mmol/L) and without (1.90 mmol/L) chloracne compared to unexposed controls (1.41
mmol/L) (Martin, 1984). TCP production workers from West Virginia who had chloracne
had statistically nonsignificant increase in mean triglyceride levels (chloracne = 1.69
mmol/L; without chloracne = 1.46 mmol/L) (Moses et al., 1984). In addition, compared to
the unexposed background population, triglyceride levels in Air Force Ranch Hands were
significantly elevated for all serum 2,3,7,8-TCDD categories (Table 7-23). Among workers
in the NIOSH study there appeared to be a small rise in triglyceride levels with increasing
serum 2,3,7,8-TCDD (Calvert et al., 1993). The mean adjusted triglyceride levels and the
percent of abnormal triglyceride values increase with increasing serum 2,3,7,8-TCDD level
(<158 femtograms/liter [fg/L], mean = 1.04 mmol/L, % abnormal = 5.7; 158-520 fg/L,
mean = 1.26 mmol/L, % abnormal = 6.1; 521-1,515 fg/L, mean = 1.23 mmol/L, %
abnormal = 6.1; 1,516-19,717 fg/L, 1.35 mmol/L, % abnormal = 1.7, p<0.05 compared
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to referents [1.15 mmol/L]). Odds ratios and 95% confidence intervals for the quartiles are
OR=0.7 (95% CI=0.5, 12.0), OR=1.1 (95% CI=0.4, 3.2), OR=0.9 (95% CI=0.3, 2.9),
and OR=1.7 (95% CI=0.6, 4.6), respectively. The authors suggest that despite this small
rise with 2,3,7,8-TCDD level, the influence of factors such as gender, body mass index, use
of beta-blocker medication, and smoking had far greater effects on lipid concentration than
did 2,3,7,8-TCDD level. Similarly, triglyceride levels in the BASF accident cohort were
similar to those in the referent cohort and not related to 2,3,7,8-TCDD level (Ott et al.,
1993b).
Triglyceride levels were not elevated in Missouri (Hoffman et al., 1986; Webb et al.,
1989) or Seveso residents (Mocarelli et al., 1986; Assennato et al., 1989) or in U.S. Army
Vietnam veterans (Centers for Disease Control Vietnam Experience Study, 1988a).
7.13.2.13. Comment
The effect of exposure to 2,3,7,8-TCDD-contaminated chemicals on cholesterol or
triglyceride levels is not consistently well defined in the available studies. It is possible the
transient elevations in total cholesterol and triglyceride levels may have occurred after high
2,3,7,8-TCDD exposure, as in the experience of the British and Czechoslovakia TCP
workers and British laboratory workers. However, this scenario does not concur with the
evidence from Seveso or among Ranch Hands. Despite their very high exposure to 2,3,7,8-
TCDD-contaminated chemicals, neither adults nor children from Seveso had lipid levels
above the referent level. In contrast, Ranch Hands continue to have elevated lipid levels
despite the extended length of time between exposure and testing. Other factors, such as
dietary fat intake, familial hypercholesterolemia, alcohol consumption, and exercise, which
also affect cholesterol and other lipid levels, may be factors that were not considered in many
of these studies.
In the 1992 phase of the Ranch Hand study, additional parameters that may affect
lipid levels were examined. It is hoped that this new information will help clarify the
increased levels observed in the Ranch Hand population.
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7.13.3. Other Gastrointestinal Disorders
A variety of gastrointestinal disorders other than liver conditions were reported among
TCDD-exposed groups. After heavy, acute, or chronic exposure, chemical workers in West
Virginia (Ashe and Suskind, 1950), West Germany (Baader and Bauer, 1951; Bauer et al.,
1961), and Czechoslovakia (Jirasek et al., 1974) consistently reported transient episodes of
right upper quadrant pain, loss of appetite, and nausea. None of the reports suggest an
etiology for these symptoms nor were the symptoms reported in later follow-up studies of
any cohorts (Suskind and Hertzberg, 1984; Moses et al., 1984; Pazderova-Vejlupkova et al.,
1981).
Three investigations of TCP production workers reported an increased prevalence of a
history of upper gastrointestinal tract ulcer across all age strata of West Virginia workers
(exposed = 20.7% vs. unexposed = 5.5%) (Suskind and Hertzberg, 1984) and all digestive
system diseases (type not specified) among workers employed in a plant in Midland,
Michigan (prevalence: exposed = 1.5% vs. unexposed = 0.5%) (Bond et al., 1983). The
factors contributing to these conditions have not been examined fully. Neither the Ranch
Hand study (Roegner et al., 1991) nor the NIOSH study (Calvert et al., 1992) found
increased risk of upper gastrointestinal tract ulcers with increasing serum TCDD level.
7.13.4. Thyroid Function
The thyroid plays an essential role in the maintenance of metabolic rate, food intake,
and differentiation and maturation of various cell types. Because many of the toxic effects of
2,3,7,8-TCDD noted in animals resemble the signs of thyroid dysfunction, researchers
considered the role of the thyroid in 2,3,7,8-TCDD toxicity (Neal et al., 1979). Some
studies found a single high dose of 2,3,7,8-TCDD resulted in decreased levels of serum
thyroxine (T4), indicating hypothyroidism, but no consistent findings were reported for
alterations in 3,5,3'-triiodothyronine (T3); researchers report decreases, no change, and
increases in levels of T3 (Bastomsky, 1977; Potter et al., 1983; Pazdernik and Kozman,
1985; Potter et al., 1986; Henry and Gasiewicz, 1987; Roth et al., 1988; Muzi et al., 1989).
Furthermore, hypothyroidism induced in rats was protective against 2,3,7,8-TCDD-induced
weight loss, immunotoxicity, and mortality (Rozman et al., 1984; Pazdernik and Kozman,
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1985). This protective effect was reversed when T4 supplements were given to these animals
(Rozman et al., 1985). A recent study by Henry and Gasiewicz (1987), however, found that
in hamsters serum T3 and T4 levels increased after 2,3,7,8-TCDD administration, putting in
question the role of the thyroid in 2,3,7,8-TCDD-induced toxicity.
These animal findings suggest that if the thyroid plays a role in human toxicity,
hypothyroidism would be manifested as a reduction in serum T4; in extreme cases of
2,3,7,8-TCDD toxicity, however, one may experience hyperthyroidism. Only two studies of
production workers examined this issue (Suskind and Hertzberg, 1984; Ott et al., 1993b).
Suskind and Hertzberg (1984) performed T4 radioimmunoassay and thyroxine-binding
globulin (TBG) tests and found no significant differences between exposed and unexposed
workers. Quantitative results were not presented. Similarly, TSH, T4, and TBG levels were
within normal range in the BASF accident workers, although TGB levels were positively
related to 2,3,7,8-TCDD levels (Ott et al., 1993b). The Ranch Hand study indicated a
reduction of T3% uptake (Table 7-24) and an increase in the mean level of thyroid-
stimulating hormone (TSH) (Table 7-25) with increasing serum 2,3,7,8-TCDD level; these
results, however, did not reach statistical significance. Among Army Vietnam veterans,
mean TSH levels, but not mean free thyroxine index (FTI) levels, were statistically
significantly higher than among non-Vietnam veterans, after adjustment for the six entry
characteristics of age and year of enlistment, race, enlistment status, general technical test
score, and primary military occupation (Table 7-24) (Centers for Disease Control Vietnam
Experience Study, 1988a). However, the percent of values that were out-of-reference range
did not differ significantly for TSH (Vietnam veterans, 1.0%; non-Vietnam veterans, 0.6%,
OR=2.0, 95% CI=0.9-4.3) and FTI (Vietnam veterans, 5.4%; non-Vietnam veterans,
4.6%, OR=1.2, 95% CI=0.9-1.5). The exposure levels were low, based on the sample for
which 2,3,7,8-TCDD was measured.
The most recent study was conducted among infants in The Netherlands (Pluim et al.,
1992). In a letter to the editor, Pluim and colleagues (1992) examined thyroid function
among term breast-fed infants in relation to the total toxic equivalents per kg of breast milk
fat (TEQ/kg) of dioxins (congeners were not specified). The authors measured T4, TBG, and
TSH levels sequentially in infants at birth, at 1 week of age, and at 11 weeks of age (Tables
7-131 06/30/94
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Table 7-24. Levels of Triiodothyronine Percent (T3%) Uptake or Free Thyroxine Index in Vietnam Veterans
Outcome
T3% uptake
Free thyroxine
index
Author
Roegner et
al., 1991
Centers for
Disease
Control
Vietnam
Experience
Study, 1988a
Population
U.S. Air Force
Ranch Hand
personnel
Unknown ^10"
Low 15-<33.3a
High >33.3a
U.S. Army
Vietnam veterans
Exposed
N
338
194
181b
2,490
Mean level
30.67
30.36
29.99
2.2C
Adjusted RR
1.14
0.93
0.45
Adjusted OR
(95% CI)
1.2
(0.9,1.5)
Unexposed
N
772
1,972
Mean level
30.66
2.2C
3
D
O
1
O
c
o
s
O
h—i
a
"Serum 2,3,7,8-TCDD in pg/g of lipid.
bp<0.05 comparison of veterans at background level with ^33.3 pg/g TCDD.
'Geometric mean.
UJ
O
-------
Table 7-25. Levels of Thyroid-Stimulating Hormone (TSH) in Vietnam Veterans, Nursing Infants, and BASF Accident Cohort
Author
Roegneretal., 1991
Centers for Disease
Control Vietnam
Experience Study, 1988a
Pluimet al., 1992
Ott et al., 1993b
Population
U.S. Air Force
Ranch Hand
personnel
Unknown ^ 10b
Low 15-;£33.3b
High >33.3b
U.S. Army
Vietnam veterans
Neonates; The
Netherlands
At birth
1 wk postnatal
1 1 wks postnatal
BASF chemical
workers
Exposed
N
338
194
181
2,490
lle
11
12
130
Mean level*
0.948
0.978
1.026C
1.6d
11.9
2.56
2.50°
1.19
(0.90)h
Adjusted RR
(95% CI)
1.45 (0.62,3.40)
0.88 (0.24,3.02)
2.15(0.80,5.79)
Adjusted OR
(95% CI)
2.0 (0.9-4.3)
1.9f
0.41
0.26
—
Unexposed
N
616
1,972
14*
15
18
i
Mean
level"
0.920
1.6"
10.4
2.93
1.81
SD
—
1.3f
0.41
0.19
—
o
o
2
3
O
c!
O
&
n
'Units: jiU/ml.
bSerum 2,3,7,8-TCDD level in pg/g of lipid.
cp<0.05 compared to low-exposure group.
''Geometric mean.
"High-exposure group: 29.2-62.7 ng toxic equivalents/kg (toxic
equivalencies [TEQ] per kg milk fat).
Standard error of the mean.
gLow-exposure group: 8.7-28 ng TEQ/kg.
Standard deviation.
'No referent values.
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7-25 and 7-26). Infants were classified in the high-exposure group if the breast milk contained
29.2-62.7 ng TEQ/kg and in the low-exposed group if the breast milk contained less than
28.0 ng TEQ/kg. At 1 week and 11 weeks postnatal, T4 and T4/TBG ratios were
significantly higher among infants in the "high-exposure group." At 11 weeks, TSH was
also higher in infants ingesting milk with the higher concentration of dioxins. The authors
suggest that exposure to high levels of dioxins either in utero or through breast milk
modulates the hypothalamic-pituitary-thyroid regulatory system of the infant (Pluim et al.,
1992).
In summary, experimental animal studies provide some evidence that thyroid may
mediate, at least in part, 2,3,7,8-TCDD toxicity. These results were based on animals
receiving a single high dose of 2,3,7,8-TCDD. In addition, studies were performed in only
two species of animals, and the effects may be species specific. The applicability of these
findings to humans, therefore, needs further evaluation.
7.13.4.1. Comment
Few human studies examined the relationship between 2,3,7,8-TCDD exposure and
thyroid function, and the results of present research are equivocal. These studies examined
individuals with relatively lower exposure to 2,3,7,8-TCDD than the animals, and exposure
was chronic. Little or no information has been reported for the effects specifically in
production workers or Seveso residents, two groups with documented high serum 2,3,7,8-
TCDD levels (Fingerhut et al., 1991a; Mocarelli et al., 1991). The data from the study of
Ranch Hands, which measured serum 2,3,7,8-TCDD levels, suggest that in adults there are
few long-term effects. Although the study of nursing infants suggests that ingestion of breast
milk with a higher total TEQ may alter thyroid function, the study is limited by small
numbers and a short observation period (Pluim et al., 1992).
The variety of ways in which 2,3,7,8-TCDD may affect thyroid function in humans
makes it difficult to predict the net outcomes in terms of individual laboratory tests. It is
possible that different relationships may exist among thyroid function indicators, depending
on the exposure level of the individual.
7-134 06/30/94
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Table 7-26. Levels of Thyroxine-Binding Globulin (TBG), Thyroxine (T4), or T4/TBG in Nursing Infants and BASF Accident
Cohort
Outcome
TBG (nmoI/L)
T4 (nmoI/L)
Author
Pluim et al.,
1992
Population
Neonates; The
Netherlands
At birth
1 wk postnatal
1 1 wks postnatal
Neonates; The
Netherlands
At birth
1 wk postnatal
11 wks postnatal
Exposed"
N
15
19
16
15
19
16
Mean level
589.5
546.2
500.7
134.3
178.7C
122.2C
Standard
deviation
30.5d
19.1
13.0
4.8d
5.5
3.0
Unexposedb
N
18
19
18
18
19
18
Mean level
520.1
532.6
519.0
122.5
154.5
111.1
Standard
deviation
27.2d
16.3
29.4
4.1d
6.3
4.0
O
o
H-•
oo
o
>— i
a
"High-exposure group: 29.2-62.7 ng toxic equivalents/kg (TEQ per kg milk fat).
••Low-exposure group: 8.7-28 ng TEQ/kg.
cp<0.05.
Standard error of the mean.
u>
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7-136
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7.13.5. Diabetes
Diabetes and fasting serum glucose levels were evaluated in cross-sectional medical
studies due to the apparently high prevalence of diabetes and abnormal glucose tolerance tests
in one case report of 55 TCP workers (Pazderova-Vejlupkova et al., 1981). In this group,
evaluated 10 years after exposure ended, approximately 50% of the subjects had either
confirmed cases of diabetes or abnormal glucose tolerance tests.
The results of later medical studies are mixed. Cross-sectional studies of workers
from Nitro, West Virginia (Suskind and Hertzberg, 1984; Moses et al., 1984), found no
difference in glucose levels between the exposed and control populations, although no
quantitative values were presented in either study. Similarly, the adjusted odds ratio for out-
of-range fasting glucose levels comparing Vietnam veterans to non-Vietnam veterans was not
statistically significant (OR=1.0, 95% CI=0.4-2.2). But a comparison of the adjusted mean
fasting glucose levels between the two groups was marginally significant (Vietnam veterans,
5.2 mmol/L; non-Vietnam veterans, 5.1 mmol/L, p<0.05) (Centers for Disease Control
Vietnam Experience Study, 1988a). Mean fasting glucose levels in the BASF accident cohort
were marginally elevated compared to the referent population and associated with current
levels of 2,3,7,8-TCDD (p=0.062) but not the back-extrapolated level (Ott et al., 1993b).
Results from the NIOSH (Sweeney et al., 1992) and Ranch Hand studies (Roegner et al.,
1991) more strongly suggest that serum 2,3,7,8-TCDD levels may be positively and
significantly related to diabetes and fasting serum glucose levels.
In the NIOSH study, the relationship between exposure to 2,3,7,8-TCDD and possible
alterations in glucose metabolism was assessed using two outcome measures: case definition
of diabetes and fasting serum glucose levels. Participants met the case definition of diabetes
if their fasting serum glucose level was 7.8 mmol/L or greater on two consecutive occasions
(National Diabetes Data Group, 1979) or if they reported a positive history of physician-
diagnosed diabetes after the date of first employment at the plant. Based on logistic
regression, the risk of diabetes was found to increase by approximately 12% for every 100
pg/g 2,3,7,8-TCDD (p<0.001) (Table 7-27). Adjusted fasting serum glucose levels in
workers and referents were also compared in a separate analysis using linear regression.
Fasting serum glucose was 10% higher in workers with 1,000 pg/g 2,3,7,8-TCDD than in
7-137 06/30/94
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Table 7-27. Multiple Logistic Regression Model for Cases of Diabetes" for TCP Production Workers
Variable
Intercept"1
Serum TCDD
Age
Body mass index
Race
0= other, 1= white
Family history of
diabetes
(0=no, l=yes)
Beta
-8.8913
0.00114
0.0776
0.0996
-1.6659
0.6957
Standard
error
1.7131
0.0003
0.0204
0.0370
0.4706
0.3641
Odds
ratiob
—
1.12
2.17
1.65
0.19
2.01
p-Valuec
0.001
<0.001
< 0.001
0.007
<0.001
0.056
"History of physician-diagnosed diabetes (diagnosed after exposure to 2,3,7,8-TCDD-contaminated
chemicals) or fasting serum glucose level of 7.8 mmol/L on 2 consecutive days.
bOdds ratios for continuous variables are based on the difference of 100 pg per gram of lipid of TCDD,
10 years of age, and 5 kg/m2 body mass index.
Cp-Values are based on Wald Chi-square statistic.
•"Number of observations = 530.
Source: Sweeney et al., 1992.
7-138
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DRAFT-DO NOT QUOTE OR CITE
workers and referents with 20 pg/g 2,3,7,8-TCDD (p<0.001) (Table 7-28). Yet, based on
the magnitude of the regression coefficients, age, and body mass index (kg/m2), both known
risk factors for diabetes appear to have greater influence on the increases in both the risk of
diabetes and elevated fasting serum glucose levels than 2,3,7,8-TCDD level.
In the Ranch Hand study, diabetic status was assessed by measuring fasting serum
glucose and 2-hour postprandial glucose and by using a case definition of diabetes. Diabetes
was defined as having a verified history of diabetes or an oral glucose tolerance test of
> 11.1 mmol/L (200 mg/dL) (Roegner et al., 1991). The analyses of all three parameters
suggest a consistent association between serum 2,3,7,8-TCDD levels above 33.3 pg/g and an
increased risk of diabetes. Adjusted relative risks comparing Ranch Hands with serum
2,3,7,8-TCDD above 33.3 pg/g and the unexposed comparison group were statistically
significantly elevated for fasting serum glucose levels and diabetes (glucose, RR=2.95,
p<0.001; diabetes, RR=2.51, p< 0.001) (Table 7-29) and for the 2-hour postprandial
glucose test (RR=2.35, p<0.035). In addition, Ranch Hand personnel meeting the case
definition for diabetes were also more likely to have earlier onset of diabetes than the
unexposed comparisons (Wolfe et al., 1992a).
7.13.5.1. Comment
The limited information on the risk of diabetes or other alterations in glucose
metabolism in humans in relation to 2,3,7,8-TCDD is not well addressed by the available
animal literature, of which there is very little. The effects of 2,3,7,8-TCDD on glucose
metabolism have been evaluated only in a few laboratory studies. Although these studies
suggest that 2,3,7,8-TCDD may alter glucose metabolism, for the most part the animal
studies do not corroborate the direction of the findings of two cross-sectional medical studies
(Wolfe et al., 1992a; Sweeney et al., 1992). Animal studies appear to contradict the human
data, suggesting that 2,3,7,8-TCDD lowers glucose levels. Several studies of rats and rhesus
monkeys showed consistent decreases in serum glucose levels after daily doses administered
over 30 days (Zinkl et al., 1973) or after a single dose of 2,3,7,8-TCDD (McConnell et al.,
1978a; Gasiewicz et al., 1980; Schiller et al., 1986; Ebner et al., 1988) (Table 7-30). In
one study glucose levels continued to drop up to 3 weeks postexposure (Gorski et al., 1990).
7-139 06/30/94
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Table 7-28. Multiple Linear Regression Model for Log Fasting Serum Glucose for TCP Production
Workers and Unexposed Referents
Variable
Intercept2
Serum TCDD
Age
Body mass index
Parameter estimate
4.2112
0.00008
0.00288
0.0066
Ratio
0.0548
0.00002
0.00064
0.0014
p-Value
0.0001
<0.001
<0.001
< 0.007
Adjusted R2 = 0.10.
"Number of observations = 508.
Source: Sweeney et al., 1992.
7-140
06/30/94
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Table 7-29. Adjusted Relative Risk (RR) for Fasting Serum Glucose Levels, Cases of Diabetes, and Mean 2-Hour Postprandial
Glucose Levels by Category of Lipid Adjusted Serum 2,3,7,8-TCDD in Ranch Hands
Category serum
2,3,7,8-TCDD (pg/g)
Unknown <10C
Low 15-<33.3C
High >33.3C
Fasting serum glucose (RR)
0.66
1.18
2.95d
Diabetes
(RR)"
0.82
1.01
2.51d
2-Hour postprandial
glucose level
(RR)b
0.88
0.60
2.35e
Tl
H
O
O
2
O
H
O
C
1
g
O
t—l
H
"Defined as having a verified history of diabetes or 2-hour postprandial glucose level of > 11.1 mmol/L (200 mg/dL).
kComparison of diabetics and normals.
cSerum 2,3,7,8-TCDD levels in pg/g of lipid.
dp<0.001.
^=0.035.
Source: Roegner et al., 1991.
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DRAFT-DO NOT QUOTE OR CITE
Table 7-30. Effects of Exposure to 2,3,7,8-TCDD on Serum Glucose Levels in Nonhuman Mammalian
Species
Author
Zinkl et al.,
1993
Gasiewicz et
al., 1980
Schiller et
al., 1986
Gorski et al.,
1990
McConnell et
al., 1978a
DeCaprio et
al., 1986
Ebner et al.,
1988
Species
CD rat
Rats
Fischer rat
Sprague-
Dawley rat
Rhesus
monkey
Hartley
guinea pig
New Zealand
rabbits
Route
oral
ip
(TPN)
ip
(chowfed)
gavage
ip
gavage
oral
ip
Dose (/x/kg)
0.1
1.0
10.1
100C
30
60
90
180°
270°
360°
125°
70
350C
0.01
0.06
0.44
1
50
Duration
Ix/day for 30 days
1 time
1 time
1 time
1 time
90 days
1 time
Percentage of
serum glucose
levels in control
animals"
91b
71
51
29d
51d
74d-e
54d-e
30d,e
43d.e
38d>e
39d,e
day 4 75M
day 8 67M
day 16 50M
day 21 31M
(decreased)b-d
NCb,e,f
NCb,e,f
NCb,e,f
NC
15 min NC
1 hour 125d'e
2 day 87
10 day SO*'
"Relative to control values.
bFemale animals.
cLethal dose.
•"Significantly different from controls.
"Male animals.
fData not displayed.
ip = intraperitoneal.
TPN = total parenteral nutrition.
NC = no change.
7-142
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Dose-related decreases were also noted in CD rats fed 0.1, 1.0, or 10.0 /ig/kg daily for 30
days (Zinkl et al., 1973). In contrast, lower daily doses of approximately 0.004 /ig/kg/day
administered to guinea pigs over a 90-day period produced no significant changes in either
glucose or insulin levels (DeCaprio et al., 1986). The results of the animal studies suggest
that glucose levels are altered, generally decreased, by short-term, high-dose exposure to
2,3,7,8-TCDD but that the response may be species specific. The diametrically opposed
findings of both human and most animal studies may, hypothetically, be due to a number of
factors, for example, the species studied, the length of the exposure and short observation
periods, the rate of insulin metabolism after 2,3,7,8-TCDD exposure, possible differential
effects of 2,3,7,8-TCDD on the various types of islet cells, and the high, usually single,
2,3,7,8-TCDD dose.
The outcome measures used in the NIOSH and the Ranch Hand studies, case of
diabetes and fasting serum glucose, did not permit a determination of the type of diabetes
involved. However, both studies show that highly 2,3,7,8-TCDD-exposed individuals may
be at a slightly greater risk for developing diabetes than individuals with background levels
of 2,3,7,8-TCDD. They also experience a greater prevalence of elevated fasting glucose
levels, which may be a precursor to conversion to a diabetic state. However, in the NIOSH
study, the traditional risk factors for diabetes—age, body mass index or weight, and family
history of diabetes—appear substantially more influential than 2,3,7,8-TCDD in the
development of diabetes.
7.13.6. Immunologic Effects
Animal toxicologic studies have demonstrated numerous immunologic effects after
exposure to 2,3,7,8-TCDD (see Chapter 4 for a more comprehensive review). In humans,
the information with which to assess the immunologic consequences of exposure is sparse.
Six epidemiologic studies and one case report have described the immunologic function of
populations exposed to 2,3,7,8-TCDD (Evans et al., 1988; Hoffman et al., 1986; Webb et
al., 1989; Centers for Disease Control Vietnam Experience Study, 1988a; Roegner et al.,
1991; Jennings et al., 1988; Reggiani, 1978; Ott et al., 1993b) and as has one study of
extruder personnel exposed to brominated dioxins and furans (Zober et al., 1992).
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Evaluation of the immunologic status in exposed residential populations has not found
a relationship between exposure and impaired status. Immunocompetence was tested twice in
44 children who were residents of the region of Seveso with the highest 2,3,7,8-TCDD
contamination and in 43 age-matched children who did not reside in the contaminated area
(Reggiani, 1978). Twenty of the exposed children had chloracne and 24 had no skin lesions.
The tests included serum immunoglobulins, complement levels, lymphocyte subpopulations,
and lymphocyte activity analysis. Although no data were presented, the authors reported that
the various measures were within normal range and that there was no difference between the
two groups.
Initial studies of Missouri residents who had the potential for exposure to 2,3,7,8-
TCDD-contaminated soil suggested that 2,3,7,8-TCDD caused depression in cell-mediated
immunity (delayed hypersensitivity), as demonstrated by a statistically significant increase in
anergy (exposed vs. nonexposed: 11.8% vs. 1.1%) (Hoffman et al., 1986). This study,
however, was limited by the exclusion from test results of 61% and 32% of the exposed and
unexposed groups, respectively. A follow-up study confirmed the presence of substantial
bias in the first study. Evans et al. (1988) retested 28 of the 50 exposed residents and 15 of
the 27 unexposed residents who did not respond (anergic) or responded weakly (relatively
weakly) to an antigen challenge in the first study. A follow-up study could not confirm the
presence of anergy. Both studies found in the exposed residents increased frequencies in
CD4/CD8 ratio of less than 1.0 (Table 7-31). No other abnormalities were noted by
Hoffman et al. (1986) (Tables 7-32 to 7-37).
In a later study of Missouri residents, Webb et al. (1989) found no clinical evidence
of immunosuppression in 40 individuals whose adipose 2,3,7,8-TCDD levels ranged from
under 20 pg/g to over 430 pg/g (top of range not given). Tests included serum
immunoglobulins, T-cell surface markers OKT3, OKT4, OKT8, OKTll, Leullc, CD4/CD8
ratio, (CD4 + CD8)/CD3, and Bl and B2 cells. In logistic regression, significant (p<0.05)
relationships were noted for IgG, %CD3, %T11, %CD8, and %CD4 + LEU8 POS,
controlling for age and sex (Tables 7-32 to 7-37).
The effect of past occupational exposure on immunologic function was examined in
18 British workers who were evaluated 17 years after accidental industrial exposure to
7-144 06/30/94
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Table 7-31. CD4/CD8 Ratios in Missouri Residents, Vietnam Veterans, and BASF Accident Cohort
Author
Roegner et al.,
1991
Centers for
Disease Control
Vietnam
Experience Study,
1988a
Hoffman et al.,
1986
Webbetal., 1989
Ottetal., 1993b
Zober et al., 1992
Population
U.S. Air Force Ranch
Hand personnel
Unknown <1Q?
Low 15-<33.3a
High >33.3a
U.S. Army Vietnam
veterans
Missouri residents
Missouri residents
<20C
20-60C
>60C
BASF accident cohort
BASF personnel
exposed to TBDD and
TBDFe
Exposed
N
126
72
72
2,490
135
16
12
12
132
21
Mean level
(SD)
1.72
1.91
1.99
1.8b
1.9 (0.8)
2.0 (0.7)
2.1 (1.0)
1.4 (0.7)
1.6 (0.94)
1.6 (0.5)
Ratio
__.
OR < Reference Range
0.9
OR > Reference Range
1.1
% Abnormal
8.2
—
—
Unexposed
N
301
1,972
142
42d
42
Mean level
(SD)
1.89
1.8b
1.9 (0.6)
1.5 (0.6)
1.5 (0.6)
Ratio
% Abnormal
6.3
—
—
O
o
z
9
O
d
3
w
o
I-H
H
tfl
O
"Serum 2,3,7, 8-TCDD level in pg/g of lipid.
""Geometric mean.
'Adipose 2,3,7,8-TCDD level in pg/g of lipid.
dFrom Zober et al., 1992.
= 2,3,7, 8-tetrabrorninated dibenzo-o-dioxins and TBDF = 2,3,7,8-tetrabrominated dibenzofurans.
-------
Table 7-32. Total Lymphocytes in 2,4,5-T Production Workers, Missouri Residents, Vietnam Veterans, and BASF Accident
Cohort
^1
I
ON
Author
Roegner et al.,
1991
Centers for
Disease Control
Vietnam
Experience
Study, 1988a
Hoffman et al.,
1986
Webbetal.,
1989
Jennings et al.,
1988
Population
U.S. Air Force
Ranch Hand
personnel
Unknown <10C
Low 15-<33.3C
High >33.3C
U.S. Army
Vietnam veterans
Missouri
residents
Missouri
residents
<20e
20-60"
>60e
2,4,5-T
production
workers exposed
17 yrs prior to
the study
Exposed
N
127
73
74
2,490
135
16
12
12
18
Mean level"
(SD)b
1,954 — d
2,011 —
2,032 —
1,973 —
2,465 (724)
2,200 (830)
2,300 (600)
2,200 (720)
1,980 (840)
Ratio
—
OR < Reference Range
1.0
OR > Reference Range
1.2
—
% Lymphocytes
32
32
28
Unexposed
N
301
1,972
142
—
15
Mean level (SD)
1,972 —
1,936 —
2,311(634)
—
2,020 (470)
Ratio
Tl
H
I
O
o
o
c
3
w
o
»
n
3
w
o
vo
"Units: counts/mm3.
bSD = standard deviation.
cSerum 2,3,7,8-TCDD level in pg/g of lipid.
d— = data not presented.
"Adipose 2,3,7,8-TCDD level in pg/g of lipid.
fFrom Zober et al., 1992.
STBDD = 2,3,7,8-Tetrabrominated dibenzo-/>-dioxins;
TBDF = 2,3,7,8-Tetrabrominated dibenzofurans.
-------
Table 7-32. (continued)
Author
Ottetal., 1993b
Zober et al. ,
1992
Population
BASF accident
cohort
BASF personnel
exposed to
TBDD* and
TBDFs
Exposed
N
133
21
Mean level"
(SD)b
1,978.3 (805)
2,179.5(678)
Ratio
% Lymphocytes
33.4 (9.4)
% Lymphocytes
33.4 (8.4)
Unexposed
N
42
42
Mean level (SD)
2,267.6 (837.5)
2,267.6 (837.5)
Ratio
% Lymphocytes
36 (12.4)
% Lymphocytes
36 (12.4)
o
3
o
o\
u>
o
9
W
8
O
-------
Table 7-33. Bl Levels in Production Workers, 2,4,5-T Missouri Residents, Vietnam Veterans, BASF Accident Cohort, and
Extruder Personnel
Author
Roegner et al.,
1991
Centers for
Disease Control
Vietnam
Experience
Study, 1988a
Webbetal.,
1989
Jennings et al. ,
1988
Population
U.S. Air Force
Ranch Hand
personnel
Unknown <10b
Low 15-33.3b
U.S. Army
Vietnam veterans
Missouri residents
<20d
20-60d
>60d
2,4,5-T production
workers exposed
17 yrs prior to the
study
Exposed
N
127
71
73
2,490
16
12
12
18
Mean level" (SD)
176 —
183 —
191 —
240C
190 (865)
189 (983)
171 (573)
210 (110)
Ratio
OR < Reference Range
1.1
OR > Reference Range
1.2
% Bl cells
9.1
8.3
7.8
—
Unexposed
N
301
1,972
—
15
Mean level" (SD)
172 —
230C —
—
160 (80)
Ratio
O
O
C!
1
§
n
M
-------
DRAFT-DO NOT QUOTE OR CITE
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c
• HI
I
^
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ti-
_4j
8
^
1
so
•M
3
Q
c/5.
%
_w
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o
i
I
s
s
-
z
e
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I
u
J
•+•*
i
,~,
s
o"
2
3
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o"
O
(^J
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1
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^
8
^
w ^*
CQ d
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fN
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1 i
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« s 5 Q
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®s FT< C"^ ^
7-149
06/30/94
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Table 7-34. CD4 Levels in Production Workers, Missouri Residents, Vietnam Veterans, and BASF Chemical Workers
Author
Roegner et
al., 1991
Centers for
Disease
Control
Vietnam
Experience
Study, 1988a
Hoffman et
al., 1986
Webb et al.,
1989
Population
U.S. Air Force
Ranch Hand
personnel
Unknown <10b
Low 15-<33.3b
High >33.3b
U.S. Army
Vietnam
veterans
Missouri
residents
Missouri
residents
<20d
20-60d
>60d
Exposed
N
127
72
72
2,490
135
16
12
12
Mean level" (SD)
867 —
945 —
929 —
1,020° —
1,021 (353)
1,084 (485)
1,198 (391)
963 (403)
Ratio
OR < Reference Range
1.0
OR > Reference Range
1.4
% Abnormal
0.7
% T4 cells
48
51
42
Unexposed
N
301
1,972
142
—
Mean level (SD)
907 —
990° —
1,033 (346)
—
Ratio
«._
—
% Abnormal
0.0
—
O
o
I
o
g
o
H
"Units: cells/mm3.
bSerum 2,3,7,8-TCDD level in pg/g of lipid.
cGeometric mean.
dAdipose 2,3,7,8-TCDD level in pg/g of lipid.
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-------
DRAFT-DO NOT QUOTE OR CITE
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7-151
06/30/94
-------
Table 7-35. CDS Levels in 2,4,5-T Production Workers, Missouri Residents, Vietnam Veterans, and BASF Accident Cohort
Author
Roegner et
al., 1991
Centers for
Disease
Control
Vietnam
Experience
Study,
1988a
Hoffman et
al., 1986
Webb et
al., 1989
Population
U.S. Air Force
Ranch Hand
personnel
Unknown <10b
Low 15-<33.3b
High >33.3b
U.S. Army
Vietnam veterans
Missouri residents
Missouri residents
<20d
20-60d
>60d
Exposed
N
126
71
73
2,490
135
16
12
12
Mean level" (SD)
485 —
465 ---
475 —
560C —
592 (223)
562 (215)e
645 (225)
807 (381)
Ratio
OR < Reference Range
1.0
OR > Reference Range
0.9
% Abnormal
1.5
% T8 cells
26
28
35
Unexposed
N
301
1,972
142
Mean level" (SD)
473 —
550 —
578 (198)
Ratio
% Abnormal
0.0
O
o
3
3
o
W
O
»
o
3
w
"Units: cells/mm3.
bSerum 2,3,7,8-TCDD level in pg/g of lipid.
'Geometric mean.
dAdipose tissue 2,3,7,8-TCDD level in pg/g of lipid.
ep<0.05.
-------
DRAFT-DO NOT QUOTE OR CITE
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BASF personnel
exposed to TBDD8
and TBDF*
« $
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N 13
7-153
06/30/94
-------
Table 7-36. IgG Levels in Missouri Residents, Vietnam Veterans, and BASF Accident Cohort
Author
Roegner et al. ,
1991
Centers for Disease
Control Vietnam
Experience Study,
1988a
Webbetal., 1989
Ottetal., 1993b
Zoberet al., 1992
Population
U.S. Air Force Ranch
Hand personnel
Unknown ^10b
Low 15-<33.3b
High >33.3b
U.S. Army Vietnam
veterans
Missouri residents
<20d
20-60d
>60d
BASF accident cohort
BASF personnel
exposed to TBDDg
and TBDF8
Exposed
N
335
190
175
2,490
16
12
12
132
21
Mean level' (SD)
1,087 —
1,122 —
1,122 —
1,078C
l,064e (273)
1,146 (193)
1,151 (223)
l,199f (226)
1,057.7 (199.0)
Ratio
—
OR < Reference Range
1.0
OR > Reference Range
1.0
—
—
—
Unexposed
N
757
1,972
—
194
42
Mean level" (SD)
1,120
1,077C
—
1,182.6 (310.0)
1,102.9(207.1)
o
O
tn
8
o
"Units: mg/dl.
bSerum 2,3,7,8-TCDD level in pg/g of lipid.
cGeometric mean.
^Adipose 2,3,7,8-TCDD level in pg/g of lipid.
•p<0.05 trend.
'Significant positive relation between IgG and current 2,3,7,8-TCDD level and back-extrapolated 2,3,7,8-TCDD level (p<0.01).
«TBDD = 2,3,7,8-Tetrabrominated dibenzo-^-dioxins;
TBDF = 2,3,7,8-Tetrabrominated dibenzofurans.
-------
Table 7-37. IgM Levels in Missouri Residents, Vietnam Veterans, BASF Accident Cohort, and Extruder Personnel
Author
Roegner et al.,
1991
Centers for Disease
Control Vietnam
Experience Study,
1988a
Webbetal., 1989
Ottetal., 1993b
Zober et al., 1992
Population
U.S. Air Force Ranch
Hand personnel
Unknown <10b
Low 15-<33.3b
High >33.3b
U.S. Army Vietnam
veterans
Missouri residents
<20d
20-60d
>60d
BASF accident cohort
Personnel exposed to
TBDD" and TBDFe
Exposed
N
335
190
175
2,490
16
12
12
132
21
Mean level*
(SD)
107 —
96 —
106 —
121C —
128 (89)
157 (57)
1 14 (44)
139.6
(65.1)
142f (52.6)
Ratio
OR < Reference Range
1.0
OR > Reference Range
1.0
—
—
—
Unexposed
N
757
1,972
—
192
42
Mean level*
(SD)
103
121C
—
134.7 (70)
114.7f
(46.5)
0
o
o
tfl
§
n
"Units: mg/dl.
bSerum 2,3,7,8-TCDD level in pg/g of lipid.
'Geometric mean.
dAdipose 2,3,7,8-TCDD level in pg/g of lipid.
TBDD = 2,3,7,8-Tetrabrominated dibenzo-^-dioxins;
TBDF = 2,3,7,8-Tetrabrominated dibenzofurans.
fp=0.04.
-------
DRAFT-DO NOT QUOTE OR CITE
chemicals contaminated with 2,3,7,8-TCDD (Jennings et al., 1988). It is not clear from the
article when occupational exposure to 2,3,7,8-TCDD ended for these workers. Exposed
workers and unexposed controls were matched for age, race, sex, smoking habit, alcohol
consumption, and percent of ideal body weight. There were no significant differences in the
levels of immunoglobulins, T and B lymphocytes, responsiveness to phytohemagglutinin A,
and in the number of CD4 and CDS counts. Three measures were found to be statistically
significantly (p<0.05) higher in workers than in controls: antinuclear antibodies (ANA) (8
workers vs. 0 controls, p<0.01) (when Hep2 cells were used as substrate but not when rat
liver cells were used), immune complexes (workers = 11 vs. 3 controls, p<0.05), and
natural killer cells (NK) (workers = 0.21 X 106/L vs. controls = 0.59 x 106/L, p<0.002)
identified by the monoclonal antibody Leu-7 (Table 7-38). No other well-conducted studies
of exposed individuals have measured these parameters. In the discussion, the authors could
not explain the physiologic basis of their findings and suggested that further research was
needed.
Among participants in the BASF accident study cohort, with the exception of natural
killer cells and helper-inducer cells, the proportions of some lymphocyte populations (B cells,
T cells, T helper cells, T suppressor cells) were lower among workers, but the distribution of
cells in referents and workers was equivalent (Ott et al., 1993b). Levels for IgA, IgG, IgM,
and complement C4 and C3 were slightly higher in workers than in the unexposed referent
population. There also appeared to be slight dose-related increases in IgA, IgG, and
complement C4 with 2,3,7,8-TCDD levels measured in October 1988 and February 1992.
IgA and IgG were related to the half-life extrapolated 2,3,7,8-TCDD levels. It must be
noted, however, that the statistically significant relationship between IgA and 2,3,7,8-TCDD
is most likely due to a case of liver cirrhosis and the association with IgG due to a liver
carcinoma.
Among the 21 extruder personnel exposed to both 2,3,7,8-TBDD and 2,3,7,8-TBDF,
of the 16 parameters tested, only complement C4 was statistically significantly (p<0.01)
associated with concentrations of 2,3,7,8-TBDD and 2,3,7,8-TBDF (Zober et al., 1992).
Borderline associations were noted between 2,3,7,8-TBDF and decreases in total lymphocyte
count (p=0.056), T-cell count (p=0.045), T-helper cell count (CD4) (p=0.045) and an
7-156 06/30/94
-------
Table 7-38. Levels of Natural Killer Cells in Missouri Residents, Vietnam Veterans, and Extruder Personnel
o
o>
Ui
o
Author
Roegner etal., 1991
Jennings et al., 1988
Zoberetal., 1992
Population
U.S. Air Force Ranch
Hand personnel
Unknown £10"
Low 15-<£33.3"
High >33.3"
2,4,5-T production
workers exposed 17
yrs prior to the study
BASF personnel
exposed to TBDDf and
TBDFf
Exposed
N
126
70
72
18
21
Mean level (SD)
455b-c
378
386
400° (210)d
280.5 (175.2)
Unexposed
N
291
15
42
Mean level (SD)
414b
590° (230)
250.8 (149.9)
O
O
1
ffl
O
X)
O
H
w
•Serum 2,3,7,8-TCDD level in pg/g of lipid.
bUnits: cpm.
cNet response.
dp<0.05.
"Units: 103/mm3.
TBDD = 2,3,7,8-Tetrabrominated dibenzo-/?-dioxins;
TBDF = 2,3,7,8-Tetrabrominated dibenzofiirans.
-------
DRAFT-DO NOT QUOTE OR CITE
increase in complement C3 (0.054), and between 2,3,7,8-TBDD and a decreased percent
lymphocyte count (p=0.054). However, with the exception of complement C3, the
associations appear to be driven by a single individual with the highest 2,3,7,8-TBDD levels
(478 pg/g of lipid) who, at the time of the study, exhibited no evidence of clinical
immunodeficiency.
Two studies extensively evaluated immunologic function in Vietnam veterans. No
significant differences were noted among U.S. Army ground troops and the comparison
population in lymphocyte subset populations, T-cell populations, or serum immunoglobulins
(Tables 7-31 to 7-37) (Centers for Disease Control Vietnam Experience Study, 1988a).
Comprehensive immunologic profiles were developed for each participant of the U.S. Air
Force Ranch Hand Study (Tables 7-31 to 7-38) (Roegner et al., 1991). Significant positive
associations were found between IgA and serum 2,3,7,8-TCDD. The authors suggest that
the rise in IgA is consistent with a subclinical inflammatory response, but the authors could
not explain the source of the inflammatory response.
7.13.6.1. Comment
The information on immunologic function in humans, children or adults, relative to
exposure to 2,3,7,8-TCDD is scarce. The available epidemiologic studies are restricted to
adults and do not describe a consistent pattern of effects among the studies. Natural killer
cells (NK) were increased among one population of 2,3,7,8-TCDD chemical workers
examined 17 years after exposure ended (Jennings et al., 1988). These findings were not
corroborated in Ranch Hands (Roegner et al., 1991), the BASF accident cohort (Ott et al.,
1993b), or workers exposed to 2,3,7,8-TBDD and TBDF (Zober et al., 1992). Dose-related
elevations in IgA were observed in Ranch Hands in relation to current levels and in the
BASF accident cohort with respect to both current and half-life extrapolated 2,3,7,8-TCDD
levels. Yet, IgA was not higher in adult Missouri residents with adipose 2,3,7,8-TCDD
levels above background (Webb et al., 1989). IgG was also significantly related to 2,3,7,8-
TCDD in the BASF accident cohort but not in Ranch Hands. While complement C3 and C4
were elevated in both the BASF accident cohort and the extruder personnel, no other study
examined these end points.
7-158 06/30/94
-------
DRAFT-DO NOT QUOTE OR CITE
The effect of acute, high exposure to 2,3,7,8-TCDD among children from Seveso was
reportedly negative within 2 years after exposure (Reggiani, 1978). Although no data have
been published illustrating the values obtained from the tests of immunologic function in
these children, the author indicates that the measured parameters were no different in the
exposed and unexposed children after two series of tests.
More comprehensive evaluations of immunologic function with respect to 2,3,7,8-
TCDD exposure are necessary to assess more definitively the relationships observed in
nonhuman species. Longitudinal studies of the maturing human immunologic system may
provide the greatest insight, particularly since animal studies have found many of their
significant results in immature animals and breast milk is a source of 2,3,7,8-TCDD and
other related compounds. Additional studies of highly exposed adults may also shed light on
the effects of long-term chronic exposures. Therefore, there appears to be too little
information to suggest definitively that 2,3,7,8-TCDD, at the levels observed, is a
immunotoxin in humans.
7.13.7. Neurologic Effects
Although there are few studies reporting neurologic abnormalities related to
2,3,7,8-TCDD exposure in adult animal models, neurologic effects are reported to have
occurred shortly after exposure in occupationally exposed individuals (Ashe and Suskind,
1950; Baader and Bauer, 1951; Bauer et al., 1961; Goldman, 1972; Jirasek et al., 1974;
Oliver, 1975; Pocchiari et al., 1979) (Table 7-39) and in Seveso residents (Filippini et al.,
1981) (Table 7-40). Previous case reports and studies found symptoms referable to the
central (CNS) and peripheral (PNS) nervous system among workers and community residents
exposed to 2,3,7,8-TCDD-contaminated materials. While human studies reveal a wide
spectrum of effects due to 2,3,7,8-TCDD (Sweeney et al., 1989), very few toxicologic
studies have focused on the nervous system. Singer et al. (1982) reviewed the following
animal studies, which examined the relationship of CNS dysfunction and 2,3,7,8-TCDD
exposure: Elovaara et al. (1977) found anomalous CNS function in some rats exposed to a
single dose of 2,3,7,8-TCDD, and Creso et al. (1978) reported the CNS symptoms of
irritability, restlessness, and increased aggression in rats administered 2,3,7,8-TCDD. No
7-159 06/30/94
-------
Table 7-39. Case Reports of Psychological and Neurologic Effects Among Individuals Exposed to 2,3,7,8-TCDD-Contaminated
Materials
Author
Ashe and Suskind,
1950
West Virginia
Suskind et al., 1953
West Virginia
Baader and Bauer,
1951
Nordheim-
Westfalan, West
Germany
Exposure
Precursor and reaction
products of TCP process
TCP,3 NaOH,b 2,4,5-T*
TCP
2,4,5-T
PCPd
TCP
Population
4 employees involved in
cleanup after 1949
explosion or worked with
equipment used prior to
explosion; studied 6 mos
and 1 yr after explosion
36 workers with chloracne
after TCP reactor explosion
25 persons exposed during
production of TCP and
2,4,5-T
1948-1953
10 men in PCP production
experimental TCP
production
Evaluation
Clinical history and
examination
Clinical history and
examination
Record review case
history
Findings
Headaches
Insomnia
Social withdrawal
Nervousness
Fatigue
Crying spells
Numbness in feet
Decreased libido, muscle
strength, sensory ability,
myelin sheath and fiber
destruction of sural nerve
Fatigue
Nervousness
Irritability
Reduced libido
Back and limb pain
Vertigo
Paresthesia
Self-reported: pain and
weakness
Paresthesia and pain in
gluteal and femoral region
r ™
>
3
6
o
1
o
o
w
g
n
ON
o
u>
o
vo
TCP = 2,4,5-trichlorophenol.
bNaOH = sodium hydroxide.
C2,4,5-T = 2,4,5-trichlorophenoxyacetic acid.
dPCP = pentachlorophenol.
-------
Table 7-39. (continued)
Author
Bauer etal., 1961
Hamburg, West
Germany
Goldman, 1972
Ludwigshafen,
West Germany
Exposure
TCP
2,4,5-T
TCP
2,4,5-T
Population
31 men in TCP and 2,4,5-T
production with residual
chloracne, continual
"neuromuscular weakness, "
vasovagal disturbances,
psychopathological
disturbances" 5 yrs after
first exam
42 BASF workers with
chloracne after TCP reactor
explosion in 1953
Evaluation
Clinical history and
examination
Hamburg-Wechsler
Test
Rorschach
Psychogram
Clinical history and
examination
Findings
Reduced libido
Headaches, dizziness
Decreased libido
Irritability
Depression
Sleep disturbances
Anorexia
Paresthesia
Tremor
Muscle weakness
Hamburg-Wechsler
indicative of acquired
decrease in mental
efficiency
Rorschach Psychogram
indicative of decreased
emotional reactivity,
slowed thinking process for
concentration
Neurasthenia
ON
o
o
z
3
i
o
-------
Table 7-39. (continued)
Author
Jirasek et al., 1974
Spolano,
Czechoslovakia
Oliver, 1975
England
Exposure
TCP
2,4,5-T
PCP, HCBC
Pure 2,3,7,8-TCDD
Population
55 workers in TCP and
2,4,5-T production
Followed: 1969-1973
3 scientists synthesis of
2,3,7,8-TCDD
Evaluation
Clinical examination
history and
examination
Not described
Findings
35 patients with
"neurasthenic or depressive
syndrome"
Subject A:
Fatigue and headache
Subject B:
Headaches
Loss of vigor and fatigue
Irritability and anger
Poor concentration
Subject C:
Loss of concentration
(apathy and fatigue)
"Loss of energy and drive"
Some difficulty sleeping
~J
H-*
ON
6
o
3
§
O
n
"HCB = hexachlorobenzene.
O
o\
OJ
o
-------
Table 7-39. (continued)
1
o\
Author
Pazderova-
Vejlupkova et al.,
1981
Spolano,
Czechoslovakia
Exposure
TCP
2,4,5-T
PCP, HCB
Population
4-year follow-up of 44
production workers
Evaluation
Clinical psychiatric
examination
Findings
Psychiatric changes at start
of "intoxication"
83% with neurotic
symptoms; neurasthenia
syndromes with depressive
component; depressive
syndrome with endogenous
component (N=55)
16% with
pseudoneurasthenia
syndrome and CNS
arteriosclerosis
3 % with no psychiatric
signs or symptoms
Psychiatric changes in
exposed individuals at
follow-up (1973-1979)
58% with neurotic
symptoms and no
depressive or anxiety
symptoms
18% with severe
pseudoneurasthenia and
signs of dementia (usually
in patients over 50 years
old but dementia occurred
in 30-year-old patient)
!>
H
1
6
o
*Z
9
o
o
t— 1
a
-------
Table 7-39. (continued)
Author
Pazderova-
Vejlupkova et al.,
1981 (cont.)
Kimmig and
Schultz, 1957b
Hamburg, Germany
Poland et al., 1971
New Jersey
Exposure
TCP
2,4,5-T
TCP
2,4,5-T
2,4-Df
HCB
TCP reactor explosion in
1960
Daily exposure 1951-1969
Population
3 1 production workers with
chloracne (1953-1954)
73 male workers in
production, maintenance,
office areas
Evaluation
Clinical history and
examination
Noted histories of
smoking, alcohol,
medications
MMPI8 given to 52
production and 17
unexposed
administrative staff
Findings
24% with no psychiatric
signs or symptoms.
Tiredness (N=3)
Headaches (N=5)
Headaches (N= 8)
Severity of acne
significantly correlated
with high score on the
mania scale of MMPI.
N^
5
d
o
z
s
o
c;
w
o
»
n
H
W
f2,4-D = 2,4-dichlorophenoxyacetic acid.
8Minnesota Multiphasic Personality Inventory.
-------
Table 7-40. Cross-Sectional Studies of Psychological and Neurologic Effects Among Residents of Missouri and Seveso Exposed to
2,3,7,8-TCDD-Contaminated Materials
Author
Webbetal., 1989
Missouri
Pocchiariet al., 1979
Seveso, Italy
Exposure
2,3,7,8-TCDD-
contaminated
waste oil 1971
TCP
TCP reactor
accident
July 1976
Study
population
68 volunteers in
State Dioxin
Registry;
estimated
exposure to 20-
100 ppb TCDD
for 2 yrs or 100
ppb for 6 months
446 residents of
Seveso and
Meda, Italy
200 workers
from Icmesa
plant
Comparison
population
36 volunteers in State
Dioxin Registry with
no history of exposure
to 2,3,7,8-TCDD
255 residents of
nearby towns without
contamination
None
Evaluation
Missouri Dioxin
Health Studies.
Progress Report
(1983)
Not detailed but
included clinical
exam and NCVb
Exam, EMGC,
NCV
Findings
No differences in
neurologic exam or in
feeling of pins-needles;
loss of sensation in
extremities, tingling in
fingers and toes,
diminished VIB 256,
diminished VIB 64,
diminished pin sensation,
diminished thermal
sensation (PRR=2%)a
1977: neurologic
damage in 6.7% of
Seveso residents and in
1.2% unexposed
residents
1978: neurologic
damage in 11.7% of
Seveso residents
workers: 4%
neuropathic clinical signs
exam, EMG, NCV
"Prevalence risk ratio.
bNerve conduction velocity.
°Electromyography.
-------
Table 7-40. (continued)
Author
Filippinietal., 1981
Seveso, Italy
Exposure
See description
above
Study
population
308 residents of
Seveso, Italy
Comparison
population
305 residents of
nearby towns without
contamination
Evaluation
Symptoms:
pain, tingling,
numbness
Sensory review
Muscle tone and
strength
NCV of ulnar and
peroneal nerves
Findings
Elevated PRR for
peripheral neuropathy in
Seveso residents with
indicators of exposure
(high GOT, ALT, AST,
or chloracne) (PRR=2.8,
95% CI= 1.2-6.5) Seveso
residents with
predisposing factors
(alcohol or inflammatory
disease) (PRR=2.6, 95%
CI= 1.2-5.6)
O
50
r"
TJ
i
O
o
H
c!
O
H
W
o\
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n
^H
3
vo
-------
DRAFT-DO NOT QUOTE OR CITE
animal studies have specifically evaluated peripheral or autonomic functions related to
2,3,7,8-TCDD exposure.
7.13.7.1. Neurobehavioral Assessments
Numerous case reports cite symptoms referable to the nervous system occurring after
acute exposure among occupationally exposed individuals (Creso et al., 1978; Ashe and
Suskind, 1950; Suskind et al., 1953; Goldman, 1972), as well as chronic exposure to
2,3,7,8-TCDD-contaminated materials (Oliver, 1975; Baader and Bauer, 1951; Kimmig and
Schulz, 1957a, b; Bauer et al., 1961; Poland et al., 1971). Symptoms include headache
(Ashe and Suskind, 1950; Bauer et al., 1961; Jirasek et al., 1974; Oliver, 1975; Kimmig and
Schulz, 1957a, b; Poland et al., 1971), insomnia (Ashe and Suskind, 1950; Suskind et al.,
1953; Oliver, 1975; Kimmig and Schulz, 1957a, b), nervousness or irritability (Ashe and
Suskind, 1950; Suskind et al., 1953; Bauer et al., 1961; Oliver, 1975), depression and
anxiety (Bauer et al., 1961; Jirasek et al., 1974), loss of libido, and encephalopathy (Jirasek
et al., 1974; Kimmig and Schulz, 1957a) (Table 7-39).
Some reports indicate that symptoms referable to the CNS and PNS may persist in
some exposed individuals for as long as 25 years (Suskind et al., 1953; Poland et al., 1971;
Jirasek et al., 1973; Jirasek et al., 1974; Creso et al., 1978; Ashe and Suskind, 1950;
Suskind et al., 1953). In 1953, Suskind et al. reported a variety of CNS-related symptoms in
36 workers from a plant in Nitro, West Virginia, who had developed chloracne and other
symptoms after exposure to contaminants subsequent to a TCP reactor explosion in March
1949 (N=ll) or during normal production process of TCP and 2,4,5-T (N=25) between
1948 and 1953 (Suskind et al., 1953). Such symptoms reported among this relatively young
group (average age = 36 years, range = 22-63 years) included fatigue (N=21), nervousness
and irritability (N=17), and decreased libido (N=13). No attempt was made to determine
whether these symptoms also occurred among exposed individuals without chloracne or
among the nonexposed plant population.
Between 1968 and 1969, Jirasek et al. (1974) "observed very closely" (by clinical
evaluation) a group of 55 workers exposed to 2,3,7,8-TCDD in the production of 2,4,5-T,
hexachlorobenzene, and pentachlorophenol. As described in the report, intoxication with
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2,3,7,8-TCDD "occurred gradually from 1965 to 1968" although further quantification of
exposure is not described. Psychiatric examination revealed the following: severe neurotic
symptoms (64%), neurasthenia syndrome with depressive component (11%), depressive
syndromes (8%), pseudoneurasthenia syndromes in patients with arteriosclerosis of the CNS
(14%), and normal psychiatric examination (3%). One patient died at age 57 years with
rapidly progressive dementia secondary to an atypical arteriosclerosis involving brain and
other organs.
Ten years after the initial examination, Pazderova-Vejlupkova et al. (1981) evaluated
the health status of 44 of the original 55 workers. They found the following on psychiatric
examination: 58% continued to have neurotic symptoms without depressive or anxiety
components, 18% developed severe neurasthenia syndromes with signs of dementia, and 24%
were normal.
Poland et al. (1971) administered the Minnesota Multiphasic Personality Inventory
(MMPI) to 52 male production workers who were exposed at that time to TCP, 2,4,5-T,
2,4-D, and other chemicals. Severity of acne correlated significantly with a high score on
the hypomania scale of the MMPI. When production workers were compared with 17
presumably unexposed administrative workers, the two groups differed on only one MMPI
symptom scale; exposed production workers scored higher on the hypochondriasis scale.
Table 7-41 describes the results of neurologic and neurobehavioral assessments of
TCP production workers, Ranch Hands, and U.S. Army Vietnam veterans.
Moses et al. (1984) found a significant excess among workers with chloracne
compared to those without lesions for the following symptoms: insomnia, decreased libido,
and difficulties with ejaculation or erection. There was no difference between the two groups
for the symptoms of fatigue, irritability, nervousness, depression, or personality changes.
In the study by Suskind and Hertzberg (1984), psychological symptoms were
evaluated by interview and peripheral nerve function by nerve conduction velocity of the
peroneal motor and sural sensory nerve fibers. The complaint of loss of libido was more
frequent among the exposed than unexposed even after stratification by age. The complaint
of nervousness, depression, or anxiety was not significantly related to exposure (16.3% vs.
11.7%) in the crude analysis even though the sample size was sufficient to detect a twofold
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Table 7-41. Cross-Sectional Studies of Psychological and Neurologic Effects Among TCP Production Workers and Vietnam
Veterans Exposed to 2,3,7,8-TCDD-Contaminated Materials
Author
Moses etal., 1984
West Virginia
Suskind and
Hertzberg, 1984
West Virginia
Exposure
TCP
2,4,5-Tb
1949 TCP
reactor
explosion
TCP
2,4,5-T
Study
population
226 workers invited
based on union
record of
employment in
production of TCP
or 2,4,5-T
Definition of
exposure = "current
or history of
chloracne"
204 workers or
maintenance
workers in TCP or
2,4,5-T dept.
Comparison
population
Workers without
chloracne (N= 109)
Conduction
velocities
163 workers never
in TCP or 2,4,5-T
Mean age of exposed
population was
younger (56.7 vs.
46.2) (p<0.0001).
Evaluation
Review of
symptoms (ROS)
obtained by
examining
physician
Interviews and
clinical
examination
Findings
Significant diff. with and
without chloracne for:
insomnia; decreased
libido, erection, or
ejaculation; no change for
fatigue, irritability,
nervousness, depression,
or personality changes.
Decreased pin prick
sensation in 18.3% with
chloracne; no decreased
pin prick sensation
without chloracne
Decreased libido was
more frequent among
exposed by Mantel-
Haenszel Chi by age.
Nervousness/depression/
anxiety not increased in
exposed (sample size
sufficient to detect a 2-
fold increase)
O
1
6
0
5*
o
H
O
§
d
cn
?0
Q
M
•^J
h->*
ON
"TCP = 2,4,5-trichlorophenol.
k2,4,5-T = 2,4,5-trichlorophenoxyacetic acid.
s
w
o
-------
Table 7-41. (continued)
Author
Sweeney et al.,
1993
Alderfer et al. ,
1992
Missouri and New
Jersey
Lathropet al., 1984
Exposure
TCP 2,4,5-T
Vietnam service
1962-1971
Study
population
280 production
workers with daily
exposure
New Jersey
1951-1969
Missouri 1968-1972
1,208 Ranch Hands
assigned to aerial
spraying herbicides
and insecticides
Republic of Vietnam
Comparison
population
261 unexposed
community residents
matched for age,
race, gender
1,238 men who flew
cargo missions in
Southeast Asia
Matched by month
of birth, race, and
occupational code
Evaluation
Symptom and
medical history;
neurologic
examination;
neurophysiologic
tests of vibration
and thermal
sensitivity;
Beck Depression
Scale; SCL-90-R
Self-report of
psychological or
emotional illness
Diagnostic
Interview
Schedule
(modified)
Self-reported
depression
Findings
No significant difference
for symptoms or any
examination variable
No significant difference
in mood disorders
No difference
Significantly more
fatigue, anger, erosion,
anxiety, closest for high
school-educated Ranch
Hands; no difference for
college-educated
Greater for Ranch Hands
O
o
o
I
w
o
o
HH
a
-------
Table 7-41. (continued)
Author
Lathropetal., 1984
(cont.)
Exposure
Study
population
Comparison
population
Evaluation
Cornell Index (self-
administered
inventory of
neuropsychiatric
symptoms)
(psychophysiologic)
MMPI
Halstead-Reitan
Battery
Wechsler Adult
Intelligence Scale
(WAIS)
Findings
After adjustment for
education 4/10 parameters
abnormal (nervousness,
anxiety, startle,
psychosomatic,
gastrointestinal system);
abnormal parameters
inversely related to
education level.
High school-educated
Ranch Hands showed
significant deficits on
scales for hypochondria,
masculinity/femininity;
mania/hypomania but
comparisons show more
denial; MMPI scores
influenced by education
(p<0.01)
No impairment in Ranch
Hands
Scores related to
educational level
d
>
*fl
i
i
d
O
g
3
n
^•^
o
•jj
o
w
Q
ffl
I
t—k
-J
o
ON
-------
Table 7-41. (continued)
Author
Roegner et al.,
1991
Exposure
Vietnam service
Study
population
720 USAF Ranch
Hands; serum
2,3,7,8-TCDD
levels were
measured (lipid
adjusted)
Comparison
population
779 who flew cargo
missions in
Southeast Asia;
serum 2,3,7,8-
TCDD levels were
measured
Evaluation
Self-report
SCL-90-Ra
Cornell Medical
Index (CMI)
Neurologic exam
Findings
No significant mental,
emotional, or sleep
disorders
No significant differences
with increasing serum
levels
Significantly higher mean
schizoid and schizotypal
scores in Ranch Hands
over 33.3 pg/g 2,3,7,8-
TCDD, but they did not
relate to similar scales in
the SCL-90-R
Overall, no consistent
relationship between
neurologic abnormalities
and TCDD level.
O
>
H
6
o
z;
3
n
vX
q
w
o
/o
Q
H
W
-J
to
aSCL-90-R: Symptom Checklist Revised.
O
o\
w
o
-------
DRAFT-DO NOT QUOTE OR CITE
increase. However, in the group over 50 years old there was a significant exposure effect
(19% vs. 6.4%) for the complaint of nervousness, depression, or anxiety. The complaint of
impotence was significantly related to exposure in the crude analysis, but after stratification
by age this effect disappeared.
The effects of exposure to 2,3,7,8-TCDD on measures of current symptoms of
depression were evaluated (Alderfer et al., 1992) as part of the NIOSH cross-sectional
medical study (Sweeney et al., 1989). Symptoms of depressed mood were measured by the
Beck Depression Inventory and the depression subscale of the Self-Report Symptom
Checklist-90-Revised (SCL-90-R). Neither serum 2,3,7,8-TCDD levels nor status as a
worker was associated with depressed mood as assessed by either the Beck Depression
Inventory or the SCL-90-R depression subscale (Alderfer et al., 1992). This finding supports
the conclusion that current serum levels of 2,3,7,8-TCDD are not associated with current
depression among a population of workers that was highly exposed to TCDD. However,
because this cross-sectional study was conducted many years after 2,3,7,8-TCDD exposure,
this analysis could not address the question of whether 2,3,7,8-TCDD is associated with past
depression that resolved before the study was performed. These findings are consistent with
those of the U.S. Air Force study (Roegner et al., 1991) of personnel who applied Agent
Orange during the Vietnam war: serum 2,3,7,8-TCDD was not associated with the
depression subscale score of the SCL-90-R after controlling for covariates.
The Air Force study conducted neurologic and psychological assessments of the
participating Ranch Hands and comparisons (Lathrop et al., 1984). In the first examination
series published in 1984 (baseline), the psychological assessment included a self-report of
psychological or emotional illness, the Diagnostic Interview Schedule (DIS), Cornell Medical
Index (inventory of psychophysiologic symptoms), MMPI, Halstead-Reitan Battery (HRB),
and Wechsler Adult Intelligence Scale (WAIS). In the 1987 follow-up examination, the
psychological assessment included an interviewer-administered questionnaire in which each
participant was asked about the occurrence of mental or emotional disorders and sleep
disorders. The presence of posttraumatic stress disorder was based on a subset of questions
from the MMPI; the WAIS IQ assessment was deleted and the Millon Clinical Multiaxial
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Inventory (MCMI) and the Symptom Checklist-90-Revised (SCL-90-R) were added (Lathrop
etal., 1987).
Results of the 1984 baseline study revealed no difference between Ranch Hands and
the comparison group for self-reported psychological or emotional illness. The DIS revealed
significantly more fatigue, anger, erosion, and anxiety for high school-educated but not
college-educated Ranch Hands. These outcomes were highly related to education. The
Cornell Medical Index found 4 of 10 parameters abnormal for Ranch Hands—startle,
psychosomatic, gastrointestinal nervousness, and anxiety. These parameters were inversely
related to education level. On the MMPI, high school-educated Ranch Hands showed
significant differences (more deficits) on subscales for hypochondria, masculinity/femininity,
and mania/hypomania, but comparisons scored higher on the subscale for denial, a finding
that might undermine the deficits noted for Ranch Hands. Again, MMPI scores were
influenced by education level (p<0.01) but not exposure level. For both the HRB and
WAIS, there was no difference between Ranch Hands and comparisons, and the scores were
related to education level.
In the 1987 reanalysis with serum 2,3,7,8-TCDD (Roegner et al., 1991), there was
no significant difference between groups or relationship with serum 2,3,7,8-TCDD levels on
reported (and verified) data on lifetime psychological illness or sleep disorders or any SCL-
90-R. Some of the MCMI parameters appeared to be related to serum 2,3,7,8-TCDD levels
(significantly higher mean schizoid and schizotypal scores and significantly lower mean
histrionic score in the group above 33.3 pg/g than in comparisons). However, these findings
were inconsistent with similar variables in the SCL-90-R and the self-reported histories.
Comprehensive neurologic and psychological assessments were conducted on
participants of the Vietnam Experience Study (Centers for Disease Control Vietnam
Experience Study, 1988a, b). The neurobehavioral tests evaluated aptitude, concept
formation and problem-solving, memory, manual dexterity, verbal skills, visuo-motor skills,
attention, and mental control. Among Vietnam veterans there was a significantly greater
prevalence of alcohol abuse or dependence (Vietnam veterans, 13.7%; non-Vietnam veterans,
9.2%; OR=1.5, 95% CI = 1.2-1.8), depression (Vietnam veterans, 4.5%; non-Vietnam
veterans, 3.2%; OR=2.0, 95% CI=1.4-2.9), and a higher prevalence of poor psychological
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status (Centers for Disease Control Vietnam Experience Study, 1988b). The poor
psychological status tended to be most prevalent in Vietnam veterans who were not white,
who enlisted before their 19th birthday, and whose enlistment test scores fell below the group
median.
In the study of residents of the Quail Run Mobile Home Park, neurobehavioral tests
evaluated reaction time, mood, memory, visuo-motor coordination, intelligence, and
indicators of psychological stress (Hoffman et al., 1986). Differences between Quail Run
residents and controls were observed in the vocabulary subtest of the WAIS (Quail Run =
34.7, controls = 41.1, raw scores; p<0.01), tension/anxiety raw score (Quail Run = 13.7,
controls = 11.1; p<0.01), and anger/hostility scale (Quail Run = 11.6, controls = 8.9;
p<0.05) of the POMS inventory, and for the depression/dejection and fatigue/inertia scales
(no data provided).
7.13.7.2. Neurologic Status
On neurologic examination, 11 of 60 West Virginia workers with chloracne exhibited
decreased sensitivity to pin prick, whereas none of the 34 subjects without chloracne had
decreased pin prick sensation (p<0.01) (Moses et al., 1984). There were no other
differences in performance of the neurologic examination noted in the text. When examined
by Suskind and Hertzberg (1984), no significant differences were noted in the conduction
velocities of either nerve fiber (sural sensory: exposed workers, mean = 42.06+0.49;
unexposed workers, mean 41.49+0.54; peroneal motor: exposed workers, mean =
41.77+0.47; unexposed workers, mean = 42.62±0.52).
Among New Jersey and Missouri TCP workers, the overall neurologic status and
peripheral nerve function were assessed for all 281 workers and 260 referents by self-
reported medical history, neurologic examination, electrophysiologic tests of nerve
conduction velocity, amplitude and latency, and vibratory and thermal threshold (Sweeney et
al., 1993). No differences in neurologic status or nerve function between workers or
referents were detected. Additionally, although the mean serum 2,3,7,8-TCDD level in the
workers was 220 pg/g, there was no relationship between neurologic function and levels of
serum 2,3,7,8-TCDD.
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The neurologic examination in the 1987 follow-up study evaluated cranial, CNS, and
PNS function in participating Ranch Hands (Lathrop et al., 1987). In general, there was no
difference in the prevalence of neurologic abnormalities in Ranch Hands and comparisons.
However, Ranch Hands with serum 2,3,7,8-TCDD levels above 33.3 pg/g tended to have a
higher proportion of individuals with abnormal coordination than the comparisons (Ranch
Hands, 2.7%; comparisons, 0.4%; adjusted RR=18.3, p<0.001) (Roegner et al., 1991).
Overall neurologic status of Army Vietnam veterans did not differ from that of non-
Vietnam veterans (Vietnam veterans, 1.0%; non-Vietnam veterans, 0.8%; OR=1.2, 95%
CI=0.6-2.3) (Centers for Disease Control Vietnam Experience Study, 1988b). Only self-
reported symptoms related to nerve disorders were significantly more prevalent among
Vietnam veterans than non-Vietnam veterans (Vietnam veterans, 8.2%; non-Vietnam
veterans, 6.5%; OR=1.2, 95% CI=1.0-1.6).
Table 7-40 describes the results of neurologic and neurobehavioral studies of Seveso
and Missouri residents. While three studies evaluated the neurologic status of residents of
Seveso (Pocchiari et al., 1979; Filippini et al., 1981; Assennato et al., 1989), no studies
evaluated neurobehavioral changes. In an effort to quantify exposure-related neurologic
disorders among Seveso residents and among workers from the Icmesa plant, two
government-sponsored screenings were conducted in 1977 and 1978 on 308 residents of
Seveso and 200 workers. Among these workers, approximately 4% (N=8) were found to
have damage to nerve fibers of multiple (unspecified) nerves, controlling for confounding
factors such as alcohol abuse, diabetes, kidney disease, and neurotoxic medication use
(Pocchiari et al., 1979). The report did not describe the extent of worker exposure to
2,3,7,8-TCDD. Other potential neurotoxic occupational exposures do not appear to have
been considered. Three workers were described as having polyneuropathies of the lower
limbs.
In 1981, prevalence risk ratios (PRR) for neuropathy were calculated separately for
the 308 Seveso residents (Filippini et al., 1981). PRRs for neuropathy were determined for
residents who exhibited clinical indication of 2,3,7,8-TCDD exposure, defined as the
presence of elevated liver enzyme levels (GGT, ALT, AST) (which are also indicative of
nonspecific insults to the liver) or chloracne, and for those who exhibited conditions that are
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risk factors for neuropathy, e.g., alcoholism, inflammatory disease, diabetes, or potential
occupational exposure to neurotoxins. Seveso residents who had clinical indication of
2,3,7,8-TCDD exposure (chloracne, or elevated liver enzymes GGT, AST, ALT) or who had
risk factors for neuropathy were found to have significantly greater prevalence of neuropathy
than residents without either manifestation (PRR exposure = 2.8, 95% CI=1.2-6.5; PRR for
possible 2,3,7,8-TCDD-predisposing factors = 2.6, 95% CI=1.2-5.6) (Filippini et al.,
1981). Additional analysis identified that individuals who met the definition for chloracne or
abnormal levels of hepatic enzymes were significantly more at risk than residents without
either condition.
Residents of Seveso who developed chloracne after the reactor release (N=193) were
invited to a series of three follow-up screenings in 1982-1983, 1983-1984, and 1985
(Assennato et al., 1989). A control group from a nearby but uncontaminated area was also
examined. Conduction velocities of the median motor, peroneal motor, and sural sensory
fibers were conducted in 1982-1983 and 1985 (Assennato et al., 1989). No increases in the
prevalence of abnormal electrophysiologic measures were observed in the chloracne group
when compared with controls without chloracne. In addition, there was no change in the
conduction velocities for each fiber from the 1982-1983 to 1985 studies.
Quail Run residents reported significantly more "numbness" or "pins and needles" in
the hands or feet (28.6%) than the controls (18.1%) (p<0.05), but there were no differences
in mean threshold scores for the more objective neurosensory tests. Residents also reported
more persistent severe headaches (Quail Run, 26.0%; control, 14.2%; p<0.05).
Participants in the study by Webb et al. (1989) did not complete neurobehavioral
tests, but they were examined by a neurologist. The results were unremarkable. Of the 38
participants, two with levels above background had abnormal pin prick sensitivity (<20
pg/g, N=2; 20-60 pg/g, N = l; >60 pg/g, N=l), three had abnormal vibration thresholds
(<20 pg/g, N=2; 20-60 pg/g, N=l; >60 pg/g, N=2), and four had abnormal reflexes
(<20 pg/g, N=2; 20-60 pg/g, N=2; >60 pg/g, N=2). The results of other components of
the neurologic examination were not reported and are assumed to be normal.
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7.13.7.3. Comment
The overall results of these case reports and epidemiologic studies demonstrate that
exposure to 2,3,7,8-TCDD-contaminated materials is associated with symptoms referable to
the central and peripheral nervous systems shortly following exposure and, in some cases,
lasting many years (Tables 7-40 and 7-41). Symptoms include fatigue, nervousness, anxiety,
and decreased libido (Ashe and Suskind, 1950; Suskind et al., 1953; Oliver, 1975; Kimmig
and Schulz, 1957a, b; Bauer et al., 1961; Jirasek et al., 1974). One case report found
mania/hypomania on the MMPI (Poland et al., 1971). These symptoms are consistent with
mood disorder. While one study reported neurasthenia with signs of dementia lasting 10 or
more years following exposure (Jirasek et al., 1974), the U.S. Air Force study of Vietnam
veterans used neurobehavioral testing but was unable to demonstrate cognitive or other
functional CNS deficits. However, this negative study did not investigate the relationship
between serum 2,3,7,8-TCDD levels and neurobehavioral deficits. NIOSH investigated
measures of depressed mood many years after exposure to 2,3,7,8-TCDD among production
workers and found no relationship between depressive symptoms and serum 2,3,7,8-TCDD
levels (Alderfer et al., 1992).
Overall neurologic status of workers, community residents, and Vietnam veterans
exposed to 2,3,7,8-TCDD and evaluated from 5 to 37 years after last exposure appears to be
normal. These data suggest that, although exposure to 2,3,7,8-TCDD may have been
extensive as in the case of the exposed workers, Ranch Hands, and Seveso residents and case
reports describe many related symptoms, the effects may have been transient. If so, studies
conducted years after the last exposure would not detect such changes. These results suggest
that, in adults, no long-term neurologic effects were caused by even high exposure to
2,3,7,8-TCDD-contaminated materials. However, there is very little information with which
to examine the effects of exposure on the developing human neurologic system.
7.13.8. Circulatory System
The relationship between human exposure to 2,3,7,8-TCDD-contaminated chemicals
and disorders of the circulatory system has been explored in a variety of reports. A number
of early case reports have described effects on the cardiovascular system among individuals
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reportedly exposed to chemicals contaminated with 2,3,7,8-TCDD. Myocarditis (Goldman,
1972), myocardial infarctions (Walker and Martin, 1979; Bauer et al., 1961), ectasia of the
coronary arteries (England, 1981), and rapidly progressive atherosclerosis (Jirasek et al.,
1974; Pazderova-Vejlupkova et al., 1981) have been reported.
Studies have described mortality from diseases affecting the circulatory system among
populations exposed to 2,3,7,8-TCDD (Bond et al., 1987; Coggon et al., 1991; Fingerhut et
al., 1991b; Zober et al., 1990; Bueno de Mesquita et al., 1993; Bertazzi et al., 1989, 1992;
Collins et al., 1993) (Table 7-42). The circulatory system includes ICD-9 codes 390-459
(International Classification of Diseases 9). The results are limited because the studies were
primarily designed to test hypotheses relating to cancer and, secondarily, to characterize
mortality from causes other than cancer, without detailed consideration of confounders or
extent of exposure to 2,3,7,8-TCDD. In addition, because the study populations have
differing age distributions, direct comparison of the SMRs is not valid. However, any trends
suggested by the direction and magnitude of the SMRs are important.
Among TCP production workers, mortality from all diseases of the circulatory system
was similar to mortality in the general population, as described by an SMR of 100, and in
workers from The Netherlands (Plant A) (SMR=98, 95% CI=65-142) (Bueno de Mesquita
et al., 1993), the United States (Nitro, West Virginia) (SMR=90, 95% CI=80-100) (Collins
et al., 1993), and Great Britain (SMR=116, 95% CI=91-146) (Coggon et al., 1991) (Table
7-42). In two studies, mortality in workers with chloracne was not significantly different
from that of the national comparison group (U.S. workers, SMR=95, 95% CI=79-113;
German workers, SMR=121, 90% CI=83-170) (Bond et al., 1987; Zober et al., 1990).
Mortality from more specific end points, such as ischemic heart disease, all heart
disease, and cerebrovascular disease, was noted in a few studies. The SMR from ischemic
heart disease was 102 (95% CI=63-158) among Dutch workers (Bueno de Mesquita et al.,
1993) and 96 (95% CI=51-164) in U.S. workers (Midland, MI) (Bond et al., 1987). And,
in an analysis that included all diseases of the heart combined among 5,200 U.S. production
workers, the SMR was 96 (95% CI=87-106) (Fingerhut et al., 1991b). Cerebrovascular
disease mortality was slightly elevated among Dutch TCP production workers (SMR=117,
95% CI=38-274) (Bueno de Mesquita et al., 1993) and increased by more than twofold in
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Table 7-42. Mortality from Diseases of the Circulatory System in Populations Exposed to 2,3,7,8-TCDD
Author
Fingerhut et al.,
1991b
Fingerhut et al.,
1991b
Zober et al., 1990
Coggonetal., 1991
Bond et al., 1987
Bondetal., 1987
Population
2,4,5-TCP and
2,4,5-T production
workers, USA
2,4,5-TCP and
2,4,5-T production
workers, USA
2,4,5-T production
workers with
chloracne, Germany
2,4,5-T synthesis or
formulation, Great
Britain
U.S. chemical
production workers:
TCP + 2,4,5-T
with chloracne
(Michigan)
U.S. chemical
production: workers
without chloracne
(TCP + 2,4,5-T)
(Michigan)
Outcome
Diseases of the
heart (ICD 390-
398, 402-404, 410-
414, 420-429)
Diseases of the
circulatory system
(ICD 401, 403,
405, 415-417, 430-
438, 440-459)
Diseases of the
circulatory system
(ICD 390-458)
Diseases of the
circulatory system
Diseases of the
circulatory system
ASHDC
Vascular lesions of
CNS
Diseases of the
circulatory system
ASHDC
Vascular lesions of
CNS
No. of
deaths
393
67
74
19
13
4
130
106
10
SMR"
96
77
121
116
102
96
208
95
110
64
95% CI
87-106
60-98
83-170b
91-146
61-159
51-146
57-539
79-113
90-133
31-117
Cohort size
5,172
5,172
127
2,239
322
2,026
Years of
follow-up
1942-1987
1942-1987
1953-1987
1975-1987
1940-1982
1940-1982
o
o
oo
o
o
o\
o3
o
§
o
a
-------
Table 7-42. (continued)
Author
Bueno de Mesquita
etah, 1993
Bueno de Mesquita
et ah, 1993
Bueno de Mesquita
et ah, 1993
Michalek et ah,
1990
Centers for Disease
Control Vietnam
Experience Study,
1988c
Fett et ah, 1987b
Population
2,4,5-TCP and
2,4,5-T production
workers, The
Netherlands
2,4,5-TCP and
2,4,5-T production
workers, The
Netherlands
2,4,5-TCP and
2,4,5-T production
workers, The
Netherlands
U.S. Air Force
Ranch Hand
personnel
U.S. Army
Vietnam veterans
Australian Vietnam
veterans; served
> 12 months
Outcome
Diseases of the
circulatory system
(ICD 390-458)
Ischemic heart
disease
(ICD 410-414)
Cerebrovascular
disease
(ICD 430-438)
Diseases of the
circulatory system
Diseases of the
circulatory system
(ICD 390-459)
Diseases of the
circulatory system
No. of
deaths
28
20
5
25
12
2011
SMR"
98
102
117
110
0.49d
1.7
95% CI
65-142
63-158
38-274
60-150
0.25-0.99
0.9-3.0
Cohort size
549
549
549
1,261
9,324
Vietnam
veterans =
19,205
Non-Vietnam
veterans =
25,677
Years of
follow-up
1955-1985
1955-1985
1955-1985
1961-1987
1965-1983
1966-1985
O
6
o
1
o
c
9
w
o
H
w
I
oo
o
o\
u3
O
-------
Table 7-42. (continued)
Author
Bertazzi et al., 1989
Bertazzi et al., 1992
Population
Residents of
Seveso, Italy,
ages 20-74,
Zone A (high
TCDD region)
Residents of
Seveso, Italy, ages
1-19 years
Outcome
Diseases of the
circulatory system
(ICD 390-459)
ISHDf
Cerebrovascular
disease
Diseases of the
circulatory system
(ICD 390-459)
No. of
deaths
11
6
2
5
0
2
SMR"
Males: 1.75d
Females: 1.89d
Males: 1.25d
Males: 3.3d
Males: not
reported
Females: 1.63d
95% CI
1.0-3.2
0.8-4.2
0.5-3.3
1.4-8.0
0.3-8.1
Cohort size
556e
306e
Years of
follow-up
1976-1986
1976-1986
aSMR = Standardized mortality ratio.
b90% confidence interval.
cAth*»rr»«nl*»mtir Vif»art Hicp-ncp.
O
>
3
1
O
0
2
3
O
oo
to
dRelative risk.
"Zone A males and females combined.
flschemic heart disease.
3
W
g
O
H
W
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DRAFT-DO NOT QUOTE OR CITE
Michigan TCP production workers with chloracne (SMR=208, 95% CI=57-539) (Bond et
al., 1987).
The SMRs for circulatory system diseases reported in the various mortality studies are
close to 100, suggesting that the "healthy worker effect" is not seen in these studies.
Generally, because employed workers are healthier than the general population, the SMR for
cardiovascular disease in employed populations tends to be lower than 100 (McMichael,
1976; Fox and Collier, 1976). The absence of a healthy worker effect, in light of the
positive animal data, suggests that more detailed analyses should be conducted for
cardiovascular outcomes in these populations.
Among 1,261 Ranch Hand personnel, mortality from circulatory disease was
nonsignificantly elevated (SMR = 110, 95% CI=60-150) compared with that of a comparison
population of 19,101 other Air Force veterans who were not exposed to herbicides (Michalek
et al., 1990). Similar nonsignificant increases in the relative mortality ratio (RMR) were
observed for circulatory diseases (RMR=1.6, 95% CI=0.8-3.2) in Australian Vietnam
veterans (N=19,205; 260 deaths) compared to 25,677 (263 deaths) non-Vietnam veterans
who only served in Australia (Fett et al., 1987b). In contrast, the unadjusted relative risk of
0.49 (95% CI=0.25-0.99) for all circulatory diseases suggested a deficit of deaths from this
cause among 9,325 Vietnam Army veterans (246 deaths) compared to 8,989 non-Vietnam
veterans (200 deaths) (Centers for Disease Control Vietnam Experience Study, 1988c).
Bertazzi and colleagues examined the mortality experience of Seveso residents ages
1-19 years (Bertazzi et al., 1992) and ages 20-74 years (Bertazzi et al., 1989) 10 years after
the contamination of the town by 2,3,7,8-TCDD-contaminated effluent. In the younger
population, two deaths from circulatory diseases occurred only in female residents
(RR=1.63, 95% CI=0.3-8.1). In the older population, circulatory disease mortality of
residents from Zone A (the most highly contaminated region) was elevated in both males
(RR=1.75, 95% CI = 1.0-3.2) and females (RR=1.89, 95% CI=0.8-4.2). In males, the
highest death rate occurred during the first quinquennium, 1976-1981 (RR=2.04, 95%
CI=1.0-4.2), and, in females, the highest death rate occurred during the second
quinquennium, 1982-1986. The authors suggest that the study was limited by the small
number of subjects and the crude measure of 2,3,7,8-TCDD exposure. The authors could
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not attribute the increased mortality from circulatory disease to 2,3,7,8-TCDD exposure but
suggested that the "high stress and pollution" imposed on the residents of Zone A may have
been a contributing factor.
Several cross-sectional medical studies have also examined the association between
2,3,7,8-TCDD exposure and effects on the cardiovascular system (Suskind and Hertzberg,
1984; Moses et al., 1984; Bond et al., 1983; Centers for Disease Control Vietnam
Experience Study, 1988a; Roegner et al., 1991). Statistically significant associations relative
to 2,3,7,8-TCDD exposure were found only in the Ranch Hand study for diastolic blood
pressure, arrhythmias detected on the electrocardiogram (ECG), and peripheral pulse
abnormalities (Roegner et al., 1991). However, there is some doubt that the significant
findings are dose related because significant increases in the mean diastolic blood pressure
were found in Ranch Hands with serum 2,3,7,8-TCDD levels from 15 to 33.3 pg/g but not
in the Ranch Hands with higher serum 2,3,7,8-TCDD levels. The adjusted odds ratios for
ECG-diagnosed arrhythmias among Ranch Hands were not entirely consistent with a dose-
response relationship. For each serum 2,3,7,8-TCDD category the odds ratios were: serum
2,3,7,8-TCDD levels <10 pg/g, OR=1.33 (95% CI=0.68-2.58, p=0.40); 15-33.3 pg/g,
OR=0.83 (95% CI=0.31, 2.23, p=0.71); and for serum 2,3,7,8-TCDD levels above 33.3
pg/g the OR was 2.34 (95% CI = 1.00-5.51, p=0.051). The proportion of individuals in this
group with arrhythmias (5.2%) was not much higher than in Ranch Hands whose serum
levels were < 10 pg/g (4.7%). Finally, relative to the comparison group, the proportion of
abnormal peripheral pulses in all Ranch Hands, regardless of serum level, was elevated.
No excess abnormalities or disorders of the circulatory system or heart were found in
several groups of TCP production workers, although their potential for exposure to 2,3,7,8-
TCDD-contaminated chemicals was high (Suskind and Hertzberg, 1984; Moses et al., 1984;
Bond et al., 1983). The prevalence of hypertension or coronary artery disease (both self-
reported), abnormal ECG findings, atherosclerotic changes (not specified) on chest X-ray, or
blood pressure elevation was not elevated in West Virginia TCP production workers (Suskind
and Hertzberg, 1984). Similarly, when Moses et al. (1984) examined TCP production
workers with chloracne and compared them with workers not affected by chloracne, they
found no increased risk for self-reported abnormal ECG, self-reported angina, or self-
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reported myocardial infarction and no difference in the physical examination of the
cardiovascular system. Finally, Bond et al. (1983) found no increased risk for self-reported
hypertension among workers involved in the production of trichlorophenol and 2,4,5-T.
7.13.8.1. Comment
The picture relating exposure to 2,3,7,8-TCDD and diseases of the circulatory system
is mixed. Animal data indicate that high doses of 2,3,7,8-TCDD affect cardiac and vascular
integrity (Allen and Carstens, 1967; Allen et al., 1977; Norback and Allen, 1973). Data
from a few animal studies suggest that relatively high doses of 2,3,7,8-TCDD cause damage
to the myocardium and heart valves in rats (Kociba et al., 1978; Buu-Hoi, 1972) and to the
arterial walls in rabbits (Brewster et al., 1987). Other research found that 2,3,7,8-TCDD
may alter cardiac function in rats and guinea pigs (Hermansky et al., 1988; Kelling et al.,
1987; Canga et al., 1988), causing reduced spontaneous and isoproterenol-induced heart
contractility; this suggests that 2,3,7,8-TCDD may increase the risk of arrhythmias among
the dosed animals. The authors postulate that 2,3,7,8-TCDD alters cyclic-AMP
concentrations, altering the responsiveness of cardiac cells to /3-adrenergic stimuli (Brewster
et al., 1987). In contrast, histopathologic changes were not observed in the cardiovascular
system of hamsters, which appears to be resistant to the effects of 2,3,7,8-TCDD at levels of
3,000 /xg/kg of 2,3,7,8-TCDD (Olson et al., 1980). Other experimental studies suggest an
association between 2,3,7,8-TCDD and alterations in lipid levels (Poli et al., 1980; Albro et
al., 1978; Bombick et al., 1984; Swift etal., 1981).
Findings from mortality and morbidity studies of production workers are not
definitive, and suggest the need for analyses more specifically examining possible effects on
the circulatory system and heart. The outcomes examined in the animal and human studies
are different. Animal studies describe morphologic and chemical changes in the vascular and
cardiac cells caused by 2,3,7,8-TCDD. On the other hand, the diseases and causes of death
in the human studies assessed the possible consequences of exposure on long-term pathologic
changes to the tissues, which cause cell and, sometimes, organ and system failure. Using the
animal data, it is possible to project the long-term consequences of exposure to the organ or
system. However, the animal studies do not account for the possibility of intervening events,
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such as the repair of tissue after exposure ends, or other events that reduce the hypothesized
end point to a much less drastic outcome.
Circulatory diseases and diseases of the heart have not been rigorously investigated as
hypothesized health outcomes of exposure to 2,3,7,8-TCDD. Therefore, mortality studies
have not considered the possible confounding effect of other variables, including smoking,
lipid levels, and other conditions, that influence circulatory diseases and disorders of the
heart. Furthermore, the mortality studies tended to examine the circulatory system as a
whole, without consideration of the possible pathophysiology of 2,3,7,8-TCDD exposure on
the various components of the system, e.g., vessels versus cardiac muscle damage. Finally,
mortality studies count only those conditions that ultimately caused the death of the
individual. Events such as myocardial infarctions, which may debilitate the individual but
not cause death, may be missed if death is caused by another circumstance. Therefore, the
effect of 2,3,7,8-TCDD on the coronary arteries might be missed because it was not coded
as the underlying cause of death on the death certificate.
With the exception of the Ranch Hand study (Roegner et al., 1991), cross-sectional
analyses of other more highly exposed groups were limited by their lack of good exposure
data and their inability to examine the relationship between serum 2,3,7,8-TCDD levels and
diseases of the circulatory system or heart. Such studies may also be limited by the fact that
they include a survivor population, as described in the introduction.
Further research would be useful to define the relationship between the pathologic end
points observed in animals after high, single doses of 2,3,7,8-TCDD and the disease
outcomes observed in humans after high, long-term exposure. To identify whether 2,3,7,8-
TCDD has an effect on the human vasculature, additional work is needed to determine
whether certain doses of 2,3,7,8-TCDD cause changes in the human vascular system; to
determine whether there are changes in the action of chemicals associated with human
cardiac muscle contraction caused by 2,3,7,8-TCDD exposure; and to assess mortality and
morbidity in individuals with potential for 2,3,7,8-TCDD exposure while carefully
controlling for other risk factors and using more accurate measures of exposure.
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7.13.9. Pulmonary Effects
Studies of long-term exposure to 2,3,7,8-TCDD in Sprague-Dawley rats (Kociba et
al., 1979; Van Miller et al., 1977), B6C3F1 mice (NTP, 1982a), Swiss-Webster mice (NTP,
1982b), and rhesus monkeys (Allen et al., 1977) have reported changes in bronchiolar or
alveolar tissue ranging from epithelial hyperplasia and metaplasia to squamous cell
carcinomas. The hyperplastic and metaplastic changes observed in exposed animals are
similar to the pathologic picture of chronic bronchitis in humans (American Thoracic Society,
1962).
Case reports have described temporary respiratory irritation (Zack and Suskind, 1980)
and tracheobronchitis (Goldman, 1972) among chemical workers exposed to 2,3,7,8-TCDD-
contaminated herbicides following industrial accidents. In addition, Baader and Bauer (1951)
reported chronic bronchitis in seven workers involved in pentachlorophenol production,
which resolved in all but two workers within 2 weeks after production was discontinued.
There is conflicting evidence from controlled epidemiologic studies regarding an
association between chronic respiratory system effects and human exposure to substances
contaminated with 2,3,7,8-TCDD. One study of workers involved in the production of TCP
and 2,4,5-T suggested that 2,3,7,8-TCDD exposure increases the risk for abnormal
ventilatory function (Suskind and Hertzberg, 1984). This study found a statistically
significantly increased risk for an abnormal forced expiratory volume at 1 second (FEVj)
(p<0.01), an abnormal forced vital capacity (FVC) (p<0.001), and an abnormal FEV,/FVC
ratio (p<0.05) among workers who were smoking at the time of the study. For workers,
the percent predicted spirometric parameters for FEVl5 FVC, and FEVj/FVC were 99.4%,
92.7%, and 76.5% and for referents, 104.4%, 97.6%, and 79.9%, respectively. The only
other study of TCP and 2,4,5-T production workers that reported ventilatory function
findings found no association between serum 2,3,7,8-TCDD levels and declines in ventilatory
function (Calvert et al., 1991). The disparity in results between the two studies may be
related to the age of the unexposed populations and potential exposures of the exposed. In
the Suskind and Hertzberg study, the exposed workers were, on average, 10 years older than
the unexposed workers. Although the authors indirectly adjusted for age by analyzing
age-adjusted ventilatory measures, it is not clear if these adjustments can completely control
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for a 10-year difference in age. In the study by Calvert et al. (1991), the difference in mean
age between the exposed and unexposed groups was 0.6 years. The second difference
involves the potential for exposure to 2,4,5-T acid dust at the plants studied. The 2,4,5-T
acid that was produced at the plant studied by Suskind and Hertzberg was finished as a
powder. At the plants studied by Calvert et al. (1991), the 2,4,5-T acid was finished as a
liquid. Therefore, the potential for exposure to 2,4,5-T acid dust was greater at the plant
studied by Suskind and Hertzberg (1984). Although we are not aware of any published
reports supporting an association between ventilatory function and 2,4,5-T acid exposure, a
respiratory burden of particles, in the absence of a specific toxic agent, can be a probable
cause of ventilatory function declines (Becklake, 1985).
The Ranch Hand study also examined the association between serum 2,3,7,8-TCDD
level and respiratory system effects (Roegner et al., 1991). This study found significant
declines in the mean FEV, and the mean forced expiratory volume (FVC) for Ranch Hands
with serum 2,3,7,8-TCDD levels above 33.3 pg/g (adjusted mean FEVj = 91.3%; mean
FVC = 87.4) compared to a nonexposed comparison group (adjusted mean FEV! = 93.5%;
mean FVC =91.7) (Roegner et al., 1991). The 2,3,7,8-TCDD-related declines were small
and were interpreted by the authors to be "subtle" and "not clinically significant." As
expected, smoking appeared to have the greater influence on lung function.
7.13.9.1. Comment
In conclusion, case reports indicate that intense acute exposure to 2,3,7,8-TCDD can
produce respiratory irritation. However, the findings from controlled epidemiologic studies
do not support an association between 2,3,7,8-TCDD exposure and chronic effects on the
respiratory system.
7.13.10. Renal Effects
There is little evidence in the animal or human data to suggest that exposure to
2,3,7,8-TCDD is related to renal or bladder dysfunction. In a single case report, a child
exposed to 2,3,7,8-TCDD after contact with soil sprayed with contaminated waste oil was
diagnosed with focal pyelonephritis (Kimbrough et al., 1977). After diagnosis and treatment,
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the condition resolved with no reported recurrence. No major renal or bladder dysfunctions
were noted among Air Force Ranch Hands (Lathrop et al., 1984, 1987; Roegner et al.,
1991) or among TCP production workers from West Virginia (Suskind and Hertzberg, 1984)
or New Jersey (Poland et al., 1971).
7.13.11. Reproductive Effects
Animal studies of reproductive effects have focused primarily on maternal 2,3,7,8-
TCDD exposure and teratogenesis (Courtney and Moore, 1971; Courtney, 1976; Giavinni et
al., 1983; Khera and Ruddick, 1973; Neubert and Dillman, 1972). Some ambiguity is
introduced into the interpretation of results because of differences in study design, dosages,
and species sensitivity. However, prenatal exposure to 2,3,7,8-TCDD has been associated
with cleft palate and kidney abnormalities in rats (Khera and Ruddick, 1973) and mice
(Courtney and Moore, 1971; Neubert and Dillman, 1972). Fetal resorption (Courtney, 1976)
and reduced fertility (Giavinni et al., 1983) have also been observed in studies of maternal
exposure.
Fewer studies have focused on the effects of dioxin on the male reproductive system
or on the results of matings in which only the males were exposed to dioxin. Studies of
male exposures have not provided evidence of changes in sperm characteristics or effects on
the offspring (Lamb et al., 1980; Murray et al., 1979). Chapter 5 contains a more detailed
account of the animal literature.
Thus, experimental research has emphasized maternally mediated reproductive effects
of 2,3,7,8-TCDD while in humans, studies of paternal exposures have predominated.
Moreover, assessment of exposed male animals has most commonly examined the effects of
2,3,7,8-TCDD on spermatogenesis, fertility, and sex organ development (Theobold and
Peterson, in press), whereas studies of human males have mainly targeted effects on
congenital malformations and recognized spontaneous abortions. Experimental research
designed to corroborate human investigations may provide critically needed data to plug the
gaps in our understanding of the mechanisms through which 2,3,7,8-TCDD exposure may
operate to produce adverse reproductive events in humans.
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The origin of concerns regarding a potential link between exposure to chlorinated
dioxins and adverse reproductive events can be traced to early animal studies reporting
increased incidence of developmental abnormalities in rats and mice exposed early in
gestation to 2,4,5-T (Courtney and Moore, 1971). This was of grave concern, as the U.S.
military's most widely used herbicide during the Vietnam conflict at that time, Agent
Orange, was composed of approximately equal proportions by weight of the n-butyl esters of
2,4-D and 2,4,5-T. The latter is contaminated by 2,3,7,8-TCDD during manufacture.
One dilemma encountered when attempting to review the epidemiologic literature
dealing with dioxins and reproductive effects is the categorization of studies of sufficient
similarity to allow for comparative analysis. These studies vary greatly in the nature
(occupational, environmental) and route (inhalation, digestion, absorption) of exposure; in the
reproductive outcomes examined (often multiple end points were considered, and case
definition differed across studies); in the assessment of parental exposure (maternal, paternal,
or both); and in the timing of exposure relative to the pregnancy.
The examination of reproductive and developmental disorders poses several challenges
to the researcher as compared with other health outcomes. First, to understand both normal
and pathologic reproduction, evaluation should include paternal and maternal, and sometimes
fetal, contributions. The recent increased interest in male-mediated reproductive toxicity
emphasizes the need to consider the couple as the unit of analysis in many reproductive study
settings.
The second challenge to researchers is the interrelatedness of the spectrum of
reproductive end points available for study. Fecundity (the capability to conceive), fertility
(the capability to produce live children), and early pregnancy loss (those conceptions that do
not survive to be recognized by usual diagnostic methods) as related to 2,3,7,8-TCDD
exposure have not been evaluated. Studies investigating early fetal loss and reduced fertility
among couples where exposure to 2,3,7,8-TCDD is an issue are not available. Clearly,
these end points affect the rates of reproductive outcomes occurring later in the reproductive
spectrum.
Another feature of reproductive effects is the changing vulnerability of the developing
organism throughout gestation. Exposure to a single teratogen throughout pregnancy may
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result in different effects at various stages of gestation. The window of susceptibility varies
among different teratogens; therefore, knowledge of the timing of exposure is critical in these
studies.
Finally, although not restricted to studies of reproductive events, care must be given
to the collection and analysis of confounding variables. These factors may need to be
obtained for both mother and father, with attention to the timing of specific characteristics,
such as changes in smoking or occupation during pregnancy. These are some points to keep
in mind as the reader reviews the studies of 2,3,7,8-TCDD and reproductive effects
presented here.
The reproductive effects of dioxins in humans were succinctly and elegantly reviewed
by Hatch in 1984 (Hatch, 1984b). In her review, Hatch employed type of exposure, i.e.,
populations that were exposed occupationally, environmentally, or through industrial accident
or military service, as the classification scheme for the research presented. The same
approach has been followed in this review. The earlier investigations of 2,3,7,8-TCDD and
reproductive effects, i.e., those conducted prior to 1984, are presented separately from the
more recent studies. The development of assays in the mid-1980s to quantitate 2,3,7,8-
TCDD in serum and adipose tissue, allowing individual measurements of exposure, warrants
this dichotomy of the research.
7.13.12. Review of the Literature Prior to 1984
7.13.12.1. Occupational Studies
Epidemiologic studies of occupational dioxin exposure and reproductive effects have
focused on potential paternally mediated effects, with exposures occurring at varying
intervals relative to conception. Townsend et al. (1982) interviewed 370 wives of employees
exposed to dioxins at the Dow Michigan Division in Midland, Michigan (63% of those
eligible for the study), and 345 control wives of Dow employees who were not exposed to
dioxin (62% of the eligible control pool). Exposure classification was determined by an
industrial hygienist familiar with the processes performed at the plant. Employees were
considered exposed to dioxin if they had been assigned for at least 1 month to specific jobs
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associated with chlorophenol processes between 1939 and 1975. All outcomes were reported
by the employee's spouse.
There was no systematic attempt to ascertain the reason(s) for the high refusal rate in
both cohorts. "Unsolicited reasons" for refusal included divorce, death of spouse, or no
pregnancies; no breakdown by cohort was provided. The possibility of differential rates of
infertility or early pregnancy loss, as reflected by the reported absence of pregnancy, was not
addressed in this study.
For the multiple end points of spontaneous abortion, stillbirth, birth defects, infant
mortality, and childhood morbidity, no significant association between dioxin exposure and
any adverse event was identified (Table 7-43). Given the very long interval when both
exposure and event could have occurred and the minimum requirement for paternal exposure
being 1 month at any time during the interval, an effect of exposure, if it existed, would
have been diluted by this approach unless the damage was "irreparable," as it was defined in
this report. The authors do not discuss whether time since last exposure was considered in
the analysis.
In another study of occupational exposure to 2,4,5-T, Smith et al. (1982) evaluated
548 professional chemical sprayers and their spouses in New Zealand, ascertaining spraying
activities from the males and reproductive histories from their spouses through a mailed
questionnaire. A group of 441 agricultural contractors and their spouses served as controls.
This study had impressive response rates of 89% for the exposed and 83% for the
nonexposed groups. The investigators noted that wives of the chemical sprayers anecdotally
reported assisting their husbands in spraying activities, some performing this task during their
pregnancy. No associations between herbicide exposure and the outcomes of spontaneous
abortion (OR=0.89, 95% CI=0.6-1.3) or congenital malformations (OR=1.2, 95% CI=0.6-
2.4) were identified.
The questionnaires were completed in 1980 and elicited information on spraying
activities and reproductive events that occurred between 1969 and 1980, a rather long period
of time for recall of these events. Reported pregnancy outcomes were categorized into three
groups based on whether the fathers had sprayed any chemicals at any time during, or prior
to, the calendar year in which the pregnancy occurred and whether 2,4,5-T had been used.
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Table 7-43. Results of Studies Examining the Effect of Dioxin on Reproductive Outcomes in Humans, 1984-1992
Author
Smith et al..
1982
Townsend et
al.. 1982
Exposed
group
548 male
pesticide
applicators
who sprayed
2.4,5-T and
other
pesticides
370 male
chemical
workers
exposed to
2.4,5-T only
and their
spouses
Control
group
441 agricultural
contractors
345 employees
not exposed to
2,4,5-T and their
spouses
Type of
exposure
Spraying of
2.4.5-T
Working with
chlorophenol
processes
Data source:
exposure/
outcome
Mailed survey/mailed
survey
Interviewer-
administered
questionnaire/
interviewer-
administered
questionnaire
Outcome
Total births
Congenital defect
Miscarriage*1
Stillbirth
# Conceptions
All fetal deaths
Stillbirth
Spontaneous
abortions
Infant deaths
Health defects
Congenital
malformation
Outcome in
exposed
(N)
1,172
13
43
3
418
50
7
43
6
32
21
Outcome in
unexposed
(N)
1,122
9
40
0
2,031
246
33
213
39
155
87
OR
1.19*
0.89
—
— °
1.02
0.97
0.96
0.82
0.93
1.08
95% CI
0.58-2.45
0.61-1.30
—
0.71-1.47
0.38-2.36
0.65-1.42
0.30-2.09
0.60-1.43
0.63-1.83
I
6
o
I
I
s
n
-------
Table 7-43. (continued)
Author
Townsend et
al., 1982
(cont.)
Mastroiacovo
et al., 1988
Stehr et al.,
1986
Hoffman and
Stehr-Green,
1989
Exposed
group
Male chemical
workers
exposed to any
dioxins
2,900 births in
Zones A, B,
and R, Seveso,
Italy, 1977-
1981
68 persons
residing in
areas of high
TCDD
concentration
154 residents
of Quail Run
Mobile Home
Park
Control
group
12,391 births in
study area
outside Zones A,
B, and R
36 persons with
no known
contact with
contaminated
soil
155 residents
nonexposed
mobile park
home
Type of
exposure
TCDD cloud
released from
chemical plant
accident
Contact with
TCDD-
contaminated
soil
Contact with
soil sprayed
with TCDD for
dust control
Data source:
exposure/
outcome
TCDD soil analysis/
Seveso Congenital
Malformations
Registry
EPA soil analyses for
TCDD/interview
EPA soil analyses for
TCDD/interview
Outcome
Total conceptuses
All fetal deaths
Stillbirth
Spontaneous
abortions
Infant deaths
Health defects
Congenital
malformation
Total birth defects
Multiple birth
defects
Syndromes
Major birth
defects
Minor birth
defects
Infertility (males)
Impotence
Infertility
(females)
Fetal deaths
Spontaneous
abortions
Congenital
malformations
Outcome in
exposed
(N)
737
100
15
85
9
52
30
137
10
5
67
70
-
-
-
-
Outcome in
unex posed
(N)
2,031
246
33
213
39
155
87
605
38
29
343
262
-
-
-
-
OR
1.03"
1.06
1.03
0.63
0.85
0.85
0.97
1.12
0.74
0.83
1.14
-
-
__
_
-
95% CI
0.78-1.37
0.54-2.09
0.77-1.39
0.27-1.39
0.60-1.21
0.53-1.35
0.83-1.13
0.63-2.02
0.33-1.63
0.67-1.04
0.92-1.42
NS
NS
NS
NS
NS
NS
0
o
o
c
o
H
m
O
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a
-------
DRAFT-DO NOT QUOTE OR CITE
3
$
as
0
.S -e
Outcome
unexpose
(N)
iig
,
h8
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s. i
E^ S
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|||
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'S m
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§£•38
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CN
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S ti ri ^
o o o o
o — o »
CJ f^ •* *^
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Total birth defects
Low birth weight
Perinatal mortality
Suspected birth
defects
<**
5 3 J
OS
0 J 13
|i !
Ill
u
o
11
> i
a1 §
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If 1
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7-195
06/30/94
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Table 7-43. (continued)
Author
Substudy #2.
CDC 1988d,
1989
Stellman et
al.. 1988
Aschengrau
and Monson.
1989
Aschengrau
and Monson,
1990
Wolfe et al.,
1992b
Exposed
group
127 offspring
of Vietnam
veterans
2,858 Vietnam
veterans
201
spontaneous
abortion cases
at Boston
Hospital for
Women
966 infants
with late
adverse
pregnancy
outcomes at
Boston
Hospital for
Women
2,533
conceptions
among 791
Ranch Hand
personnel
Control
group
94 offspring of
non-Vietnam
veterans
3,933 non-
Vietnam veterans
1.119 full-term
births at Boston
Hospital for
Women
998 normal term
infants at Boston
Hospital for
Women
2,074
conceptions
among 768 non-
Ranch Hand
personnel
Type of
exposure
Vietnam
military service
Vietnam
military service
Vietnam
military service
Vietnam
military service
Spraying/
handling of
Agent Orange
Data source:
exposure/
outcome
Military records/self-
report and hospital
record verification
Survey/
survey
Military records/
hospital records
Military records/
Hospital records
Serum TCDD levels/
hospital and medical
records
Outcome
Cerebrospinal
malformations
Difficulty
conceiving
Time to
conception
Birth weight
Spontaneous
abortion
Spontaneous
abortion
Total birth defects
> 1 Major
malformation
Minor
malformation
Stillbirths
Neonatal deaths
Total birth defects
£10*
15-^33.3
>33.3
Outcome in
exposed
(N)
26
349
4.4 years
-
231
8
55
18
11
5
3
202. 11
293.1
193.8
Outcome in
unex posed
(N)
12
363
4.4 years
-
195
44
656
151
189
51
36
208.0
OR
1.8
1.2
-
-
1.3
0.9
1.3
1.8
0.9
1.5
1.2
0.96m'n
1.58
0.92
95% CI
0.8-4.0
NS
NS
NS
1.4-2.0
0.42-1.9
0.9-1.9
1.0-3.1
0.5-1.7
0.4-3.9
0.2-4.2
0.69-1.34
1.10-2.27
0.64-1.32
o
o
O
o
XJ
o
-jj
1—*
VO
-------
Table 7-43. (continued)
Author
Wolfe et al..
1992b (cont.)
Exposed
group
Control
group
Type of
exposure
Data source:
exposure/
outcome
Outcome
Nervous system
anomalies
^lO*
15- £33.3
>33.3
Respiratory
system anomalies
£10*
15-^33.3
>33.3
Digestive system
anomalies
<10k
15-<33.3
>33.3
Genital anomalies
33.3
Urinary system
anomalies
•no*
15- £33.3
>33.3
Musculoskeletal
deformities
£10"
15-£33.3
>33,3
Outcome in
exposed
(N)
O.O1
5.7
13.2
7. 11
5.7
4.4
21.31
34.5
17.6
3.51
51.7
13.2
14.21
34.5
22.0
120.61
143.7
105.7
Outcome in
unexposed
(N)
3.1
2.0
24.51
18.31
12.21
134.61
OR
—
1.88°-°
4.37»-°
3.5n-°
2.83
2.17
0.83
1.30
0.64
0.19
2.92
0.72
1.16
2.88
1.82
O.W*
1.08
0.76
95% CI
0.20-18.3
0.87, 21.8
0.49-25.0
0.26-31.4
0.20-24.0
0.31-2.23
0.48-3.51
0.21-1.91
0.03-1.43
1.29-6.61
0.21-2.46
0.37-3.63
1.07-7.79
0.63-5.22
0.59, 132
0.68-1.71
0.48-1.21
o
o
z
9
O
c
s
w
g
H
W
-------
Table 7-43. (continued)
Author
Wolfe et al.,
1992b (cont.)
Exposed
group
Control
group
Type of
exposure
Data source:
exposure/
outcome
Outcome
Anomalies of the
skin
<10"
15-<33.3
>33.3
Circulatory
system and heart
anomalies
iSlO1
15-<33.3
>33.3
Outcome in
exposed
(N)
17.7'
34.5
8.8
14.2'
46.0
8.8
Outcome in
unexposed
(N)
21.41
16.3
OR
0.76
1.83
0.46
0.85
2.16
0.32
95% CI
0.25-2.26
0.72-4.70
0.11-1.95
0.27-2.69
0.81-5.74
0.04-2.52
O
o
2
o
H
O
w
o
&
o
w
VO
oo
"Relative risk.
bRate: 86/1.000 births in applicators; 93/1,000 births in agricultural contractors (controls).
"Adjusted for mother's age at time of birth, birth control methods, labor and delivery complications, medical conditions and medications during
pregnancy, smoking and alcohol use during pregnancy, high job risk, and gravidity.
dRate: Zone A: 0; Zone B: 57.5/1.000 births; Zone R: 45.1/1,000 births; zone non-ABR: 48.8/1,000 births.
•90% CI.
fZones A and B.
•Zones A and B vs. non-ABR.
'Controls matched on maternal age, race, hospital of birth, plurality, and year of birth.
"Odds ratio adjusted for veteran's age at birth, year of entry in army, enlistment status, general technical test score, military occupational specialty,
years between entry and birth, maternal age, and gravidity.
JChildren born after the father was stationed in Southeast Asia (SEA).
''Logit (p) = /30 + 01dl + /Jjdj + ftd3, where p = probability of an adverse reproductive outcome; dt. d2, d3 are indicators for the dioxin
categories: Unknown (Ranch Hands with up to 10 pg/g current dioxin), Low (Ranch Hands with more than 15 pg/g of lipid and up to 33.3 pg/g
of lipid current dioxin), and High (Ranch Hands with more than 33.3 pg/g of lipid current dioxin).
'Rate of abnormals.
"Unadjusted.
"Adjusted analysis not statistically significant.
"No adjusted analysis: total defects < 10.
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There were 1,122 births reported among the 441 control spouses and 1,172 among the
548 spouses of the exposed sprayers. If all other factors were considered equivalent in both
cohorts (maternal age, socioeconomic status, and maternal smoking histories were shown to
be similar), then there appear to be 220 fewer births observed in the exposed group than
might be expected.
This study is limited by the lack of information on the total number of conceptuses
and the high probability of exposure to chemicals other than 2,4,5-T. While exposure levels
were not quantitated in this study, a later study of a subset from this same population was
conducted to estimate 2,3,7,8-TCDD exposure in this group.
In 1988, nine pesticide applicators with the greatest number of years and months per
year of pesticide application were selected for serum 2,3,7,8-TCDD analysis (Table 7-44).
The mean serum 2,3,7,8-TCDD level, adjusted for total lipids, among the cases (53.3 pg/g)
was 10 times that of age-matched controls (5.6 pg/g). However, exposure in the
reproductive study was based on self-reports of pesticide application, and research has
demonstrated repeatedly that self-reports do not correlate with documented serum 2,3,7,8-
TCDD levels (Needham et al., 1991). Therefore, it would be helpful if serum levels could
be obtained from a subset of those applicators with lower self-reported exposures.
A clinical epidemiology study conducted in 1979 examined workers involved with the
manufacture of 2,4,5-T between 1948 and 1969 in Nitro, West Virginia (Suskind and
Hertzberg, 1984). All active and retired plant employees exposed to the 2,4,5-T process
during that 22-year interval comprised the eligible pool of "exposed" subjects. The control
group consisted of current and former plant employees who were never associated with the
2,4,5-T process, according to company records. The response rates for these cohorts were
61% (N=204) and 46% (N=163), respectively.
A reproductive history was obtained from these male employees during an interview
and clinical examination. No attempt was made to verify reports of live births, infant deaths,
miscarriages, birth defects, and stillbirths with the spouses or through medical records.
There were no significant differences in rates of any adverse outcome by exposure status.
Given the poor response rates, crude measure of exposure, and lack of verification of
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Table 7-44. 2,3,7,8-TCDD Levels (pg/g of Lipid) for Selected Populations
Author
Mocarelli et al.,
1991
Patterson et al.,
1986b
Smith et al.,
1992
CDC, 1988
Kahn, 1988
Study population
Seveso, Italy
residents:
10 Zone A
10 former Zone A
10 non-ABR Zone
39 Missouri residents
with history of
TCDD exposure
57 Missouri residents
with no known
TCDD exposure
9 New Zealand
pesticide applicators
9 controls
646 Vietnam ground
combat troops with
service in heavily
sprayed areas
97 non-Vietnam
veterans
10 "heavily exposed"
Vietnam veterans
10 Vietnam veterans
with "little or no"
exposure
7 non- Vietnam
veterans
Specimen
Serum
Adipose tissue
Serum
Serum
blood (per lipids)
adipose tissue
blood (per lipids)
adipose tissue
blood (per lipids)
adipose tissue
Range
828 - 56,000
1770 - 10,400
nd-137
2.8 - 750
1.4-20.2
3.0- 131.0
2.4- 11.3
nd-45
nd- 15
-
-
-
Mean
19,144
5,240
79.7
7.4
53.3
5.6
4.2
4.1
46.3
41.7
6.6
5.1
4.3
3.2
Median
14,000
4,540
17.0
6.4
37.6
9.3
3.8
3.8
25.1
15.4
5.3
5.4
3.9
3.5
O
O\
Tl
H
O
O
I
O
w
O
»
n
H
W
-------
Table 7-44. (continued)
Author
Schecter et al. ,
1989
Rang, 1991
Roegner et al.,
1991
Phuong, 1989b
Study population
26 Vietnam veterans
36 Vietnam veterans
79 Non-Vietnam
veterans
80 Civilians
872 Ranch Hands
1060 Controls
Vietnamese
Populations:
9 OB/GYN patients
from a South
Vietnam hospital
Specimen
adipose tissue
adipose tissue
serum
adipose tissue
Range
nd- 11
-
-
-
0-617.8
0-54.8
nd - 103
Mean
5.8
13.4
12.5
15.8
-
-
23
Median
-
10.0
11.4
11.8
12.8
4.2
11.3
to
o
o
z
o
H
s
o
HH
H
W
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DRAFT-DO NOT QUOTE OR CITE
paternally reported reproductive histories, this study was not likely to detect an association
between 2,3,7,8-TCDD and reproductive events if one existed.
7.13.12.2. Environmental Studies
The problem of documentation of exposure is perhaps of even greater concern in
studies of subjects environmentally exposed to dioxins that are lacking individual 2,3,7,8-
TCDD measures. The route of exposure (inhalation, ingestion, absorption), length and
intensity of exposure, and the timing of the exposure are more difficult to estimate in free-
living populations compared with workers in an occupational setting.
Selection bias is also a critical concern in environmental studies because there is no
equivalent to company records in the occupational setting that defines the population at risk.
Proximity of residence to a contaminated site was generally the best option available for
identifying the population at risk of exposure. Issues such as length of time at that residence,
the amount of time the subject spent in the home, and the occurrence of the contamination
episode relative to the time spent at home may have greatly affected the degree of exposure.
Volunteer bias, which is inherent in studies that rely on subjects responding to
publicized requests for participation, is an additional concern in studies of this nature.
Moreover, it is extremely difficult to conduct epidemiologic investigations under crisis
situations such as the industrial accident that occurred in Seveso, Italy. Given these
limitations, the efforts of these investigators have provided valuable impetus to the
refinement of study designs and the development of more sophisticated techniques to explore
this issue.
7.13.12.3. The Seveso, Italy, Dioxin Accident of 1976
The 1976 accident at a chemical plant near Meda, Italy, in which effluent containing
2,3,7,8-TCDD among other contaminants was released, has been well described (Pocchiari,
1979; Reggiani, 1978). Following the reactor release, an extensive surveillance system was
initiated to monitor the health of the exposed population. Rates of reproductive events were
examined by "zones" of exposure, as described above.
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The Seveso Congenital Malformations Registry enrolled all live births and stillbirths
occurring January 1, 1977, through December 31, 1982, to women residents of Zones A, B,
R, and non-ABR (Mastroiacovo et al., 1988). A total of 15,291 births (live and still) and
742 birth defects were recorded: Zone A, 26 births and no birth defects; Zone B, 435 births
and 25 birth defects (57.5/1,000 births); Zone R, 2,439 births and 110 birth defects
(45.1/1,000 births); Zone non-ABR, 12,391 births and 605 birth defects (48.8/1,000 births).
Birth defects were confirmed by medical records. Relative risk estimate for total defects
comparing Zones A+B with Zone non-ABR is 1.2 (90% CI=0.88-1.64) (Table 7-43). The
rate of birth defects occurring early in the observation period (first quarter of 1977) did not
differ from the rate of birth defects occurring later in the observation period (data not
shown), and there were no discernible patterns of defects within or between the exposure
groups.
To document an increase in the rate of an event, sound baseline information is
required. No reliable information regarding rates of birth defects in this area was available
until the establishment of the Seveso Congenital Malformations Registry 6 months after the
exposure incident. Thus the potential contribution of 2,3,7,8-TCDD to the congenital
malformation rate cannot be separated from improved case ascertainment.
The influence of both spontaneous and induced abortions on the birth defect rate is
likewise unknown. As with congenital malformations, reliable background data for
spontaneous abortion in the study area were not available. The impact of induced abortions
sought after the accident is difficult to estimate, particularly since abortion was illegal in Italy
in 1976. The birth rate for the entire study area declined between 1976 and 1980, and the
small number of conceptions available for study limits the power of these studies to detect an
association between 2,3,7,8-TCDD exposure and spontaneous abortion and birth defects.
Another adverse outcome that has not been described in most epidemiologic studies is
the effect of 2,3,7,8-TCDD on the chromosome. Pocchiari et al. described one study in
which 30 therapeutic and four spontaneous abortions from the Seveso area were examined
(Pocchiari, 1979). There were no indications of mutagenic, teratogenic, or embryotoxic
effects that could be attributed to 2,3,7,8-TCDD exposure. However, it was difficult to
determine maternal exposure status for any of the cases. A more recent study examined the
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association between 2,3,7,8-TCDD and cytogenetic abnormalities in fetuses from abortions
induced shortly after the Seveso accident (Tenchini et al., 1983). The frequency of aberrant
cells and the mean number of aberrations per damaged cell were significantly higher in
exposed versus unexposed fetuses. However, the paper omits important data, including
whether the fetus was conceived before or after 2,3,7,8-TCDD exposure, how long after
exposure the abortion occurred, the zone in which the mother and father resided at the time
of conception and pregnancy, and tissue 2,3,7,8-TCDD concentration in the fetuses. No
other studies have examined fetal tissue in this manner to corroborate these data. This
approach offers exciting possibilities for future research, however, as genetic techniques in
the area of DNA-adduct analyses become more widely available.
7.13.12.4. Studies of Exposure to Agent Orange by Military Veterans and Vietnamese
Civilians
The problem of exposure documentation has also been a highly controversial issue in
studies of potential exposure to Agent Orange in Vietnam and adverse health effects among
Vietnam veterans and residents of Vietnam and their offspring. An early study among the
Vietnamese population encompassed a 10-year period in Vietnam from 1960 to 1969 (Cutting
et al., 1970). Exposure was dichotomized into pre- or light-spraying years from 1960 to
1965 and heavy-spraying years from 1966 to 1969. A total of 480,087 births, 16,166
stillbirths, 2,866 hydatidiform moles, and 2,355 congenital malformations of all types were
examined in this study. Pregnancy outcome data were collected from 22 hospitals.
Increases in the rates of stillbirths, molar pregnancies, and congenital malformations
were noted in the coastal plain and delta areas following heavy spraying, although the authors
emphasized the slight downward trend observed for all outcomes in the countrywide rates.
Several biases in this sampling approach severely limit the interpretation of the study's
findings. The births examined were not representative of the births in the country during
that period. In addition, the hospital records were incomplete, and transport of the mothers
to the selected hospitals resulted in uncertainty regarding maternal residence during the
pregnancy.
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A second investigation conducted in Vietnam used the HERBS tape (military records
detailing Agent Orange spraying missions) covering the period from 1965 to 1971 to
determine maternal exposure status according to area of residence (Kunstadter, 1982).
HERBS data were matched to hospital records indicating the date of birth to imply
"estimated date of conception" and to maternal residence at birth to imply "potential for
maternal exposure." Birth outcome data were collected from hospital records, which as in
the above study were subject to inaccuracies and incompleteness.
No association between spraying of Agent Orange and any type of birth defect or
perinatal mortality was noted. Although cleft lip defects increased in proportion to other
malformations during the heavy spraying period, the total number of birth defects declined
and continued to decrease after spraying activities ceased.
Finally, Australian investigators examined the relationship between service in Vietnam
during 1962-1972 and birth defects (Report to the Minister for Veterans' Affairs, 1983).
Cases were infants born with any of a defined set of congenital malformations in any of 34
hospitals in New South Wales, Victoria, and the Australian Capital Territory from 1966 to
1979. Control infants were matched to cases on maternal age, hospital, and time of birth,
yielding 8,500 matched pairs for analysis.
Fathers of case and control infants were matched against a list of members who had
served in the Australian Army during the specified time period. No associations were
detected for Vietnam service and total birth defects (OR = 1.02, 95% CI=0.8-1.3) or for any
of the approximately 100 birth defects examined. Length of service in Vietnam, time
between deplanement and conception, and Vietnam service prior to and following conception
were considered in the analysis.
A critical point that should be emphasized regarding the Australian study is the
assessment of service in Vietnam as the exposure of interest. The author clearly stated that
investigations had indicated that exposure to herbicides was "infrequent and probably very
low in Australian troops in Vietnam; the study does not exclude possible effects of herbicides
in situations of substantial exposure" (Report to the Ministry for Veterans' Affairs, 1983).
A series of unpublished studies conducted by Vietnamese investigators was reviewed
by Hatch (1984a) and should be mentioned in this review. While these reports also have
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limitations, including incomplete background data for rates of reproductive events and
sparsity of epidemiologic details provided in the studies, the assessment of Vietnamese
populations offered the opportunity to evaluate pregnancies with various patterns of parental
exposure.
Investigations conducted in northern Vietnam assessed pregnancies with no maternal
exposure to spraying activities; paternal exposure was presumed to have occurred only when
the father had performed military service in the south.
Studies of couples in southern Vietnam compared reproductive outcomes observed in
sprayed versus nonsprayed areas and represent potential associations between either maternal
and/or paternal herbicide exposure and risk of adverse pregnancy outcome.
Three studies examined presumed paternal herbicide exposure and birth defects among
pregnancies in northern Vietnam. Lang and Van compared the frequency of birth defects
during the period 1975-1978 as a function of the father having served in south Vietnam
(Hatch, 1984a). Among 2,547 offspring whose fathers had never served in the south, the
birth defect rate was 6 per 1,000 (N=15). Among 511 offspring whose fathers had served
in the south, the rate was 29 per 1,000 (N=15). Similar results were obtained in a study by
Lang et al. at unspecified agricultural and handicraft cooperatives in northern Vietnam
(Hatch, 1984a). The congenital malformation rate among "exposed" pregnancies (paternal
service in southern Vietnam) was 23-26 per 1,000 (71 or 82 of 3,147). In comparison, the
rate among the unexposed pregnancies was 5 per 1,000 (10 of 2,172).
In what is perhaps the most stringent of the Vietnamese studies by Can et al., 40,064
women from three rice-growing districts in northern Vietnam were assessed (although few
details are provided on the method of selection) (Hatch, 1984a). All of the women were
married and pregnant at least once during the war and had no history of tuberculosis,
syphilis, or malaria or of using antibiotics or hormones during pregnancy. "Detailed"
interviews were conducted by physicians and midwives, and district health records were
consulted in an attempt to validate reported pregnancies and outcomes. Only pregnancies
conceived during the conflict were considered in this study.
There were a total of 121,993 pregnancies among the 29,041 women whose spouses
were "nonexposed" and 32,069 pregnancies among the 11,023 women whose spouses had
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served in the south. Service in south Vietnam was associated with an increased rate of
spontaneous abortion (p=0.05). The increased rate of congenital malformations among the
exposed pregnancies (0.6%) compared with the unexposed (0.4%) was marginally significant
(p=0.10).
In attempt to confirm this finding, these investigators conducted a case/control study
in which they examined a random sample of 61 families of children who had survived with a
birth defect. A control group of 183 families of normal children matched on maternal age,
number of deliveries, living environment, and age was selected. Forty-nine percent of
children with a birth defect had a father who had served in south Vietnam compared with
21% of the children without malformations, yielding an odds ratio of 3.6.
Additional unpublished studies conducted in Vietnam by Vietnamese investigators
were reviewed by Constable and Hatch (1985). The reader is referred to this source for
details of the studies. It was concluded that studies of presumptive paternal "herbicide"
exposure prior to or at conception were suggestive of a relationship with congenital
malformations. The evidence for an association with spontaneous abortion was "less
convincing," and for molar pregnancies no relationship appeared to exist. Those studies that
examined both maternal and paternal "herbicide" exposure were also suggestive of a relation
to birth defects as well as spontaneous abortion, stillbirths, and molar pregnancies. Follow-
up studies by these Vietnamese investigators supported these earlier findings (Huong et al.,
1989; Phuongetal., 1989b).
In the next section, the impact of assays to measure individual levels of 2,3,7,8-
TCDD is discussed, and studies of 2,3,7,8-TCDD exposure and reproductive effects that
have been published since 1984 are presented.
7.13.13. Review of the Literature From 1984 to 1992
During the interval since 1984, assays that had been developed to measure TCDD in
serum and adipose tissue were being tested and refined. Several investigators have since
used these assays in (usually small) subsets of their study populations to describe exposure to
2,3,7,8-TCDD in their total sample and also in attempts to validate their assumptions
regarding magnitude of exposure for their study subjects. Subsets were generally selected to
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represent subjects designated as "high" versus "low" exposure by the study investigators,
utilizing various information sources including self-reports, company or military records, etc.
Table 7-44 presents the results of the exposure analyses (Mocarelli et al., 1991; Patterson et
al., 1986b; Smith et al., 1992; Centers for Disease Control Veterans Health Studies, 1988;
Fingerhut et al., 1991a; Schecter et al., 1989; Kahn et al., 1988; Kang et al., 1991; Roegner
et al., 1991; Phuong et al., 1989b).
These data illustrate wide variability in groups presumed to have been exposed to
2,3,7,8-TCDD at levels above background (<20 pg/g). For example, the mean and median
serum levels of Vietnam ground combat troops with service in areas heavily sprayed with
Agent Orange did not exceed the levels found in the U.S. general population. There is
evidence for significantly higher exposure to 2,3,7,8-TCDD among certain subgroups of
Vietnam veterans (Kahn et al., 1988; Roegner et al., 1991) as well as residents of Vietnam
(Phuong et al., 1989b), Seveso (Mocarelli et al., 1991), and Missouri (Patterson et al.,
1986b), and occupational groups (Fingerhut et al., 1991a; Beck et al., 1989).
It is also important to note that with the exception of the Ranch Hand study, these
subsets were selected to describe 2,3,7,8-TCDD exposure in the total study sample and not
for an examination of the relationship between 2,3,7,8-TCDD and reproductive events. In
addition, the data from the Ranch Hand population indicated a serum 2,3,7,8-TCDD half-life
of 7.1 years (Pirkle et al., 1989), but serum samples in this group were collected and
analyzed at 11- and 15-year intervals following exposure. Questions regarding the impact of
initial dose, age, gender, and pregnancy itself on half-life in humans remain unanswered at
this time.
7.13.14. Environmental Studies
7.13.14.1. The Times Beach, Missouri, 2,3,7,8-TCDD Episode
In 1971, a waste oil dealer in Missouri disposed of waste sludge containing
approximately 29 kg 2,3,7,8-TCDD by mixing it with waste oils as a dust control spray,
which was subsequently distributed throughout the state. General media announcements from
health officials were made, urging persons potentially exposed to 2,3,7,8-TCDD to
participate in a survey and health screening process. People were warned of their potential
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exposure by virtue of their residence, employment, or engagement in recreational activities in
the contaminated sites. A pilot study was initiated in 1983, in which a subset of those
persons who had responded to the media announcements was assessed (Stehr et al., 1986).
Approximately 800 completed questionnaires were screened to select participants; it was not
clear if this number represents all of the questionnaires that had been returned up to that
time.
This phase of the study was intended to identify potential problems for future
investigations. Persons determined to be at "high" versus "low" risk for dioxin exposure,
based on their completed surveys, were selected. "High risk" was defined as either (1)
reported residence or occupation in areas with TCDD levels between 20 and 100 ppb for at
least 2 years or in areas with TCDD levels > 100 ppb for at least 6 months, or (2)
participation in activities requiring close contact with soil in areas with TCDD concentrations
as described for similar periods of time. "Low-risk" persons were determined to have had
no access to or "regular high-soil-contact activities" in any known contaminated area.
Controls were frequency matched with the high exposed group on type of exposure site, age,
sex, race, and socioeconomic status.
Sixty-eight "high-risk" persons (83% of those eligible) and 36 "low-risk" persons
(90% response) were evaluated through physical, neurological, and dermatological
examinations, laboratory tests, and interviews.
Information on reproductive outcomes was obtained during the interview administered
to the subjects "or their nearest relative." None of the outcomes observed differed
significantly by exposure status (Table 7-43), although high-risk women had a later mean age
at menarche (p=0.06). A total of 30 births were available for assessment in this sample.
The following year, a more intensive study was undertaken to test the results found in
the pilot study (Hoffman et al., 1986). The exposed group consisted of residents of the
Quail Run Mobile Home Park in Gray Summit, Missouri, where TCDD levels in the soil
were measured at up to 2,200 ppb. Data on 95 of the approximately 207 households in the
park were available; 154 persons (74%) who were both "eligible and interested" agreed to
participate. The unexposed group consisted of 155 residents of a nonexposed trailer park,
representing 77% of those "both eligible and interested" in participation.
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The protocol was similar to that employed in the pilot study described above. The
authors reported that "no differences were found between the exposed and unexposed groups
in the frequency of reproductive disorders or adverse pregnancy outcomes, such as fetal
deaths, spontaneous abortions, and children with congenital malformations." No sample
sizes or statistical test results for any of the outcomes were reported. Clearly, these studies
of Missouri residents were not designed to investigate the relationship between dioxin
exposure and reproductive outcomes, and not very much can be learned about this association
from these results.
In a retrospective cohort study by Stockbauer et al. (1988), the association between
2,3,7,8-TCDD and reproductive outcomes was examined among residents of contaminated
areas in Missouri. All live births and stillbirths that occurred in the nine residential areas
identified as contaminated with TCDD during the period 1972-1982 were identified through
Missouri vital statistics records. A total of 402 births were examined from six residential
areas. TCDD levels in the soil from these six areas ranged from 241 to 2,200 ppb. 2,3,7,8-
TCDD levels were not reported for the three areas where no births occurred during this
interval, which would have been of interest to note.
A reference group of 804 unexposed births, matched for maternal age and race,
hospital and year of birth, and plurality, was selected from the vital statistics records.
Medical records were abstracted to ascertain birth defects in the matched sets. In addition,
the births were linked to a statewide birth defect register that had been recently initiated.
Data on several potentially confounding variables were obtained from birth certificates.
The exposed mothers tended to be less educated, had more children, were more likely
to be in the extremes of the prepregnant weight distribution, and were more likely to smoke
cigarettes. Statistical testing for these differences was not reported. Moreover, statistical
adjustment for these potential confounders was performed only in the birth weight analyses,
although there was little change in the birth weight risk ratios after adjustment.
Increased risk ratios were reported in the exposed group for infant death (OR=2.0),
fetal death (OR = 1.6), and perinatal death (OR =1.3), low birth weight (AOR=1.5), and
several subcategories of malformations, although none of these findings achieved statistical
significance (Table 7-43).
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Two approaches were used to determine a dose-response relationship. In the first, the
study data set was matched against the Missouri central listing of dioxin-exposed persons,
which yielded 98 of the exposed mothers and none of the control mothers. These 98 women
were then dichotomized into "high" (N=20) or "low" (N=78) exposure groups. High
exposure was defined as residence for at least 6 months in areas with > 100 ppb TCDD in
the soil or >2 years in areas with TCDD levels from 20 to 100 ppb. Low exposure was
defined as residence in areas with similar TCDD levels as described above but for less than
the required time period or residence at sites with 1-19 ppb TCDD. No evidence for a dose-
response relationship was observed with this analysis.
The second approach categorized the births into two different intervals relative to the
spraying of the soil with the TCDD-contaminated sludge in 1971 to 1973. When births in
the 1972-1974 period were compared with births from 1975 to 1982, the authors reported
that "the only birth outcomes with higher risk ratios in the earlier time period were very low
birth weight, birth defects, and major birth defects." None of these observations were
statistically significant, but sample sizes and testing results were not provided in the paper.
The authors acknowledged the possibility of exposure misclassification and also the
"modest" power of the study to detect associations due to the small sample size.
7.13.14.2. Studies of Vietnam Experience in Ground Troops and Ranch Hands Published
1984-1992
7.13.14.2.1. Evaluation of exposure
Evidence from earlier studies to determine if Agent Orange exposure increased the
risk of adverse pregnancy outcomes among Vietnam veterans has been described as "sparse,
sometimes off the point, sometimes conflicting..." (Hatch and Stein, 1986). The general
dissatisfaction with these studies had a common factor: the lack of a valid measure of dioxin
exposure. Once the assays to document individual 2,3,7,8-TCDD exposure became available
and served as the gold standard for assessing the exposure assumptions made by study
investigators, this concern regarding exposure misclassification was shown to be justified.
Until 1992, when the first study to examine reproductive outcomes among Vietnam veterans
based on individual exposure measurements of 2,3,7,8-TCDD was published, this remained
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the major criticism of the research. However, even though 2,3,7,8-TCDD serum levels were
available for reanalysis of the 1984 data of the Ranch Hand study, the lack of a good
estimate of 2,3,7,8-TCDD level at the time of conception continues to instill uncertainty
about the findings.
7.13.14.2.2. Exposure indices
7.13.14.2.2.1. The CDC case-control study's exposure opportunity index. In the early
1980s, the CDC conducted a large case-control study that examined the relationship between
service in Vietnam and risk of congenital malformations (Erickson et al., 1984). Although
service in Vietnam was the major exposure variable, two additional measures of exposure
were assessed. Vietnam veterans were asked if they believed they had been exposed to
Agent Orange. In addition, an exposure opportunity index (EOI), developed by the Army
Agent Orange Task Force, assigned a score estimating the likelihood of exposure based on
places and times of Vietnam service. These scores ranged from minimal (1) to high (5)
opportunity for Agent Orange exposure.
The CDC then reported on a study, using the serum assay as the standard, to evaluate
the validity of both self-reported exposure to Agent Orange and use of military records to
fomulate the EOI used in the case-control study of military service in Vietnam and birth
defects. The results revealed a poor correlation (no correlation coefficient was provided)
between both of these exposure estimates and serum TCDD levels (Centers for Disease
Control Veterans Health Study, 1988). In addition, the distributions of serum 2,3,7,8-TCDD
levels were nearly identical (median = 3.8 ppt) among 646 ground combat troops who had
served in heavily sprayed areas compared with 97 veterans who had never served in
Vietnam.
In a separate study, Kahn et al. (1988) measured serum 2,3,7,8-TCDD levels in 10
Vietnam veterans who reported that they had handled Agent Orange "regularly" while in
Vietnam, 10 Vietnam veterans with little or no exposure to Agent Orange, and 27 Vietnam-
era veterans. The levels of serum 2,3,7,8-TCDD among those men who had handled Agent
Orange were significantly elevated (median = 25.1 pg/g blood fat) compared with the
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Vietnam control veterans (median = 5.3 pg/g) and the Vietnam-era veterans (median = 3.9
pg/g) (p<0.01).
7.13.14.2.2.2. The Baseline Ranch Hand study's exposure index. In 1984, the U.S. Air
Force released preliminary results of a 20-year study designed to examine the personnel
responsible for conducting the aerial spraying of herbicides in Vietnam (Lathrop et al.,
1984). The analyses in this baseline study were based on cohort status, i.e., Ranch Hands
("exposed") versus controls ("nonexposed"). In an attempt to determine a link between
exposure and clinical end points, an exposure index was developed to estimate individual
2,3,7,8-TCDD exposure. The exposure index developed for this study was defined as the
product of a TCDD weighting factor and the gallons of TCDD-contaminated herbicides
sprayed during the veteran's tour divided by the number of Ranch Hands sharing his duties
during his tour.
In 1988, a report describing a U.S. Air Force and CDC collaborative pilot study
utilizing the serum 2,3,7,8-TCDD assay among 200 Air Force Ranch Hand personnel was
published. It was noted that the Ranch Hand personnel had significantly higher serum
2,3,7,8-TCDD levels than controls (Wolfe et al., 1988). These data also indicated that the
Ranch Hands as a whole were not as highly exposed to 2,3,7,8-TCDD as compared with the
NIOSH cohort (Piacitelli et al., 1992) and the Seveso population (Mocarelli et al., 1991).
In 1992, the report describing reproductive outcomes among Ranch Hand personnel
became available (Wolfe et al., 1992b). In addition to outcome verification, serum 2,3,7,8-
TCDD levels were measured in a sample of Ranch Hands (N=791) and the comparison
population (N=942). A comparison of the exposure index used in the baseline Ranch Hand
study with individual 2,3,7,8-TCDD levels revealed "considerable misclassification" among
the study subjects.
As a result of these investigations, it became clear that the likelihood of exposure
misclassification in studies of the relationship between 2,3,7,8-TCDD and reproductive
events, without direct measures of individual exposure, casts considerable doubt as to the
validity of the findings.
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An equally important finding resulting from these investigations was the inability of
the exposure indices to classify exposure status in individuals (Needham et al., 1991).
2,3,7,8-TCDD analyses in small subsets of the total study sample, selected on the basis of
presumed exposure as defined by an "exposure index" (either military or government
records, self-reports, or 2,3,7,8-TCDD levels in soil samples), have demonstrated that the
exposure classification schemes employed in these studies were not valid. Therefore, studies
that defined exposure as paternal military service in Vietnam should be evaluated without
inference to 2,3,7,8-TCDD exposure.
7.13.15. Study Results
7.13.15.1. Atlanta Congenital Defects Program Study
In 1984, the CDC released the findings from the large case-control study that
examined the relationship between service in Vietnam and risk of congenital malformations
(Erickson et al., 1984). Case-group babies were infants with serious structural
malformations born between 1968 and 1980 and registered with the Metropolitan Atlanta
Congenital Defects Program (MACDP). Of 7,133 eligible cases, maternal interviews were
obtained for 4,929 (69%). Control babies were selected from Georgia vital statistics records
and were frequency matched to cases on race and year and hospital of birth. Of 4,246
eligible controls, maternal interviews were obtained for 3,029 (71%). Paternal interview
rates were 56% and 57%, respectively.
While response rates were similar by case status overall, among nonwhites
significantly more cases were not interviewed. The major reason for nonparticipation was
the inability to locate subjects rather than subjects refusing to enroll.
Among the children with congenital malformations, 428 (9%) were fathered by
Vietnam veterans and 4,387 (91%) were fathered by non-Vietnam veterans; identical
percentages were noted among the control infants. The odds ratio for service in Vietnam and
birth defects of any type was 0.97 (95% CI=0.83-1.14). Odds ratios were also calculated
for 96 separate categories of birth defects, with no significant associations observed.
The EOI developed for this study (as described above) also was not associated with
total birth defects. However, significant associations were observed for spina bifida, cleft lip
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with or without cleft palate, neoplasms, and coloboma. With the exception of the last defect,
all three also showed evidence of a dose-response relationship.
For the analysis that examined potential associations between self-reported Agent
Orange exposure and birth defects, the frequency of exposure among fathers and each type of
malformation was compared with the frequency among fathers and all other defects in an
attempt to reduce recall bias. Twenty-five percent (N=74) of the Vietnam veterans reported
that they believed they were exposed to Agent Orange during their service in Vietnam. The
analysis of self-reported exposure and birth defects was negative on all counts, in contrast to
the EOI analysis, which found significant associations as described above. However, the
small numbers of cases of many individual defects resulted in a virtual lack of power to
detect any associations for these specific defects.
7.13.15.2. CDC Vietnam Experience Study
As part of a separate, large, multifaceted study mandated by Congress, the CDC
evaluated the risk of service in Vietnam and adverse reproductive outcomes (Centers for
Disease Control Vietnam Experience Study, 1988d, 1989). The Vietnam Experience Study
protocol involved two phases. In the first phase, a random sample of male veterans who met
eligibility criteria related to military service was selected for a telephone interview.
Of the eligible "exposed" group, i.e., those veterans who had served in Vietnam,
84% (N=7,924) agreed to participate, and 84% (N=7,364) of the nonexposed (those who
had not served in Vietnam) were enrolled. Of these 15,288 veterans who completed the
telephone interview, a random sample was selected for the second phase, which consisted of
a comprehensive medical examination. For this phase, the response rates were 75%
(N=2,490) for the Vietnam veterans and 63% (N= 1,972) for the non-Vietnam veterans
group.
During the telephone interview, veterans were questioned about miscarriage, induced
abortion, ectopic pregnancy, live births, stillbirths, birth defects, as well as leukemia and
other childhood cancers, and major health problems or impairments during the first 5 years
of life.
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A preliminary analysis of the interview data revealed that the Vietnam veterans had
reported 40-50% more birth defects than the non-Vietnam veterans. In addition, a difference
between the cohorts for certain subcategories of malformations, including spina bifida, cleft
lip with or without cleft palate, and hydrocephalus, was noted. Therefore, a substudy was
conducted to compare the rates of total birth defects in Vietnam and non-Vietnam veterans as
these events were recorded on hospital records.
The eligible sample for this substudy consisted of 2,282 veterans who had not yet
received their medical examinations, as it would be easier to collect the additional
information required and permission to obtain hospital records for their offspring from this
group. Hospital records were obtained for 92% (N= 1,791) of the offspring of Vietnam
veterans and 91% (N= 1,575) of the offspring of non-Vietnam veterans.
When birth defects were identified from hospital records, there was no association of
Vietnam service with total, major, minor, or suspected birth defects. From the telephone
interview data, the odds ratio for Vietnam service and total birth defects was 1.3 (95 %
CI = 1.2-1.4); in the hospital records substudy, the odds ratio was 1.1 (95% CI=0.7-1.8). It
was concluded that this finding supported the explanation of differential reporting in the
telephone interview.
The odds ratios for selected categories of birth defects calculated from both the
telephone interview and hospital records study are presented in Table 7-45. The rate of birth
defects increased for both cohorts when malformations were identified in the medical
records.
The extent of differential reporting between the two cohorts has also been described
(Centers for Disease Control Vietnam Experience Study, 1989). Overall, the authors
concluded that agreement between veterans' reports and hospital records for the presence of a
birth defect was "relatively poor" for both cohorts. Positive predictive value, sensitivity, and
the kappa statistic were slightly lower among Vietnam veterans (Table 7-46). It was further
stated that there was no evidence of selection bias or participation bias in this substudy
because no differences were noted between cohorts in health histories and demographic or
military covariates, and the participation of both groups of veterans was high. However, the
subjects in this substudy were selected from the group of veterans who completed the
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Table 7-45. Odds Ratios for Selected Categories of Birth Defects for the Telephone
Interview and Hospital Records Study in the Vietnam Experience Study, 1989
Defect Category
Total
Telephone
Rate
(per 1,000)
Vietnam
veterans
64.6
Interview
Rate
(per 1,000)
Controls
49.5
OR
1.3
95% CI
1.2-1.4
Hospital Records Study
Total
Major
Minor
72.6
28.5
32.4
71.1
23.5
34.3
1.0
1.1
1.00
0.8-1.4
0.7-1.8
0.7-1.5
Adapted from the Centers for Disease Control Vietnam Experience Study (1988d).
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Table 7-46. Results of the Misclassification Analyses for Birth Defects in the Hospital
Records Substudy, Vietnam Experience Study, 1989
Vietnam veterans
PPV:
NPV:
Sensitivity:
Specificity:
% Agreement:
Kappa index:
24.8%
95.2%
27.1%
94.7%
90.6%
20.9%
Non- Vietnam
PPV:
NPV:
Sensitivity:
Specificity:
% Agreement:
Kappa index:
veterans
32.9%
95.8%
30.3%
96.2%
92.4%
27.6%
PPV = Positive predictive value.
NPV = Negative predictive value.
Adapted from the Centers for Disease Control Vietnam Experience Study (1989).
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examination. The response rate for this phase of the study was not high, as noted above,
with 75% of Vietnam veterans and only 63% of the non-Vietnam veterans participating in the
medical examination. While characteristics of the subset who agreed to undergo physical
examination did not differ from the telephone interview sample, no data on those who
refused the exam and no reasons for refusing to participate were provided in this paper.
Adjusted odds ratios for two additional outcomes verified in the hospital records
substudy, low birth weight (AOR=1.1, 95% CI=0.8-1.4) and perinatal mortality
(AOR=1.6, 95% CI=0.8-3.1), did not differ by Vietnam service status. A race-specific
analysis of total, major, and minor defects showed an increased risk for total and minor birth
defects among black veterans. The adjusted odds ratio was 3.3 (95% CI= 1.5-7.5) for total
defects and 2.9 (95% CI= 1.1-8.0) for minor malformations. This finding is based on very
small numbers, however, and on multiple occurrences of two minor defects in two families.
From data obtained during the telephone interview, the adjusted odds ratio for
Vietnam service and spontaneous abortion was 1.3 (95% CI = 1.2-1.4). Although an excess
among Vietnam veterans was noted across all three trimesters, only the association in the
first trimester was significant. There was no attempt to confirm this end point by using
hospital records. No significant differences were observed for the reproductive outcomes of
induced abortion (AOR = 1.0, 95% CI=0.9-1.2), stillbirths (AOR=0.9, 95% CI=0.7-1.1),
ectopic pregnancies (AOR=1.0, 95% CI=0.7-1.2), or childhood cancers (OR=1.5, 95%
CI=0.8-2.8).
In addition to the reported excess of total birth defects in the telephone interview,
Vietnam veterans also reported more neural tube defects and hydrocephalus than the non-
Vietnam veterans. A second substudy was undertaken to examine the increase in
cerebrospinal malformations (CSMs). In this substudy, an attempt was made to obtain
hospital records for all of the offspring identified in the telephone interview as meeting one
of the following criteria: (1) offspring with a CSM as reported by a veteran, (2) those with a
reported condition that suggested a CSM, or (3) all reported stillbirths.
Of the 403 offspring reported to have a CSM, 109 were ineligible, 14% (N=58)
because of conception prior to father's military service, and 12.6% (N=51) were classified
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as miscarriages. Again, the issue of the impact of spontaneous abortion on rates of birth
defects is introduced.
The CSM substudy was limited by a poor response rate among the non-Vietnam
veterans. The sample thus consisted of 127 offspring of Vietnam veterans (82.5%) and 94
children of the non-Vietnam veterans (67%). Compared with fathers who participated in the
CSM study, nonparticipants were more likely to be nonwhite, less educated, unmarried,
younger at the time of their child's birth, and have lower general technical test scores.
There were also differences in paternal covariates between the participating cohorts that
could confound the findings. Compared with Vietnam veterans, non-Vietnam veterans were
better educated, had higher general technical test scores, were more likely to be married,
were older when their child was born, were less likely to consume alcohol, and were more
likely to have had a nontactical primary military occupational specialty (MOS) in the Army,
and to have served in both the later and earlier time periods.
The number of CSMs reported by the veterans that were verified by hospital records
was examined separately by stillbirth and live birth status. Among the reported stillbirths,
five CSMs were documented among children of Vietnam veterans and six among the non-
Vietnam veterans. Ten of these 11 CSM had not been reported by the veterans. Positive
predictive values derived from this analysis were 6.5% for Vietnam veterans and 8.1% for
non-Vietnam veterans.
Among the live births, 21 of 49 reported CSM cases were noted on hospital records
for the Vietnam veterans; 6 of the reported 20 cases were observed among the non-Vietnam
veterans group, yielding positive predictive values of 42.9% and 30%, respectively.
Tables 35 and 36 in the unpublished technical report of the study list the fathers'
descriptions of the birth defects in their offspring obtained through the telephone interview by
cohort (Centers for Disease Control Vietnam Experience Study, 1989). It was intriguing to
note that among Vietnam veterans, for 22 of the 55 reported cases of birth defects, the
hospital record finding was "none." In five additional cases, the hospital record finding was
"none of the nervous system." Of these 27 nondocumented reports of birth defects, 3 cases
were reported as having died within the first year of life. Death during the first year of life
was also reported for one of the eight unverified CSMs in the non-Vietnam veterans cohort.
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In summary, the CDC investigators concluded that for "most reproductive and child
health outcomes studied, Vietnam veterans were more likely to report an adverse event than
were non-Vietnam veterans." In the two substudies conducted to compare rates that were
identified through hospital records, no significant differences in adverse outcomes between
the two cohorts were determined. However, the ability of this study to address the issue of
2,3,7,8-TCDD exposure and reproductive outcome is severely limited. The question of bias
still remains in the two substudies. In addition, exposure was defined as service in Vietnam,
which does not provide much insight into the question of the reproductive toxicity of 2,3,7,8-
TCDD.
7.13.15.3. American Legion Study
The relation of self-reported exposure to Agent Orange and reproductive outcomes
was part of a study conducted among 6,810 American Legionnaires who had served during
the Vietnam war (Stellman et al., 1988). Information was obtained through a questionnaire
mailed to 2,858 veterans (42%) who had served in Southeast Asia and 3,933 veterans (58%)
who had served elsewhere. No association between Agent Orange exposure and difficulty
with conception, time to conception of the first child, or infant birth weight was observed.
However, the proportion of spontaneous abortion was significantly higher among the spouses
of veterans who served in Vietnam (7.6%) compared with controls (5.5%) (p<0.001).
These figures were well below the background rate for recognized spontaneous abortion (15-
20%) in the general population.
The significance of these findings is limited by the lack of verification of self-reported
exposure, the low response rate, the lack of outcome verification through medical records,
and the selection of veterans from the American Legion organization, as they may not be
representative of all veterans who served during the Vietnam conflict.
7.13.15.4. Boston Hospital Study
A case-control study to investigate the relationship between paternal military service
in Vietnam and risk of spontaneous abortion was conducted at Boston Hospital for Women
(Aschengrau and Monson, 1989). Cases identified through hospital records were
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spontaneous abortions <27 weeks gestation that occurred between July 1976 and February
1978 (N=201). Paternal identifying information from the hospital birth records was linked
with national and state military records to identify those fathers who had served in Vietnam.
Frequency of service in Vietnam was compared for cases and all full-term live birth controls
(N=l,119) born at the hospital during this time period. No association was detected
(OR=0.88, 95% CI=0.42-1.86).
The same investigators conducted a subsequent study that expanded the outcomes
examined to include late adverse pregnancy outcomes (Aschengrau and Monson, 1990).
Case infants identified through hospital records included 857 congenital malformations, 61
stillbirths, and 48 neonatal deaths that occurred at the hospital during August 1977 and
March 1980. "Exposure" was defined by using the same method as in the previous study.
Frequency of paternal service in Vietnam was compared for cases and 998 normal term
infant controls. Again, no associations with any of these later adverse outcomes were
detected (Table 7-43).
7.13.15.5. The Ranch Hand Study
7.13.15.5.1. Baseline study -1984. This initial report of the health of Ranch Hand
personnel used cohort status (Ranch Hand vs. comparisons) as the basis for evaluating effects
and exposure. This group of exposed veterans included those who served in Vietnam during
1962-1965, when Herbicides Purple, Pink, and Green were sprayed. These herbicides had
higher TCDD concentrations (33 ppm, 66 ppm, and 66 ppm, respectively) than Herbicide
Orange with 2 ppm TCDD (Lathrop et al., 1984).
The protocol consisted of a comprehensive personal and family health questionnaire
and a physical examination, including an in-depth laboratory analysis. The response rates for
each phase of the protocol were quite different both within and between cohorts.
Participation in the questionnaire phase was 97% (N= 1,174) for the Ranch Hands and 93%
(N=956) for controls. In the physical examination phase, participation dropped to 87%
(N= 1,045) for the Ranch Hands and 76% (N=773) for controls.
Nonresponders were "on the average" younger than participants. Ranch Hand
enlisted personnel had higher participation rates than officers, and black Ranch Hand officers
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had lower participation rates than nonblack officers. The difference in the response rates for
the physical examination phase of the protocol was ascribed partially to the active
encouragement of the Ranch Hand Association for participation and the intense media
coverage that the study received. The authors stated that the "majority" of reasons given for
nonparticipation were "no time-no interest" and passive refusal.
The reproductive outcomes evaluated in this phase of the study were ascertained
through questionnaires obtained from both the veterans and their spouses or partners. A total
of 7,399 conceptions were analyzed in this report. There were 3,293 conceptions among
1,174 Ranch Hands and 4,106 among the 1,531 controls.
No significant differences were reported for four "measures of fertility": (1) the
number of childless marriages, (2) the number of couples having achieved their desired
family size, (3) the number of childless marriages per total number of marriages, and (4) the
number of conceptions per years spent together, which included nonmarital relationships.
The fertility analysis was performed on the total number of conceptions reported and was not
adjusted for any confounding variables. Moreover, exposure in this analysis was defined by
a simple dichotomy of Ranch Hands versus controls.
To examine the relationship between Ranch Hand status and spontaneous and induced
abortion, stillbirths, and live births, exposure was stratified by pre- and post-Southeast Asia
(SEA) service. The unadjusted analysis indicated that Ranch Hands had increased
spontaneous abortion rates in both pre-SEA duty (p=0.06) and post-SEA duty (p=0.13).
The report qualified this by stating that these inferences based on analyses that were not
adjusted for "key factors affecting pregnancy outcome are of questionable value," although
no similar qualification was given for the fertility analysis. After adjustment for maternal
age, smoking, and alcohol use and paternal age, no significant difference was observed for
spontaneous abortion.
Among the live births with complete data obtained to allow for adjustment of
cofactors, no difference in risk of prematurity was noted. However, the estimate of
gestational age was based on parental report, which is not a sensitive measure of gestational
length, and it was not clear whether prematurity was analyzed as only a dichotomous variable
(<37 weeks, >37 weeks) or as a continuous variable. No analyses for birth weight
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differences by exposure status were performed, which was unfortunate given the controversy
regarding the finding of lower birth weight among infants exposed to PCDDs and PCDFs
(Yen et al., 1989; Rogan et al., 1988).
Unadjusted analyses were conducted to examine the relationship between exposure
and neonatal death, infant death, physical handicaps, birth defects, and learning disabilities.
These analyses were stratified by pre- and post-SEA service periods. The results indicated
that Ranch Hands were more likely to report physical handicaps (p=0.07), birth defects
(p=0.08), and neonatal deaths (p=0.02) in the post-SEA analysis. After adjustment for the
maternal and paternal covariates described above, the relationship with birth defects achieved
statistical significance (p=0.04); the other relationships were not statistically significant.
Twelve of the 76 birth defects reported to have occurred among the Ranch Hands
after post-SEA service were skin anomalies (ICD Code 757). When these anomalies are
excluded, this relationship is no longer statistically significant (p=0.14) although "still of
interest."
Finally, semen samples from Ranch Hands (N=560) and controls (N=409) were
analyzed for sperm count and morphology. The response rates for this parameter were
72.5% and 76.5%, respectively, although some of the samples submitted were ineligible for
analysis because of prior vasectomies and orchiectomies. Linear regression techniques
examined sperm count as a continuous variable and percentages of sperm with abnormal
morphology as dependent variables. Independent variables were age and exposure to
industrial chemicals. No differences in either parameter were identified.
This finding contrasts with the semen analysis results obtained among 324 Vietnam
veterans and 247 non-Vietnam veterans in the Vietnam Experience Study (Centers for
Disease Control Vietnam Experience Study, 1988a). That analysis indicated that Vietnam
veterans had significantly lower sperm concentrations (OR=2.3, 95% CI = 1.2-4.3), below
the clinical reference value (20 million cells/mL), than the non-Vietnam veterans. In
addition, Vietnam veterans had a significantly lower average proportion of "normal" sperm
heads. These analyses were adjusted for six covariates, although industrial chemicals were
not among them.
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7.13.15.5.2. The Ranch Hand study - reproductive outcomes 1992. The significant
association between Ranch Hand status and birth defects found in the previous study was
sufficiently troubling to launch a massive project to verify all reported conceptions and
pregnancy outcomes through medical record abstraction. In addition, in 1987 serum 2,3,7,8-
TCDD levels were obtained from a subset of Ranch Hands and controls. In 1992, the Air
Force released the results of the first study that examined the relationship between direct
measure of individual serum 2,3,7,8-TCDD levels and verified reproductive outcomes (Wolfe
et al., 1992b). A total of 4,607 conceptions were examined in this study; 2,533 were
contributed by 791 Ranch Hands, and 2,074 were contributed by 768 controls.
Ranch Hand personnel were shown to have significantly higher 2,3,7,8-TCDD levels
compared with the controls. The median values were 12.8 pg/g and 4.2 pg/g, respectively.
The 98th percentile for Ranch Hands was 166.4 pg/g; for controls, 10.4 pg/g. 2,3,7,8-
TCDD levels were determined in 1987. These results were used to estimate initial doses
received during the veterans' tour in Southeast Asia but not the 2,3,7,8-TCDD level at the
time of conception.
The fertility analysis performed in the earlier study was not repeated according to
level of serum 2,3,7,8-TCDD, which was a disappointing omission. There was a significant
variation in the association between 2,3,7,8-TCDD and miscarriage with time since SEA tour
(< 18.6 years or > 18.6 years) and time of conception (pre- or post-SEA tour) among Ranch
Hands with current 2,3,7,8-TCDD levels > 10 ppt (p=0.01) (Table 7-47). This was
attributed to the low miscarriage rate among the pre-SEA Ranch Hands with current 2,3,7,8-
TCDD levels >33.3 pg/g lipids. In examining post-SEA conceptions only, a linear trend
can be seen for spontaneous abortions and increasing 2,3,7,8-TCDD levels among Ranch
Hands who had "late tours" in SEA, i.e., less than or equal to 18.6 years had elapsed
between their tour of duty and current 2,3,7,8-TCDD levels. The opposite trend is noted in
Ranch Hands with "early tours," i.e., more than 18.6 years had elapsed between the end of
duty and the 1987 blood draw. It was concluded that 2,3,7,8-TCDD did not affect the rates
of miscarriage because it seemed "implausible that dioxin would act differently in the two
groups."
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Table 7-47. Rates of Miscarriage (per 1,000) by Pre- and Post-Vietnam Tour Status and
Time Since Tour of Duty, Among 1,475 Ranch Hands with >10 pg/g Serum Dioxin,
Ranch Hand Study, 1992a
Miscarriage rate (no./n)
current dioxin
Time of
conception
Pre-tour
Post-tour
Time since tour
(years)
<18.6
>18.6
<18.6
>18.6
10-14.9
Pg/g
142.0
(23/162)
123.9
(14/113)
92.1
(7/76)
237.3
(14/59)
15-33.3
Pg/g
146.8
(32/218)
159.4
(33/207)
136.6
(22/161)
198.6
(29/146)
>33.3
Pg/g
48.8
(2/41)
166.7
(16/96)
168.5
(15/89)
121.5
(13/107)
p-Value
0.014b
"Adapted from Wolfe et al., 1992b.
bComparison of pre- and post-tour data.
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An alternative explanation might be that there is a relationship, but it cannot be
detected by this type of analysis. To evaluate the relationship between 2,3,7,8-TCDD level
and spontaneous abortion, 2,3,7,8-TCDD level at the time of conception must be considered.
Assuming a half-life of 7 years in humans (Pirkle et al., 1989), it would seem reasonable,
for example, to assume that the two groups of Ranch Hands with 2,3,7,8-TCDD levels of
10-14.9 pg/g of lipids with post-SEA conceptions may have had very different 2,3,7,8-
TCDD levels at the time their children were conceived. This is possible because the early
tour veterans had more time to decrease their body burden of 2,3,7,8-TCDD before their
bloods were drawn in 1987 than did their late tour counterparts.
Table 7-43 illustrates the risk estimates for reproductive outcomes in the Ranch Hand
study. Interestingly, the only statistically significant associations between 2,3,7,8-TCDD and
adverse events (total birth defects, genital anomalies, and urinary system anomalies) occurred
among Ranch Hands with 2,3,7,8-TCDD levels of 15-33.3 pg/g of lipids and not among
those in the >33.3 pg/g group.
The report stated that the "expected dose-pattern" for 2,3,7,8-TCDD and total adverse
reproductive outcomes (miscarriage, tubal pregnancy, other noninduced abortive pregnancy,
or stillbirth) is the "linear one in which the highest anomaly rate occurs at the highest levels
of dioxin." This statement raises at least two questions. If a linear response is assumed,
might this imply that very early pregnancy losses occur at the highest 2,3,7,8-TCDD levels,
so that the conceptus would not survive long enough to be clinically recognized? Or, are
very early pregnancy losses and clinically recognized spontaneous abortions two separate
entities with different thresholds? Such a scenario has been suggested to explain changes in
spontaneous abortions observed after exposure to radiation in Hiroshima (Miller and Blot,
1972).
These questions are of interest because the rate of each of these end points may
directly affect the rates of all subsequent reproductive outcomes that are available for
examination. The miscarriages assessed in this study are most likely late spontaneous
abortions, as 99.6% of reported miscarriages were verified through medical records.
An analysis in which the 1987 dioxin levels are used to estimate dioxin level at time
of conception would be a worthwhile effort. If a relationship between paternal 2,3,7,8-
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TCDD level and adverse reproductive outcome does exist, this may help determine the dose-
response pattern of the relationship.
No evidence was found to support an association between 2,3,7,8-TCDD and total
adverse outcomes. These findings should be viewed with caution in view of the above
comments concerning the unexplored area of events early in gestation.
Overall, there was little convincing evidence to support an association between birth
weight, examined as both a continuous variable and dichotomized (<2,500 g or >2,500 g),
and paternal 2,3,7,8-TCDD level. Analyses that adjusted for covariates included maternal
and paternal age, maternal alcohol use and smoking, and race of the father. No assessment
of 2,3,7,8-TCDD and prematurity was reported.
The potential association between cohort status and birth defects was examined for all
defects combined and 12 additional categories of malformations. The only categories with
sufficient numbers of verified post-SEA cases to detect a relative risk of 2 were total birth
defects (229 cases among 1,045 Ranch Hands and 289 cases among 1,602 controls) and
musculoskeletal deformities (132 cases among Ranch Hands and 180 among controls).
A significant variation was observed in the association between total birth defects
(p=0.03), defects of the respiratory system (p=0.03), and urinary system abnormalities
(p=0.04) by Ranch Hand versus control status with time of conception (pre- or post-SEA).
All of these findings were due to a lower rate among Ranch Hands in the pre-SEA
conceptions and a higher rate among the post-SEA conceptions for the Ranch Hands.
Analyses of birth defects by 2,3,7,8-TCDD level did not find any "consistent
patterns" to support an association. For example, among children of enlisted flying and
enlisted ground personnel, children of Ranch Hands with 2,3,7,8-TCDD levels <10 pg/g
lipids had higher rates (433 per 1,000 and 317 per 1,000) than children of controls with
background 2,3,7,8-TCDD levels < 10 pg/g lipids (229 per 1,000). However, rates of
children of enlisted ground personnel with 2,3,7,8-TCDD levels >33.3 pg/g lipids were not
significantly elevated. Again, these analyses were not based on 2,3,7,8-TCDD level at time
of conception. Moreover, if the higher 2,3,7,8-TCDD levels were related to early pregnancy
loss, these results would make more biological sense, as the abnormal conceptuses due to
2,3,7,8-TCDD exposure would have been lost before the pregnancy was recognized.
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There was also a significant association between 2,3,7,8-TCDD levels and neonatal
death (OR=5.5, 95% CI = 1.5-20.7). Insufficient numbers (N=13) precluded the calculation
of an adjusted odds ratio for this finding.
Finally, no association was detected between 2,3,7,8-TCDD level and either sperm
count or percentage of abnormal sperm. These analyses were based on semen samples that
had been collected in 1982.
7.13.15.6. Reproductive Hormones
In laboratory rats, 2,3,7,8-TCDD has been related to decreased testosterone levels
with evidence that dioxin decreases testosterone synthesis (Kleeman et al., 1990; Mebus et
al., 1987; Moore and Peterson, 1988; Moore et al., 1985).
A reported symptom of men who were exposed to 2,3,7,8-TCDD-contaminated
materials as a result of daily exposure and industrial accidents is reduced libido (Baader and
Bauer, 1951; Bauer et al., 1961; Suskind et al., 1953). Two independently conducted
studies of West Virginia TCP workers noted that exposed study subjects also reported this
condition approximately 50% more often than either the unexposed controls or individuals
without chloracne (Moses et al., 1984; Suskind and Hertzberg, 1984). Endocrine studies or
evaluations of conditions or situations that may lead to a reduction in libido were not
conducted.
In the NIOSH study of TCP production workers, questions regarding libido were not
asked; however, reproductive hormone levels were measured and related to serum 2,3,7,8-
TCDD levels. In linear regression analyses, serum 2,3,7,8-TCDD was positively and
significantly related to serum levels of luteinizing hormone (LH) and follicle-stimulating
hormone (FSH) and inversely related to total testosterone after adjustment for potential
confounders (p<0,05) (Egeland et al., 1994). The prevalence of abnormally low
testosterone was two to four times higher among workers with serum 2,3,7,8-TCDD levels
of 20-75 pg/g (OR=3.9, 95% CI = 1.3, 11.3), 76-243 pg/g (OR=2.7, 95% CI=0.9, 8.2), or
>244 pg/g (OR=2.1, 95% CI=0.8, 5.8) than among unexposed referents (4.8%) (mean
serum 2,3,7,8-TCDD = 7 pg/g). Workers in these same serum 2,3,7,8-TCDD quartiles had
a higher prevalence of abnormally high LH than workers with serum 2,3,7,8-TCDD levels of
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244 pg/g to 3,400 pg/g, but the differences between each serum 2,3,7,8-TCDD category and
referents were not significant.
The Ranch Hand veterans study provides the only other human data available that
evaluated the relationship between serum 2,3,7,8-TCDD and testosterone (Roegner et al.,
1991). Ranch Hand veterans with current serum dioxin levels exceeding 33.3 pg/g were
reported to have a lower mean total serum testosterone level (515.0 ng/dL) than the
nonexposed comparison group (525.2 ng/dL), but the difference was statistically
nonsignificant. No association was observed with FSH and LH. This same exposure group
had a nearly fourfold excess of unspecified testicular abnormalities.
Testosterone, FSH, and LH were also measured in U.S. Army veterans and non-
Vietnam veterans (Centers for Disease Control Vietnam Experience Study, 1988a). No
significant differences in hormone means were noted between the two groups. Additionally,
the proportions of values outside the reference range were also similar.
7.13.15.7. Comment
The human data offer some evidence of alterations in male reproductive hormone
levels associated with substantial occupational exposure to 2,3,7,8-TCDD. The results
support the animal literature in which dioxin-related effects have been observed on the
hypothalamic-pituitary-Leydig-cell axis and on testosterone synthesis.
7.14. NONCANCER EFFECTS OF INGESTION OF RICE OIL CONTAMINATED
WITH POLYCHLORINATED DIBENZOFURANS, QUATERPHENYLS, AND
BIPHENYLS IN JAPAN (YUSHO) AND TAIWAN (YU-CHENG)
This section briefly reviews the noncancer effects observed in Yusho (Japan) and Yu-
Cheng (Taiwan) victims, individuals exposed, by ingestion, to large concentrations of
compounds structurally related to dioxins, namely polychlorinated dibenzofurans,
quaterphenyls, and biphenyls. The history of each incident, the chemicals in question, and
levels of exposure are described in this chapter. In addition, other reviews have summarized
the numerous papers dedicated to Yusho and Yu-Cheng (Lii and Wong, 1984; Kuratsune,
1989; Rogan, 1989).
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Reports describing effects among individuals who ingested the contaminated rice oil
both in Taiwan and Japan are limited to acute rather than chronic effects. Studies have not
comprehensively evaluated long-term effects despite the facts that almost 25 years have
passed since the Yusho incident and 15 years since the Yu-Cheng incident and that serum
levels of some contaminants are available for both populations. Recent epidemiologic studies
have concentrated on the development of offspring of Yu-Cheng mothers. These children
were exposed in utero at the time the contaminants were ingested or were conceived after the
poisoning and were exposed to residual contaminants transplacentally or through breast milk
(Chen et al., 1992; Guo et al., 1992; Lai et al., 1993; Hsu et al., 1993; Guo et al., 1993).
7.14.1. Acute Effects in Adults and Children Directly Exposed to Contaminated Rice
Oil
In both groups, the most notable acute effects are dermatologic and neurologic signs
and symptoms of fatigue, headaches, and gastrointestinal distress (nausea, vomiting,
abdominal pain) (Kuratsune, 1989; Rogan, 1989).
7.14.1.1. Yusho
The initial recognition of Yusho occurred in 1968. As of 1983, a total of 2,060
individuals were identified as part of the Yusho population (Masuda et al., 1985). Five years
after exposure ended, the mean concentrations of PCBs in the adipose tissue, liver, and blood
of Yusho cases were 1.9 ppm, 0.08 ppm, and 6.7 ppb (Masuda et al., 1985), respectively,
which were about twice the levels in the control group. Adipose tissue levels of PCDFs
ranged from 6 to 13 ppb (Masuda et al., 1985). Sixteen years after exposure, mean PCQ
level in adipose tissue of Yusho cases was 207 ppb, approximately 100 times the level in
Japanese controls (Kashimoto et al., 1985).
In addition to the major health effects, other possible outcomes were examined.
Effects observed shortly after exposure included elevated triglyceride levels and effects on
female reproductive hormones noted by changes in menstrual and basal body temperature
patterns and lowered excretion of estrogens and pregnanediol in exposed women (Kuratsune,
1989). However, fertility and other measures of reproductive function were not evaluated.
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Evidence of chronic bronchitis and respiratory infections still remained 14 years after
exposure ended (Nakanishi et al., 1985). However, more than 10 years postexposure, PCB
levels were not related to levels of serum triiodothyronine (T3), thyroxine (T4), and
thyroxine-binding globulin (TBG) (Murai et al., 1987). Although the liver is the suspected
target organ for halogenated hydrocarbons and marked proliferation in the endoplasmic
reticulum was observed, clinical evidence of liver damage, such as alterations in liver
enzymes or liver disease, was not observed (Kuratsune, 1989).
Dermatologic effects were the most evident signs, characterized by hyperpigmentation
of the nails, gingivae, and face and by nail deformities, horny plugs, comedones, acneform
eruptions, cysts, and other abnormal keratotic changes (Urabe and Asahi, 1985). Acneform
eruptions were observed on the face, cheeks, auricles, retroauricular areas, inguinal regions,
and external genitalia (Urabe and Asahi, 1985). More than 80% of Yusho cases experienced
one or more dermatologic effects (Kuratsune, 1989), which diminished in severity over time
(Urabe and Asahi, 1985).
Ophthalmologic effects were characterized by swelling and hypersecretion of the
meibomian glands and pigmentary changes of the conjunctiva (Kuratsune et al., 1972). More
than 80% of Yusho cases exhibited ocular changes, which, in some cases, appeared to persist
15 years after exposure ended (Kuratsune, 1989).
Thirty percent of the cases reported having at least one symptom consistent with
neurologic involvement, such as limb parasthesia and spasms, weakness, headaches, and
fatigue (Kuratsune, 1972). As summarized by Kuratsune (1989), Kuriowa et al. (1969)
found mostly sensory deficits, identified through slowed nerve conduction velocities in 23
cases. Follow-up of these cases indicated that the neurologic symptoms disappeared over
time; however, conduction velocities were not repeated.
A number of studies examined the immune status of Yusho cases (Kuratsune, 1989).
Significant decreases in mean IgA and IgM and increases in IgG were noted in 28 cases
tested in 1970 (p<0.05) (Nakanishi et al., 1985). Within 2 years, means levels of all three
immunoglobulins returned to normal. Small increases in the percentage of CD4 cells, small
decreases in the percentage of CDS cells, and enhanced lymphocyte stimulation were also
noted in Yusho cases (Nakanishi et al., 1985).
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Studies of offspring of Yusho cases have been limited to descriptions of effects on
newborns exposed in utero. An early description of 13 children born to exposed mothers
noted two stillborn infants, one of whom was diffusely and deeply hyperpigmented (Rogan,
1982). Neonates described in other reports were darkly pigmented and had marked
secretions of the conjunctival palpebra, gingival hyperplasia, hyperkeratosis, calcification of
the skull, low birth weight, and natal teeth (Yamashita and Hayashi, 1985). The abnormal
pigmentation disappeared after 2 to 5 months. No other physical abnormalities (neurologic,
cardiovascular, or malformations) were identified.
7.14.1.2. Yu-Cheng
The initial recognition of Yu-Cheng occurred in 1979. As of 1983, approximately
2,000 individuals were found to have been exposed to the contaminated rice oil. Within the
first year of exposure, mean serum PCB, PCDF, and PCQ levels for 15 cases were 60 ppm
(range 4-188 ppm), 0.14 ppb (range <0.005-0.27 ppb), and 19.3 ppb (range 0.9-63.8),
respectively (Kashimoto et al., 1985). Analysis of PCB levels in 1980-1981 in 165 cases
(mean 38 ppb, range 10-720) (Rogan, 1989) and in 1985 in 32 cases (mean 15.4 ppb, range
0.6-86.8) (Lundgren et al., 1988) suggested that some PCBs were being eliminated. It is not
clear from the reports if the samples were drawn from distinctly different individuals or
included some of the same individuals.
The ophthalmologic and dermatologic changes observed in Yu-Cheng cases were very
similar in character and anatomical distribution to those noted in Yusho cases (Lii and Wu,
1985). In 89 cases followed for up to 17 months, dermatologic conditions of 38% of the
cases improved, 54% remained the same, and 7% showed deterioration of their conditions
(Lii and Wong, 1984).
Like Yusho cases, Yu-Cheng cases examined within 2 years of exposure for nerve
function exhibited slowing of sensory nerve conduction. They also exhibited motor nerve
slowing and mixed deficits (Chen et al., 1981, 1983, 1985; Chia and Chu, 1984). Twenty
percent of a population of 27 individuals also had abnormal EEGs (Chia and Chu, 1984).
However, the authors suggest that any correlation between PCB exposure and the abnormal
EEGs may be spurious due to low PCB levels in the cerebrospinal fluid (0.5-2.3 ppb)
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(measured in four subjects), despite much higher blood PCB levels of 48-64 ppb. A sample
of 28 individuals with peripheral neuropathy in 1980 was reexamined in 1982 and was found
to have normal EEGs and some recovery of sensory and motor nerve conduction velocity
(Chia and Chu, 1985).
In 1981, immunologic function was assessed on different subsets of Yu-Cheng cases
and summarized by Lii and Wong (1984). In 30 cases compared with unexposed controls,
both IgA and IgM were significantly decreased, while IgG did not differ from controls. In
this same group, percentages of active T cells and T cells were significantly increased
(p<0.05), while total lymphocyte count and percentage of B cells were unchanged.
Significant increases in helper T cells (T4) but not suppressor T cells (T8) were also
observed. In another group of cases, response to lymphocyte-stimulating mitogens was
mixed and the findings unclear. In 143 cases, reaction to streptococci antigen appeared to be
significantly (p<0.05) depressed relative to controls.
Alterations in porphyrin levels and liver enzymes have been identified as acute
reactions to exposure to halogenated polycyclic hydrocarbons, including PCBs. Porphyrin
levels were measured in two exposed groups (Chang et al., 1980; Gladen et al., 1988). In
1980, statistically significant elevations in 24-hour urinary excretion of uroporphyrin
(exposed = 41.23 /*g ± 24.56; unexposed = 13.57 /*g ± 11.76, p<0.01) and a-
aminolevulinic acid (exposed = 1.002 mg + 0.600; unexposed = 0.715 ± 0.337, p<0.05)
were noted among 69 subjects (Chang et al., 1980). Coproporphyrin and porphobilinogen
levels were increased in the exposed group but were not significantly elevated. The second
study group was composed of 75 children born between June 1978 and March 1985 to
mothers who ingested contaminated rice oil (Gladen et al., 1988). Spot urines were collected
in 1985. Mean total porphyrin (exposed = 95.2 ng/L; unexposed = 80.7 /xg/L) and
coproporphyrin (exposed = 72.4 /*/L; unexposed = 59.8 /x/L) excretion was elevated in the
exposed, possibly due to extremely high levels (>200 /xg/L) in eight exposed children and
two controls (Rogan et al., 1988). However, no porphyria cutanea tarda, a severe form of
porphyria, was observed in either group of children. Moderate, but statistically significant,
increases were observed in AST and ALT levels in 23 cases tested 1 year after exposure (Lii
and Wong, 1984). LDH and bilirubin levels were not significantly elevated. As in Yusho
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cases, triglyceride levels were significantly increased by approximately twice the level in
unexposed controls.
7.14.1.3. Effects Observed in Offspring of Yu-Cheng Cases
A concerted effort has been made to evaluate the overall development of children
exposed prenatally to contaminated rice oil ingested by their mothers. Two studies are
currently being conducted: one to evaluate children and controls born between 1978 and
1985, and the second, to evaluate children born in 1985 and later. The purpose of the two
studies is to determine if there are differences among children exposed in utero by the
mothers' ingestion of the contaminated rice oil or shortly afterward and children conceived 6
or more years after direct exposure ended (Chen et al., 1992; Chen et al., 1993).
In terms of musculoskeletal development, several studies have documented delays and
abnormalities (Yu et al., 1991; Rogan, 1989; Guo et al., 1993; Chen et al., 1993). In one
of the first studies conducted in 1985, Rogan and colleagues (1988) examined 117 children
born since the mothers' exposure in 1979 and 107 unexposed controls. In this study, babies
of exposed mothers were consistently smaller and shorter at birth than controls and had
similar characteristics: natal teeth, neonatal conjunctivitis, and pigmentation. Exposed
mothers reported a mean birth weight 479 g lower than that reported by control mothers; no
validation of these reports using medical records was undertaken. As older children, they
exhibited a variety of signs and symptoms: fragile chipped teeth and gum hypertrophy,
pigmented and deformed fingernails and toenails, and abnormal lung auscultation. In this
same study, neurologic developmental assessments were also conducted to evaluate
development (Yu et al., 1991). Forty-nine percent of Yu-Cheng children compared with
22% of controls were developmentally delayed in 32 of 33 developmental milestones, 12%
had clinical evidence of developmental or psychomotor delays compared with 2% of controls,
and 7% of Yu-Cheng children versus 3% of controls had speech problems. These delays
were noted at all ages and persisted over 2 years of testing. Delay was greater in children of
smaller size and in children who had exhibited neonatal symptoms of intoxication.
In 1991, the musculoskeletal development of 56 Yu-Cheng children (age range 6-10
years) and their matched controls was again assessed. Only children born first after the
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mother's exposure were shorter in stature and had decreased lean muscle mass and soft tissue
content (Quo et al., 1992). A subsequent examination of 110 Yu-Cheng children and 108
controls, ages 8 through 14 years, found Yu-Cheng girls to be significantly shorter than
controls, matched to the exposed children by age, sex, maternal age, parents' combined
educational level, occupation, and neighborhood (Guo et al., 1993). As measured by the
Tanner scale, sexual maturation was not slower in Yu-Cheng boys or girls. However, Yu-
Cheng boys aged 11-14 had significantly shorter penis length, but testicular and scrotal
development did not differ from the controls. Penile length was not related to sexual
development as measured by the Tanner scale.
With a validated and standard battery of tests, cognitive and behavioral development
of Yu-Cheng offspring was studied yearly from 1985 through 1991. Throughout the testing
period, Yu-Cheng children scored consistently lower in the Standford Binet IQ (SB-IQ) and
4-5 points lower than controls (with the same matching criteria as the above study) in three
subscales of the Wechsler Intelligence Scale for Children, Revised (WISC-R): verbal IQ
(VIQ), performance IQ (PIQ), and full-scale IQ (FIQ) (Chen et al., 1992; Lai et al., 1993).
Yu-Cheng children are also reported to exhibit more health, habit, and behavior problems as
reported by parents responding to the Rutter's scale and to manifest higher activity based on
teachers' responses to the Teacher's Activity Check List (Hsu et al., 1993).
Analysis of physical and cognitive development began in October 1991 of 104
children whose mothers were exposed and 109 children whose fathers but not mothers were
exposed and of three matched controls born after 1985 (Guo et al., 1993). Like children
born before 1985, the later-born children were shorter in stature and lower in weight than
controls, although the authors indicate that the differences have disappeared. Yu-Cheng
children are reported to have higher activity levels but do not have temperament, physical,
habit, or behavioral problems. Overall, scores on all tests in paternally exposed children
were similar to those of the controls. However, maternally exposed children scored lower on
the SB-IQ and on all subscales of the WISC-R.
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7.14.1.4. Comment
Data from Yusho and Yu-Cheng strongly implicate direct ingestion of contaminated
rice oil with numerous acute effects of the skin and peripheral and central nervous systems.
The data on immunologic function suggest possible effects, but the numbers of subjects in the
various studies were too small to determine an exposure-response relationship. Similarly,
data on elevated triglyceride and liver enzyme levels are from a small number of cases and,
therefore, the relationship between exposure and the effect is unclear. Furthermore, there
are few data to evaluate the long-term effects of these very high exposures. Because many
of the 4,000 individuals who were exposed in these two episodes were children, longitudinal
studies would be invaluable in assessing the long-term health effects of these exposures.
One difficulty in evaluating the various reports relating to Yusho and Yu-Cheng is the
inability to determine if the effects are generalizable to the entire exposed population. It is
the impression of this reviewer that many of the cases reported in the literature were those
that exhibited the severest signs and symptoms and probably had the highest body burden of
contaminants. Some reports included small numbers of cases and controls, relative to the
size of the exposed population, and others had no controls. Finally, while much work has
gone into determining the severe acute effect, it would be interesting to know what chronic,
age, and gender-specific effects are now being exhibited in the approximately 4,000
individuals directly exposed to the contaminated rice oil.
Another difficulty is the presence of several chlorinated hydrocarbons in the
contaminated oil, which results in uncertainty as to which contaminant or combination of
contaminants is responsible for the noted effects. The data on the offspring of exposed Yu-
Cheng mothers and fathers are fascinating and disturbing. It is apparent that parental
exposure-more specifically, maternal exposure~to PCBs, PCDFs, and PCQs is directly
linked to in utero exposure of the fetus, affecting cognitive, selective musculoskeletal, and
possibly, sexual development of the offspring. The ongoing assessment of development in
the Yu-Cheng children is imperative to determine the long-term consequences of these
exposures on the children's future quality of life and to emphasize the lesson that parental
exposures greatly affect the fetus. Prospective studies of fertility and reproductive
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experience in these offspring would provide insight into possible intergenerational effects of
exposure to these compounds.
7.15. SUMMARY
The data presented in this chapter describe nonmalignant effects in epidemiologic
studies of populations with the potential for exposure to chemicals contaminated with 2,3,7,8-
TCDD. The purpose of this review is to highlight the salient results of the studies and to
assess whether the observed effect was related to exposure to 2,3,7,8-TCDD.
In summary, based on the results of two or more studies, recent evidence suggests
that chloracne, elevated GGT levels, an increased risk of diabetes, and altered reproductive
hormone levels (luteinizing hormone, follicle-stimulating hormone, and testosterone) appear
to be long-term consequences of exposure to 2,3,7,8-TCDD (Table 7-48). In contrast,
multiple studies show possible acute effects but few chronic exposure-related effects for
dermatologic end points other than chloracne, such as eye lid cyst, hypertrichosis,
hyperpigmentation, actinic keratosis, and Peyronie's disease; for liver diseases such as
cirrhosis, liver enlargement, and hepatic enzyme levels (LDH, AST, ALT, and D-glucaric
acid) other than GGT; and for porphyrias and renal, neurologic, and pulmonary disorders.
Circulatory and heart diseases, reproductive outcomes, immunologic disorders, lipid levels,
and thyroid function require further study before their respective relationships to 2,3,7,8-
TCDD can be more definitively assessed. Studies are under way in several locations that
may yield useful information regarding these outcomes.
In the best of circumstances when reviewing studies, it would be ideal if all studies
examined the same end points in the same manner, had sufficient statistical power to detect
truly positive findings, had good estimates of extent of exposure, and had consistent
exposure-response relationships. In the absence of ideal situations, epidemiologists examine
the evidence of studies using "six tenets of judgment" (Hatch and Stein, 1986; Hill, 1965) to
assess the collective wisdom of the study results. These tenets are: temporality (sequence of
events), degree of exposure, strength, consistency and specificity of association, and biologic
plausibility.
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Table 7-48. Summary of Effects Observed in Humans
System
Dermatologic
Liver
Liver enzymes
Urinary porphyrins
Lipids
Cholesterol
Triglycerides
Other GI
Thyroid
Diabetes
Immune
Neuro
Circulatory
Pulmonary
Renal
Reproductive
Reproductive hormones
Chromosome
Acute
conjunctivitis
red and irritated eyes
blepharitis
temporary enlargement
t GOT
t AST
f ALT
t D-glucaric acid excretion
+ /-"
PCT
uroporphyrin
urobilinogen
coproporphyrin
t
t
t
+\- RUQC pain
loss of appetite
nausea
t T4
t T4/TBG in some studies
no data
no data
+ /- i libido
t irritability
t nervousness
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In evaluating many of the studies that examined the relationship between serum
2,3,7,8-TCDD, there are several common threads that bear noting. They will be discussed
first to avoid repetition throughout the summary.
In terms of temporality, all studies reviewed in this chapter were conducted after the
presumed exposure occurred. Some of the studies were completed shortly after the
exposure, as in Seveso (Mocarelli et al., 1986; Filippini et al., 1981); others were conducted
many years after the groups' last exposure to evaluate the more chronic health outcome
(Suskind and Hertzberg, 1984; Moses et al., 1984; Bond et al., 1987; Roegner et al., 1991;
Assennato et al., 1989; Calvert et al., 1991, 1992, 1993; Alderfer et al., 1992; Sweeney et
al., 1990; Egeland et al., 1994; Sweeney et al., 1992; Webb et al., 1989; Ott et al., 1993b).
One dilemma in assessing the effect of past exposures is ascertaining whether an effect
observed many years postexposure is due to the exposure itself or to an exposure or event
that occurred during the intervening period. Another problem is determining what, in the
analysis, the investigator considered the most important of the possible confounding
exposures. Finally, restricting examination of events to those that occurred after the
exposure does not in and of itself satisfy this time order criterion. Several factors must be
considered, such as the half-life of the contaminant in the body and the concentration at the
time of the event. Consistency in the results of similarly designed studies of 2,3,7,8-TCDD-
exposed populations should help strengthen the conclusion of an effect or no effect.
Determination of the extent of exposure throughout the studies was varied. When the
risk of disease increases with the dose or gradient of exposure, the evidence for causation is
strengthened. It should be emphasized that there are many possible dose-response patterns,
which may result in different threshold levels for different end points. Due to the exposure
misclassification bias present in most of the dioxin research, with the exception of a few
studies, it is not valid to attempt to determine dose-response relationships. To summarize,
four studies evaluated the relationship between nonmalignant effects and body levels of
2,3,7,8-TCDD: the Ranch Hand study of U.S. Air Force personnel (Roegner et al., 1991);
the study of 50 Missouri residents (Webb et al., 1989); the evaluation of the BASF accident
cohort (Ott et al., 1993b); and the NIOSH study of 281 TCP production workers (Sweeney et
al., 1989, 1990, 1993; Calvert et al., 1991, 1992, 1993; Alderfer et al., 1992; Egeland et
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al., 1994). In 1988, the workers in the NIOSH study had the highest serum levels (mean =
220 pg/g) in these four studies.
In two populations, Seveso and U.S. Army veterans (Mocarelli et al., 1991; Centers
for Disease Control Veterans Health Studies, 1988), serum 2,3,7,8-TCDD levels were
measured, but the levels were not used to examine dose-response relationships. In the
Veterans Health Studies, more than 99% of the serum 2,3,7,8-TCDD levels of the sample of
both Vietnam and non-Vietnam veterans were at the background level (4 pg/g). Therefore,
comparisons were made between the two groups as a whole.
Serum levels of Seveso residents were obtained for a small proportion (N=20) of the
total number of residents of Zone A within 1 year of the reactor release (Mocarelli et al.,
1991). The data suggest that the levels may be related to a number of factors, including age
(younger children were outside at the time of the release), whether the resident was inside or
outside, ingestion of local produce, number of days of residence in the area after the release,
to name a few. The data suggest that the potential for substantial exposure was high for
individuals residing in the area. The range of levels in the 20 Zone A residents was 820
pg/g to 56,000 pg/g (median = 7,400 pg/g).
The majority of the remaining studies examined the differences between individuals
identified as exposed or unexposed or with or without chloracne. Most of these studies did
not evaluate other parameters that might explain differences in effects between exposed and
unexposed, for example, the length of exposure. However, one study assessed dose-response
relationships based on a statistical algorithm of intensity and highest dose of TCDD exposure
(Bond et al., 1989).
In terms of the magnitude (or strength) of the association, this criterion refers to the
degree to which the measure of association (e.g., odds ratio or relative risk) exceeds the null
value of 1. The stronger the association between exposure and effect, the more convincing is
the argument for causation. There is no definitive cutpoint to numerically define a
meaningful measure of association. Other factors, such as the prevalence of the exposure in
the population, affect the significance of the measure. Because so many adverse health
conditions are multifactorial in etiology, a general rule of thumb is that a relative risk less
than 2 renders a cause-effect relationship less likely.
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A critical element that should always accompany the effect measure is a confidence
interval. Placement of an interval around the measure enables quantitation of the result for a
more meaningful interpretation. An odds ratio of 30 is quite impressive, but if the 95%
confidence interval is 0.9-200, the magnitude of the association is less impressive.
If an association between a factor and a disease is demonstrated across a variety of
studies employing different designs and different populations (consistency), then the argument
for causation is strengthened. Replication of an association under different conditions
decreases the likelihood that confounding is responsible for the observed association.
Consistency is a powerful criterion for causation but only when "the variables under test
(exposure, outcomes) are similar enough" to justify the comparison of the various studies'
findings (Hoffman et al., 1986).
It should also be determined a priori that each study included in the critical evaluation
process is in adherence to basic epidemiologic principles governing study design and
analysis. Deficient studies with suspect results should be excluded. While this is not to
imply that such studies have no worth, as invaluable information has often been derived from
these studies that improve on subsequent examinations of the issue, they have no place in the
evaluation process. Unfortunately, in studies of 2,3,7,8-TCDD and effects in humans, the
probability of exposure misclassification forces exclusion of much of the research to date.
Specificity refers to the uniqueness of the association between a factor and an
outcome. If the relationship were absolute, then only factor X would be related to only
effect Y. It is indeed rare to encounter this type of association, which renders this criterion
generally less useful in the evaluation process.
Finally, according to the criterion of biological plausibility, the observed association
between exposure and effect should be consistent with existing theory and information from
other scientific disciplines. Certainly one would feel more secure in the causation debate if
the biological basis for an observed association can be explained. However, biological
implausibility may simply reflect gaps in existing scientific knowledge that could explain the
relationship.
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7.15.1. Effects Having a Positive Relationship With Exposure to 2,3,7,8-TCDD
The following section describes those end points for which there is good evidence
from two or more studies suggesting an effect of exposure to 2,3,7,8-TCDD.
7.15.1.1. Chloracne
7.15.1.1.1. Temporality. Chloracne is one of the best known of the medical consequences
of exposure to 2,3,7,8-TCDD-contaminated substances. In general, it has been observed in
most incidents where substantial exposure has occurred, particularly among TCP production
workers (Goldman, 1972; May, 1973; Bleiberg et al., 1964; Bond et al., 1987; Suskind and
Hertzberg, 1984; Moses et al., 1984; Zober et al., 1990) and Seveso residents (Reggiani,
1978; Caramaschi et al., 1981; Ideo et al., 1985; Mocarelli et al., 1986; Assennato et al.,
1989). As previously stated, chloracne appears within several weeks to months from the
time of exposure, often resolving after discontinuation of exposure (Moses et al., 1984;
Suskind and Hertzberg, 1984), although for some it may remain for extended periods after
exposure ended (Moses et al., 1984).
7.15.1.1.2. Degree of exposure, consistency of the association. The amount of exposure
necessary for development of chloracne has not been resolved, but studies suggest that high
exposure (both high acute and long-term exposure) to 2,3,7,8-TCDD increases the likelihood
of chloracne, as evidenced by chloracne in TCP production workers and Seveso residents
who have documented high serum 2,3,7,8-TCDD levels (Beck et al., 1989; Fingerhut et al.,
1991a; Mocarelli et al., 1991; Neuberger et al., 1991) or in individuals who have a work
history with long duration of exposure to 2,3,7,8-TCDD-contaminated chemicals (Bond et
al., 1989). The absence of substantial chloracne in U.S. Army Vietnam veterans whose
mean serum 2,3,7,8-TCDD levels were at background (4 pg/g) (Centers for Disease Control
Vietnam Experience Study, 1988d) and U.S. Air Force Ranch Hands whose serum 2,3,7,8-
TCDD levels fell intermediate to those of workers and Army Vietnam veterans (Roegner et
al., 1991) suggests that there is a higher incidence of the disorder among those with higher
serum 2,3,7,8-TCDD levels.
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7.15.1.1.3. Strength of the association. In earlier studies, chloracne was considered to be a
"hallmark of dioxin intoxication" (Suskind, 1985). However, only in two studies were risk
estimates calculated for chloracne. Both were studies of different cohorts of TCP production
workers (Suskind and Hertzberg, 1984; Bond et al., 1989); one group was employed in a
West Virginia plant, the other in a plant in Michigan. Of the 203 West Virginia workers,
52.7% (p<0.001) were found to have clinical evidence of chloracne, and 86.3% reported a
history of chloracne (p<0.001) (Suskind and Hertzberg, 1984). None of the unexposed
workers had clinical evidence or reported a history of chloracne. Among the Michigan
workers, the relative risk for cases of chloracne was highest for individuals with the longest
duration of exposure (>60 months; RR=3.5, 95% CI=2.3-5.1), those with the highest
cumulative dose of TCDD (based on duration of assignment across and within 2,3,7,8-
TCDD-contaminated areas in the plant) (RR=8.0, 95% CI=4.2-15.3), and those with the
highest intensity of 2,3,7,8-TCDD exposure (RR=71.5, 95% CI=32.1-159.2) (Bond et al.,
1989).
7.15.1.1.4. Specificity of the association. Chloracne is associated with exposure to other
polyhalogenated chemicals, including dibenzofurans, PCBs, naphthalenes, and others (Taylor,
1979). The likelihood of exposure to other polyhalogenated chemicals in the populations
studied is probably low, particularly among the Seveso children, whose exposure was to TCP
reactant effluents that were primarily contaminated with 2,3,7,8-TCDD. The issue is more
relevant in chemical workers, who by virtue of their occupation, have the potential for
exposure to other chemicals. Yet, much of the documented chloracne appeared shortly after
TCP reactor releases (Ashe and Suskind, 1950; Goldman, 1972; May, 1973) or during TCP
or 2,4,5-T production (Bond et al., 1989), suggesting that 2,3,7,8-TCDD was the
chloranegenic agent.
7.15.1.1.5. Biologic plausibility. Animal studies have been effective in describing the
relationship between 2,3,7,8-TCDD and chloracne, particularly in rhesus monkeys (McNulty,
1977; Allen et al., 1977; McConnell et al., 1978). Subsequent to exposure to 2,3,7,8-
TCDD, monkeys developed chloracne and swelling of the meibomian gland, a modified
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sebaceous gland. The histologic changes in the meibomian gland are physiologically similar
to those observed in human chloracne (Dunagin, 1984).
In summary, the evidence provided by the various studies convincingly states what is
already presumed, that chloracne is a common sequela of exposure to 2,3,7,8-TCDD. More
information is needed to determine the level and frequency of 2,3,7,8-TCDD exposure
needed to cause chloracne and whether personal susceptibility plays a role in the etiology.
Finally, it is important to recall that the absence of chloracne does not imply lack of
exposure (Mocarelli et al., 1991).
7.15.1.2. Gamma Glutamyl Trans/erase (GGT) Levels
7.15.1.2.1. Temporality, degree of exposure, and strength and consistency of association.
There appears to be a consistent pattern of increased GGT levels among individuals exposed
to 2,3,7,8-TCDD-contaminated chemicals. Elevated levels of serum GGT have been
observed within a year after exposure in Seveso children (Caramaschi et al., 1981; Mocarelli
et al., 1986) and 10 or more years after cessation of exposure among TCP and 2,4,5-T
production workers (May, 1982; Martin, 1984; Moses et al., 1984; Calvert et al., 1992) and
among Ranch Hands (Roegner et al., 1991). All of these groups had a high likelihood of
substantial exposure to 2,3,7,8-TCDD. In addition, for those studies that evaluated dose-
response relationships with 2,3,7,8-TCDD levels, the effect was observed only at the highest
levels or categories of 2,3,7,8-TCDD.
In contrast, although background levels of serum 2,3,7,8-TCDD suggested minimal
exposure to Army Vietnam veterans, GGT was increased, at borderline significance, among
Vietnam veterans compared to non-Vietnam veterans (Centers for Disease Control Vietnam
Experience Study, 1988a). In addition, despite the increases observed in some occupational
cohorts, other studies of TCP production workers from West Virginia or Missouri residents
measured but did not report elevations in GGT levels (Suskind and Hertzberg, 1984; Webb et
al., 1989).
7.15.1.2.2. Specificity. In clinical practice, GGT is often measured because it is elevated in
almost all hepatobiliary diseases and is used as a marker for alcoholic intake (Guzelian,
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1985). In individuals with hepatobiliary disease, elevations in GOT are usually accompanied
by increases in other hepatic enzymes, e.g., AST and ALT, and metabolites, e.g., uro- and
coproporphyrins. Significant increases in hepatic enzymes other than GOT and metabolic
products were not observed in individuals whose GGT levels were elevated 10 or more years
after exposure ended, suggesting that the effect may be GGT specific. However, in the
Seveso male children and those with chloracne (both sexes), ALT was significantly increased
concomitantly with GGT within 1 year of the reactor release (Mocarelli et al., 1986;
Caramaschi et al., 1981). In a longitudinal analysis, both enzymes returned to normal levels
within 5 years after exposure (Mocarelli et al., 1986; Assennato et al., 1989).
These data suggest that in the absence of increases in other hepatic enzymes,
elevations in GGT are associated with exposure to 2,3,7,8-TCDD, particularly among
individuals who were exposed to high 2,3,7,8-TCDD levels. The fact that investigators
observed a decline in enzyme levels in Seveso children but a continued elevation in TCP
workers may reflect differences in how exposure occurred (i.e., acute but high doses in
Seveso versus continuous or frequent long-term, medium to high doses in TCP workers) or
differences in the metabolism of the maturing versus mature system, or a bit of both.
7.15.1.2.3. Biologic plausibility. The animal data with respect to 2,3,7,8-TCDD-related
effects on GGT are sparse. Statistically significant changes in hepatic enzyme levels,
particularly AST, ALT, and ALK, have been observed after exposure to 2,3,7,8-TCDD in
rats and hamsters (Gasiewicz et al., 1980; Kociba et al., 1978; Olson et al., 1980). Only
one study evaluated GGT levels (Kociba et al., 1978). Moderate but statistically
nonsignificant increases were noted in rats fed 0.10 ^g/kg 2,3,7,8-TCDD daily for 2 years,
and no increases were observed in control animals.
Among humans, increased levels of GGT may suggest activity such as cholestases,
liver regeneration, or drug or xenobiotic metabolism. In human adults, 2,3,7,8-TCDD is
stored in the adipose tissue and has a half-life of approximately 7 years (Pirkle et al., 1989).
Continued GGT activity in adults with serum 2,3,7,8-TCDD levels many times over
background levels may reflect continuous, low-level metabolism of 2,3,7,8-TCDD.
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In summary, GGT is the only hepatic enzyme found in a number of studies to be
chronically elevated in adults exposed to high levels of 2,3,7,8-TCDD. The consistency of
the findings in a number of studies suggests that the finding may reflect a true effect of
exposure but for which the clinical significance is unclear. Long-term, pathologic
consequences of elevated GGT have not been illustrated by excess mortality from liver
disorders or cancer or in excess morbidity in the available cross-sectional studies.
It must be recognized that the absence of an effect in a cross-sectional study, for
example, liver enzymes, does not obviate the possibility that the enzyme levels may have
been increased concurrent to the exposure but declined after cessation. The apparently
transient elevations in ALT levels among the Seveso children suggest that hepatic enzyme
levels other than GGT may react in this manner to 2,3,7,8-TCDD exposure.
7.15.1.3. Diabetes and Fasting Serum Glucose Levels
7.15.1.3.1. Strength and consistency of association. The epidemiologic evidence for an
increased risk of diabetes and for an elevated prevalence of abnormal fasting serum glucose
levels is based on two cross-sectional medical studies conducted at least 15 years and up to
37 years after last exposure to 2,3,7,8-TCDD-contaminated chemicals (Sweeney et al., 1992;
Wolfe et al., 1992a) and supporting evidence from a third study of the BASF accident cohort
(Ott et al., 1993b). All three studies found that individuals at the highest serum 2,3,7,8-
TCDD levels had a slight but statistically significant or borderline significant increased risk
for developing diabetes or having an elevated fasting serum glucose. Further study is needed
to corroborate the findings of these three cross-sectional studies.
7.15.1.3.2. Specificity. Diabetes mellitus is a heterogeneous disorder that is a consequence
of alterations in the number or function of pancreatic beta cells responsible for insulin
secretion and carbohydrate metabolism. Depending on its type, diabetes has been attributed
to endogenous factors such as genetic predisposition, to autoimmune processes, and to
exogenous factors such as viral infections (Yoon et al., 1987) and chemical exposures,
notably a rat poison (Miller et al., 1978) and some medications (Wilson and LeDoux, 1989),
environmental toxins (Diabetes Epidemiology Research International, 1987), age, obesity,
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reduced physical exercise, diet, socioeconomic status, increased insulin resistance by the beta
cells, and possibly parity (Pareschi and Tomasi, 1987).
Both the NIOSH and Ranch Hand studies used strict criteria to identify cases. In the
NIOSH study, the relationship between exposure to 2,3,7,8-TCDD and possible alterations in
glucose metabolism was assessed using two outcome measures: case definition of diabetes
and fasting serum glucose levels. Participants met the case definition of diabetes if their
fasting serum glucose level was 7.8 mmol/L or greater on two consecutive occasions
(National Diabetes Data Group, 1979) or if they reported a positive history of physician-
diagnosed diabetes after the date of first employment at the plant. In the Ranch Hand study,
diabetic status was assessed by measuring fasting serum glucose and 2-hour postprandial
glucose and using a case definition of diabetes. Diabetes was defined as having a verified
history of diabetes or an oral glucose tolerance test of > 11.1 mmol/L (200 mg/dL) (Roegner
et al., 1991). Both studies used acceptable definitions for epidemiologic studies of diabetes.
In the NIOSH study, the definition may exclude some individuals who would have otherwise
been classified as diabetic with the glucose tolerance test. However, there should be no
differential bias in identifying cases in workers or the unexposed referents.
7.15.1.3.3. Biologic plausibility. The effects of 2,3,7,8-TCDD on glucose metabolism have
been evaluated only in a few laboratory studies (Zinkl et al., 1973; McConnell et al., 1978a;
Gasiewicz et al., 1980; Schiller et al., 1986; Ebner et al., 1988; Gorski et al., 1990).
Although these studies suggest that 2,3,7,8-TCDD may alter glucose metabolism, for the
most part the animal studies do not corroborate the direction of the findings of two cross-
sectional medical studies (Wolfe et al., 1992a; Sweeney et al., 1993). Long-term feeding
studies to evaluate the relationship between glucose levels, the development of diabetes, and
2,3,7,8-TCDD dose would be helpful in assessing the effect of exposure on the physiologic
integrity of the islet cells. In addition, such studies may identify other factors that may affect
either directly or indirectly the function of the islet cells.
In summary, both the NIOSH and Ranch Hand studies show that highly 2,3,7,8-
TCDD-exposed individuals may be at a slightly greater risk for developing diabetes and
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experience a greater prevalence of elevated fasting glucose levels, which may be a precursor
of conversion to a diabetic state. However, in the NIOSH study, the traditional risk factors
for diabetes-age, body mass index or weight, and family history of diabetes-appear
substantially more influential than 2,3,7,8-TCDD in the development of diabetes. These
factors could not be evaluated in the data presented in the publicly available material.
7.15.1.4. Reproductive Hormones
7.15.1.4.1. Strength and consistency of association. Levels of reproductive hormones have
been measured with respect to exposure to 2,3,7,8-TCDD in three cross-sectional medical
studies. Testosterone, LH, and FSH were measured in TCP and 2,4,5-T production workers
(Egeland et al., 1994) and in Army Vietnam veterans (Centers for Disease Control Vietnam
Experience Study, 1988d), and testosterone and FSH in Ranch Hands (Roegner et al., 1991).
The risk of abnormally low testosterone was two to four times higher in exposed workers
with serum 2,3,7,8-TCDD levels above 20 pg/g than in unexposed referents (Egeland et al.,
1994). The risk in the Ranch Hands, who had lower exposures than the production workers,
was 1.3 times (OR = 1.3) that of the comparison group (Roegner et al., 1991). No
significant associations were found between Vietnam experience and altered reproductive
hormone levels (Centers for Disease Control Vietnam Experience Study, 1988d). Only the
NIOSH study found an association between serum 2,3,7,8-TCDD level and increases in
serum LH.
7.15.1.4.2. Specificity. The NIOSH study excluded from analysis participants who had
conditions that might have influenced gonadotropin and/or testosterone levels: history of
prostate cancer, thyroid or other hormone usage, or liver cirrhosis. Similarly, in the Ranch
Hand study, individuals with orchiectomies or who were taking testosterone medication were
excluded from the analysis of testosterone; no participants were excluded from the analyses
of FSH. The CDC study of Vietnam veterans did not describe the exclusions.
7.15.1.4.3. Biologic plausibility. In rats, 2,3,7,8-TCDD has been shown to decrease
testosterone levels (Moore et al., 1985; Moore and Peterson, 1988; Mebus et al., 1987)
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through a decrease in testosterone synthesis (Kleeman et al., 1990) or by decreasing the
production of pregnenolone from cholesterol (Ruangwies et al., 1991). In addition, 2,3,7,8-
TCDD has been shown in rats to reduce the responsiveness of the pituitary to testosterone
(Bookstaff et al., 1990a) and of the Leydig cells to LH stimulation (Moore et al., 1991).
The findings of the NIOSH and Ranch Hand studies are plausible given the
pharmacological and toxicological properties of 2,3,7,8-TCDD. A mechanism responsible
for the effects may involve the ability of 2,3,7,8-TCDD to influence hormone receptors.
The aromatic hydrocarbon (Ah) receptor to which 2,3,7,8-TCDD binds resembles steroid
hormone receptors in both structure and mode of action. Studies suggest that 2,3,7,8-TCDD
modulates hormone receptors, including estrogens (Romkes and Safe, 1988; Romkes et al.,
1987), prolactin, and its own Ah receptor (Poland and Glover, 1980; Morrow et al., 1986).
However, the effect of 2,3,7,8-TCDD on testosterone receptors has not been evaluated.
In summary, the results from both the NIOSH and Ranch Hand studies are limited by
the cross-sectional nature of the data and the type of clinical assessments conducted.
However, the available data provide evidence that alterations in human male reproductive
hormone levels are associated with serum 2,3,7,8-TCDD.
7.15.2. Possible Acute Effects of Exposure to 2,3,7,8-TCDD
The following section reviews end points that were described in groups shortly after
exposure to 2,3,7,8-TCDD but were not observed as chronic effects in studies conducted
many years after exposure ceased. Also reviewed are end points observed as long-term
effects in single studies.
7.15.2.1. Dermatologic Conditions Other Than Chloracne
Dermatologic conditions other than chloracne, such as hyperpigmentation,
hypertrichosis, and eyelid cysts, have been related to exposure to 2,3,7,8-TCDD in early
case reports (Ashe and Suskind, 1950; Suskind et al., 1953; Bleiberg et al., 1964; Poland et
al., 1971; Bauer et al., 1961; Goldman, 1972; Jirasek et al., 1974; Oliver, 1975). However,
these conditions may have been acute effects of 2,3,7,8-TCDD exposure that resolved over
time or may be residual effects of chloracne because they appear to occur more frequently in
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individuals with persistent chloracne (Suskind and Hertzberg, 1984). These conditions were
not observed in studies in which the cohorts were examined years after cessation of
exposure, in individuals with the potential for high exposure, or in those with high adipose or
serum 2,3,7,8-TCDD levels (Moses et al., 1984; Webb, 1989; Roegner et al., 1991).
Actinic keratosis, Peyronie's disease, and basal cell carcinoma may not be associated
with 2,3,7,8-TCDD. All three conditions were observed in only one study (Suskind and
Hertzberg, 1984; Lathrop et al., 1984) and were not observed in studies of individuals with
similar potential for exposure (Ott et al., 1993b).
7.15.2.2. Liver Enzymes Other Than GGT and Hepatomegaly
A number of studies reported elevated liver enzymes, particularly AST and ALT,
among individuals who were being exposed at the time of the measurement (May, 1973) or
whose exposure was within a few years of the measurement (Jirasek et al., 1974; Mocarelli
et al., 1986; Caramaschi et al., 1981). Follow-up studies or longitudinal analyses of exposed
cohorts suggest that the increase in enzyme level resolves over time (Mocarelli et al., 1986;
Assennato et al., 1989; Pazderova-Vejlupkova et al., 1981; May, 1982). In studies of
exposed populations tested many years after exposure ceased, levels of AST and ALT were
within normal range (Calvert et al., 1992; Webb et al., 1989; Roegner et al., 1991; Suskind
and Hertzberg, 1984; Moses et al., 1984).
D-glucaric acid was tested in a number of 2,3,7,8-TCDD-exposed populations as an
indicator of enzyme induction (Ideo et al., 1985; Martin, 1984; Roegner et al., 1991; Calvert
et al., 1992). Shortly after the TCP reactor release, D-glucaric acid levels in Seveso
children were elevated (Ideo et al., 1985). No other studies of exposed groups tested 5 to 37
years after exposure ceased found elevations of this enzyme (Martin, 1984; Roegner et al.,
1991; Calvert etal., 1992).
These data suggest that certain hepatic enzymes are increased as a response to high,
exogenous exposure to 2,3,7,8-TCDD. Once the exposure ends, the enzyme levels seem to
decrease over time, as observed in the Seveso populations (Mocarelli et al., 1986; Ideo et
al., 1985). Additional evidence of the acute nature of AST, ALT, and D-glucaric acid
elevations is demonstrated by the lack of such increases in studies of highly exposed groups
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conducted long after exposure ceased (Calvert et al., 1992; Roegner et al., 1991; Martin,
1984).
As in the case of ALT, AST, and D-glucaric acid, hepatomegaly appears to be a
condition reported in case reports after high exposure to 2,3,7,8-TCDD-contaminated
chemicals, particularly among TCP production workers examined after a TCP reactor
explosion and among Seveso residents (Ashe and Suskind, 1950; Suskind et al., 1953;
Jirasek et al., 1974; Reggiani, 1980). Later studies conducted after exposure ceased failed to
find excess dose-related hepatomegaly in the exposed populations (Bond et al., 1983; Suskind
and Hertzberg, 1984; Moses et al., 1984; Calvert et al., 1992; Centers for Disease Control
Vietnam Experience Study, 1988a; Roegner et al., 1991; Webb et al., 1989; Hoffman et al.,
1986). However, the absence of an effect in cross-sectional studies does not confirm the lack
of an effect in the past.
7.15.2.3. Pulmonary Disorders
Early case reports suggest that exposure to 2,3,7,8-TCDD chemicals may cause
temporary respiratory irritation (Zack and Suskind, 1980) and tracheobronchitis (Goldman,
1972). The data from two cross-sectional medical studies provide weak evidence of slightly
decreased lung function among exposed individuals (Suskind and Hertzberg, 1984; Roegner
et al., 1991). In these studies, the effects may be due more to smoking (Roegner et al.,
1991) or to a substantial age difference between the exposed and unexposed groups (Suskind
and Hertzberg, 1984). One study of highly exposed TCP production workers found no
relationship between serum 2,3,7,8-TCDD levels and chronic obstructive pulmonary disease,
bronchitis, or decreased pulmonary function (Calvert et al., 1992).
In conclusion, case reports indicate that intense acute exposure to 2,3,7,8-TCDD can
produce respiratory irritation. However, the findings from controlled epidemiologic studies
conducted many years after exposure do not convincingly support an association between
2,3,7,8-TCDD exposure and chronic effects on the respiratory system.
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7.15.2.4. Neurologic Disorders
The results of case reports and epidemiologic studies demonstrate that exposure to
2,3,7,8-TCDD-contaminated materials is associated with symptoms referable to the central
and peripheral nervous systems shortly following exposure and, in some cases, lasting many
years (Filippini et al., 1981; Ashe and Suskind, 1950; Moses et al., 1984). Overall,
however, the neurologic status of workers, community residents, and Vietnam veterans
exposed to 2,3,7,8-TCDD and evaluated from 5 to 37 years after last exposure appears to be
normal (Centers for Disease Control of Vietnam Experience Study, 1988a; Lathrop et al.,
1984; Sweeney et al., 1993). The data suggest that, although exposure to 2,3,7,8-TCDD
may have been extensive as in exposed workers, Ranch Hands, and Seveso residents, the
effects described in case reports may have been transient (Filippini et al., 1981; Lathrop et
al., 1984; Centers for Disease Control Vietnam Experience Study, 1988a, b; Assennato et
al., 1989; Alderfer et al., 1992; Sweeney et al., 1993). The findings of recent studies
suggest that in adults there are no long-term neurologic effects caused by even high exposure
to 2,3,7,8-TCDD-contaminated materials, but there is very little information with which to
examine the effects of exposure on the developing human neurologic system.
7.15.2.5. Porphyrias
In rats and mice, exposure to 2,3,7,8-TCDD has been clearly shown to alter
porphyrin metabolism (Goldstein et al., 1973; Smith et al., 1981; Jones and Chelsky, 1986;
DeVerneiul et al., 1983; Cantoni et al., 1981; Goldstein et al., 1982). Whether 2,3,7,8-
TCDD is associated with porphyrin changes in humans, particularly porphyria cutanea tarda
(PCT), is a subject of unresolved debate. It has been suggested that the PCT and elevated
urinary porphyrins observed in the New Jersey and Czechoslovakian workers during the
years of operation of the plants were the result of exposure to hexachlorobenzene, which was
produced at the same time as TCP (Pazderova-Vejlupkova et al., 1981; Jones and Chelsky,
1986). These statements have not been corroborated with strong studies. In the follow-up
studies, urinary porphyrin levels of these TCP production workers were not elevated
(Pazderova-Vejlupkova et al., 1981; Poland et al., 1971) or did not differ from levels in the
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control group (Calvert et al., 1993). Doss et al. (1984) also described transient elevations in
coproporphyrins among 22 Seveso residents exposed to 2,3,7,8-TCDD.
Because 2,3,7,8-TCDD is a potent porphyrigen in rats and mice, it has been of
interest to determine whether exposures may have contributed to the observed changes in
porphyrin levels in human populations. The NIOSH study could not address the question of
etiology or transient porphyria, but it did not find porphyria in highly exposed workers many
years after their occupational exposure.
7.15.3. Effects For Which Further Research Is Needed
The following section describes end points for which the animal data have
demonstrated exposure-related effects, but the human data are inconclusive and require
further study.
7.15.3.1. Diseases of the Circulatory System
In general, the results of the cohort mortality studies of TCP production workers were
remarkably similar. For all of the studies, the standardized mortality ratios for diseases of
the circulatory system (ICD-9: ICD 390-459) were approximately 100, meaning that the
death rate in the exposed population was nearly the same as that in the general population,
controlling for age, race, gender, and calendar year (Fingerhut et al., 199Ib; Zober et al.,
1990; Bueno de Mesquita et al., 1993; Bertazzi et al., 1989, 1992; Collins et al., 1993;
Bond et al., 1989; Coggon et al., 1991). None of the SMRs above 100 were statistically
significantly elevated.
Mortality from circulatory system diseases among Ranch Hands (SMR = 110, 95%
CI=60-150) (Michalek et al., 1990) and Australian Vietnam veterans (RR=1.7, 95%
CI=0.9-3.0) (Fett, 1987b) was nonsignificantly elevated. There was a deficit of deaths from
this cause among U.S. Army Vietnam veterans compared to non-Vietnam veterans (Centers
for Disease Control Vietnam Experience Study, 1988c). Elevated mortality from circulatory
diseases among Seveso residents is considered by the authors to be the result of
environmental stresses and possibly other risk factors rather than exposure to 2,3,7,8-TCDD
(Bertazzi et al., 1989). In addition, a comparison of Zone A residents (2,3,7,8-TCDD-
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contaminated region) with Zone B residents (less contaminated region) found the circulatory
disease mortality rate to be higher in the residents of the less contaminated area.
Diseases of the circulatory system, particularly heart disease, are the leading causes of
death among populations of most developed nations. Leading risk factors include age,
cigarette smoking, elevated lipid levels, obesity, hypertension, diabetes, and physical
inactivity (Smith et al., 1992). Among the studies that examined mortality from circulatory
system diseases, none directly adjusted SMRs for known risk factors or attempted to evaluate
jointly the contribution of known risk factors and 2,3,7,8-TCDD to the mortality rates
(Fingerhut et al., 1991b; Zober et al., 1990; Bueno de Mesquita et al., 1993; Bertazzi et al.,
1989, 1992; Collins et al., 1993; Bond et al., 1989; Coggon et al., 1991). Therefore, given
the strong contribution of these risk factors, it is not possible to rule out physical and
personal risk factors in the etiology of diseases of the circulatory system and heart in these
populations. However, the absence of a "health worker effect" for these causes of death
suggests that future research be directed specifically at the relationship between circulatory
and heart disease and exposure to 2,3,7,8-TCDD.
Cross-sectional morbidity studies have not found increases in the prevalence of
circulatory or heart disease among TCP workers, Ranch Hands, or U.S. Army Vietnam
veterans (Suskind and Hertzberg, 1984; Bond et al., 1987; Moses et al., 1984; Centers for
Disease Control Vietnam Experience Study, 1988a). In some cross-sectional studies, risk
estimates were adjusted for some risk factors, depending on the study (Suskind and
Hertzberg, 1984; Poland et al., 1971; Moses et al., 1984; Bond et al., 1983; Centers for
Disease Control Vietnam Experience Study 1988a; Roegner et al., 1991). Ranch Hands
were the only group to experience marginal differences in diastolic blood pressure,
arrhythmias, and peripheral pulse abnormalities after adjusting for selected risk factors
(Roegner etal., 1991).
The animal data suggest that at high levels of 2,3,7,8-TCDD, the vascular system,
cardiac muscle, and valves and function may be affected by exposure (Kociba et al., 1978;
Buu-Hoi et al., 1972; Brewster et al., 1988; Hermansky et al., 1988; Kelling et al., 1987;
Canga et al., 1988). However, with the exception of the long-term feeding study (Kociba et
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al., 1978), the exposures in animals were single, high doses and the human exposures
(except Seveso) were chronic, medium to high doses.
In summary, the animal studies suggest that 2,3,7,8-TCDD causes pathologic changes
that may lead to later circulatory system disease. However, long-term studies of mature and
aged animals have not been carried out to evaluate these hypotheses and to correlate the
results of the animal with the human studies. Few epidemiologic studies were designed to
control for many of the risk factors known to cause circulatory system and heart disease, but
a consistent absence of the health worker effect for circulatory disorders and heart disease in
numerous mortality studies suggests the need for additional research in this area. These
studies should also include methods to quantify subject exposure to 2,3,7,8-TCDD.
7.15.3.2. Reproductive Effects
7.15.3.2.1. Consistency. A variety of study designs, including case-control, ecologic,
cross-sectional, and historical cohort designs, have addressed the issue of 2,3,7,8-TCDD and
reproductive effects in humans. Unfortunately, the different criteria for case definitions
across studies make it difficult to compare the results. In addition, the method of case
ascertainment for certain end points influences the rate of events observed. The Vietnam
Experience Study substudies of veteran-reported birth defects compared with those identified
through hospital records demonstrated that rates of self-reported outcomes differed by
exposure status. Moreover, predictive value of self-reported events was poor in both
cohorts. In contrast, rates of birth defects in the Ranch Hand study were similarly reported
by the Ranch Hands and controls. Both groups underreported 7% of birth defects in children
conceived prior to their SEA tour and 14% after their tour of duty.
7.15.3.2.2. Strength. With the exception of the finding that Vietnam veterans were more
than twice as likely to have low sperm concentrations (OR=2.3, 95% CI=1.2-4.3), no effect
measure greater than 2 was noted in any of these investigations. This is not surprising,
given the limitations of the studies, particularly with regard to exposure misclassification.
Therefore, the trends across these studies carry more import than "statistically significant"
results.
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7.15.3.2.3. Temporality, dose-response. While these studies have restricted inclusion of
reproductive events to those that occurred after exposure to 2,3,7,8-TCDD was suspected, no
study has evaluated 2,3,7,8-TCDD levels at the time of the outcome. Determination of a
2,3,7,8-TCDD dose-response relationship with adverse reproductive outcomes would not be
valid unless individual 2,3,7,8-TCDD levels were available. Elucidation of a dose-response
pattern may be derived using data from the Ranch Hand study to calculate 2,3,7,8-TCDD
levels at the time of the reproductive event, which would be an important contribution toward
understanding this phenomenon. However, with regard to spontaneous abortion, this analysis
would not be able to address those losses occurring early in gestation.
7.15.3.2.4. Biological plausibility. A growing body of animal research described in
Chapter 5 lends biological plausibility to the association between dioxin and most of the
reproductive end points evaluated in these studies, with the notable exception of molar
pregnancies. There is growing evidence that dioxin affects testis and accessory gland weight,
testicular morphology, spermatogenesis, and fertility in males. A model for a paternally
mediated dioxin effect on congenital malformations has not been reported. Among female
animals, the primary reproductive end points that have been examined include decreased
fertility and pregnancy loss.
The mechanism by which 2,3,7,8-TCDD causes adverse reproductive and
developmental effects has not been well described, although considerable insight has been
gained from research focusing on the Ah receptor. While the Ah receptor has been linked
with birth defects in several mouse strains, it appears that the mechanism of effect may be
dependent on the outcome evaluated, as well as other dioxin congeners to which the
population is exposed. Clearly, these relationships in humans have not been adequately
investigated.
The discovery in the Times Beach, Missouri; CDC; and Ranch Hand studies that self-
reported dioxin exposure and exposure indices developed from the analyses of 2,3,7,8-TCDD
in soil and military records to estimate individual exposure are poorly correlated with serum
2,3,7,8-TCDD levels was a critical turning point in understanding the inconsistencies of the
research to date regarding 2,3,7,8-TCDD and effects. Thus, because of the likelihood of
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exposure misclassification in those studies lacking direct measures of exposure, the findings
have been severely limited.
7.15.3.3. Specific Reproductive End Points
1. Spontaneous abortions. Miscarriages were investigated in several studies with
different designs and varied patterns of parental exposure. Events were generally ascertained
by self- or spousal-report. When case ascertainment was through medical records, such as in
the Ranch Hand study or the Vietnamese investigations, the events are by definition restricted
to those miscarriages that were clinically recognized.
Research in the area of early pregnancy loss indicates that 30-50% of all conceptions
are lost prior to or during implantation (Hertig et al., 1959). The rate of loss between
implantation and expected first menstrual period ranges from 22% to 30% (Wilcox et al.,
1988; Sweeney et al., 1988). Thus it is clear that restriction of the examination of
pregnancy loss to those events that are ascertained through self-reports or medical records
results in a large proportion of the outcome of interest being missed. In studies of
environmental factors and spontaneous abortion in which information is lacking concerning
the conditions surrounding conception, "the conflation of different doses with different effects
can mislead" (Kline et al., 1989). Because of these discrepancies, it would not be
meaningful to pool the results of the research on the association between dioxin exposure and
miscarriage to judge the "consistency" of the association.
Overall, it must be acknowledged that the data compiled to date are inadequate to
address this issue. To simply enumerate and compare the number of "positive" versus
"negative" studies to ascertain consistency in the research would be inappropriate. The
reasons for this have been described above in detail, with emphasis on the high (40-50%)
exposure misclassification that has been documented in the majority of these investigations,
the small sample sizes evaluated, lack of data on dioxin levels at the time of conception, and
the unknown impact of early pregnancy loss on identification of birth defects. The animal
and human evidence for a 2,3,7,8-TCDD-pregnancy loss relationship is sufficiently
suggestive to warrant further investigation. Several studies of various designs and
populations have demonstrated weak but consistent associations (Reggiani, 1978; Hatch,
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1984b; Constable and Hatch, 1985; Huong et al., 1989; Phuong et al., 1989a; Stellman et
al., 1988), whereas others have not (Townsend et al., 1982; Smith et al., 1982; Cutting et
al., 1970; Kunstadter, 1982; Report to the Minister for Veterans' Affairs, 1983; Stockbauer
et al., 1988; Erickson et al., 1984; Centers for Disease Control Vietnam Experience Study,
1989). (These studies include those that should be restricted to the assessment of military
service in Vietnam and reproductive events.) The Ranch Hand study leaves several questions
unanswered, including the determination of a dose-response level at the time of conception
and adverse reproductive outcomes, as well as the impact of early pregnancy losses on rates
of spontaneous abortion and birth defects that survive long enough to be "counted."
2. Congenital malformations. The confusing evidence regarding the relationship
between dioxin exposure and birth defects results not only from the same limitations
described above for the studies of miscarriage but also from the lack of power to evaluate
specific types of malformations. To increase the power to detect a potential relationship, the
studies have combined all birth defects together and calculated an odds ratio for total birth
defects. Given emerging evidence for etiologic heterogeneity among subgroups of birth
defects (Khoury, 1989), it is possible that this approach might dilute the effect measure.
These studies also should be stratified by type of parental exposure, i.e., paternal,
maternal, or both. Biologically plausible mechanisms for birth defects resulting from
paternal exposure to toxic substances have not been well researched. It has been known for
many years that temporary infertility may occur after exposure of human males to certain
toxic substances. However, animal research suggests that spermatogenesis is "particularly
resilient" after exposure to these chemicals (Pearn, 1983). If dioxin were related to
malformations among the offspring conceived after paternal service in Vietnam, it is implied
that the effect must occur premeiotically. Some animal studies have found that
spermatogonia and spermatocytes (premeiotic spermatogenic cells) were able to repair DNA
after exposure to toxic agents, whereas spermatids and spermatozoa did not have this
capability (Lee and Dixon, 1978).
A few studies (Hatch, 1984a; Constable and Hatch, 1985; Huong et al., 1989;
Phuong et al., 1989a) have suggested an association, including those investigations of the
Yusho and Yu Cheng incidents (Rogan, 1982; Yen et al., 1989; Rogan et al., 1988). Many
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studies have failed to find a relationship between dioxin and birth defects (Townsend et al.,
1982; Smith et al., 1982; Cutting et al., 1970; Kunstadter, 1982; Stockbauer et al., 1988;
Erickson et al., 1984; Centers for Disease Control Vietnam Experience Study, 1989;
Mastroiacovo et al., 1988). Again, however, in view of the serious limitations of these
studies of 2,3,7,8-TCDD and reproductive events, it must be concluded that the relationship
between paternal dioxin exposure and congenital malformations remains unknown. However,
if military service in Vietnam is the exposure of interest, there is little evidence to support an
association with birth defects.
3. "Miscellaneous" end points. Additional reproductive outcomes that were
evaluated in a subset of the studies include molar pregnancies (in the Vietnamese studies),
infant birth weight, neonatal and infant death, and child cancer and mortality. Mainly
because of small sample sizes, it is difficult to reach conclusions regarding neonatal, infant,
and child mortality and childhood cancers. However, the increased risk for neonatal death
observed in the Ranch Hand study, the only study with individual TCDD levels, should be
further investigated. Available evidence does not support an association between paternal
dioxin level and low birth weight (Wolfe et al., 1992b; Centers for Disease Control Vietnam
Experience Study, 1989); a maternally mediated effect with documentation of TCDD
exposure has not been examined.
There was evidence for a maternally mediated effect of dioxin exposure and birth
defects in the Vietnamese studies. An attempt to confirm these findings by using recent
advances in exposure measurement would be a worthwhile effort.
The Vietnam Experience Study found a significant relationship between service in
Vietnam and sperm abnormalities, while the Ranch Hand study did not confirm these results
when exposure was defined by both cohort status and 2,3,7,8-TCDD levels. However, the
data on alterations in male reproductive hormone levels associated with occupational
exposure to 2,3,7,8-TCDD emphasize that further research in these areas is required.
In conclusion, the research to date has been successful in resolving some confusion
surrounding the conflicting evidence for an association of dioxin exposure and various
reproductive end points in humans. High occurrence of exposure misclassification,
differences in case definitions across studies, and small sample sizes have severely limited
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the power of these studies to address these questions. Additional research that includes a
measure of dioxin level at the time of conception for both the father and mother is necessary
if the effect of dioxins on the spectrum of reproductive outcomes is to be understood.
7.15.3.4. Immunologic Effects
Information on immunologic function in humans relative to exposure to 2,3,7,8-
TCDD is scarce. The available epidemiologic studies do not describe a consistent pattern of
effects among the studies. Two studies of German workers, one exposed to 2,3,7,8-TCDD
and the other to 2,3,7,8-tetrabrominated dioxin and furan, observed dose-related increases of
complements C3 or C4 (Zober et al., 1992; Ott et al., 1993b). Other studies of exposed
groups have not examined complement components to any great extent.
More comprehensive evaluations of immunologic function with respect to 2,3,7,8-
TCDD exposure are necessary to assess more definitively the relationships observed in
nonhuman species. Longitudinal studies of the maturing human immunologic system may
provide the greatest insight, particularly since animal studies have found significant results in
immature animals and human breast milk is a source of 2,3,7,8-TCDD and other related
compounds. Additional studies of highly exposed adults may also shed light on the effects of
long-term chronic exposures. Therefore, there appears to be too little information to suggest
definitively that 2,3,7,8-TCDD, at the levels observed, is an immunotoxin in humans.
7.15.3.5. Lipids
Animal studies indicate that 2,3,7,8-TCDD is associated with generally increased
serum cholesterol and serum triglyceride levels. The effect of exposure to 2,3,7,8-TCDD-
contaminated chemicals on lipids is not consistent in the available epidemiology studies.
Elevations in total cholesterol and triglyceride levels were reported after high 2,3,7,8-TCDD
exposure in TCP workers (Pazderova-Vejlupkova et al., 1981; Martin, 1984) and laboratory
workers (Oliver, 1975). Despite their very high exposure to 2,3,7,8-TCDD-contaminated
chemicals, neither adults nor children from Seveso had lipid levels above the referent level.
Risk factors such as dietary fat intake, familial hypercholesterolemia, alcohol consumption,
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and exercise, which also affect cholesterol and other lipid levels, may be factors that were
not considered in these studies.
Ranch Hands continue to have elevated lipid levels despite the extended length of time
between exposure and testing. In the 1992 phase of the Ranch Hand study, additional
parameters that may affect lipid levels were examined. Analysis of the lipid data collected
during the NIOSH study also will contribute to the information on lipid levels in relation to
serum 2,3,7,8-TCDD levels.
7.15.3.6. Thyroid Function
Many effects of 2,3,7,8-TCDD exposure in animals resemble signs of thyroid
dysfunction. In the few human studies that examined the relationship between 2,3,7,8-
TCDD exposure and thyroid function, the results are mostly equivocal (Centers for Disease
Control Vietnam Experience Study, 1988a; Roegner et al., 1991; Suskind and Hertzberg,
1984). However, TBG appears to be positively correlated with current levels of 2,3,7,8-
TCDD in the BASF accident cohort (Ott et al., 1993b). Little information on thyroid
function has been reported for production workers and none for Seveso residents, two groups
with documented high serum 2,3,7,8-TCDD levels. A study of nursing infants suggests that
ingestion of breast milk with a higher total TEQ may alter thyroid function, but the study is
limited by the small size and a short observation period (Pluim et al., 1992). Because the
products of the thyroid control many aspects of metabolism, a better understanding of the
relationship between 2,3,7,8-TCDD and thyroid function is important, especially regarding
the developing human.
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30: 2825-2830.
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Smith, A.M.; Fisher, D.O.; Pearce, N.; Chapman, C.J. (1982) Congenital defects and miscarriages among New
Zealand 2,4,5-T sprayers. Arch. Environ. Health 37: 197-200.
Smith, A.H.; Patterson, D.G., Jr.; Warner, M.L.; MacKenzie, R.; Needham, L.L. (1992) Serum 2,3,7,8-
tetrachlorodibenzo-p-dioxin levels of New Zealand pesticide applicators and their implications for
cancer hypotheses. J. Natl. Cancer Inst. 84: 104-108.
Stehr, P.A.; Stein, G.; Falk, H.; Samson, E.; Smith, S.J.; Steinberg, K.; Webb, K.; Ayres, S.; Schramm, W.;
Donnell, H.D.; Gedney, W.B. (1986) A pilot epidemiologic study of possible health effects associated
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Stellman, S.D.; Stellman, J.M.; Sommer, J.F. (1988) Health and reproductive outcomes among American
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Stockbauer, J.W.; Hoffman, R.E.; Schramm, W.F.; Edmonds, L.D. (1988) Reproductive outcomes of mothers
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immunological tests. Br. J. Ind. Med. 49: 532-544.
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8. DOSE-RESPONSE MODELING
8.1. INTRODUCTION
The current effort to reevaluate the risk of exposure to dioxins is being termed a
Biological Basis for Risk Assessment. The underlying premise is that this may be a special
case for a nonmutagenic, receptor-mediated carcinogen. The goal of this reassessment is to
consider more mechanistically based models that are sufficiently credible to the scientific
community.
Most of the information presented in the Introduction is found in more extensive
detail in the other background chapters. It is useful to summarize the salient features of
those papers that have an impact on the development of dose-response models so that readers
of this chapter will be able to evaluate the scientific foundation on which the dose-response
models presented in this chapter are based.
2,3,7,8-TCDD is the most potent form of a broad family of xenobiotics that bind to
an intracellular protein known as the Ah receptor. Other members of this family include
halogenated hydrocarbons such as the PCBs, naphthalenes, and dibenzofurans, as well as
nonhalogenated species such as 3-methylcholanthrene and /8-naphthaflavone. The biological
properties of dioxins have been investigated extensively in over 5,000 publications and
abstracts since the identification of TCDD as a chloracnegen (Kimming and Schultz, 1957).
Much of the biological activity of TCDD appears to follow the rank order binding affinity of
the congeners and analogs to the Ah receptor (AhR). This rank order holds for toxic
responses such as acute toxicity and teratogenicity and for changes in concentration of several
hepatic proteins including the induction of cytochromes P-450IA1 and IA2 and the
modulation of the estrogen receptor and epidermal growth factor receptor (EGFR). The
relationship between AhR binding and carcinogenicity of TCDD is less clear. However,
TCDD is a carcinogen in several strains of laboratory mice and rats, and the tumor sites
include liver, thyroid, and the respiratory tract, as well as others. There is considerable
evidence that TCDD does not interact directly with the DNA to cause mutations. The study
most utilized for the cancer risk assessment of TCDD is that of Kociba et al. (1978). These
authors reported an increase in hepatocellular carcinomas and hepatomas, along with
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decreases in several endocrine tumors in female rats. Male rats were remarkably less
susceptible to the hepatocarcinogenic actions of TCDD. However, there is no striking sex
specificity in the hepatocarcinogenic actions of TCDD in mice.
The overall hypothesis of TCDD action, put forth by several groups, is based on the
transcription of specific genes exemplified by the cytochrome P-450IA1 gene. The biological
basis for this approach is outlined in Chapter 2, Mechanism(s) of Action. Although
substantial gaps in our knowledge remain, it is accepted by most researchers that most if not
all cellular responses to TCDD require interaction between TCDD and the Ah receptor. The
binding of TCDD to AhR is similar, although not necessarily identical, to the interaction of
many steroid hormones with their intracellular receptors, as pointed out by Gustafsson in a
series of articles drawing comparisons between the AhR and the glucocorticoid receptor
(Poellinger et al., 1986, 1987; Cuthill et al., 1988). DeVito et al. (1991) have also drawn an
analogy between the action of TCDD and estradiol at their respective receptor proteins.
The binding of TCDD to AhR is reversible. However, subsequent events seem to
reduce the likelihood of dissociation of the ligand:receptor complex. One such event that has
been recently studied is the association of the ligand:receptor complex with another
macromolecule, the so-called ARNT (AhR nuclear transport) protein (Hoffman et al., 1991).
There may be a family of ARNT proteins that differ by cell types, which could account, in
part, for the diversity of actions of TCDD in different tissues. The association of ARNT
with the ligand-bound receptor induces some physical changes in the complex, which tends to
reduce dissociation of the ligand and favors the movement and/or retention of the complex
into the nucleus. Overall, the relationship between TCDD concentration and nuclear AhR-
TCDD concentration appears to be linear, indicating that, at low ligand concentrations,
ARNT is not a rate-limiting factor. In the case of transcriptional activation of the CYP1A1
gene, the AhR-ARNT-TCDD complex (activated TCDD complex) associates with specific
elements in the genome called the xenobiotic (or dioxin) responsive elements (XREs or
DREs). The association of the activated TCDD complex with the DRE is also reversible
(Gasiewicz et al., 1991), and there is recent evidence that at least two DREs need to be
occupied to transcriptionally activate the CYP1A1 gene (Chapter 2, Mechanism(s) of
Action). The structure and amino acid sequence of the AhR protein have been reported
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(Burbach et al., 1992; Ema et al., 1992). Each AhR appears to bind one molecule of
TCDD, and at low concentrations of ligand (i.e., when [ligand] < < [receptor]), the binding
of TCDD to AhR is likely to be linearly related to [TCDD].
Much of the sequence of events is analogous to the steroid receptors and the
respective genomic response elements. This similarity helps us in proposing biological
models of TCDD action and risk assessment. The steroid hormones and their receptors
belong to a multigene family that includes the thyroid hormone receptors, oncogene products,
glucocorticoids, mineralcorticoids, vitamin D, retinoids, androgens, estrogens, and progestins
(Chapter 2, Mechanism(s) of Action; Evans et al., 1988). Biologically, these are all
multipotent agents that induce a range of cellular responses in different organs, many at
extremely low concentrations. They share a nuclear location for the transduction of
ligand:receptor action, and their common mechanism of action is the regulation of gene
expression (Jensen, 1991). Within the family of known receptors from these agents, there is
considerable sequence homology and a common basic structure, consisting of a ligand-
binding domain and a DNA-binding domain. The biological activity of these receptors is
varyingly regulated by metals and by phosphorylation state. Some-but not all-hormone
receptors may interact with chaperone-type proteins, subsequent to ligand binding, which
transduce conformational changes and other events critical to nuclear translocation and DNA
binding. The ARNT protein functions in this fashion (Hoffman et al., 1991). Other
receptors are associated with so-called heat shock proteins or proteins that must be shed to
transform the liganded receptor into a DNA-binding form, and the DNA-binding domain of
some receptors contains zinc finger loops, although this does not seem to be the case for the
Ah receptor.
The steroid hormone receptors, sometimes as liganded dimers, move to the nucleus
and regulate gene transcription through specific DNA sequences near the target genes. These
events are complex because of interactions between the liganded receptor and nuclear
proteins that function as transcription factors by binding to other DNA sites upstream of the
hormone responsive elements and the regulated gene. These transcription factors may
regulate the binding affinity of the steroid hormone receptor itself to DNA (Muller et al.,
1991). Additional complexity is introduced by the interactions among steroid hormone
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receptors, at the genetic level, and by the effects of hormone upon the number,
conformation, and localization of receptors. Down- and up-regulation of receptor gene
transcription and receptor synthesis may also be involved in cell-level modulation of steroid
hormone action. As pointed out by Muller et al. (1991), every step in the signal transduction
pathway of these hormones, from receptor gene transcription to ligand: receptor: DNA action,
is likely to be regulated by both dependent and independent factors. Moreover, nongenomic
effects of steroid hormones occur rapidly and independent of nuclear translocation of
liganded hormone receptors or in the presence of protein synthesis inhibitors (Roth and
Grunfield, 1985).
Attempts to model the steps involved in signal transduction have examined events
step-by-step as well as the overall set of reactions from entrance of hormone into the cell to
cellular response. Of interest in this report is the information that may be available
concerning the overall dose-response relationship for steroid hormones. The highly complex
cascade of biological events that intervenes from hormone entrance to cell response may
modulate hormone action in the following ways: It may amplify cell response, in the way
that second messengers for membrane-associated receptors (such as neurotransmitter
receptors) appear to amplify molecular signaling; it may transduce response in a manner
proportional to concentration of hormone (that is, linearly); or it may introduce dampening
into the response network. Amplification of signal transduction implies that, at some stage in
a multistep process, more than one event is triggered as a consequence of one preceding
event. Dampening implies that, at some stage, more than one event must occur before the
next event is triggered. Linear transduction implies that the relationship of event to event,
for all steps, is invariant.
In considering these possible dose-response relationships, it is likely to be important
to distinguish among endogenous and exogenous ligands for the same steroid hormone
receptor, particularly if the two types of ligands differ in rates of turnover (degradability) or
affinity for the receptor. We are hampered in our inferences for the dioxins because the
endogenous ligand(s), if present, has not yet been identified, and thus we are not certain if
TCDD is more or less stable than this ligand, if its affinity is higher or lower than an
endogenous ligand, or if an endogenous ligand would act as an agonist or antagonist for
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dioxin-like effects. With respect to stability, it is unlikely that an endogenous ligand would
be as stable as TCDD. Most endogenous ligands for steroid hormone and other receptors are
rapidly cleared, either by compartmentation (as with neurotransmitter reuptake processes) or
by enzymatic degradation, as with steroids. With respect to kinetics of binding of TCDD, its
in vivo affinity for the receptor is extremely high, in the 10~n range. If the affinity for the
natural ligand is even higher, then it is likely that the overall relationships between natural
ligand and receptor are even stronger than those we may explicate for TCDD; if it is lower,
then it would be an unusual member of the steroid hormone family. Of course, differences
in affinity, if these exist, may not influence the overall kinetics of the dose-response
relationship as much as differences in the number of events required to trigger the reaction
from step to step.
Evaluation of dose-response relationships for receptor-mediated events requires
information on the quantitative relationships between ligand concentration, receptor
occupancy, and biological response. Roth and Grunfeld in The Textbook of Endocrinology
(1985) state:
At very low concentrations of hormone ([H] <
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is important to note that the data in Figure 8-1 are plotted on a semilog scale. If the same
data are plotted arithmetically (Figure 8-2), then the shape of the dose-response curve readily
conveys the linear relationship between receptor occupancy and biological response at lower
concentrations and saturation at higher concentrations.
Such a simple proportional relationship does not explain the diversity of biological
responses that can be elicited by a single hormone utilizing a single receptor. For example,
low concentrations of insulin produce much greater effects on fat cells than on muscle cells.
These differences are due to tissue- and cell-specific factors that modulate the qualitative
relationship between receptor occupancy and response. Similarly, it is expected that there
are markedly different dose-response relationships for different effects of TCDD.
Coordinated biological responses, such as TCDD-mediated increases in cell proliferation,
likely involve other hormone systems, which means that the dose-response relationships for
relatively simple responses (i.e., CYP1A1 induction) may not accurately predict dose-
response relationships for complex responses such as cancer. As we gain more
understanding of the entire sequence of events responsible for TCDD-mediated toxic effects,
we will enhance our ability to more accurately predict dose-response relationships. The
mechanism(s) responsible for qualitative and quantitative diversity in receptor-mediated
responses will be discussed in more detail in Section 8.7, Knowledge Gaps.
Cancer is a multistep, multistage disease in which several operationally defined events
have been described primarily on the basis of assay systems developed to detect these events.
These events are initiation, fixation, promotion, and progression, as shown in Figure 8-3.
Although this schematic arrays these events in a linear progression, it is important to realize
that there are multiple pathways by which a cell may progress through these stages from an
early alteration in gene structure or expression to the expansion of a clone into a tumor. The
early alteration in gene structure or expression is often referred to as initiation. Structural
damage to DNA, through alkylation or deletion, is an example of initiation. This initial
event is immortalized, or stabilized, in the daughter cells of the initiated cell through the
processes of fixation. Promotion is the enhanced growth of the cell population with fixed
genetic damage; promotion may be supported by hormones and other modifications in cell
growth and proliferation. Progression is a term used to describe additional alterations in
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(arithmetic scale)
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gene structure or expression, such as second mutations for colon cancer, that appear to be
necessary in the growth of the clone into a clinical end stage. These events are not
necessarily ordered in this sequence, nor is it clear that distinctly different events-genotoxic
and epigenetic—are involved in each stage. Some of these events are often defined by the
test systems used to assay for their occurrence: For instance, initiation is often equated with
mutation, such as the mutations that are detectable in in vitro bacterial assays of the Ames
type. Promotion is often equated with a positive result in an experimental paradigm of
sequential treatment of animals with a strong mutagen, followed by a putative "promoter."
Dioxin is an operational promoter, as defined in assay systems of skin and liver in
mice and rats, respectively (Pilot et al., 1987; Clark et al., 1991; Lucier et al., 1991). Since
many hormones are promoters (Pilot and Dragan, 1991; Lucier, 1992), it is nol surprising
lhat dioxin has these properties as well. The interactions of dioxin with hormones are
complex (see below), and in cancer bioassays there is evidence for interactions among dioxin
and sex hormones, in that the rates of liver tumors in rodents are much higher in females
than in males (National Toxicology Program [NTP]), and ovariectomy suppresses dioxin
promotion of diethymitrosamine (DEN)-initiated liver tumors (Lucier et al., 1991). Dioxin is
also a complete carcinogen; that is, il induces increased lumor yields in experimental animals
not pretreated with strong mutagens (Kociba et al., 1978; NTP, 1982). Dioxin is not a
mutagen in in vitro systems commonly used to detect mutation through DNA damage. There
is some evidence for in vivo clastogenicity of DNA (increased chromosomal breaks) in
animals exposed to dioxin (Slohs el al., 1990). These data have presenled challenges lo Ihe
application of general models for cancer risk assessment, which are based on assumptions of
mutagenesis as a fundamenlal mechanism for chemical- or radiation-induced cancer.
The general approach of the U.S. EPA to regulation of carcinogens is to use the
Armilage-Doll model of carcinogenesis (Figure 8-4). In Ihis model, the movement of cells
from one slage lo the other is assumed to be due to a sequence of mutations similar to the
step of initiation/fixation discussed above. As with any mathematical model, specific forms
must be chosen for the rate constanls that define the process. The EPA formulation of this
model assumes the mutation rates [/*,(d,t)] are a linear function of dose and are constant over
time. These assumptions result in a cumulative tumor incidence rate that can be
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8-11
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approximated by a polynomial function of dose with coefficient q; for dose', i=0,l,2,...k. In
the low-dose region, the risk is dominated by the linear term in the polynomial (q^. EPA
generally uses a 95% upper-confidence limit (q! on the linear term of this formulation of the
multistage model for cancer risk assessment. This model, using the 95% upper-confidence
limit on the linear term, is referred to as the linearized multistage (LMS) model. However,
this choice has not been predicated solely on the correctness of a K-mutation process of
cancer development. The linearized mathematical properties of the multistage model can be
appropriate for a larger class of mechanisms: Dose-response behavior, which is linear at low
dose, may have upward curvature in the intermediate range and show a downward curvature
or saturation of response at high dose. In particular, arguments that a compound's action is
additive to background biological processes lead to a linear response at low dose under rather
general conditions (Crump et al., 1976). Therefore, for practical modeling purposes, it is
important to address whether biological knowledge about the action of a carcinogen can fit
the general dose-response shape predicted by the linearized multistage model.
Our knowledge of dioxins may challenge some of the simplifying assumptions that
have been the biological basis for cancer risk assessment models. While dioxins do not
possess the properties of an operationally defined "genotoxin" (using in vitro tests of
mutagenicity), dioxins clearly affect gene expression (see below and also Whitlock chapter).
These gene-level effects appear to involve a transcription regulator, the so-called dioxin
receptor, which when liganded to dioxin and translocated to the nucleus is associated with
specific recognition elements upstream of target genes whose transcription may be up-
regulated or down-regulated by the dioxin receptor. It is difficult to fit these events into our
current conceptual model of chemical carcinogenesis.
For other lexicological end points such as terata, organ toxicity, acute toxicity, etc., a
threshold has often been assumed primarily as a matter of policy. For these end points,
safety factors or uncertainty factors have been used to estimate no-effect exposure levels.
This threshold approach is used by the World Health Organization to set acceptable daily
intakes (ADIs) for direct and indirect food additives. EPA now uses the term reference
dose. In distinction, EPA policy assumes the dose-response curve for excess carcinogenic
risk to be linear through dose zero. Several mechanisms could generally lead to this form of
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response, including direct mutational activity of the chemical agent and/or additivity to
background rate of tumor formation (Portier, 1987). Since TCDD does not bind covalently
to DNA and must exert its effects through receptor action, this default position must be
carefully reexamined.
8.1.1. Introduction to Modeling for TCDD
There are several models under consideration at the present time, ranging from very
simple to complex. It is obvious that the biology governing the toxicity of TCDD, beyond a
few initial critical events, is not straightforward. These critical events, the first of which is
binding to the Ah receptor, are generally response-independent. The response-dependent
events are species-, gender-, organ-, tissue-, and perhaps cell-specific. If the binding to the
AhR is essential but not sufficient for effects to occur, then the dose-response curve for this
event (as well as the rate equations) should be a better predictor of biological action than
dose as long as the dose-response curves for these subsequent actions are functionally similar
to the receptor binding curves. In general, the available data indicate receptor involvement is
necessary for most if not all low-dose actions of TCDD. Since the AhR has been detected in
virtually all cells but all cells do not respond to TCDD, there must be other factors that are
necessary for TCDD action. The roles of these cell-specific factors must be elucidated
before there is a complete understanding of TCDD action. However, a model may be
developed for specific end points by using available data and reasonable assumptions.
Several important factors have been generally accepted. One, TCDD is a member of
a class of xenobiotics (and probably natural products) that is nonmutagenic, binds to a
cellular receptor, and alters cell growth and development. Two, a significant amount of
information is available for estimating risks from exposure to this compound and the default
position of directly applying the LMS model as a function of dose needs to be reevaluated.
Three, the biology of receptor-mediated events should be included in any modeling exercise
for TCDD. The goal of the modeling is to use as much data as possible to reduce these
uncertainties and to identify the areas where data gaps exist.
One difficulty with a novel, albeit biologically based, approach is that it is replacing
paradigms (safety factors and LMS models) upon which the U.S. Government's risk
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assessments have been based. There is no a priori reason to believe that a model based on
greater experimental evidence will be more or less conservative than the LMS model.
However, basing the modeling on a mechanistic understanding of the biochemistry of TCDD-
induced toxicity should increase our confidence in the resultant risk estimates. As previously
stated by Greenlee and collaborators (1991):
Neither the position taken by U.S. EPA or by Environment Canada (and
several other countries such as Germany and the Netherlands) is based on any
detailed mechanistic understanding of receptor-mediated interactions between
dioxin and target tissues. Biologically-based strategies use knowledge of the
mechanistic events in the various steps in the scheme for risk assessment.
Interspecies extrapolation strategies would be conducted based on how these
mechanistic steps vary from species to species. There are numerous steps that
can be examined mechanistically, and fairly ambitious programs have been
proposed to examine the mechanistic details of many or most of these
individual steps. More focused risk assessment approaches are also being
proposed based on examination of individual steps believed to be critical in
establishing the overall shape of the dose-response curve for the induction of
tumors (or other toxic end points) by dioxin.
This chapter examines several end points and focuses on dose-response models for
cancer in laboratory animals and humans. It also evaluates the use of biomarkers of TCDD
action as surrogates for modeling receptor-mediated events. In addition, the chapter presents
a section titled Knowledge Gaps. Critical examination of this section can lead to new
experiments that will add to our knowledge of TCDD action. Increased information on key
molecular, cellular, and tissue-specific events will be important to validate a new risk
paradigm for TCDD and perhaps other receptor-mediated nonmutagenic toxicants. This
chapter focuses on cancer risk modeling. We review some critical data on noncancer end
points, but we do not attempt to model them. These end points are clearly important when
considering the public health risk of dioxins. However, the lack of knowledge on molecular
action and molecular dosimetry limits our ability to propose mechanistically based
mathematical models of noncancer end points at this time.
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Mathematical modeling can be a powerful tool for understanding and combining
information on complex biological phenomena. The development and use of mechanistically
based mathematical models are illustrated by Figure 8-5. The beginning point is generally a
series of experiments studying a xenobiotic agent. The experimental results (data) can
indicate a mechanism, which leads to the creation of a mathematical model. The model is
used to make inferences that are then validated against the existing knowledge base for the
agent and effect under study. This can then lead to new experiments, which support further
laps through the model development loop. On each pass through the loop, the model either
gains additional validation by predicting the new experimental results or it is modified to
encompass new results without sacrificing its base in previous results. In either case,
subsequent loops through the model generally increase our confidence in accepting (or
rejecting) a final model (although it may be difficult or impossible to quantify this
confidence).
Confidence in any one model is not only dependent on the information available for
that compound but also supported by the information available on other systems that act
similarly and for which models have already been developed. In the case of TCDD, the
modeling of effects will be greatly enhanced by existing information on the receptor-based
systems and general work in physiologically based pharmacokinetic models and in tumor
incidence modeling.
There is no one model development loop for any given compound or effect. Instead,
there are unusually numerous pathways leading to the development of a mechanistic model.
In modeling the effects of TCDD, there are many smaller model development circles that
make up the larger overall model. For example, a mechanistic approach to TCDD-induced
carcinogenicity must include models of exposure, tissue distribution, tissue diffusion, cellular
biochemistry, cellular action, tumor incidence, and cancer mortality. At each stage and for
each model, data must be collected and understood in order for the model to be valid and
acceptable as a tool for understanding the observed effects and for predicting the effects of
TCDD outside of the relatively limited range of experimental or epidemiological findings.
The use of mechanistically based modeling to extrapolate risks of exposure patterns
and doses outside the range of the data is in its infancy. Even though there may be high
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Experimental
Evidence
Experimental
Design
Model
Development
Model
Inference
Figure 8-5. Developing a mechanistically based mathematical model
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confidence in the ability of the model to predict experimental results, there could be low
confidence in the ability of the model to predict outside the range of data. The use of
models in risk assessment thus demands a careful scrutiny of the behavior of the model under
a variety of exposure scenarios. This scrutiny has not generally been applied in science to
the use of mechanistic models. It is important to note that mechanistic modeling has a role
to aid in explaining and understanding experimental results, separate from its use in risk
assessment. Our confidence in the methods used in mechanistic modeling will differ with
use. For a historical perspective, it is important to recognize that this is not the first loop
through the cycle of mechanistic modeling for carcinogenesis. Early risk assessments based
on tolerance distributions used the then current understanding of carcinogenesis to develop
statistical models that could be used for risk estimation (Mantel and Bryan, 1961). Later use
of the LMS model was based on an understanding of carcinogenic mechanisms. Thus, this
current exercise is a part of the process of mechanistically based modeling for risk
assessment purposes.
In any realistic and practical modeling exercise, the major components of the model
revolve around the estimation of model parameters utilizing statistical tools. These tools
range from very simple techniques, such as estimating a mean, to extremely complicated
approaches, such as estimation via maximizing a statistical likelihood. The estimation of
parameters is not done in a vacuum but is tied to the data available to characterize the model.
The way in which data are used to estimate those model parameters is the major component
in determining the confidence placed in any mathematical model. Fundamentally, sufficient
data need to be available to show that the model accurately represents at least what is known
of biological processes that are associated with toxic events.
In modeling biological phenomena, the data can be divided into four broad categories,
as shown in Table 8-1. At the top are effects observed in the whole animal. Examples of
data in this category are survival of the organism, ability to reproduce, and overall function
of the organism (e.g., behavioral data). Going from whole organism to tissue/organ system
responses to cellular responses down to biochemical responses in the cell, the levels of data
are increasingly specific and reductionist. The data range from the bottom of Table 8-1,
which is extremely mechanistic and deals with the interactions between molecules, to the data
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Table 8-1. Examples of Levels of Information Available for Estimating Parameters
in Dose-Response Modeling
Organism
Tissue
Cell
Biochemical
Morbidity
Mortality
Fertility
Improper development/function
Hyperplasia
Hypertrophy
Tumorigenesis
Chemical distribution/disposition
Mitosis
Cell death
Cytoarchitectural pathology
Gene expression
Protein levels
Receptor binding
Adduct formation
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at the top of Table 8-1, which is often no more than counting bodies. All of this information
is relevant and must be incorporated into a mathematical model aimed at understanding the
specific biological response.
Mathematical models that incorporate parameters that are mechanistic in nature do not
automatically constitute "mechanistic models." The types of data available for the model and
the method by which these data are incorporated into the model determine if a model is truly
"mechanistic," that is, soundly based on the biology rather than simply a curve fit to the
same data.
There are two basic ways in which biological effects can be estimated. The first and
most common approach is a "top-down" approach. In the top-down approach, data on the
effect of interest (e.g., carcinogenicity) are modeled directly by applying statistical tools to
link the observed data (e.g., tumor incidence data from a carcinogenicity experiment) to a
model (e.g., the multistage model of carcinogenesis). This approach is extremely powerful
in its ability to describe the observed results and to generate hypotheses about model
parameters and the potential effects of changes in these parameters. Where this modeling
approach begins to lack credibility is in its ability to predict responses outside the range of
the data currently being evaluated. Even when the model being applied to the data is
mechanistic in the sense that the model parameters are tied to some mechanism for the toxic
effect (e.g., mutation rates and molecular effects), without direct evidence concerning the
value for this parameter or even evidence supporting the particular structure of the model,
one is basically left with a curve fit to the data. The historical application of the LMS model
in risk assessment has been in this fashion.
True mechanistic modeling must be viewed in a different fashion. In this case, the
model structure and the parameters in the model are derived in a "bottom-up" fashion. The
mechanistic parameters in the model are estimated directly from mechanistic data rather than
from effects data or data one level higher in the hierarchy of data illustrated in Table 8-1.
The goal of true mechanistic modeling is to explain all or most known results relating to the
process under study in a way that is reasonable in its biology and soundly rooted in the data
at hand. In this case, biological confidence in predictions from the model would be much
higher than that from the "curve fitting" approach.
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In practice, it is generally impossible to completely eliminate curve fitting from
mechanistic modeling. At some point in the modeling process, gaps must be filled to relate
the mechanistic effects to the observed toxic effects. It is generally at this point that some
amount of curve fitting is necessary. Although not technically mechanistic modeling, this
combined approach is preferred to simple curve fitting when inferences outside of the range
of the toxic effects data are desired.
This is not intended to infer that with mechanistic modeling we can get a precise
estimate of risk of a toxic effect outside the range of the data or even necessarily a more
precise estimate of risk than with curve fitting alone. Without data, the statistical issue of
the accuracy of a prediction cannot be easily addressed. Thus, while there may be greater
biological confidence in extrapolated results, it is unlikely that an increased statistical
confidence can be demonstrated. However, for each level and type of data, there are ranges
of exposure beyond which it is impossible to demonstrate an effect given the practical
constraints on the experimental protocols. In general, effects can be demonstrated at lower
exposures for data at the bottom of Table 8-1 compared to the data at the top. If this is the
case, there may be both increased biological confidence in extrapolated results and increased
statistical accuracy. This is also not intended to infer that models derived through curved
fitting should always be given less weight. Many of the advances in biology and science are
due to the application (either formally or informally) of a model to data at a higher level in
order to generate hypotheses about effects at lower levels. The major difference between the
application of a "curve-fit" model in basic biology and that in risk estimation is that in basic
biology one is creating hypotheses which at some point can be tested. In risk estimation, it
is unlikely that one will ever be able to validate extrapolated risk estimates.
Many side issues are also related to the use of this model development loop in trying
to understand a biological mechanism. One of considerable importance is experimental
design. For mechanistic modeling aimed at risk assessment, we are just beginning to
understand the types of experiments that may assist us. In general design situations, one
would have a mechanism in mind, qualitatively describe that mechanism and form the
structure of a mechanistic model, and make educated guesses about the parameters of this
model, then use the quantitative model to develop experimental designs that are optimally
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relevant to characterizing the mechanism. For the purpose of risk estimation, this basic
outline holds. There are also some simple design rules that are not required but would aid in
the extrapolation of results to doses outside the observed response range and to humans from
animals. A few of these design points are listed in Table 8-2.
TCDD can be considered as a prototype for exploring and examining the ability of
mechanistic modeling to improve the accuracy of quantitative risk assessment. The database
for a mechanistic modeling approach to TCDD is very extensive and contains a considerable
amount of information on low-dose behavior. In addition, there is some concordance
between human data and experimental evidence in animals. On the other hand, some aspects
of the mechanism by which TCDD induces its effects, such as binding to the Ah receptor,
have not been modeled extensively, and thus we are in only the first few loops through the
model development cycle shown in Figure 8-5. Because of this, several competing
mechanistic theories may agree with the existing data, adding to the uncertainty in any
projected risk estimates. This outcome is inevitable for the application of the technology of
mechanistic modeling to a new area. To reiterate an earlier point, mechanistic modeling can
aid in explaining and understanding experimental results, beyond its proposed use in risk
assessment. Our confidence in the methods used in mechanistic modeling will differ
depending on its use. As we know more about the limitations of current data and current
methods for the application of mechanistic models to risk estimation, we can improve
experimental designs and significantly improve the process. Since this is an early attempt
along these lines, we must be cautious in coming to a judgment concerning the overall utility
of mechanistic modeling as an important tool for risk assessment.
In the National Academy of Sciences report Risk Assessment in the Federal
Government: Managing the Process (National Research Council, 1983), "dose-response
assessment" referred to the process of estimating the expected incidence of response for
various exposure levels in animals and humans. Tissue response is not always directly
related to exposure. This can be due to saturation and activation of metabolic pathways
(Hoel et al., 1983); influence of competing pathways having different efficiencies of action
for the parent compound and/or its key metabolites; and factors such as cytotoxicity,
mitogenesis, or endocrine influences that can radically modify the homeostatic state of the
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Table 8-2. Design Considerations for Risk Assessment Purposes
Aspects of risk estimation
Mechanism
Species extrapolation
Other
Design points to consider
Multiple times of observation
Multiple doses
Multiple ages at exposure
Pharmacokinetics
Multiple species
Both sexes
Tissue concentrations (including blood)
System clearance
Essential interactions with endocrine
systems
Multiple end points for individual
animals
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tissue. These complex interactions can result in markedly nonlinear dose response;
nonlinearities could lead to risk estimates that may be greater or less than the risk derived
from a linear model. Because of the potential for nonlinearities, it is essential to distinguish
between exposure level and dose to critical tissue or cell when modeling risks from exposure
to xenobiotics. It is also essential that we understand the quantitative relationship between
target tissue dose and changes in gene expression. This is especially true when extrapolating
to low doses and extrapolating across species.
For dioxin, the abundance of data on many levels allows one to create a collection of
models that include the determination of the quantitative relationship between dioxin exposure
and tissue concentration, tissue concentration and cellular action, cellular action and tissue
response, and finally tissue response and host survival (Portier et al., 1984). This portion of
the reevaluation of dioxin risks entails the description and development of mechanistically
based mathematical models of the effects of TCDD. This includes a discussion of the
extrapolation of tissue dosimetry and response from high-dose exposures to those expected at
much lower exposure based on empirical relationships used to derive explicit, though
incomplete, biologically based mechanistic models of the events involved in the toxic action
of dioxin.
Biological modeling is the process of developing mathematical descriptions of the
interrelationships among the mechanistic determinants of those toxic events resulting from
exposure to TCDD. Research with dioxin has focused on biological responses at the levels
of organization shown in Table 8-1 (i.e., biochemical, cellular, tissue, and organism). At
each level of organization, we focus on the mechanisms responsible for these observations.
Mechanism refers to the critical biological factors that regulate occurrence, incidence, and
severity of a particular factor and the nature of the interrelationships among these factors.
The details of the mechanisms of interaction differ markedly for the various levels of
biological organization with specific determinants of the behavior at each level driving the
creation on an appropriate quantitative model.
For dioxin, the mechanisms of three processes are of primary interest: (1) the
dosimetry of dioxin throughout the body and specifically to target tissues; (2) the molecular
interactions between dioxin and tissues, emphasizing the activation of gene transcription and
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increases in cellular protein concentrations of specific growth-regulatory gene products and
specific cytochromes; and (3) the progressive tissue-level alterations resulting from these
interactions that lead, eventually, to toxicity. The modeling process involves identification of
the mechanistic determination of the dose-response continuum through experimentation and
the encoding of these processes in mathematical equations. The extent to which model
predictions coincide with experimental results is a test of the validity of the model structure.
After validation, the model can be used for risk assessment. In addition to their use in risk
assessment, these models have importance for aiding in the design of future research, both in
terms of a basic understanding of dioxin toxicity and further risk estimation.
The following sections discuss the mechanistic biological modeling for dioxin with
regard to dosimetry, induction of gene transcription, and tissue response, especially those
associated with hepatic carcinogenesis. This modeling effort follows a natural progression
related to the kind of information available at the time at which the model was developed.
We will begin with a review of tissue concentration followed by modulation of protein
concentrations and tissue response.
8.1.2. Dose Delivery, Tissue Modeling, and Biochemical Modeling (see Appendices A
and B for complete manuscripts by Kohn et al., 1993 and Andersen et al., 1993b)
Tissue dosimetry encompasses the absorption, distribution, metabolism, and
elimination of dioxin from tissues within the body. The determinants of dioxin dosimetry in
the body include physicochemical properties (e.g., diffusion constants, partitioning constants,
kinetic constants, and biochemical parameters for metabolism) and physiological parameters
(e.g., organ flows and volumes). The mathematical structure that describes the
interconnections among these determinants constitutes a mathematical model for the tissue
dosimetry of dioxin. The model used in the context of this reassessment is a physiologically
based pharmacokinetic (PB-PK) model. These types of models describe the pharmacokinetics
of dioxin with a series of mass-balance differential equations. These models have been
validated in the observable response range for numerous compounds in both animals and
humans, making them extremely useful for risk assessment, especially for cross-species
extrapolation. In addition, they aid in extrapolation from one chemical to other structurally
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related chemicals since many of the components of the model can be deduced for structurally
related compounds. The development of PB-PK models is discussed for general use
(Gerlowski and Jain, 1983) and for use in risk assessment by Clewell and Andersen (1985).
In brief, a PB-PK model consists of a series of compartments that are based on the
anatomy and physiology of the test animals; hence, the term PB-PK. The time course of
behavior in each compartment is defined by an equation containing terms for input and loss
of chemical. For example, if Ct represents the concentration of compound in a tissue (()
and CB the concentration of compound in blood (B), one of the simplest relationships one
might use is:
" 'w CB - r/B C, - rm C, W
where dC(/dt represents the change in the concentration in the tissue over time (t), rB/ is the
rate (per unit concentration) of the movement of the compound from blood to tissue, r/B the
rate from tissue to blood, and rm the rate of metabolism in the tissue. Equations of this form
will be used in mass balance modeling of the pharmacokinetic processing of TCDD.
Several PB-PK models have been developed for dioxin and related chemicals (see
Chapter 1, Disposition and Pharmacokinetics, for a brief overview). PCBs have been
extensively studied (Lutz et al., 1977, 1984; Matthews and Dedrick, 1984). King et al.
(1983) modeled the kinetics of 2,3 7,8-TCDF in several species, and Kissel and Robarge
(1988) proposed a human PB-PK model.
The development of PB-PK models for TCDD began with work by Leung et al.
(1988) in mice. This model was extended to Sprague-Dawley rats by Leung et al. (1990a)
and to 2-iodo-3,7,8-trichlorodibenzo-p-dioxin in mice (Leung et al., 1990b). Since many of
the regulatory standards for dioxin have been based on a finding of hepatocarcinogenicity in
female Sprague-Dawley rats, we will focus on the model by Leung et al. (1990a) in this
strain and species.
The Leung et al. (1990a) PB-PK model contains five tissue compartments including
blood, liver, fat, slowly perfused tissue, and richly perfused tissue. This early model is
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blood flow limited, a condition that is appropriate when membrane diffusion is much more
rapid than blood flow to the tissue. Thus, in this PB-PK model, the tissue/tissue blood
compartments are lumped together as a single compartment in which the effluent venous
blood concentration of TCDD is equilibrated with the tissue concentration. The model
includes a TCDD-binding component in blood described by a linear process with an effective
equilibrium between the bound and free TCDD given by a binding constant, K^. It also
includes binding of TCDD to two classes of protein in the liver: one corresponding to the
high-affinity, low-capacity Ah receptor and the other to a lower affinity, higher capacity
microsomal protein (CYP1A2), which is inducible by TCDD. In the formulation of this PB-
PK model by Leung et al. (1990a), both types of binding proteins are explicitly defined using
an instantaneous correspondence between occupancy and induction using separate binding
capacities and dissociation constants for each protein. These binding reactions are modeled
via Michaelis-Menten equations.
In the Leung et al. (1990a) model, the tissue storage capacity depends on the partition
coefficients (assumed to be linear with concentration) and the specific protein binding.
Dioxin is very lipophilic and is found in higher concentrations in liver than would be
expected based on partition coefficients. The incorporation of a term for specific binding of
dioxin to a liver protein used by Leung et al. (1988) is a modification over earlier models for
these lipophilic compounds.
In various studies, dioxin has been administered by intravenous administration,
intraperitoneal injection, oral feeding or intubation (gavage), or subcutaneous (sc) injection.
In the PB-PK modeling framework, intravenous injection can be described by starting the
integration with an initial mass equal to the dose in the blood compartment. Oral intubation
and sc injection can be modeled as if they adhere to first-order uptake kinetics with dioxin
appearing in the liver blood after oral administration and in the mixed venous blood after sc
injection. Feeding was modeled by Leung et al. (1988, 1990a) as a zero-order input on days
that dioxin was included in the diet. These descriptions of the routes of uptake are clearly
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not defined in specific physiological terms; they are empirical attempts to estimate an overall
rate of uptake of TCDD into the PB-PK model. This is one area in which additional
research could improve dose-response modeling for TCDD. Efforts to provide more
biological details concerning the physiological basis of absorption across these various
membranes, including intact skin, would prove valuable for exposure assessments with
dioxin. With the iodinated analog, 2-iodo,3,7,8-trichlorodibenzo-p-dioxin, the estimated rate
constant for oral absorption was considerably larger in induced (0.15/hour) than in naive
animals (0.4/hour). The physiological basis of this change is unknown.
With many volatile organic chemicals there are convenient in vitro methods for
estimating partition coefficients (Gargas et al., 1989). For TCDD and other highly
lipophilic, essentially nonvolatile compounds, there are no reliable in vitro methods and these
constants have to be estimated from measurements of tissue and blood concentrations in
exposed animals. This leads to a difficulty in differentiating between specific-tissue binding
and the partitioning to the tissue. Leung et al. (1990a) overcame this problem by assuming
binding occurred only in the liver and that the liver partition coefficient was the same as the
kidney. This permitted estimation of the relative binding capacities and affinities of specific
hepatic proteins. The predictions from this modeling exercise prompted a series of
experiments to examine the nature of these binding proteins in mice (Poland et al.,
1989a, b).
Metabolic clearance was modeled as a first-order process. In the mouse with the
iodo-derivative, dioxin pretreatment at maximally inducible levels caused a threefold increase
in the rate of metabolism. There is no evidence to suggest an increase of metabolism in the
rat for TCDD. Chapter 1, Disposition and Pharmacokinetics, discusses pathways for TCDD
metabolism.
Finally, Leung et al. (1990a) kept all physiological parameters (e.g., flow rates, tissue
weights) constant over the lifetime of the animal.
Dioxin and dioxin analogs have dose- and time-dependent kinetics in both rodents
(Kociba et al., 1976; Poland et al., 1989a; Abraham et al., 1988; Rose et al., 1976;
Tritscher et al., 1992) and humans (Pirkle et al., 1989; Carrier, 1991). For single- and
short-duration exposures, as the exposure level increases, the proportion of total dose found
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in the liver increases. For chronic exposures, there appears to be a linear relationship
between dose and tissue concentration in the gavage study of Tritscher et al. (1992), but this
may be simply an inability to observe nonlinearity in liver in intermediate dose ranges. The
Leung et al. (1990a) model adequately predicts the tissue concentrations observed by Rose et
al. (1976) but did considerably worse at predicting the results observed by Kociba et al.
(1976), underpredicting concentrations at the lowest dose by a factor of 3.2 and
overpredicting concentrations at the highest dose by a factor of 2. The data of Abraham et
al. (1988) and Tritscher et al. (1992) were not available at the time this model was
developed, but at least for the data of Tritscher et al. (1992), this model has been shown to
overpredict the tissue concentrations (Kohn et al., 1993).
As mentioned earlier, the default position of the EPA in estimating risks from
exposure to xenobiotics involves the use of a model that produces risk that is proportionate to
dose for low doses (low-dose linearity). Thus, in discussing the models and submodels that
form a basis for a mechanistic model for TCDD, we will focus on aspects of the model that
could lead to nonproportional response for low environmental doses. The model of Leung et
al. (1990a) predicts slight nonlinearity between administered dose and tissue concentration in
the experimental dose range. In the low-dose range, the model predicts a linear relationship
between dose and concentration. (They argue, however, that tissue dose alone should not be
used for risk assessment for TCDD due to the large species specificity in the ability of
TCDD to elicit toxicity). They instead suggest that use of time-weighted receptor occupancy
linked with a two-stage model of carcinogenesis as a better approach to risk estimation. The
time-weighted receptor occupancy predictions derived from the Leung et al. (1990a) model
are linear in the low-dose region, reaching saturation in the range of high doses used to
assess the toxicity of TCDD. Comparison of this proposed effective dose with the two-stage
model of carcinogenesis is presented later in this chapter.
Looking at one aspect of modeling TCDD's effects, Portier et al. (1993) examined the
relationship between tissue concentration and the modulation of the three liver proteins by
TCDD in intact female Sprague-Dawley rats. The proteins studied included the induction of
two hepatic cytochrome P-450 isozymes, CYP1A1 and CYP1A2, and the reduction in
maximal binding to the EOF receptor in the hepatic plasma membrane. The modulation of
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these proteins is believed to be mediated through TCDD binding to the Ah receptor. Then,
as described in earlier chapters, through a series of alterations in the receptor-dioxin
complex, transport to the nucleus, binding to transcriptionally active recognition sites on
DNA, activation of gene transcription, and alterations in gene MRNA products, CYP1A1
and CYP1A2 are induced. Reduction in maximal binding to the EGF receptor requires
additional protein interactions.
General empirical models have been developed for the regulation of gene expression
(Hargrove et al., 1990). This modeling approach includes mRNA production by a zero-order
process and first-order degradation. Activation alters one or both of these rates. The
production of protein is assumed to be directly related to mRNA concentration. A more
specific pharmacodynamic model has been described to account for the induction of tyrosine
aminotransferase (TAT) activity by the corticosteroid prednisolone (Nichols et al., 1989). In
this induction model, the input prednisolone concentration is specified by the measured time
course of prednisolone in plasma. Prednisolone binding to receptor is specified by
association and dissociation rate constants. The prednisolone receptor binds DNA with a
specified association rate constant, and the bound receptor recycles to cytosol with a
transport time, T (effective compartment transport times are included to account for delays
between interaction with DNA and the appearance of TAT activity). A power function can
describe a nonlinear relationship between the concentration of prednisolone receptor and the
production rate of protein. The actions of prednisolone and maintenance of its tissue
concentration are much more short-lived than those of dioxin, and the modeling period of
interest is only on the order of several hours to a day instead of days, weeks, or months as
with dioxin.
The important relationships presented here are the association of dioxin with the Ah
receptor and the association of the dioxin receptor complex with DNA. As described above,
Leung et al. (1988) modeled the induction of CYP1A2 as being due to a basal amount of
protein plus an additional amount of protein resulting from binding of TCDD to the Ah
receptor. The extent of induction was calculated as instantaneously related to percent
occupancy of the Ah receptor via a Michaelis-Menten type relationship. Changes in
CYP1A1 and EGF receptor binding were not modeled by Leung et al. (1990a).
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Portier et al. (1993) modeled the rate-limiting step in the induction of CYP1A1 and
CYP1A2 following exposure to TCDD using a Hill equation. Hill equations are commonly
used for modeling ligand-receptor binding data. This equation allows for both linear and
nonlinear response below the maximal induction range. A complete discussion of Hill
kinetics and other models for ligand-receptor binding is given by Boeynaems and Dumont
(1980); examples of the use of Hill kinetics for ligand-receptor binding include the
muscarinic acetylcholine receptors (Hulme et al., 1981), nicotinic acetylcholine receptors,
opiate receptors (Blume, 1981), the Ah receptor (Gasiewicz, 1984), estrogen receptors
(Notides et al., 1985), and glucocorticoid receptor (Sunahara et al., 1989). As a direct
comparison to what was done by Leung et al. (1990a), it is interesting to note that the Hill
model can be thought of as a very general kinetic model that includes standard Michaelis-
Menten kinetics when the Hill exponent is 1. Portier et al. (1993) modeled the reduction in
maximal binding to the EOF receptor also following Hill kinetics, but with TCDD reducing
the binding from the maximum level from control valves. For all three proteins, proteolysis
was assumed to follow Michaelis-Menten kinetics. The proposed models fit the data in the
observable response range.
The major purpose of the paper by Portier et al. (1993) was to emphasize the
importance of endogenous protein expression on the shape of the tissue concentration/
response curve. For each protein, they considered two separate models. In the first, the
additional expression of protein induced by TCDD is independent of the basal level
expression. Such a mechanism is similar to that used by Leung et al. Under this model
protein expression is given by the equation:
. „
at p
where P is the concentration of protein in the liver, Bp is the basal rate of production of
protein, Vm is the maximal level of induction of protein by TCDD, Kd is the apparent
dissociation constant for binding (in the rate-limiting step), C is the concentration of TCDD
in the tissue, Vp is the maximal rate of proteolysis, Kp is the proteolysis rate constant, and n
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is the Hill exponent. When the Hill exponent is estimated to be an integer, the estimate of n
can be interpreted as corresponding to the effective number of binding sites that must be
occupied for the effect of the binding reaction to be expressed. When the Hill exponent is
not an integer, other molecular meanings apply. These are discussed in detail in Boeynaems
and Dumont (1980).
The second model Portier et al. considered was one in which the basal expression of
these proteins was due to a ligand that competed with TCDD for binding sites. This leads to
equations of the form:
Vm (C+E)n Vp P
Kd + (C+E)n Kp
[3]
where E refers to the concentration of this ligand in units of TCDD binding-affinity
equivalents. Under steady-state conditions, equations (2) and (3) are simplified (Portier et
al., 1993).
Using these simpler formulas, they see virtually no difference between the
independent and additive models in the observable response range, even to the point of
getting almost equal Hill coefficients in the two models for all three proteins. In the low-
dose range where risk extrapolation would occur, the models differed depending on the value
of the Hill coefficient.
In all cases, the additive model resulted in low-dose linearity. This is expected,
since, under the additive model, each additional molecule of TCDD adds more ligand to the
pool available for binding and thus increases the concentration of protein. Similar
observations have been made with regard to statistical (Hoel, 1980) and mechanistic (Portier,
1987) models for tumor incidence. For CYP1A1, the Hill exponent was estimated to be ~2.
When the Hill exponent is > 1, the independent model yields a nonlinear dose-response that
is concave (threshold looking). For CYP1A2, the Hill exponent was estimated to be -0.5.
When the Hill exponent is estimated to be < 1, dose-response is again nonlinear, but in this
case it is convex, indicating greater than linear increases in response for low doses. Finally,
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for the EOF receptor, the Hill exponent was approximately 1, in which case the two models
are identical.
Thus, even though these two basic models show almost identical response in the
observable response region, their low-dose behavior is remarkably different. If either
CYP1A1 or CYP1A2 levels had been used as dose surrogates for low-dose risk estimation,
the choice of the independent or additive model would make a difference of several orders of
magnitude in the risk estimates for humans. Using CYP1A1 as a dose surrogate, the
independent model would predict much lower risk estimates than the additive model. For
CYP1A2, the opposite occurs. For EOF receptor, there would be no difference.
Andersen et al. (1993b) modified the model of Leung et al. (1990a) to include Hill
kinetics in the induction of CYP1A1 and CYP1A2 and to use a diffusion limited approach to
the development of a PB-PK model as compared to the blood flow limited approach used by
Leung et al. Diffusion limited modeling is preferred when diffusion into a tissue is less
rapid than blood flow to a tissue. In the model used by Andersen et al. (1993b), each tissue
has two subcompartments, the tissue blood compartment and the tissue itself. Free TCDD
flows into the tissue blood compartment and, from there, diffuses into the tissue. There is
no direct relationship between effluent venous concentrations and tissue concentration in this
diffusion limited model. For TCDD, the diffusion limited approach is preferred due to the
compound's potentially slow diffusion into the liver from blood (Kohn et al., 1993).
Binding of TCDD to the Ah receptor was modeled in a fashion identical to that used
by Leung et al. (1990a). The concentration of CYP1A2 was modeled as before using a
steady-state model, in this case, with Hill kinetics instead of a Michaelis-Menten model. The
resulting equation is identical to that used by Portier et al. (1993) for the independent
induction of CYP1A2 except that they related this to the concentration of Ah receptor/TCDD
complex rather than the concentration of TCDD in the liver. Since they assume binding of
TCDD to the Ah receptor follows Hill kinetics with a Hill coefficient of 1 (Michaelis-Menten
kinetics), the model of Andersen et al. (1993b) approaches the independent induction model
of Portier et al. (1993) for low doses.
The induction of CYP1A1 was modeled as a time-dependent process as in equation
(2), again utilizing TCDD bound to the Ah receptor rather than tissue concentration of
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TCDD. Most of the physiological constants and many of the pharmacological and
biochemical constants used in the Leung et al. (1990a) model were changed for the Andersen
et al. (1993b) model to correspond to Wistar rats. The parameters in the model were
optimized to reproduce tissue distribution and CYPlAl-dependent enzyme activity in a study
by Abraham et al. (1988) and liver and fat concentrations in a study by Krowke et al.
(1989). For the longer exposure regimens and observation periods, changes in total body
weight and the proportion of weight as fat compartment volume were included via piecewise
constant values (changes occurred at 840 hours and 1,340 hours).
Andersen et al. (1993b) noted that the liver/fat concentration ratio changes as dose
changes due to an increase in the amount of binding protein in the liver. For high doses in
chronic exposure studies, this introduces a nonlinearity into the concentration of TCDD in
the liver. In the low-dose region, because the Hill coefficient for CYP1A2 and for the Ah
receptor are equal to 1, the liver concentration as a function of dose is still effectively linear
(i.e., small doses of TCDD will bind to the Ah receptor increasing the amount of Ah
receptor/TCDD complex, which then induces additional production of CYP1A2, which can
bind to free dioxin). In the observable response range, there is a slight nonlinearity in the
concentration of TCDD in the liver as a function of dose under chronic exposure (Andersen
et al., 1992). This nonlinearity in the dose region of 1 to 100 ng/kg/day does not agree with
the findings of Kociba et al. (1976) and Tritscher et al. (1992) for chronic exposure in
Sprague-Dawley rats. The plateau in total liver concentration predicted by the model of
Andersen et al. (1993b) does occur in the data of Kociba et al. (1976) and Tritscher et al.
(1992), in the range of 100 ng/kg/day consistent with the 87 ng/kg/day predicted by
Andersen et al. (1993b). However, the changes in liver/fat ratio observed by Andersen et al.
(1993b) and supported by human evidence (Carrier, 1991) are a necessary part of the
modeling for TCDD.
Finally, with regard to risk estimation, Andersen et al. (1993a) compared the
induction of CYP1 Al and CYP1A2, the concentration of free TCDD in the liver, and the
total concentration of TCDD in the liver to tumor incidence (Kociba et al., 1976) and the
volume of altered hepatic foci (Pitot et al., 1987). Using a biweekly dosing regimen, TCDD
was injected intramuscularly in female rats for 6 months. The concentration of TCDD in the
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liver and the concentration of induced protein were integrated over time to get summary
measures of internal exposure. They concluded that tumor promotion correlated more
closely with predicted induction of CYPlAl than the other integrated quantities. No formal
measure was used to support this observation. Thus, without a formal measure of
agreement, it is difficult to know if CYP1A2 induction can be rejected as correlating with
tumor response or focal lesion volume. This is discussed later in this chapter. Finally, the
choice of an independent induction model for CYPlAl and a Hill coefficient > 1 leads to
nonlinear low-dose behavior. If the promotional effects of TCDD follow a similar
mechanism, the risk from exposure at low doses will be negligible. For risk assessment, it is
important to know if an additive model also fits these data and agrees with the promotional
effects of TCDD since such a model will have different low-dose behavior than the
independent model.
Kohn et al. (1993) expanded upon the model of Leung et al. (1990a) to include Hill
kinetics, a restricted flow-limited PB-PK formulation, and an extensive model of the
biochemistry of TCDD in the liver. The goal of the model was to explain TCDD-mediated
alterations in hepatic proteins in the rat, specifically considering CYPlAl, CYP1A2, Ah
receptor, EGF receptor, and estrogen receptor over a wide dose range. In addition, the
model describes the distribution of TCDD to the various tissues, accounting for both time
and dose effects observed by other researchers. The PB-PK models developed by Leung et
al. (1990a) and Andersen et al. (1993b) relied on several single-dose data sets (Rose et al.,
1976; Abraham et al., 1988) and were validated against dosimetry results from longer term
subchronic and chronic dosing regimens (Kociba et al., 1976, 1978; Krowke et al., 1989).
These and more recent studies in which female Sprague-Dawley rats received TCDD
(Tritscher et al., 1992; Sewall et al., 1993) were used by Kohn et al. (1992) to model the
pharmacokinetics and induction of gene products in this sex and species. Among the data
reported by Tritscher et al. (1992) and Sewall et al. (1993) were (1) concentrations of TCDD
in blood and liver and (2) concentrations of hepatic CYPlAl and CYP1A2 and EGF receptor
in the hepatocyte plasma membrane. Kohn et al. (1993) refer to their model as the NIEHS
model. The tissue dosimetry for the NIEHS model was validated against the single dose and
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chronic dosing regimen experiments employed by Leung et al. (1990a) and Andersen et al.
(1993b).
In the biochemical effects portion of the NIEHS model, the binding of TCDD to the
Ah receptor is modeled using explicit rate constants instead of dissociation equilibrium
constants (equation 2 with Bp = 0). However, larger dissociation rates (kj, Iq) were used
leading to a formulation of the amount of TCDD-Ah-receptor complex similar to that used by
Leung et al. (1990a) and Andersen et al. (1993b) Many of the other binding reactions in the
model were handled similarly (e.g., TCDD binding to CYP1A2 and TCDD bound to blood).
This is simply a numerical trick to avoid the necessity of solving for the concentration of
TCDD in the liver using the mass conservation relationship described in Leung et al.
(1990a).
The physiology described in the NIEHS model is dependent on the body weight of the
animal. Body weight changes as a function of dose and age were recorded by Tritscher et
al. (1992) and directly incorporated into the model via a smooth function. Tissue volumes
and flows were calculated as an allometric formula based on recent work by Delp et al.
(1991). To allow the model to fit the data of both Rose et al. (1976) and Tritscher et al.
(1992), the NIEHS model includes loss of TCDD from the liver by lysis of dead cells where
the rate of cell death (and the resulting lysis) was assumed to increase as a hyperbolic
function of the cumulative exposure in the liver to unbound TCDD. No information
regarding the rate of TCDD from lysed cells is available; therefore, this feature of the
NIEHS model predicts a net contribution of TCDD clearance by TCDD-induced cell death.
In the biochemical effects portion of the NIEHS model, the Ah-receptor/TCDD
complex up-regulates four proteins; CYP1A1, CYP1A2, the Ah receptor, and transforming
growth factor-a (see Figure 1 in Appendix A). For all four proteins, synthesis and
degradation rates are defined explicitly. Changes in CYP1A1, CYP1A2, and the Ah receptor
are compared to data on these concentrations. The induction of EGF-like peptides is deduced
from observations on human keratinocytes (Choi et al., 1991; Gaido et al., 1992) and is
quantified based on an assumed interaction with the EOF receptor. However, TCDD-
mediated induction of TGF-a has not been clearly demonstrated in liver (see Appendix A).
Constitutive rates of expression for CYP1A2, Ah receptor, and EGF receptor are assumed
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independent (equation 2) at the induced expression. This has no effect on low-dose rate
extrapolation since the Hill coefficients for the induction of these proteins by the Ah
receptor/TCDD complex were estimated to be 1. Induction of CYP1A1 was assumed to be
based on additive induction (equation 3), but again the Hill exponent was estimated to be 1,
leading to low-dose linearity under either model equation (2 or 3). Thus, the NIEHS model
found that the induction of all gene products appears to be a hyperbolic function of dose
without any apparent cooperativity (i.e., the value of the Hill exponent, n, in equation 2, is
estimated to be 1). The discrepancy in the estimates of the Hill exponents between this
model and the other models discussed (Portier et al., 1993; Andersen et al., 1993b; Kedderis
et al., 1992) is probably related to the inclusion of induction of the Ah receptor in the
NIEHS model.
In the NIEHS model, the Ah-receptor/TCDD complex down-regulates the estrogen
receptor. It is assumed that the estrogen receptor-estrogen complex synergistically reacts
with the Ah receptor/TCDD complex to transcriptionally activate gene(s) that regulate
synthesis of an EGF-like peptide. This term was introduced to partially account for the
observation of reduced TCDD tumor-promoting potency in ovariectomized females as
compared to intact female rats (Lucier et al., 1991). This mechanism, although supported by
some data (Clark et al., 1991; Sunahara et al., 1989), is speculative (Kohn et al., 1993) (see
Appendices A and B).
There are basically three levels of complexity of PB-PK models for the effects of
TCDD. First is the traditional PB-PK model by Leung et al. (1988) with the added
complexity of protein binding in the liver. The next level of complexity is the model by
Andersen et al. (1993b) using diffusion limited modeling and more detailed modulation of
liver proteins. Finally, there is the model of Kohn et al. (1993) with extensive liver
biochemistry. All three models have biological structure and encode hypotheses about the
modulation of liver proteins by TCDD. However, for gene expression, all three models fall
in between curve fitting and mechanistic modeling. In their derivation, the parameters were
estimated using dose-time-response data for protein concentrations and enzyme activity,
which are a direct consequence of gene expression. This constitutes curve fitting at this
level. However, the structure of the models is derived from qualitative information on the
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effects of TCDD; the PB-PK model, and even the biochemical model of Kohn et al. (1993),
predict protein concentrations using data sets that were not included in the original derivation
of the model and that were derived from designs other than those used to characterize the
original model. This constitutes a mechanistic validation of the original models and places
these exercises in the realm of partial curve fitting and partial mechanistic modeling.
In terms of low-dose risk estimation, all three models have limitations. The Leung et
al. (1988) model fails to reproduce the tissue concentration data from Kociba et al. (1976)
and Tritscher et al. (1992). This is probably due to the high concentration of liver-binding
protein (CYP1A2) predicted by this model.
The Andersen et al. (1993b) and Kohn et al. (1993) models use Hill kinetics to
describe at least some of the binding reactions. Hill's equations are based on a molecular
scheme of interaction in which it is assumed that there are n binding sites for the ligand and
that the reaction is stable in only two states: either completely unoccupied or fully occupied.
This implies that the association process is a (n+l)-molecular reaction. In the application of
this model to experimental concentration curves, very little molecular meaning can be
attributed to the resulting model due to the flexibility of this function with regard to dose-
response shape. In addition, there is no unequivocal relationship between an estimated value
of n and the existence of molecular interactions between binding sites. For example, two
binding sites may exist, but binding to one site produces a small effect, while binding to both
sites produces a much greater effect. This would lead to a noninteger value for n when
curve fitting. Considering the importance of the Hill coefficient in terms of low-dose
extrapolation (Portier et al., 1993) and considering its limitations in terms of biological
understanding of the sequence of molecular events involved in induction (Andersen et al.,
1993b), caution must be used when extrapolating to tissue dose regions outside of those
examined directly in the experiment.
Some of the mechanistic assumptions in these models are speculative. Many of the
binding and induction equations related to the Ah receptor/TCDD complex are encoded in
equations, but their exact nature and level of control at the molecular level are unknown.
This is true of CYP1A1, CYP1A2, the Ah-receptor, the estrogen receptor, and EGF-like
peptides. Also, the reduction in EOF receptor by internalization described in the model by
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Kohn et al. (1993) represents just one mechanism for its depletion. It is also possible that
the synthesis or degradation of this protein may be under direct control of the Ah receptor,
although TCDD does not alter mRNA levels for the EOF receptor in either human
keratinocytes (Osborne et al., 1988) or mouse liver (Lin et al., 1991), and EOF receptor
does seem to move from the plasma membrane to the cell interior following TCDD exposure
in female rats (Sewall et al., 1993).
It should be noted that the mechanistic models have the advantage of suggesting
experimental strategies for pursuing the hypothesis of action of TCDD. These models
propose specific mechanisms, which can be tested in the laboratory as a means of validation
of this model. For the purposes of risk estimation, one must be careful to recognize that
these models do not necessarily impart added confidence in low-dose risk estimates because
the mechanistic links between TCDD-mediated changes in gene expression and toxic
responses are not known.
8.2. TOXIC EFFECTS
8.2.1. Modeling Liver Tumor Response for TCDD
In this section, we have decided to model liver tumor response for TCDD because of
the availability of quantitative data on this end point. It should be remembered that dioxin
induces increased tumor yield at several sites other than liver (see Chapter 6, Carcinogenicity
of TCDD in Animals), and in humans dioxin exposure is associated with increased mortality
due to lymphomas, soft tissue sarcomas, and lung and stomach cancer (see Chapter 7,
Epidemiology/Human Data). We do not imply therefore by our selection of liver tumors that
this is the exclusive site of dioxin carcinogenicity. It may be that similar events occur in
other sites or that the events discussed below—such as estrogen interaction—are unique to rat
liver.
Long-term carcinogenicity studies in rodents have shown that TCDD is a potent,
complete carcinogen (Huff et al., 1991). The highest increase in yield of tumors in TCDD-
treated animals as compared to controls was in female rodents. As discussed above, there is
no evidence for conventional mutagenicity or DNA binding by TCDD. While TCDD clearly
alters gene expression, it appears to act through a hormone-like receptor that functions as a
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transcription regulator of specific genes. Moreover, TCDD clearly interacts with several
endogenous hormones at the molecular and organ level. In liver, there is clearly an
interaction between TCDD and estrogen. As noted above, the most sensitive animal for
cancer response is the female rodent. Ovariectomy also reduced the ability of TCDD to
"promote" or produce tumors in female rats pretreated with the mutagen DEN (Lucier et al.,
1991). These results are complex and, as noted above, may be relevant to liver only.
Our overall approach models the tumor response observed in rodents and compares
that modeled response to some of the biomarkers of exposure discussed in the section on the
biochemistry of TCDD. This approach will deviate from the pure mechanistic modeling
outline discussed in the introduction. Due to limitations in the data available for
characterizing the models we will employ, some of the parameters used in this modeling
exercise had to be obtained directly from the tumor incidence data. These parameters are
then compared to the relative changes we would have expected using the biomarkers of effect
and exposure that seem reasonable for this end point. Thus, this exercise falls in between
curve fitting and pure mechanistic modeling. Two basic types of data will be used to find
parameter estimates for the two-stage model of carcinogenesis: (1) data on the number and
size of focal lesions in the liver and (2) data on the incidence of liver tumors in a 2-year
feeding study.
The carcinogenicity data we will use are from a 2-year feeding study in male and
female Sprague-Dawley rats (Kociba et al., 1978). For female rats, the study used 86
animals in the control group and 50 animals per group in the three treated groups given doses
of 1, 10, and 100 ng/kg/day. The original pathology of the study recorded significant, dose-
related increases in tumor incidence in the lung, nasal turbinates, hard palate, and liver. The
original liver pathology has been reviewed several times, most recently by a group convened
by Sauer (1990). The data we will concentrate on in this analysis is the incidence of liver
adenomas and carcinomas (combined) based on the most recent pathology review. A
summary of these data is presented in Table 8-3.
There was a substantial reduction in survival in all experimental groups (including
controls) during the course of the study. Other studies have shown that correcting for this
drop can result in as much as a twofold change in the low-dose risk estimates (Portier et al.,
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Table 8-3. Administered Dose, Tumor Response, and Number at Risk of
Hepatocellular Neoplasms in Male Sprague-Dawley Rats From Carcinogenicity
Experiments of Kociba et al. (1978) Using the Pathology Review of Sauer (1990)
Administered dose (ng/kg/day)
Number with neoplasm
Number on study
Survival-adjusted number at risk
Lifetime tumor risk
0.0
2
86
57
0.035
1
1
50
34
0.029
10
9
50
27
0.333
100
18"
50
31
0.581
"All neoplasms in all groups were hepatocellular adenoma with the exception of four
carcinoma in this high-dose group.
bUsing the "poly-3" survival adjustment suggested by Portier and Bailer (1989).
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1984). A simple correction for survival differences (Portier and Bailer, 1989) was applied to
these data to present the risk summaries given in Table 8-3. In the analysis that follows, a
more rigorous statistical approach was employed.
The data on the number and size of premalignant lesions are from an initiation-
promotion study in female Sprague-Dawley rats (Tristcher et al., 1992). This study utilized
91 animals distributed into 10 groups, 5 of which received initiation doses of 175 mg DEN
per kg body weight and 5 of which received 1 ^cL saline per gram body weight as a control.
One saline group and one DEN group received no TCDD as a control for the remaining
groups. Animals were exposed to TCDD by gavage once every 2 weeks in doses that
averaged out to 3.5, 10.7, 35.7, and 125 ng/kg/day, respectively. Exposure continued for
30 weeks, after which the animals were sacrificed. Serial sections of the liver were stained
for placenta! glutathione-S-transferase (POST). PGST-positive foci were counted and their
areas recorded. Table C-l in Appendix C summarizes these data. More detailed discussions
of the experimental design and focal lesion data are given in Tristcher et al. (1992) and
Maronpot et al. (1993).
8.2.2. Tumor Incidence
In recent years, there has been a resurgence in interest in refining the mechanistic
representation of mathematical models of carcinogenesis. With few exceptions, the
mathematic modeling of carcinogenesis at the cellular level has on the use of the multistage
model. Theoretical discussions on these models began in the mid-20th century (Arley and
Iverson, 1952; Fisher and Holloman, 1951; Nordling, 1953). The first practical application
of models from this class was done by Armitage and Doll (1954). One major failure of the
Armitage-Doll model is a lack of growth kinetics of the cell populations (Armitage and Doll,
1957; Neyman and Scott, 1967; Moolgavkar and Venzon, 1979). Several researchers
proposed a second model, the two-stage model, which is illustrated in Figure 8-6.
The two-stage model assumes that carcinogenesis is the result of two separate
mutations, the first resulting in an intermediate cell population and the second resulting in
malignancy. Cells in the normal and intermediate populations are allowed to expand in
number via replication or reduce in number due to death or differentiation. There are
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N>
n
Normal
Cells
U
M-N-I
Intermediate
Cells
i a,
^I-M
Malignant
Cells
1
g
n
Figure 8-6. A two-stage model of carcinogenesis
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different mathematical approaches to this model since several groups have proposed the same
model but used different mathematical developments to predict tumor incidence from this
model (Armitage and Doll, 1957; Neyman and Scott, 1967; Moolgavkar and Venzon, 1979;
Greenfield et al., 1984). In the application of the two-stage model that follows, the
mathematical development of this model by Moolgavkar and Venzon (1979) and subsequent
development of this model will be used.
The two-stage model (Figure 8-6) has six basic rates that must be estimated. These
are:
1. j8N = birth rate for cells in the normal state.
2. dtf = death/differentiation rate for cells in the normal state.
3- MN-I = rate at which mutations occur adding cells to the intermediate state.
4. ft = birth rate for cells in the intermediate state.
5. 5j = death rate for cells in the intermediate state.
6- MI-M = rate at which mutations occur adding cells to the malignant state.
To apply this model to dioxin, or any other chemical carcinogen, requires estimates
of these rates as they change over dose and time. A mechanistic approach to this would be
to incorporate some of the relative changes in proteins seen in the NIEHS model directly into
the two-stage model as rate changes in these parameters. Considering the complexity and
novelty of this approach, it is left as a research topic. Instead, we will apply this model
directly to tumor incidence data, focal lesion data, and cell-labeling data comparing the
resulting parameter estimates to predicted dose-surrogates from the models of Leung et al.
(1988), Andersen et al. (1993b), and Kohn et al. (1993).
This is not the first application of TCDD data to the two-stage model. An application
of this model to TCDD was presented by Thorslund (1987). Thorslund treated the effects of
TCDD as a direct promoter having an effect only on the birth rate of intermediate cells
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in the two-stage model. The number of normal cells was assumed constant (this is equivalent
to setting /3N = SN = 0 in the model in Figure 8-6). Two parametric models of the change
in fa as a function of dose were used, one model having a single parameter (a first-order
kinetic or exponential model) and the second based upon two parameters (a log-logistic
model). The parameters in the exponential two-stage model were estimated from the tumor
incidence data of Kociba et al. (1978) and validated by goodness-of-fit, cell-labeling data,
and species/sex/strain extrapolations. The slope parameter in the log-logistic two-stage
model was chosen to be 1, 2, or 3 based on slopes observed in other biological systems.
The remaining parameters in this model were estimated from the Kociba et al. (1978) data.
The liver tumor responses from the Kociba et al. (1978) study are given in Table 8-3
using the most recent pathology review of the liver sections (Sauer, 1990). Shown are the
number of animals with tumor (row 2), number of animals placed on study (row 3), a
survival-adjusted number of animals at risk (row 4), and the survival-adjusted lifetime tumor
probability (row 5 which equals the entry in row 2 divided by the entry in row 4).
There are a variety of mathematical formulations that could be used to derive a tumor
incidence rate under the two-stage model. As has been assumed by other authors (Portier
and Kopp-Schneider, 1991; Moolgavkar and Luebeck, 1992), we assume:
1. All cells act independently of all other cells;
2. The rates in the two-stage model (Figure 8-6) are constant over the lifespan of
the animal;
3. The tumor incidence rate corresponds to the rate of appearance of the first
malignant cells; and
4. BN = SN = 0 (no change in the number of normal cells).
All four of these assumptions are likely to be violated for most chemicals. In most
tissues, there is a homeostatic feedback system to control the number of cells in the tissue.
No such system can be assumed here since it results in a mathematic formulation that is
either intractable or has yet to be developed. For the large pool of normal cells in the liver,
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this is unlikely to have an effect, but for the small number of intermediate cells (at least for
early times), this could have a small effect on tumor incidence. This issue cannot be
resolved without further research. Assumption (2) is clearly violated based on the behavior
of the PB-PK models presented earlier. Time-dependent changes in tumor incidence could
be incorporated into the modeling and will be at a later time. Finally, the kinetics of cell
growth for malignant clones have been studied (Moolgavkar and Luebeck, 1992), and
assumption (3) was found to have a moderate impact on the tumor incidence rates. This will
be investigated further at a later time.
The exact tumor incidence rate was used to avoid potential bias from the routinely
used approximation (Kopp and Portier, 1989). Thorslund (1987) used this approximation in
his analysis. The methods outlined by Portier and Kopp-Schneider (1991) employing the
Kolmogorov backwards equations were used to derive the exact tumor incidence rate.
It has been suggested that the cells in PGST-positive lesions correspond to the
initiated mutated cells in the two-stage model of carcinogenesis. In the following analyses,
we have assumed this to be the case, but it is important to keep in mind that the number of
enzyme-altered cells may not correspond to mutation rate. If this is true, it is possible to
apply the methods of DeWanji et al. (1989) to data from Lucier and colleagues (Tristcher et
al., 1992) to analyze the growth characteristics of these cells (Moolgavkar et al., 1990).
This can be combined with the tumor incidence data to produce a combined model that
explains both sets of data. This is done below. Analyses of the tumor incidence data alone
and the PGST-positive foci data alone are given in Appendix C.
The most parsimonious two-stage model that agrees with the tumor incidence data and
the focal lesion data (this is based on a combination of likelihood ratio testing and plots of
the model versus the data) is a hyperbolic function in dose for the effect of TCDD on the
rate of transformation from the normal state to the intermediate state, is hyperbolic in dose
for the birth rate of intermediate cells, is constant in the ratio between the death rate and the
birth rate for intermediate cells, and uses a power function in dose on the rate of mutation
from the intermediate state to the malignant state. The resulting model is given as:
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*»-'(dose) =
dose
15]
^(dose) = p/3(dose) [6]
/i!_M(dose) = n2+n3 dose'*4 C71
where XQ jt0=3.216 x 10"4, XQ /xv=6.856 x 10"4, /ik= 19.75, B0=3.029 x 10'2, By=4.296 x
10-3, Bk=2.75, p =0.00302, ^2= 1.496 x 10'7, ^=3.082 x 10'9, j*4=2.56, and dose is
expressed in ng/kg/day. This analysis also assumes that the average rodent liver has 6xl08
normal hepatocytes. The fit of this model to the number and size of focal lesions from the
studies of Lucier et al. is illustrated in Figures 8-7 and 8-8. It is clear from these plots that
this fit is adequate; this is supported by analyses based on the magnitude of the likelihood
(see Appendix C). The fit of this model to the tumor incidence data of Kociba et al. (1976)
(PWG, 1990) is shown in Figure 8-9.
It is possible to compare the parameters estimated for the two-stage model to
predictions from the PB-PK models to try to locate a reasonable mechanistic link between the
two classes of models, to aid in species extrapolation, and to help guide us in choosing the
most appropriate curvature for low-dose extrapolation. An overall comparison of the dose
surrogates is given in Appendix D. What is illustrated in Appendix D is that most
reasonable dose surrogates correlate well with the dose-response relationships given above
and in Appendix C. For the purposes of finding a most reasonable model for low-dose risk
assessment, we will use two specific dose surrogates from Appendix C.
In the NIEHS PB-PK model (Appendix A), it is hypothesized that induction of
CYP1A2 could lead to an increase in the metabolism of estrogens to catechol estrogens
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30
§
• i-H
t/5
a
9.
(N
C
20-
t/3
X)
O
Dose (ng/kg/day)
Figure 8-7. Fit of the two-stage model to the number of focal lesions from the data of Maronpot
et al. (1993).
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0.0 ng/kg/day
•9
MI a«
Area of Focal Lesion (mm2)
3.5 ng/kg/day
I
Area of Focal Lesion (mm2)
10.7 ng/kg/day
o.«r an
Area of Focal Lesion (mm )
35.7 ng/kg/day
Area of Focal Lesion (mm2)
125 ng/kg/day
Area of Focal Lesion (mm2)
Figure 8-8. Fit of the two-stage model to the size distribution for focal lesions from the data
of Maronpot et al. (1993) where the smooth line is derived from the model and the step
function results from the data.
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10 ng/kg/day
Control
1 ng/kg/day
100 200 300 400 500 600 700
Age (days)
Figure 8-9. Fit of the TS model to the data of Kociba et al. (1978) using parameters for
/3j(d), and 5j(d) estimated from the focal lesion data.
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(Graham et al., 1988) and that further activation of these catechol estrogens can lead to cell
damage (for example, via free oxygen radicals) and eventually to mutations. Thus, the area
under the time-concentration curve for induction of CYP1A2 could serve as a useful dose
surrogate for the mutational effects of TCDD expressed as equation (4). Note that this
relationship is speculative and that the application of dose surrogates to low-dose risk
estimation is subject to a number of untestable assumptions. This is discussed in detail in
Appendix D.
In this case, the shape of the response for the area under the time-concentration curve
for the induction of CYP1A2 parallels the hyperbolic effect given above. A better approach
to this analysis would be to use the instantaneous concentration of CYP1A2 as it changes
over time to directly modify the mutation rate. However, at this time, this approach is
impractical due to inadequate mathematical development and a lack of computer algorithms
and programs for addressing the problem. However, the parallel curves for CYP1A2 and the
mutation rate indicate that dramatic changes in low-dose response by going to the
time-varying dose surrogate is unlikely.
Kohn et al. (1993) also provided a potential mechanism for the proliferative effects of
TCDD on the cells in the intermediate state. For this process, they propose a mechanism
based on the incorporation of the EGF receptor in an activated state in the cell interior rather
than on the cellular membrane. However, the relationship between the area under the time-
concentration curve for the internalized EGF receptor does not correlate well with the
hyperbolic dose response seen in the growth characteristics of the intermediate cells in the
two-stage model given by equation (5) above. This could be due to a number of factors:
incorrect two-stage model specification, incorrect specification in the NIEHS model, or a
complex relationship between internalized EGF receptor and mitosis of the intermediate cells.
Either way, the simple dose surrogates expressed in Appendix D are inappropriate for the
parameterization of the two-stage model given above. Note that for both the mutation rate
and the mitotic rate, dose-response behavior is adequately described by a linear function in
the low-dose region.
The details of a complete analysis of the two-stage model and the Kociba et al. (1978)
data are given in Appendix C. The fit of the model given by (7) above to the data by Kociba
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et al. (1978) is shown in Figure 8-9. The fit is adequate and is not statistically significantly
different from the best-fitting two-stage model, which is nonparametric for /i!_M(d) and uses
the parameters derived from the focal lesion analysis. It is significantly different from the
best-fitting two-stage model in which all parameters are derived from the tumor incidence
data alone (see Appendix C).
Because /*4> 1, the relationship between dose and the mutation rate between the
intermediate state and the malignant state is sublinear (similar to threshold behavior). As
discussed in Appendix D, none of the dose surrogates for the NIEHS model exhibit this type
of behavior. In Andersen et al. (1993b), induced CYP1A1 has this type of behavior (this is
also seen in Portier et al., 1993). The difference between the Andersen et al. findings and
those of Kohn et al. (1993) are discussed in Appendix D and will not be repeated here.
Also, as explained in Appendix D, these correlations provide very little verification of an
actual mechanistic link between the dose surrogate and the tumor incidence rate.
Even with this nonlinearity in MI-M(
having first converted all administered doses into the dose surrogate, s, using CA(d).
Extrapolation would then be done in units of CYP1A2 induction, and extrapolation would
result in a predicted induction of CYP1A2, say eA. To estimate exposure doses in humans
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s
Probability of Tumor
d
O
t/5
O>
o
01
to
o
Ul
10
oo
o
Ul
o
Ul
U)
o
I.
I
Probability of Tumor
o
to
o
*».
o
en
o
CO
I
•S
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would then require knowledge of the induction of CYP1A2 by TCDD in humans, say
sH=CH(d). The predicted human exposure dose (presence of induction) would be
dH=CH"1(eA). It is important to reemphasize here that enzyme induction is not equated to
toxic response.
When multiple dose surrogates are used, the equations become more difficult due to
the multivariate nature of the relationship. For example, if we were using CYP1A2
induction as a dose surrogate for fi^.i(d) and removal of EGF receptor from the plasma
membrane as a dose surrogate for B^d), our functional forms would have to be concerned
with the changes in both of these proteins as a function of dose. However, the basic idea
remains the same, and inversion of the formulae is used to estimate safe exposure levels.
Since there are, as yet, no comprehensive PB-PK models for humans, this approach could
not be used here. Instead, the results of the PB-PK model are used to support dose-response
relationships seen for the tumor incidence data and the focal lesion data.
The final model resulting from the combined analysis of the tumor incidence data and
the focal lesion data has several shortcomings. One serious problem concerns the size of the
intermediate cell population after 2 years of exposure to TCDD. The model does an
excellent job of describing the size and number of focal lesions in the liver at 31 weeks; but,
by 2 years, since the same rates seen in the first 31 weeks are assumed constant for 2 years,
the model predicts impossibly large foci in the liver (the model predicts that the mean
number of intermediate cells is two times larger than a normal rodent liver in the high-dose
group). The data available for this analysis consist of single time point data so there is no
data-driven way in which to estimate changes in the model parameters (mutation, birth and
death rates) over time. Arbitrary changes could be used; however, these are unlikely to
change the shape of the dose-response curve for MI-M(^) as lon§ as me changes are applied
proportionately across all treatment groups.
Additional data will be required to support any improvements in this model.
Additional focal lesion data (size and number) at a time point other than 31 weeks would be
very useful in terms of looking at time-dependent changes in the rates in this model.
Considering the size and number of saline-treated animals with foci at 31 weeks, it would
probably be best to have information on focal lesions at a later time. Because the
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proliferation rates from the focal lesion data match the observed labeling index data from
Maronpot et al. (1993), the simplest change that will lead to reasonable numbers of initiated
cells is a change in the ratio of the death rate to the birth rate for initiated cells. As the
animal ages, this rate will have to converge to 1 or less for the expected number of initiated
cells to have a reasonable size at 2 years. If the change over time is proportionately the
same over time in all dose groups, it should have little or no effect on the shape of /*i.M(d);
the only effect should be on the magnitude of this mutation rate (increasing it). Any
dose-dependent changes in this ratio may alter the dose-response shape of /xx.M(d) and hence
also alter the low-dose estimates of risk.
One final note on the final version of the two-stage model fit to the tumor incidence
data and the focal lesion data: In the low-dose region, the dose-response curve for added
risk from exposure to TCDD can be approximated by the formula risk=0.024 dose where
dose is in units of ng/kg x bw/day. The formula can be inverted to estimate a safe exposure
dose from these data. For an added risk of 1 in 1 million in the animals, the calculated
exposure dose is 41.7 fg/kg x bw/day. This is a point estimate of exposure dose; confidence
intervals would reduce this value.
We have described modeling as a continuing process. One must remember that, as a
continuing process, modeling allows us to test certain hypotheses developed from a particular
model in hopes of designing experiments that can either support or refute the model.
Modeling also allows us to determine how well a particular hypothesis, derived from
biological evidence, can explain the observed phenomena. One of the models presented here
has analyzed two data sets and has presented a hypothesis that dioxin increases, in a dose-
dependent manner, the rate at which normal cells become initiated cells, thereby increasing
the number of initiated cells. If the model is correct, then dioxins are acting to increase the
mutation rate in hepatocytes, thus increasing the number of initiated cells. In essence,
dioxins are acting as indirect mutagens.
The impetus in developing a biologically based dose-response model for dioxin was
that there is no evidence that dioxins are direct mutagens and that the carcinogenic actions of
dioxins were due to their actions as receptor-mediated tumor promoters. In theory, tumor
promoters do not alter the mutation rate and act only on previously initiated cells. In
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addition, the actions of a tumor promoter are reversible. If we compare this theory to that of
the model's prediction of dioxin's actions, we are left with several contradictions. Either the
structure of the model has led us astray or the actions of dioxins increase the mutation rate
through indirect mechanisms and that these effects are not detectable in short-term assays.
There is consensus that dioxins are not directly mutagenic, and there are very few
data that demonstrate that dioxins damage DNA directly and/or are mutagenic in short-term
assays. (See Chapter 6 for detailed discussion.) However, a recent study suggests that the
short-term assays may not be sensitive to the potential indirect mutagenic effects of dioxins.
Yang et al. (1992) have reported that exposing nontumorigenic immortalized human
keratinocytes to TCDD for 2 weeks transforms these cells into tumorigenic cells. Exposures
for less than 2 weeks do not result in transformation of this cell line. These experiments
indicate that exposure to dioxins can result in permanent changes in the phenotype of human
cells. While the mechanism by which this occurs is undetermined, it is clear that these
permanent alterations require prolonged exposures in order to be detected. These permanent
changes are contradictory to the hypothesis that dioxins act solely on the promoter stage in
the process of carcinogenesis.
If dioxins were to increase mutation rates through indirect mechanisms, what actions
of dioxins would increase the likelihood of a mutational event? One possibility is that by
increasing CYP1A1 and CYP1A2 activity, dioxins increase the production of reactive
intermediates. Evidence to support this hypothesis is that CYP1A2 metabolizes estradiol to
the catechol estrogen 2-hydroxyestradiol. Catechol estrogen formation may lead to DNA
damage and may be responsible, in part, for the carcinogenic effects of estrogens.
Dioxins may also have indirect mutagenic actions through increased hepatocyte
proliferation. Cells are constantly exposed to reactive intermediates that may cause DNA
damage. Inappropriate stimulation of cell proliferation, by dioxins, may produce an apparent
increase in the number of initiated cells in a tumor promotion study by forcing cells with
nonlethal DNA damage to replicate prior to the repair of the damage. This could lead to
fixed mutations in the genome and clonal expansion of the transformed cells, resulting in an
increase in preneoplastic foci.
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Increased transcription of a gene can also lead to potential mutagenic events.
Glucocorticoids can produce a variety of genetic rearrangements through epigenetic
mechanisms. If cells transfected with a plasmid containing a glucocorticoid response element
and a recorder gene are exposed to dexamethasone, a variety of mutagenic events such as
inversions, deletions, and insertions in the plasmid occur. In addition, antisense mRNA can
also be detected following exposure to dexamethasone. This study demonstrates that
increased transcription of the recorder gene by dexamethasone increased the likelihood of a
mutational event in the recorder gene through epigenetic mechanisms. Dioxins act in a
similar manner to dexamethasone and other steroid hormones. Both classes of chemicals
bind to intracellular receptors, and the activated receptor binds to response elements on
DNA, resulting in an increase in gene transcription. It is possible that continued exposure to
dioxins will increase the likelihood of a mutagenic event at a gene that is responsive to
dioxins due to the increased rate of transcription. Since dioxins alter growth properties of a
variety of different cell types such as hepatocytes (Lucier et al., 1991), skin (Stohs et al.,
1990), and embryonic palate (Abbott and Birnbaum, 1991), it is possible that some of the
genes that can be affected by this mechanism would be involved in the regulation of cellular
proliferation.
Chemical carcinogenesis is often divided into stages of initiation, promotion, and
more recently progression (reviewed in Pitot and Dragan, 1991). As the process of
carcinogenesis develops, the progression stage is an irreversible stage that is associated with
increase in karyotypic instability, which is associated with the development of aneuploid
malignant neoplasms (Pitot and Dragan, 1991). Thus, during the carcinogenic processes
there are steps postinitiation that result in karyotypic instability. It is possible that dioxins
are acting not only as promoters but also as progressors, and the mechanisms described
above may provide some insight into how dioxins act.
8.2.2.1. A Potential Alternative Model for Promotion of Carcinogenesis by Dioxin
PB-PK models permit estimation of tissue concentrations of dioxin and of various
protein products regulated either directly or indirectly by dioxin and the Ah receptor. In
developing a quantitative mechanistic model of the tumor responses to dioxin, it is necessary
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to propose some relationship between the concentration of these protein factors and the
critical cellular responses—mutation rates and cell proliferation rates—leading to tumor
induction. One hypothesis, consistent with dioxin's primary mechanism as a promoter, is
that dioxin, directly through effects on mitostimulatory hepatic growth factors, such as
transforming growth factor-a, causes enhanced cell replication in normal and/or initiated
cells. In some of the cancer modeling in this chapter, this approach has been followed at
least indirectly by assuming that the reductions in cell surface EGFR concentrations are an
indication of enhanced levels of an activating ligand for this receptor and that the reduction
in EGFR is a surrogate for increased cell proliferation rates.
Pilot et al. (1987) examined promotion of altered hepatic foci by dioxin in rats that
were initiated with a DEN dose of 10 mg/kg following 70% partial hepatectomy. These
altered hepatic foci are believed to be precursor lesions for overt hepatic malignancies. In
this study, doses used were equal to the bioassay doses used by Kociba et al. (1978) with an
additional low dose, 0.0001 /tg/kg/day. One important aspect of this study and the original
dioxin bioassay is the dose range covered, 100-fold in the cancer bioassay and 1,000-fold in
the promotion study. Pitot et al. (1987) found that the promotion dose-response behavior
was U-shaped; the two lower doses gave a lower response for volume of liver as foci or total
number of observed foci than did either the control or high-dose groups. This U-shaped
dose-response behavior was also observed in the liver tumor incidence in the bioassay study,
although the differences between control incidence and incidence at the lower doses were not
statistically significant. These observations indicate a more complex relationship between
dioxin exposure and response than expected for dose surrogates, whether dioxin
concentrations, cellular EGFR concentrations, or protein product concentrations, that increase
(or decrease) monotonically with dioxin concentration.
Although less marked than with dioxin, phenobarbital also had a U-shaped dose-
response behavior for hepatic promotion when examined with the same experimental protocol
over a dose range of 0.001 to 0.05% in the diet (Pitot et al., 1987). Jirtle and coworkers
(1991) have examined the biology of phenobarbital promotion in detail. They propose that
phenobarbital initially causes a mitogenic stimulus with proliferation of pericentral
hepatocytes. The liver responds to the persistent proliferative signal by increasing levels of
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mitoinhibitory growth factor, TGF-/31, and periportal hepatocytes stain intensely for TGF-j81.
Down-regulation, presumably without activation, of EGFR also occurs in phenobarbital-
treated livers. The hallmark of promotion in this hypothesis is the selection provided by the
persistent increases in TGF-/31 for initiated cells with particular alterations that decrease their
responsiveness to these inhibitory growth factors. Consistent with this mechanism, the
putative preneoplastic hepatocytes express significantly less TGF-/31 than do the periportal
hepatocytes (Jirtle and Meyer, 1992). The suppression in growth of the normal cells in the
tissue then is the critical alteration leading to promotion and really is more of a homeostatic
response than a direct response of the cell to dioxin. In this model of hepatic promotion, the
increases in mitoinhibitory factors serve to maintain the liver at constant size even in the
continued presence of proliferative stimuli provided by dioxin, phenobarbital, or other
xenobiotics that might share similar mechanisms of action. The clonal expansion of initiated
cells results from a growth advantage because of their relative insensitivity to the
antiproliferative environment produced by the promoter (Jirtle et al., 1991).
This general mechanism of promotion may also occur with dioxin (Mills and
Andersen, 1993) and could give rise to the U-shaped dose-response curve seen by Pilot et al.
(1987). Initiated cells might have differential sensitivity to apoptosis mediated by TGF-/81
(Oberhammer et al., 1992) and to proliferation induced by other factors regulated by dioxin.
If the apoptotic rates were increased at lower dose than doses that enhance proliferation, a U-
shaped curve would be expected. Another possibility is that clones observed in control
(PH/DEN; 0 /*g dioxin/kg/day) rats are derived from different precursor cells than are those
clones arising in livers of rats treated with higher daily doses of dioxin (>0.01 /Kg/kg/day).
In the control rats no net increase in TGF-/31 is expected. The initiated cells that become
foci in these rats are potentially responsive to the antiproliferative activity of TGF-/31 and, as
the dose of dioxin is increased, the death rates of these cells increase; fewer clones grow out
to become observable foci, and a lesser volume of the liver is occupied by these clones. At
higher dioxin doses, tissue TGF-/31 is increased and the selection environment favors growth
of a new population of initiated cells that are TGF-jSl nonresponsive. At intermediate doses,
the primary response is inhibition of the first type of initiated cells with a decrease in yield
of observable clones. At these doses there is not yet any increase in the appearance of
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clones arising from the second population of initiated cells. This behavior would lead to a
U-shaped dose-response curve. Early efforts to parameterize these promotion models
(Andersen et al., 1993a) indicate that the model assuming a different cellular origin of clones
in control versus high-dose dioxin rats appears the more likely of these two hypotheses at this
time.
8.2.3. Other Effects: Mammary/Uterine/Anticancer End Points
There are several lines of evidence that suggest that TCDD is anticarcinogenic in
some organs of experimental animals. The most compelling data are from the Kociba
bioassay (Kociba et al., 1978), which reported a dose-dependent decrease in several
endocrine tumors in females. The most remarkable changes occurred in the incidence of
carcinoma of the breast and uterus and all tumors of the pituitary. The decrease in these
tumors coincided with the increase in hepatic tumors in the same animals. Tumor inhibition
has also been reported in skin initiation-promotion experiments if TCDD administration
precedes the initiator. This TCDD-induced inhibition occurs even if the complete initiator-
promoter protocol is carried out following the TCDD administration (DiGiovanni et al.,
1977; Marks etal., 1981).
Mechanistic explanations for these observations are incomplete. Several laboratories
have attempted to elucidate the molecular events leading to the decrease in tumors
particularly in the breast and uterus. Three lines of evidence are under great scrutiny at this
time, including (1) alteration of the estrogen receptor, (2) alteration of estrogen metabolism,
and (3) alteration of binding of estrogen and estrogen receptor at the DNA level.
Down-regulation of the estrogen receptor has been reported by laboratories of Gallo
(Gallo et al., 1986; DeVito et al., 1990), Safe (Romkes et al., 1987; Astroff and Safe,
1988), and Gierthy et al. (1988). This down-regulation occurs in a rank order relationship
with the affinity of TCDD structural analogs to bind to the Ah receptor (Harris et al., 1990)
and parenthetically with the ability of the compounds to induce cytochrome P-450IA1. There
is differential regulation of the estrogen receptor (ER) by TCDD in uterus and liver (liver
being more sensitive), and the regulation of the uterine ER disappears in the adult animal
(DeVito et al., 1992). Collectively, these observations suggest that there are at least two
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mechanisms involved in ER regulation by TCDD. Gierthy has supported the argument that,
in MCF-7 cells, the alteration of estrogen metabolism following the induction of cytochrome
P-450 isozymes is the explanation for the decreased activity of the ER. Safe has reported
that the ED50 for decreased ER in MCF-7 mice cells is several orders of magnitude below
the ED50 for induction of 7-ethoxyresurofin-O-deethylase (EROD) in these same cells
(Zacharewski et al., 1991). In vivo studies suggest that the uterine response is more
responsive to changes in circulating estradiol (which may or may not hold up for breast
tissue), but the hepatic ER response to TCDD is independent of circulating estradiol levels
(DeVito et al., 1992). Another possibility is that TCDD induces enzymes that qualitatively
change in situ estrogen metabolism and that these novel metabolites (or families of
metabolites) alter estrogen action. Recent evidence suggests the presence of a 15-alpha
metabolite of estradiol in liver microsomes, from TCDD-treated females, incubated with
estradiol. This observation is similar to that of Gierthy et al. (1988) in MCF-7 cells.
Another possible mechanism of the decrease in breast cancer is that the binding of the
activated AhR complex to a DRE on the DNA is in the upstream region of the ER gene or is
in the vicinity of the binding region of the activated ER. Several lines of evidence also
support this hypothesis since there is ample evidence that other steroid receptor complexes
modify the DNA binding of different members of the steroid receptor superfamily
(Gustafsson et al., 1987; Cuthill et al., 1987).
The last area to be discussed here is the role of the pituitary in the regulation of the
endocrine system and in particular the relationships between pituitary-hypothalmus and the
estrogen receptors. The ER in liver and uterus is controlled by the pituitary (Lucier et al.,
1981), and growth restores the albation of hypophysectomy. Hypophysectomy has little
effect on TCDD induction of P-450, but TCDD does decrease ER in growth hormone-
restored animals (DeVito et al., 1991). Interestingly, Peterson et al. (See Chapter 5,
Developmental and Reproductive Toxicity) have made similar observations for regulation of
testosterone and testicular function by TCDD.
The importance of this subsection is at least twofold. First, the regulation of the
estrogen receptor is one of the most sensitive non-P-450 markers of TCDD exposure in
immature females, and this regulation coincides in rank-order fashion with decreases in
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mammary, uterine, and pituitary tumors in TCDD-treated female Sprague-Dawley rats.
Second, the removal of the ovaries (the primary source of estradiol) inhibits the formation of
hepatic tumors in rats by TCDD (Lucier et al., 1991) without altering the ability of TCDD to
induce cytochrome P-450IA1 (DeVito et al., 1992). However, ovariectomized animals
developed lung tumors in a two-stage model in rats, whereas no lung tumors were detected in
intact rats subjected to the same two-stage model (Clark et al., 1991). Taken together, these
findings indicate that estrogens can either enhance or protect against TCDD-mediated cancer
in a site-specific manner and that there is likely more than one mechanism for TCDD's
carcinogenic action.
8.2.4. Noncancer End Points
Previous risk assessments have focused primarily on cancer as the most important and
sensitive end point. This assumption has recently been questioned. For example, lead is
carcinogenic in experimental paradigms, yet it is the neurotoxicity that drives the risk
assessment. Past risk assessments of TCDD and its congeners have also focused on cancer
as the primary toxic end point, although it produces adverse effects in a wide variety of
tissues and cells. It is possible that the immunological, reproductive, or developmental
toxicities of TCDD are just as sensitive and important in the risk assessment process. For
noncancer end points, risk assessments traditionally have used the safety factor method to
estimate risk. Biologically based mathematical models for noncancer end points have not
been extensively utilized and are not as developed as are cancer risk models. The
development of biologically based models requires that the responses are well characterized,
tissue doses have been established, and sufficient data are available to propose a mechanistic
model. Many of the toxic effects of TCDD are well characterized with respect to dose-
response relationships, time-course relationships, species differences, and the magnitude for
the effects. Qualitative and quantitative evaluation of dose-response relationships for
noncancer end points is presented in Appendix E. This is a valuable exercise because it
looks at consistencies and inconsistencies among species strain and exposure regimen. It
helps to identify data gaps and provides a road map for future studies that will enable
biologically based risk assessments for noncancer end points. The development of
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biologically based mechanistic models requires more extensive data sets on the putative
mechanisms through which TCDD produces its toxic effects. However, it is important to
note that many of the same molecular events involved in TCDD-mediated cancer may also be
involved in the production of noncancer end points such as alterations in transforming growth
factor-2 (TGF-2), EOF receptor, and estrogen receptor. Therefore, as we learn more about
the mechanisms of TCDD-mediated, noncancer effects, we may be able to readily apply
cancer mechanistic models to other toxic effects.
8.2.5. Neurological and Behavioral Toxicity
The neurotoxic effects of the dioxins and related compounds have not received much
attention, in comparison to other target organs, despite a number of clinical and
semianecdotal reports of neurotoxic signs and symptoms in exposed humans (see, for
instance, Ashe and Suskind's (1985) reports on the Monsanto workforce; Jirasek et al., 1974;
Poland et al., 1971). There are some reports on neurochemical changes in animals
associated with exposures to PCBs and phenoxyacetic acids (Tilson et al., 1979). PCB
exposure also induced motor dysfunctions in some but not all mice (Tilson et al., 1979),
suggestive of effects on basal ganglia (circling and spinning).
Recently, Seegal and coworkers (1990) have reported significant effects of certain
PCBs on brain chemistry, specifically on aminergic pathways (norepinephrine, dopamine,
and serotonin). The structure-activity relationships of these effects suggest that they are not
associated with the Ah receptor, since it is the noncoplanar, low-chlorinated, nondioxin-like
PCBs that are neuroactive. These results are consistent with a report by Silbergeld (1992)
that the Ah receptor was not detected in neurons although it was measurable in glia.
Dioxins may be neurotoxic through indirect actions that affect nervous system
function and development. Some of the most exciting information in this area has been
published by Peterson and coworkers. They have reported that very low level, single-dose
exposures of pregnant rats results in offspring with significant alterations in sexual behavior,
characterized as a feminization of male rats (Mably et al., 1992a, b, c). It is well known
that the fetal endocrine milieu is critical to the development of the mammalian brain,
particularly the sexually dimorphic nuclei of the hypothalamus and cortex (Becker et al.,
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1992). There are critical periods of vulnerability during which small alterations in levels of
testosterone during gestation have profound and longlasting effects on brain and behaviors.
Mably et al. (1992a, b, c) and Peterson et al. (1992) have reported that these prenatal
exposures to dioxin result in altered patterns of sexual behaviors in male rats. It is likely
that the behavioral effects in this model are consequent to endocrine effects that impair the
sexual "imprinting" of the brain, rather than direct neurotoxic actions of dioxin on the
developing brain. Whether females are as sensitive to these intrauterine effects is not
presently known; since the male must be differentiated over embryofetal development from
the primordial female phenotype, it is possible that the male is more sensitive to these
endocrine-mediated effects. The consequences of these exposures for behaviors other than
sexual are also not known; however, other studies indicate that alteration in intrauterine
imprinting can affect motor activity and learning in males. These findings open up important
new areas of toxicological research on the dioxins.
8.2.6. Teratological and Developmental
8.2.6.1. Cleft Palate
Dioxins produce structural malformations and developmental toxicity in several
species. Considerable information is becoming available on mechanisms of cleft palate
formation, and it may be possible to construct mechanistic models for this effect. In mice,
increases in the incidence of cleft palate are well-characterized phenomena (Birnbaum et al.,
1987a, b, 1991). The doses required to produce cleft palate in mice are well below doses
that produce maternal toxicity or fetal mortality. In the normal developing palate, the
peridermal medial epithelial cells cease to express EGF receptor, decrease cell proliferation,
and eventually undergo programmed cell death while the basal cells differentiate into
mesenchyme, allowing the left and right palate to fuse. Temporal changes in the expression
of EGF receptor, EGF, TGF-alpha, TGF-beta 1, and TGF-beta 2 are critical for the fusion
of the palate. Experimental evidence indicates that changes in expression of these factors,
induced by TCDD, results in cleft palate formation. The medial epithelial cells of cultured
mouse embryonic palates exposed to TCDD express EGF receptor, incorporate [3H]-
thymidine, and differentiate into a stratified squamous oral-like epithelium in a dose-
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dependent manner (Abbott and Birnbaum, 1989). Changes in medial epithelial cell
differentiation are associated with increased EOF receptor, TGF-beta 1, and TGF-beta 2 and
decreased TGF-alpha levels (Abbott et al., 1992).
The use of cultured embryo palates (Abbott and Birnbaum, 1991, 1990a; Abbott et
al., 1989) has (1) led to a greater understanding of the mechanism of TCDD-induced cleft
palate and (2) enabled researchers to compare TCDD-induced biochemical changes in palate
tissue of several species. In vitro observations found that the human and rat palates are
sensitive to the cleft palate formation through the same mechanisms seen in mice: changes in
growth factors (i.e., EOF and TGFs) that are involved in the mechanism of altering
programmed cell death in the medial epithelial cells of the palate. The response to TCDD in
mouse palate cultures was about 100 to 1,000 times more sensitive than the response in
human or rat palate cultures.
The available data provide substantial information to develop a qualitative model
through which TCDD induces cleft palate. The induction of cleft palate in mice by TCDD is
mediated through the Ah receptor. TCDD binds to the Ah receptor in the medial epithelial
cells, and the activation of the Ah receptor initiates a cascade of events that increases TGF-
beta 1 mRNA and protein, increases TGF-beta 2 and EGF receptor protein levels, and
decreases TGF-alpha protein levels (Abbott et al., 1992). These changes alter the normal
signaling pathways in the medial epithelial cells. In control animals, the interaction between
these signaling pathways results in the programmed cell death of the peridermal medial
epithelial cells. The alterations in growth factor regulation by TCDD result in continued
proliferation of the peridermal medial epithelial cells and the redifferentiation of the basal
epithelial cells to stratified squamous oral-like epithelial cells, which subsequentially prevents
the fusion of the palate (Abbott and Birnbaum, 1989).
This preliminary model for the induction of cleft palate by TCDD requires better
characterization of several steps. Structure-activity relationships indicate that the Ah receptor
is involved, but there is limited evidence that the Ah receptor is present in the medial
epithelial cells of the developing palate. Cytosolic fractions of embryonic palatal shelves do
contain an Ah receptor, but which cells are expressing the Ah-receptor is undetermined
(Dencker and Pratt, 1981). It is presently unknown if the increases in TGF-beta 1 mRNA is
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mediated by the interaction of the Ah receptor with a DRE directly activating transcription of
the TGF-beta 1 gene or if the increases in TGF-beta 1 mRNA are due to the initiation of a
cascade of cytosolic or plasma membrane events mediated through the Ah receptor. Further
research is indicated into the interaction of the growth factors and their specific role in palate
formation. Because many of the data used to formulate this model are from studies using
cultured palates, the development of quantitative models would require dose-response data for
the in vivo alterations of these growth factors by TCDD, which are unavailable at this time.
Cleft palate in rats (Schwetz et al., 1973; Couture et al., 1989) and hamsters (Olson
et al., 1990) is induced at doses that result in significant maternal toxicity and fetal mortality,
and maximal induction of cleft palate is between 10% and 20%; however, in the mouse, cleft
palate can reach 100% incidence before any fetal mortality or maternal toxicity is
demonstrated. These data indicate that the mouse is extremely sensitive to this response. In
vitro studies indicate that humans may be much less sensitive than mice to TCDD-mediated
increases in cleft palate, so it is plausible that in humans cleft palate may occur only after
high exposures.
8.2.6.2. Hydronephrosis
In mice, hydronephrosis is also produced by TCDD following prenatal exposure at
doses that do not produce fetal mortality (Couture-Haws et al., 1991). Postnatal exposure
prior to day 4 can also produce hydronephrosis in mice (Couture et al., 1989). The
hydronephrosis induced by TCDD is due to occlusion of the ureter by epithelial cells (Abbott
and Birnbaum, 1990b). Increased proliferation of the epithelial cells by TCDD is associated
with increased EOF receptor. Hydronephrosis has not been reported in any other species at
doses that do not result in significant fetal mortality (Birnbaum et al., 1991).
Mice are the only species in which TCDD produces frank terata at doses that are not
fetotoxic. At present, there is no evidence that indicates humans are as sensitive as mice to
these effects. The only available data comparing the sensitivity of fetal tissue demonstrate
that human and rat fetal tissues are equally sensitive to the effects of TCDD (Birnbaum,
1991). These data suggest that sublethal exposure to TCDD may not result in frank terata of
the kidney.
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8.2.6.3. Thymic Atrophy
Prenatal exposure to TCDD produces thymic atrophy in all species tested and occurs
at doses well below those that cause maternal or fetal toxicity (Birnbaum, 1991). Thymic
atrophy occurs at similar doses in rats, guinea pigs, and hamsters exposed prenatally despite
a 5,000-fold difference in the LD50 in the adult animals (Olson et al., 1990). The sensitivity
and interspecies consistency of this response indicate that prenatal exposure to TCDD may
result in thymic atrophy in humans. The mechanism of thymic atrophy has not been
elucidated sufficiently to incorporate into a biologically based mechanistic model. Current
research has focused on the TCDD-induced alterations in thymocyte development and their
role in immunotoxicity.
8.2.7. Immunotoxicity
Although considerable research has focused on immunotoxicology of the dioxins (see
Chapter 4, Immunotoxicity), we are not at present able to develop or test a biologically based
model for purposes of risk assessment. A major obstacle to this undertaking is uncertainty as
to the outcome to be modeled. Susceptibility to infection or impairment of graft versus host
response could be proposed as the outcome for risk assessment, but not all studies have used
these responses as end points. Moreover, this may not be a sensitive indicator of immune
function. Alterations in biological markers of disease in animals or humans are not known.
Some of the measurable biochemical responses of animals to TCDD may reflect successful
response to xenobiotic challenge rather than impairment in functional integrity. Thus,
changes in immunoglobulins may reflect immune competence rather than dysfunction. Our
inability to define outcome is not unique to immunotoxicology; the continuing controversies
over the definition of acquired immunodeficiency syndrome reflect scientific uncertainty in
this area. NIEHS has proposed a tier approach to the identification of potential
immunotoxicants (Luster et al., 1992), and dioxin certainly tests positive in this system.
Our knowledge of basic immunobiology makes it difficult for us to integrate our
findings on dioxin into an overall biologically based schematic of events. We do not know
the quantitative relationships between a change in intercellular signaling and cell-mediated
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responses, although we know these events are fundamentally related. Many events in the
immune system appear to have complex interactions, with biphasic relationships. Thus we
have no quantitative context in which to develop predictive associations between events
affected by dioxin and other events in immune system response.
We place high priority on improving our ability to develop risk assessment methods
for immunotoxicants, not limited to dioxin. As noted above, progress has been made on
developing a consensus approach to the hazard identification of potential immunotoxicants,
but as yet there are no methods for using dose-response data from such tests to develop
quantitative risk assessments. Dioxin may be a prototype compound for developing such
methods, and research should be directed toward designs that encompass many different
events in immunology from early molecular and cellular events to whole animal response to
immune challenge, in order to facilitate the overall evaluation of end points. Moreover, in
such designs, sufficient dose ranges should be used to assist in the statistical evaluation of
proposed models and to compare animal and human responses.
In clinical and epidemiological studies, much data collection should be done; given
the accessibility of circulating lymphocytes and other markers in blood, it should be possible
to increase our confidence in interspecies comparisons by examining the same parameters in
exposed animals and well-characterized human populations. Because of the reported
sensitivity of the developing organism to immunotoxic effects of dioxin, a priority should be
placed on obtaining data on immunologic function in children with documented exposures to
dioxins or related compounds. Clinical studies need to be well controlled and conditions of
testing and sample collection carefully described in order to facilitate such comparisons.
8.2.8. Reproductive Toxicity
8.2.8.1. Female Reproductive Toxicity
Several studies have demonstrated that TCDD affects female reproductive function in
mice, rats, and monkeys. TCDD reduces fertility, litter size, and uterine weights. TCDD
also alters menstrual and estrus cycling in monkeys, mice, and rats. Uterine weight and
menstrual/estrus cycling are regulated by estrogens. These data indicate that TCDD has
antiestrogenic effects that result in decreased reproductive functioning.
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The antiestrogenic actions of TCDD could be mediated either by changes in
circulating estradiol, qualitative changes in estrogen metabolism, or through decreases in
estrogen receptors. In mice, TCDD does not alter serum estradiol levels, and the
antiestrogenic actions of TCDD are associated with decreases in uterine estrogen receptors
(DeVito et al., 1992). Similarly, TCDD decreases rat hepatic and uterine estrogen receptor
(Romkes et al., 1987) but does not affect serum estradiol levels. Structure-activity studies
suggest that the Ah receptor mediates the down-regulation of the estrogen receptor. The
estrogen is down-regulated by TCDD in several breast cancer cell lines (Safe et al., 1992a).
TCDD also decreases estrogen receptors in Hepa Iclc? cells but not in mutant cell types that
do not express a high affinity form of the Ah receptor nor in cells that do not accumulate
activated Ah receptors in their nucleus (Zacharewski et al., 1991). These studies provide
further evidence that the Ah receptor is involved in the down-regulation of the estrogen
receptor.
One possible mechanism for the antiestrogenic actions of TCDD is that TCDD binds
to the Ah receptor in the target tissue and through a cascade of events decreases the amount
of estrogen receptor in the cell, thus inhibiting the actions of estrogens. The down-regulation
of the estrogen receptor by TCDD could be mediated either by decreased transcription of the
estrogen receptor gene or possibly through nontranscriptional mechanisms. At present it is
unclear how TCDD down-regulates the estrogen receptor other than it is mediated through
the Ah receptor.
An alternative mechanism by which TCDD inhibits estrogenic actions is through
increases in estradiol metabolism. Following TCDD exposure, estradiol metabolism is
increased 100-fold in MCF-7 cells (Spink et al., 1990). Microsomal hydroxylation of
estradiol is increased twofold to fourfold in rats treated with TCDD (Graham et al., 1988).
The role of estrogen metabolism in the antiestrogenic actions of TCDD remains to be
determined. While there is more evidence supporting the role for the down-regulation of the
estrogen receptor mediating the antiestrogenic actions, further studies are required to
determine the extent of estradiol metabolism in vivo following TCDD treatment.
Since TCDD alters immune function and a variety of growth factor pathways and in
general is acting like a potent and persistent environmental hormone, research is needed to
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determine the relationships, if any, to endocrine-related disorders in women such as
endometriosis, osteoporosis, and cancers of the reproductive tract.
8.2.8.2. Male Reproductive Toxicity
When administered to adult rats, TCDD decreases testis and accessory sex organ
weights, decreases spermatogenesis, and reduces fertility (Moore et al., 1985; Moore and
Peterson, 1988; Bookstaff et al., 1990a). These effects are associated with decreases in
plasma testosterone (Moore et al., 1985). The decreases in circulating androgens are due to
decreased testicular responsiveness to luteinizing hormone and increased pituitary
responsiveness to feedback inhibition by androgens (Moore et al., 1989, 1991; Bookstaff et
al., 1990a, b; Kleeman et al., 1990). Although the antiandrogenic effects occur within 24
hours, the doses required to produce these effects are overtly toxic and decrease food intake
and body weight. The high doses needed demonstrate that the antiandrogenic effects are not
very sensitive effects. However, epidemiological studies have demonstrated decreased
testosterone in workers exposed to dioxin-like compounds (Dioxin92).
In contrast to the adults, the developing male reproductive system is very sensitive to
the effects of TCDD. In rats, prenatal exposure to TCDD results in decreases in sex organ
weight, impairs spermatogenesis and luteinizing hormone secretion, and demasculinizes and
feminizes sexual behavior (Mably et al., 1992a, b, c). Maternal doses as low as 0.064 /ig/kg
can produce these effects, indicating that the developing male reproductive system is one of
the most sensitive end points for the toxic effects of TCDD. The alterations in male
reproductive development are associated with decreases in testosterone levels. Sufficient
testosterone levels are critical for sexual differentiation of the central nervous system; thus
the decreases in testosterone levels may, in part, account for the TCDD-induced reproductive
alterations in male rats. It is also possible that these effects are due in part to alterations in
tissue sensitivity to testosterone (Mably et al., 1992c).
In summary, there is ample evidence that noncancer end points are extremely
sensitive to the toxic effects of TCDD. The available data do not provide enough
information to develop biologically based mechanistic models for all noncancer end points.
For some of the noncancer effects of TCDD there is sufficient evidence for which proposed
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mechanism may be modeled. Experimental evidence on cleft palate formation and male
reproductive toxicity provides sufficient evidence to propose qualitative models that can be
developed into mechanistic models. However, for immunotoxicity, thymic atrophy,
neurobehavioral toxicity, and female reproductive toxicity, the mechanisms by which they
occur are unknown and in some cases the target tissues remain undetermined. Furthermore,
few if any of the molecular events beyond ligand binding to the Ah receptor are understood
for these effects. The only information we have to develop mechanistic models is dose-
response relationships. Future studies that better characterize target tissues and the molecular
mechanisms underlying these events are indicated. Because of the importance of generating
reliable estimates of the risk for noncancer effects, the development of biologically based
dose-response models for these effects is an urgent research need.
8.3. COMPARATIVE END POINTS/QUALITATIVE COMPARISONS
There has been considerable discussion concerning the relative sensitivity of the
several lexicological end points modified by exposure to TCDD. Many scientists and
regulators have stated that carcinogenicity may not be the most sensitive end point.
However, before one can compare the magnitude of these various effects, one must
understand the limitations imposed by both statistics and biology on these types of relative
potency comparisons. This section briefly outlines some of the problems with discussing
relative potency across different end points. It provides no simple solutions to these
problems. These solutions will take a combination of research (in both statistics and biology)
and management decisions where research cannot yet be used to fill the knowledge gaps in
the comparison.
There are two basic problems with the comparison of toxic potency across end points.
These are (1) comparison of measures of mortality (or serious threats to mortality) to
measures of morbidity or to measures of biochemical modifications and (2) statistical issues
concerning the power to detect effects and the inclusion of background responses. We will
discuss each of these in detail using TCDD as an example of the problems involved in this
undertaking.
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We will consider these issues in reverse order starting with the statistical problem
first. Statements such as the one above sometimes reflect a statistical artifact due to the
nature of the biological assay being analyzed. To illustrate this point, consider the data given
in Table 8-4. This table considers four different responses to TCDD: liver cancer in female
mice and rats, immunological changes in male mice, changes in protein concentrations in
female rats, and rates of terata in mice. One practical way to address potency across end
points is to rely on significant difference from control for each treatment group. On the
basis of a statistical test, it is clear that the most sensitive end point is an effect on the
immune system where a response was detected at a dose of 0.644 ng/kg. Moreover, the end
point being measured (plaque-forming cell response) has a greater numerical value than do
other responses (such as tumor incidence), with a small variability allowing for much greater
statistical power in obtaining a response. Also, the lowest dose for induction of CYP1A1
shows a significant effect on immune responses; it is impossible to know if lower doses
might not have been significant also. Thus the location of a no-observed-effect level (NOEL)
is dependent on the statistical properties of the end point being studied and the sensitivity of
a particular response.
One practical way would be to consider the relative change in response over
background in the TCDD-treated groups. For example, the drop in plaque-forming cell
(PFC) response from control to low dose is 6% or -0.06 relative units of response per
0.322 units (pg/kg/day) increase in dose, or a slope of 0.0186. This is ~3 times greater
than the relative increase in CYP1A1 (slope of 0.068) and about one-tenth the relative
increase in liver tumors in female mice (National Toxicology Program, 1982). Even more
dramatic is the infinite change in response from control (0.00) to low dose (0.019) for cleft
palate in the Abbott and Birnbaum (1989) study. Thus, using this measure of relative
potency, cleft palate is the most sensitive end point, with liver cancer second and the
immunological response third. Thus, a second practical measure of potency, changes relative
to background, paints a very different picture than does the use of statistical p-values and is
sensitive to zero response in the control population. In addition, one has to scale between
relative drops and relative increases, the one bounded by 0 and the other unbounded
(although for practical purposes, this is easily dealt with). Finally, one could also use
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Table 8-4. Toxic End Points Database
End Point
Liver cancer in female mice
Liver cancer in female rats
Plaque-forming cells per 10* viable spleen cells
Concentration of cytochrome P450IA1 in
microsomal protein form hepatocytes (pmol/mg)
Concentration of cytochrome P450IA2 in
microsomal protein form hepatocytes (pmol/mg)
Cleft palate in mice fetus
Dose*
0.0
1.4
7.1
71.0
0.000
1.0
10.0
100.0
0.00
0.322
0.644
1.191
3.091
0.0
3.5
10.7
35.7
125.0
0.0
3.5
10.7
35.7
125.0
0.0
6000.0
9000.0
12000.0
15000.9
Response
0.04 lc (3/73)"
0.120(6/50)
0.125(6/48)
0.234 (11/47)°
0.05 1« (3/58.4)h
0.029 (1/34.0)
0.272 (9/33.1)'
0.607 (19/31/3)'
777±88
731±190
438 ±96'
118±46'
65 ±15'
12.9±11.3
56.4±26.7'
111.5±30.3e
181.4±18.4'
293.3±17.1'
63.5±38.4
88.3±23.0
161.0±55.7'
193.1 ±60.2"
297.4±88.3C
0.000 (0/159)
0.019 (2/107)
0.213 (26/122)'
0.505 (50/103)'
0.777 (84/108'
Reference
NTP (1982")
Kocibaetal., 1978f
Davis and Safe, 1988'
Tritscher et al., 1992
Tritscher et al., 1992
Abbott and Birnbaum, 1989
*In ng/kg/day
bGavage dosing
'Probability of getting a tumor prior to the end of the study
''Number with tumor/number examined for the tumor
'Significantly different from control response (p<0.05)
tCDD in diet
Survival-adjusted using poly-3 adjustment of Portier and Bailer (1989)
"•Number with tumor/poly-3 adjusted number at risk of tumor
'Single intraperitoneal injection
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absolute (rather than relative) change over background, but then one would be faced with the
problem of determining how a change of one unit in PFC response relates to a change of one
unit in CYP1A1 response.
The only practical manner in which a relative comparison of potency can be made is
in terms of mortality. In this way, all responses are in the same units and have a common
control response in terms of the background mortality in the population. However, this
approach is currently infeasible. There are both technical and practical considerations that
must be ironed out before an approach of this type can be applied. On the technical side,
there are issues concerning life expectancy versus incidence of death. For example, suppose
TCDD increased resorptions by 5% at some chosen dose. This would represent a 5%
increase in mortality for potential fetuses and an overall loss of 5% of the total available
number of animal lifetimes (resulting in a 5% drop in life expectancy). Suppose also that
this same dose of TCDD increased mortality from cancer so that, by the end of the study, an
additional 20% of the animals have died. Suppose also that this increase in mortality is late
in life so that the overall drop in life expectancy is only 5%. Thus, on one scale (life
expectancy), the two end points, resorptions and cancer, produce the same results, whereas
on another scale, mortality by age, they are different. It is unclear which measure is most
appropriate for ranking the observed effects of TCDD. On the more practical, biological
side of the issue, one must relate increases in tumor incidence to changes in mortality,
modifications in immune response and/or protein concentrations to changes in mortality, etc.
The information needed to model these relationships is currently unavailable.
This is not to say that statements concerning the relative importance of certain end
points for toxicity from exposure to TCDD cannot be addressed. However, caution must be
used in this endeavor, and one must be careful to explain the methods by which the relative
potencies were established. When this is done, the gaps in our knowledge become obvious
and future research can be better directed to fill those gaps.
8.4. RELEVANCE OF ANIMAL DATA FOR ESTIMATING HUMAN RISKS
The reliability of using animal data to estimate human risks has been questioned, and
this issue is especially important for TCDD. We know there are unusually wide species
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differences in acute toxic responses to TCDD, but we do not know if such wide differences
exist for carcinogenic and other toxic effects. However, we do know that the rank order of
species differences in acute effects does not predict the rank order for all other toxic effects.
For example, mice appear to be considerably more sensitive than rats to the teratogenic and
immunotoxic effects of TCDD, but we do not know if dose-response relationships for
immunotoxic effects in humans resemble those for rats, mice, or neither.
Although the human data are limited, it does appear that animal models are, in
general, appropriate for estimating human risks, keeping in mind that for some responses,
where wide species differences exist, the relative placement of human responses may not be
possible at this time. However, humans contain a fully functional Ah receptor (Lorenzen and
Okey, 1991; Manchester et al., 1987; Cook and Greenlee, 1989), and many of the
biochemical effects produced by TCDD in animals also occur in humans. Data on effects of
TCDD and its analogs in humans are based on in vitro (i.e., in culture) as well as
epidemiological studies. A comparison of the effects of CDDs and CDFs on laboratory
animals versus humans is given in Table 8-5. In vitro systems such as keratinocytes or
thymocytes in culture have clearly shown that human cells possess Ah receptors, and they
respond similarly to cells derived from rodents. Several reports in the literature suggest that
exposure of humans to dioxin and related compounds may be associated with cancer at many
different sites, including malignant lymphomas, soft tissue sarcomas, hepatobiliary tumors,
hematopoietic tumors, thyroid tumors, and respiratory tract tumors (Bertazzi et al., 1989,
1993; Fingerhut et al., 1991; Manz et al., 1991; Zober et al., 1990; Saracci et al., 1991).
These studies are evaluated in Chapter 7, Epidemiology/Human Data, including discussion of
confounding factors and strength of evidence. There is growing evidence from human cancer
studies that TCDD is a multisite carcinogen, which is not unexpected if we assume that
TCDD is acting like a potent and persistent hormone agonist/antagonist. Likewise, TCDD is
a multisite carcinogen in animals (Lucier, Chapter 6).
Several noncarcinogenic effects of chlorodibenzodioxins (CDDs) and
chlorodibenzofurans (CDFs) show good concordance between laboratory species and humans
as well. For example, in laboratory animals, TCDD causes altered intermediary metabolism
manifested by changes in lipid and glucose levels. Consistent with these results, workers
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Table 8-5. Similarities Between Laboratory Animals and Humans in Biological
Effects of TCDD
Effect
Laboratory
animals
Human or
human cells
In vitro
Presence of Ah receptor
Enzyme induction
Altered pattern of growth and
differentiation
Immunosuppression
Choracnogenic response
+
+
+
+
+
+
+
+
+
In vivo
Presence of Ah receptor
Enzyme induction
Altered lipid metabolism
Immune effects
Cancer
Reproductive effects
Teratogenic effects
Altered epithelial cell
differentiation
Tumor promotion
+
+
+
+
+
+
+
+
+
+
+
+
+/-
+
+/-
+/-
?
?
Source: Silbergeld and Gasiewicz, 1989
The + indicates a clear association while +/- indicates conflicting or unclear
associations; the ? indicates that not enough is known on the effects of TCDD on
the system to evaluate.
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exposed to TCDD 7 to 8 years previously during the manufacture of trichlorophenol showed
elevated total serum triacylglycerides and cholesterol with decreased high density lipoprotein
(Walker and Martin, 1979). Recently, the results of a statistical analysis of serum dioxin
analysis and health effects in Air Force personnel following exposure to Agent Orange were
reported (Wolfe et al., 1990). Significant associations between serum dioxin levels and
several lipid-related variables were found (percent body fat, cholesterol, triacylglycerols, and
HDL). Another interesting finding of these studies was a positive relationship between
dioxin exposure and diabetes, possibly the first report of such an association.
The human to experimental animal comparison is confounded by at least two factors:
(1) For most toxic effects produced by dioxin, there is marked species variation. An outlier
or highly susceptible species for one effect (i.e., guinea pigs for lethality or mice for
teratogenicity) may not be an outlier for other responses. (2) Human toxicity testing is based
on epidemiological data comparing "exposed" to "unexposed" individuals. However, the
"unexposed" cohorts contain measurable amounts of background exposure to CDDs, CDFs,
and dioxin-like PCBs. Also, the results of many epidemiological studies are hampered by
small sample size, and in many cases the actual amounts of dioxin and related compounds in
the human tissues were not examined. However, based on the available information, it
appears that humans respond to CDDs and CDFs like most experimental animals.
There is also relatively good concordance in the biochemical/molecular effects of
TCDD between laboratory animals and humans. Placentas from Taiwanese women exposed
to rice oil contaminated with PCBs and CDFs have markedly elevated levels of CYP1A1
(Lucier et al., 1987). Comparison of these data with induction data in rat liver suggests that
humans are at least as sensitive as rats to enzyme-inductive actions of TCDD and its
structural analogs (Lucier, 1991). Consistent with this contention, the in vitro
EC50 for TCDD-mediated induction of CYPlAl-dependent enzyme activities is ~1.5 nM
when using either rodent or human lymphocytes (Clark et al., 1992). However, binding of
TCDD to the Ah receptor occurs with a higher affinity in rat cellular preparations compared
to humans (Lorenzen and Okey, 1991). This difference may be related to the greater lability
of the human receptor during tissue preparation and cell fractionation procedures (Manchester
et al., 1987). In any event, it does appear that humans contain a fully functional Ah receptor
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(Cook and Greenlee, 1989) as evidenced by significant CYP1A1 induction in tissues from
exposed humans, and this response occurs with similar sensitivity as observed in
experimental animals.
One of the biochemical effects of TCDD that might have particular relevance to toxic
effects is the loss of plasma membrane EGF receptor. There is evidence to indicate that
TCDD and its structural analogs produce the same effects on the EGF receptor in human
cells and tissues as observed in experimental animals. First, incubation of human
keratinocytes with TCDD decreases plasma membrane EGF receptor, and this effect is
associated with increased synthesis of TGF-a (Choi et al., 1991; Hudson et al., 1985).
Second, placentas from humans exposed to rice oil contaminated with polychlorinated
dibenzofurans exhibit markedly reduced EGF-stimulated autophosphorylation of the EGF
receptor, and this effect occurred with similar sensitivity as observed in rats (Lucier, 1991;
Sunahara et al., 1989). The magnitude of the effect on autophosphorylation was positively
correlated with decreased birth weight of the offspring.
In summary, animal models are reasonable surrogates for estimating human risks.
However, it must be kept in mind that the animal to human comparison would be
strengthened by additional mechanistic information, especially the relevance of specific
molecular/biochemical changes to toxic responses. It is also important to note that the
mechanism of carcinogenesis (sequence of molecular events) may be quite different at
different sites. For example, the mechanism responsible for TCDD-mediated lung cancer
appears to be different from that responsible for liver cancer (see Chapter 6, Carcinogenicity
of 2,3,7,8-TCDD in Animals).
8.5. HUMAN MODELS
8.5.1. Introduction
Unlike animal data where recent studies have allowed modeling for dosimetry,
induced proteins, cell proliferation, and toxic effects, human data are very sparse. With
regard to toxic effects, Chapter 7 presents recent evidence suggesting TCDD's effects on
human reproduction, neurotoxicity, diabetes, and cancer. From a modeling viewpoint, male
reproduction, diabetes, and thyroid cancer appear to be good candidates for modeling from a
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"bottom-up" mechanistic approach, since TCDD's effects on male serum testosterone levels,
the insulin receptor, and thyroid hormones have been documented. These, however, remain
efforts for the future. The focus of this section will be on cancer, specifically liver cancer,
respiratory cancer, and all cancers combined. There are two reasons for this. First, the
emphasis on liver cancer in EPA's history of TCDD regulation demands it, and it is a logical
sequence to Section 8.3. Second, the recent epidemiology evidence for respiratory cancer,
soft tissue sarcoma, and all cancers combined suggests that dioxin is a human carcinogen.
Modeling for these cancers in humans, however, requires and receives different
approaches than has been presented in earlier sections of this chapter. The approach used for
liver cancer is one involving interspecies toxic equivalent liver doses, with extrapolation from
a rat to human no-observed-added-effect level (NOAEL). The modeling approach used for
the human epidemiology data for lung cancer and all cancers combined involves estimating
human intake dose associated with cancer response and curve fitting both additive and
multiplicative risk models to the data.
8.5.2. Modeling Toxic Effects in the Liver
One of the intrinsic features of PB-PK descriptions is that interspecies extrapolations
can be attempted with reasonable confidence in the result, assuming that relevant changes are
made in the configuration of the model to allow for changes in physiology, metabolism, and
protein binding. The limiting factor with any modeling description is, however, the
availability of relevant data sets. This is particularly true when attempting to model the
pharmacokinetics of dioxins and dioxin-like chemicals in people. The use of a physiological
pharmacokinetics description to analyze human TCDD exposure data was first attempted by
Kissel and Robarge (1988). In this work the authors used a fugacity approach to examine the
elimination of TCDD from humans using data derived from estimates of background
exposure and tissue levels, half-lives from Ranch Hand veterans, and a self-exposure
experiment by Poiger (Poiger and Schlatter, 1986).
This fugacity-based model attributes the distribution of TCDD as a simple partitioning
process with expected dose-independent linear kinetics. In both rats and mice, as noted
earlier, a disproportionate amount of TCDD is found in the liver with increasing dose. This
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phenomenon cannot be described by simple solubility/partitioning alone. The dose-dependent
liver-to-fat concentration ratio is a good indicator of this trend. In fact, analysis of data from
people exposed to CDFs, congeners of TCDD, from consumption of contaminated oil shows
a higher liver: fat ratio with increasing body burden (Carrier and Brodeur, 1991). More
complex models need to be constructed to account for the nonlinearities in dioxin disposition.
Carrier and Brodeur (1991) constructed a toxicokinetic model for halogenated
polycyclic aromatic hydrocarbons (HPAHs) in humans. This model is not a classical PB-PK
model. The analysis begins with the observation that the tissue distribution of dioxin-like
HPAHs in people and animals is body burden dependent. As the body burden increases
(Cbody), the proportion of that body burden associated with the liver (F^ increases toward a
maximum value (F^. This has been previously described for rats (Abraham et al., 1988).
The data on rats (Kociba et al., 1976), monkeys (McNulty et al., 1982), and people, using
toxic equivalency factor conversions (Kuroki and Masuda, 1978), were examined using an
empirical, Michaelis-Menton type saturable binding isotherm:
Liver fraction (F,J = F,^ C^ / (K,, +
Considering the body burden as a surrogate for liver concentration, this equation can be
loosely interpreted as the induction of binding species in the liver as dose increases. Indeed,
analyses showed that the Kj was very similar for people and experimental animals, possibly
indicative of similar protein-induction dynamics in various species. With different dioxin-
like isomers F,,^ and Kj vary; this can be thought of as changes due to different binding
affinities.
This empirical model fits the observed data in various species; however, it is a fitting
exercise and not an examination of underlying biology. The model is not physiologically
based. It in effect examines the steady-state condition of a two-compartment model,
consisting of the liver and "the rest of the body." The terms C^y and F^ are difficult to
interpret in biological terms. C^y represents a body burden of chemical rather than a tissue
concentration, and the term for maximum liver concentration F,,^ is derived empirically.
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In its present form the model assumes that at very low doses the hepatic fraction is
zero. This is unlikely due to the partitioning of dioxin into the liver and the low-dose
binding characteristics of the Ah receptor and CYP1A2 in the liver. Also in this description
the metabolism is modeled as saturable at maximal induction of liver TCDD sequestration.
Given the experimental data in many species (Chapter 1, Disposition and Pharmacokinetics),
this seems unlikely. The analysis of human dioxin kinetics does show that the dose-response
curves for and induction of hepatic-binding species for dioxin in rodents and people appear to
be very similar. If one assumes that humans and rodents are also similarly sensitive in the
toxic responses to TCDD, based on liver concentrations, some simple exposure calculations
can be made. Interspecies comparisons of the daily intake of TCDD needed to reach
equitoxic concentrations in the liver have been estimated by Carrier (1991). These are
presented in Table 8-6. It is important to note that a number of assumptions are made to
make these comparisons, and considerable uncertainty exists.
The integration of the toxicokinetic description put forward by Carrier, and the
accompanying data sets, with the more physiological approaches described for rats by
Andersen et al. (1993b) and Kohn et al. (1993) will provide an opportunity to further
investigate the determinants of disposition of TCDD in humans.
The prediction for Carrier's model of 40 pg/kg-day for a no-effect dose for cancer is
~ 7,000 times as high as that currently used by EPA but only 6 to 7 times higher than that of
other western countries (Kociba, 1991). However, this model assumes that humans are as
sensitive as rats to a given tissue burden of dioxin, and it does not account for possible
interindividual variation among humans. Also, it is specific for liver, so it does not predict
cancer responses in other tissues such as lung or any other effects. It also does not attempt
to predict biochemical effects such as enzyme induction, which occurs at much lower doses
than 40 pg/kg/day (Vanden Heuvel et al., 1994).
8.5.3. Lung Cancer and All Cancers Combined
Data from four recent epidemiology retrospective cohort occupational studies provide
evidence of the human carcinogenicity of dioxin. All showed increased mortality from
respiratory cancer; the two largest (Saracci et al., 1991; Fingerhut et al., 1991) showed
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Table 8-6. Rat and Human Comparison of Daily TCDD Intakes and Body and Liver
Concentration for Equitoxic Response
Parameter
Daily Intake
(TCDDeq)
Total body concentration
(TCDD eq/body)
Liver concentration
(TCDD eq/kg)
Low intake
Raf
1 ng/kg
61 ng/kg
540 ng/kg
Human
40.7 pg/kg
70 ng/kg
540 ng/kg
High intake
Rat5
100 ng/kg
1.45 /tg/kg
24 /ig/kg2
Human
1 ng/kg
1.2 Mg/kg
24 Mg/kg2
Source: Carrier, 1991.
•Rat NOAEL (Kociba et al., 1978).
bMarked liver toxicity; tumors in rats.
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increases in mortality from soft tissue sarcoma, and three (Fingerhut et al., 1991; Zober et
al., 1990; Manz et al., 1991) showed increased mortality from all cancers combined. In the
Fingerhut et al. (1991) study all three cancer type results showed significance only for the
high-exposure, long latent period subcohort. The largest study, with 18,000 total workers
(Saracci et al., 1991), showed no increase in overall cancer mortality, but those authors have
not presented the data allowing for a latent period. Furthermore, the Saracci et al. (1991)
study, unlike the other three, provides no way to quantitatively estimate TCDD exposure to
their cohort. Based on this lack of information, this modeling exercise will be restricted to
the other three cohorts.
Three other recent cohorts were not included in this analysis for various reasons.
Kuratsune et al. (1988) reported increased lung cancer in male victims (standard mortality
ratio [SMR]=3.3, based on eight cases) of the Yusho PCS and CDF contamination rice
poisonings. Although there are serum measurements and 37 TEF estimates available for this
cohort, there was no actual TCDD in the contaminants. Thus, this cohort also will not be
used in the modeling effort here. Collins et al. (1993) reported increased mortality for both
lung cancer and all cancers combined for a subcohort of 122 U.S. workers who developed
chloracne following exposure to dioxin at a chemical plant during a 1949 accident. Their
analysis, however, attributes this increase to co-exposure to 4-aminobiphenyl. Since that
chemical plant is included in the Fingerhut et al. (1991) cohort, it will not be included in this
analysis. The Seveso, Italy, community cohort is also not included in this analysis because
of the limited observation period following the 1976 accident. All of these studies are
discussed in much greater detail in Chapter 7, Epidemiology/Human Data.
The largest of the three studies used here is the Fingerhut et al. (1991) study of
>5,000 U.S. workers from 12 U.S. plants producing chemicals contaminated with TCDD.
Of 1,520 workers exposed to TCDD-contaminated processes for at least 1 year with a 20+
year latency, mortality was significantly increased for both respiratory cancer (SMR=142;
95% C.I. 103-192) and for all cancers combined (SMR=146; 95% C.I. 121-176). A
similar-sized cohort with less than 1-year exposure with a 20+ year latency showed no
increase in either all cancers or respiratory cancers. Manz et al. (1991), in a smaller cohort
of 1,148 men in a herbicide manufacturing plant in Hamburg, Germany, also found increased
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mortality from lung cancer (SMR=141; 95% C.I. 95-201) and all cancers combined
(SMR=124; 95% C.I. 100-152). Cancer mortality increased both among groups with
increased duration of exposure and among groups with suspected highest levels of exposure.
In the smallest of these recent studies, Zober et al. (1990) studied three subcohorts totaling
250 workers with potential exposure to dioxin during an industrial accident in 1953. Of the
127 who developed either chloracne or erythema, and who were considered among the most
highly exposed, for those with a 20+ year latent period, mortality from all cancers
(SMR=201; 95% C.I. 122-315) and from lung cancer (SMR=252; 95% C.I. 99-530) was
both statistically increased. Furthermore, the increase in total cancer deaths in all these
studies does not appear to be due totally to the increase in respiratory deaths. The SMRs for
all cancer deaths not including lung cancer remain statistically significant in all three studies.
These findings are supported by recent animal evidence from Lucier et al. (1991) who
found lung tumors in ovariectomized female Sprague-Dawley rats but not in intact female
rats following administration of TCDD. Increased lung tumors are also seen in the Kociba et
al. (1978) study with female Sprague-Dawley rats but not with male rats. Other animal data
support the tumor-promoting ability of dioxin in the liver and skin (Pitot et al., 1980).
Based on the evidence of lung cancer and all cancers combined, a quantitative analysis
of dioxin's cancer potency will be modeled from the three epidemiology studies. All three
studies attempted to verify dioxin levels in samples of their working cohorts, although in all
cases the subjects were tested decades after exposure ended. Thus, with the limited
information available, assumptions must be made about the representativeness of both these
sampled subjects and the dose-response models used to estimate risk. The details are
presented below.
8.5.3.1. Dose-Response Models
The following analysis provides maximum likelihood and 95% lower confidence limits
of incremental cancer risk based on the cancer death response in the lung and all cancers
combined in the three recent cohort studies (Fingerhut et al., 1991; Zober et al., 1990; Manz
et al., 1991). Both additive and relative risk models are used. This type of analysis has
been used previously with epidemiologic studies in several EPA health assessments (e.g.,
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methylene chloride, nickel, and cadmium). For this report the analyses will be done both
separately for each study and for all studies combined. A description of the models follows.
8.5.3.1.1. Excess or additive risk model. This model follows the assumption that the excess
cause-age-specific death rate at age t due to dioxin exposure, ht(t), is increased in an additive
way by an amount proportional to exposure at some age t-k. In mathematical terms, this is:
where Xt.k is the exposure at age t-k, and 0, the parameter to be estimated, is the
proportional increase. The total cause-age-specific rate h(t) is then additive to the
background cause-age-specific rate ho(t) as follows:
h(t) = ^(t) + ht(t)
For an individual i observed from toj until age t;, the cumulative death rate expected is:
'i
H,(t) = £ S(t)h0(t) +
where S(t) = survival function to age t. For dioxin, with a long half-life, we can let k=0.
For Nj individuals in exposure group J, the expected number of deaths is:
E. = £ H,(t) = EO; + ft£ X{ • W, = EOJ + /SX^
i-l i-1
where Ej is the total number of expected cancer deaths in the observation period from the
group exposed to average exposure Xj, E^ is the expected number of cancer deaths due to
background causes (lifetable "expected" rates), Wj is the number of person-years of
observation for the jth exposure group, and the parameter & represents the slope of the dose-
response model. To estimate 0, the observed number of cause-specific deaths in group j, Oj}
is assumed to be distributed as a Poisson random variable with expected value Ej. The
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parameter estimate, b, can be tested for being significantly >0. A statistically significant
result is evidence of an additional cancer effect due to dioxin exposure.
Under the above assumptions, the solution by maximum likelihood proceeds as
follows: The likelihood equation is
N
L =
where N = the number of separate exposure groups. The maximum likelihood estimate
(MLE) of the parameter p* is obtained by taking the first derivative of the log likelihood
equation, setting it equal to 0 and solving for b:
= 0
The asymptotic variance for the parameter estimate b is:
where b is the MLE. This variance can then be used to obtain approximate 95% upper and
lower bounds for p*. Lifetime incremental cancer risk estimates for continuous exposure are
estimated by multiplying b by 70 if X is in units of lifetime continuous exposure (i.e.,
lifetime average daily dose [LADD]).
8.5.3.1.2. Multiplicative or relative risk model. This model follows the assumption that the
background cause-age-specific rate at any age t is increased in a multiplicative way by an
amount proportional to the cumulative dose up to that age. In mathematical terms this is:
h(t) = no(t)(l + pXJ
As above, summing over the observed and expected experience yields, for each exposure
group,
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EJ/EOJ = i + px.
Again, to estimate 0, the observed number of cause-specific deaths, Oj, assumed to be a
Poisson random variable, is substituted for Ej. Following the same procedure as above, the
MLE, b, is the solution to
^j~ = E -^XJ + (°j > (i
with asymptotic variance
If Xj is in units of LADD, then lifetime incremental risk estimates per unit under this model
are obtained by multiplying b by the background lifetime cause-specific risk of death, P0.
The values P0 are derived using lifetable methods for competing risks and 1973-1977 U.S.
death rates. For lung and all cancers combined, these are 0.038 and 0.185, respectively.
8.5.3.2. Exposure and Dose Estimates
Exposure estimates are derived from serum dioxin levels in workers sampled long
after exposure ended and extrapolated backward using a first-order model for elimination
(U.S. EPA, 1994) with a biological half-life of 7.1 years (Pirkle et al., 1989). In humans,
dioxin deposits primarily in adipose tissues at normal exposure levels, although body
deposition dose dependency has been shown in animal studies (see Chapter 2, Mechanism(s)
of Action, and Section 8.5.2). In those rat studies, liver/fat concentration ratios increase
with increasing dose, due to TCDD induction of the binding protein P4501A2 in the rat
liver. However, although there is evidence that dioxin causes some human liver toxicity (Di
Domenico and Zopponi, 1986) and that the dioxin-like PCB and dibenzofuran compounds can
also cause liver cancer in humans (Kuratsune et al., 1988) (see Chapter 7), there is no direct
evidence that TCDD induces cancer in human liver. Thus, although Table 8-6 presents rat-
to-human liver concentration toxic equivalents for various rat exposures, this section uses
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only human data. In humans, all estimates suggest that adipose tissue is the major storage
compartment. Schlatter (1991) estimated a liver/fat concentration ratio of about 1/6. With a
body adipose tissue fat weight of 15% to 20% and a liver weight of 2.5%, over 90% of
stored dioxin will be in adipose tissue.
Direct initial exposure to the lung in these studies is also difficult to estimate. Both
inhalation and skin absorption are the equally likely routes of initial exposure, but the
exposure scenario cannot be distinguished. Di Domenico and Zapponi (1986) estimated that
-50% to 90% of TCDD exposure to the Seveso residents following the 1976 accident
occurred via the dermal route, but they assumed 100% dermal and inhalation absorption. A
more likely 1% to 10% dermal and 75% inhalation absorption estimate (U.S. Environmental
Protection Agency, 1985) would project that the inhalation route provided the major TCDD
exposure. To further complicate the situation, the cohorts discussed below are all
occupational so that both dermal and inhalation exposure are highly likely.
The data on body concentration levels in the three studies are presented in Table 8-7.
Fingerhut et al. (1991) measured serum levels adjusted for lipids in a sample of 253 of the
workers from 2 of the 12 plants approximately 20 years after last known exposure. They
found a highly statistically significant correlation (r=0.72; p< 0.0001) between the logarithm
of number of years of exposure to processes involving TCDD contamination and the
logarithm of individual TCDD serum levels. Based on this correlation, they divided the
sample into a high-exposure group (defined as those exposed more than 1 year) and a low-
exposure group (those exposed < 1 year). The mean TCDD level of the low-exposure group
was 69 ppt, while that of the high exposure group was 418 ppt. Among the 176 sampled
workers last exposed >20 years before, those with under 1 year of exposure (n=81) had a
mean level of 78 ppt, and those with over 1 year of exposure (n=95) had a mean level of
462 ppt.
For the Zober et al. (1991) study, the serum-level data are based on a nonrandom
sample of 28 survivors tested 32 years after the 1953 accident. These subjects were then
classified into three groups by scenario of high (Cl), medium (C2), or low (C3) chance of
TCDD exposure, with the mean (median) levels of 60 (24.5), 25 (9.5), and 25 (8) ppt,
respectively. An alternative breakdown (B) (see Table 8-7) by the 16 sample subjects who
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Table 8-7. Measured Serum TCDD Levels and Estimated Levels at Time of Last Occupational Exposure to TCDD, Based on First-Order Elimination Kinetics and
a Half-Life for Elimination of 7.1 Yean
Study
Fingerhut et al., 1991
Zober et al., 1991
Breakdown A
Breakdown B
Sample
Surviving cohort of workers from Plants 1 and 2 tested
approximately 20 yean after last occupational exposure
Exposed >1 year, all
Exposed >1 year, & 20-year latency
Exposed <1 year, all
Exposed < 1 year, fe 20-year latency
Sample of survivors tested 32 years after the accident
High-exposure scenario
Medium-exposure scenario
Low-exposure scenario
Either chloracne or erythema
Neither chloracne nor erythema
Sample
size
253
119
95
134
81
28
10
7
11
16
12
Concentration at test time
1 year, since estimate not given for & 20-year latency.
(continued on the following page)
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exhibited either chloracne or erythema versus those 12 who did not yields estimates of 50
(15) and 25 (5.8) ppt.
For the Manz et al. (1991) study, serum levels were measured in 48 unselected
members of the cohort who had what the authors believed to be similar exposures to TCDD
as the cohort. Workers were classified into three groups, having either high-, medium-, or
low-exposure opportunities, and interviews with these 48 members led to a division of 37
into the high group (mean 296 ppt, median 137 ppt) and 11 into the medium- and low-
exposure groups combined (mean 83 ppt, median 60 ppt).
Also included in Table 8-7 are measured serum levels of U.S. veterans of Operation
Ranch Hand and a sample of 100 U.S. men from the general population. The mean U.S.
estimate of 5 ppt is identical to that reported from four controls in the German population
(Schecter et al., 1988). The Fingerhut et al. (1991) referent controls had a mean level of 7
ppt.
Table 8-7 also presents estimates of median TCDD concentrations at time of last
exposure for the median levels of various cohorts, based on first-order elimination kinetics,
assuming a 7.1 year half-life. To be consistent with the requirements of the model,
background levels of 5 ppt are subtracted from each median before back extrapolation. The
formula used is:
Ct = C0e-k->
where Ct = concentration at time of measurement, C0 = estimated concentration at time of
last exposure, kg = elimination constant (per year) = 0.098, and t = years since last
exposure. For the Zober (1991) and Manz (1991) studies and the Fingerhut short-exposure
subcohort, these concentrations C0 can be considered to be from short-term exposure, and
average fat concentrations can be calculated from the formula:
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T
C = 1
where T = length of study (see Table 8-8). For the Fingerhut long-exposure subcohort with
an average exposure period of 6.8 years, average concentration during the exposure period is
estimated as 50% of the calculated C0 = 1770 ppt. Since the average length of followup for
this subcohort is 30 years, the total time from start of exposure to end of exposure is
estimated as 9 (30-21) years. This leads to a total concentration x time for this subcohort of
(Table 8-8):
21
9 X 850 + f 1770e-ttdt = 24,750 ppt - years
Table 8-8 presents calculations of equivalent exposure estimates to convert from the dose
metric of total fat concentration X time to intake dose. The process is to (1) calculate the
continuous lifetime average daily uptake dose that will produce an equivalent total
concentration X time and (2) to estimate the intake (oral) lifetime average daily dose
(LADD) that would result in the continuous uptake dose.
To calculate (1), the assumptions of steady state discussed in Chapter 6, Volume II of
U.S. EPA (1994) appear appropriate. These lead to their eq. 6-11, which is:
D = [InA] Vf Cf,M
1
where D = uptake dose, Vf = volume of distribution of fat, Cf>M = steady-state
concentration of dioxin in fat, and tin = 7.1 x 365 = 2,591.5 days
To calculate D, set Vf = 14L and first calculate Cf>M from age 21 to the age at the
end of study, which will yield the same average concentration for each of the subcohorts
described. For the Fingerhut et al. long-exposure subcohort, the equivalent Cf)gs is
(24,750/42 years) = 589 ppt. Constant daily uptake, D, for this subcohort is then 30.2
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Table 8-8. Estimates of Lifetime Average Daily Dose for Oral Intake Equivalence for TCDD Based on Total Concentration x Time Equivalence. Estimates of
TCDD Concentration Adjusted for Background at Time of Sampling and Back-Calculated Using First Order Elimination Kinetics, by Cohort
Study
Fingerhut et al. (1991) £
20-year latency
Exposed >1 year
Exposed <1 year
Zoberetal., 1991
Breakdown A
High
Medium
Low
Breakdowns
Either chlorance or
erythema
Neither chloracne nor
erythema
Manx et al., 1991
High-exp. scenario
Medium- and low-exp.
scenario
Estimated Median
Fat Concentration
at Time of Last
Exp. Adjusted for
Background'
ppt Year
1770 (1966)
ISO (1966)
450 (1953)
105 (1953)
100 (1953)
230 (1953)
20 (1953)
2750 (1954)
1150 (1954)
Average Time
from Start of
Exposure to End
of Study
Yean
30-
28*
34"
34"
34"
34"
34"
37"
37"
Average Age
at End of
Study
Years
63"
56C
70"
70"
70"
70"
70-
70"
70"
Total
Cone. X
Tune
ppt - Years
24,750
1,430
4,430
1,030
980
2,260
200
27,300
11,422
Average
Concentration
From Start of
Exposure to
End of Study
ppt
790
51
130
30
29
67
6
738
309
Equivalent
Average
Cone, for
Exposure
from Age
21 to Age
at End of
Study
ppt
589
41
90
21
20
46
4
557
233
Constant
Daily Uptake
for
Equivalent
Average
Cone.
pg/kg-day
31.5
2.2
4.9
1.1
1.1
2.5
0.2
30.0
12.5
Lifetime
Average
Daily Dose
for Oral
Intake
Equivalence
pg/kg-day*
63.0
4.4
9.7
2.3
2.2
5.0
0.4
60.0
25.0
oo
O
O\
u3
O
• From Table 8-9. Person Years/Cohort Size + 20 years.
b Estimated from study descriptions.
c From Fingerhut et al., 1991.
d Estimated from study descriptions and comparison with Fingerhut et al.
* Assumes 50% absorption by oral route.
* From Table 8-7, dates in parentheses are estimated dates of last exposure.
-------
DRAFT-DO NOT QUOTE OR CITE
pg/kg-b.w. Daily intake, or calculation (2), assuming 50% absorption from dietary, is then
estimated as 63 pg/kg-b.w.
8.5.3.3. Calculation of Risk Estimates
Table 8-9 presents the LADDs from Table 8-8, the estimated relative risks, and
sample size information for the various subcohorts. Whenever the data could be found, the
subcohorts with at least 20-year latency are presented, in order to coincide as closely as
possible with the Fingerhut et al. (1991) cohort. For the Manz et al. (1991) study, no data
on person-years at risk are available, so only the relative risk model could be used to
estimate risk. For the other studies, both models could be used. The data are shown in
Figure 8-11 and indicate trends with increasing LADDs for all three studies and for both
respiratory cancers and all cancers combined.
U.S. EPA practice for presenting risk estimates based on human data has been to use
point estimates or maximum likelihood estimates, rather than upper-limit risk estimates.
Calculations of the incremental unit risk estimates for lung cancer and all cancers
combined are presented in Tables 8-10 and 8-11, respectively, for each of the three cohorts
separately and all cohorts combined, for both the additive and multiplicative risk models.
The results show statistically significant estimates of the slope parameter for the Fingerhut
(1991) study, the Manz (1991) study, and all studies combined. Although the slope estimates
for the Zober (1990) study are greater than those for the Fingerhut (1991) and Manz (1991)
studies, the cohort is smaller and statistical significance is seen only for all cancer deaths
combined in the subcohort with chloracne or erythema. Since the Fingerhut (1991) data
provide the bulk of the weight, the estimates from the combined studies are closer to those
based on the Fingerhut (1991) study alone than to the others.
Also shown in Tables 8-10 and 8-11 are estimates of the lifetime incremental cancer
risk for 1 pk/kg-day LADD intake. These are derived by substituting the MLE estimates of
B back into the age-specific hazard rates and deriving lifetime incremental risk estimates
based on lifetable probabilities with competing risks (for practical purposes the procedure
described in the table footnotes produces nearly the same results). For lung cancer these unit
risk estimates range from SxlO"4 to 2xlO~2 (pg/kg-day)'1, with the estimates for all studies
8-93 06/30/94
-------
Table 8-9. Estimated Lifetime Average Daily Doses and Relative Risks by Individual Study Cohort
Study
Fingerhut et al.
(1991)
Zober et al.
(1990)
Manz et al.
(1991)'
Cohort
Exposed > 1 year all
Exposed < 1 year all
Exposed >1 year,
a 20-year latency
Exposed <1 year,
a 20-year latency
(^20-year latency)
High-exposure
scenario
Medium-exposure
scenario
Low-exposure
scenario
Either chloracne or
erythema
Neither chloracne nor
erythema
(^20-year latency)1*
High-exposure
exposure scenario
Medium- and low-
exposure scenario
LADD
(pg/kg-day)
63.0
4.4
9.7
2.3
2.2
5.0
0.4
60.0
25.0
Cohort size
Ni
5,172
1,520
1,516
57
74
81
109
103
96C
200*
Person
years
116,748
15,136
12,299
673
563
676
1,209
703
No data
No data
Respiratory
cancer deaths
Obs. Exp. Rel. Risk
96 84.5 1.13
43 30.2 1.42
19 18.4 1.03
3 1.25 2.52
2 1.03 1.94
0 0.97 0
5 2.09 2.39
0 1.16 0
N = 1,148
30 21.3 1.41
All cancer deaths
Obs. Exp. Rel. Risk
265 229.9 1.15
114 78.0 1.46
48 46.8 1.02
7 4.20 1.67
8 3.37 2.38
1 3.39 0.29
2.01
14 6.96
0.50
2 4.00
(Entry before 1955)
18 8.5 2.11
33 23.1 1.43
OO
0
O
6
n
o
VO
•West German reference controls used for consistency with Fingerhut and Zober studies.
""Assumed because of entry before 1955; actual data unavailable.
'Entry before 1955.
-------
Figure 8-11. Relative Risks of Lung Cancer and All Cancer Mortality in Three Recent
Studies of Workers Exposed to TCDD, by Estimated LADD Equivalence.
2.5
Fingerhut •
Zober •
Manz •
Lung Cancer
All Cancers
00
V)
£
0)
1
0)
oc
2
1.5
0.5 •
60
H
70
Lifetime Ave. Daily Dose Equiv.
pg/kg-day
Figure 8-11. Relative risks of lung cancer and all cancer mortality in three recent cohort
studies of workers exposed to TCDD by estimated LADD equivalence
-------
Table 8-10. Calculation of Incremental Unit Cancer Risk Estimates and 95% Lower Limits for Both the Additive and Relative Risk Models Based on the Lung
Cancer Deaths Response in the Fingerhut, Zober, and Manz Studies
00
Study
Fingerhut et al., 1991
Zober et al., 1990
Manz et al., 1991"
All studies combined
Model used
Additive
Multiplicative
Additive
H-M-L scenario
Chloracne or not
Multiplicative
H-M-L scenario
Chloracne or not
Multiplicative
Additive'
Multiplicative'
Parameter
estimates
ba
1.3X10-2
6.730
2. ISxlO'1
4.45x10-'
126.55
258.13
11.67
1.38xlO-2
7.862
Asymptotic
variance estimates
4.69xl05
11.83
5.28X10'2
1. 30x10-'
16476.6
43519.2
53.98
4.73xlO-5
9.90
p-Valuc for slopeb
0.02
0.02
0.17
0.11
0.16
0.11
0.06
0.02
0.01
Lifetime incremental cancer risk per 1
pg/kg-day LADD intake
Lower 95* limit
7.6xlO-5
4.1xlO-J
0
0
0
0
0
8.8xlO-J
l.OxlO-1
likelihood
estimate'
4.7x10^
2.6x10-*
7.falOJ
1.6xlOJ
4.8x10*
9.8rtOJ
4.4x10"
4.8x10"
3.0x10"
'Estimates for additive risk model given in (ng/kg-day)"1. Estimates for multiplicative risk model given in (ng/kg-day) '/h^t) where hg(t) = background age-specific hazard rate.
bOne-sided test based on asymptotic variance.
°MLE for additive risk model approximated by multiplying b by 35 since risk before age 35 is close to zero. MLE for multiplicative risk model approximated by multiplying
b by P0=0.038.
''Dataset in Table 8-9 with estimated LADD = 40 pg/kg-day.
"Fingerhut and Zober (chloracne) cohorts only.
^ingherhut, Zober (chloracne), and Manz. See also footnote d for Manz.
b
o
1
o
s
o
vo
-------
DRAFT-DO NOT QUOTE OR CITE
I
ll
I*
1
>9
H
:-i
•i|
.§ ^
Is.
1
esti
«,
1 a
i
&.
1
38
6 d
28
d o
PS
•i H
a SB
D
S
s
I '
X
.0
.9
o S I
111
9 ™ >»
IS-0
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•C 8 •§•
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.S o
si? f
in?*j
•i § 1111
1 I S H S 8
ij.ljj5
VfltkSl
8-97
06/30/94
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DRAFT-DO NOT QUOTE OR CITE
combined between 3x10^ and 5x10"* (pg/kg-day)"1. For all cancers combined the range of
MLE estimates is between IxlO"3 and 8xlO~2 (pg/kg-day)"1 with the estimates based on all
studies combined between 2xlO"3 and 3xlO'3 (pg/kg-day)"1.
These estimates from all studies combined (both lung cancer and total cancers) range
from SxlO-4 to 3xlO'3 (pg/kg-day)-1. They exceed the upper-limit estimate of 1.6X1Q-4
(pg/kg-day)-1 previously derived by EPA (U.S. EPA, 1985) based on the total cancer
response in the female Sprague-Dawley rat in the Kociba et al. (1978) lifetime feeding study
and the LMS model. Using the same Kociba (1978) study and LMS model but with the liver
histopathology rereadings from a recent reanalysis (Sauer and Goodman, 1992) and original
Kociba readings for the other tumor sites, the upper-limit estimate is O.SxlO"4 (pg/kg-day)-1.
Based on the LMS model and only the liver tumors (Sauer and Goodman reanalysis), the
upper-limit estimate is 0.5x10"* (pg/kg-day)-1. This compares closely with the MLE estimate
(for rats) of 0.24 (pg/kg-day)-1 provided by the two-stage model in Section 8.2.2, which uses
the same liver tumor pathology, together with additional liver foci data. Using a default
(body-weight)* power for rat-to-human conversion, the two-stage MLE estimate becomes 0.9
(pg/kg-day)'1. These estimates are shown in Table 8-12.
8.5.3.4. Low-Dose Deviation From Linearity
Based on the LADDs and the relative risk estimates presented in Table 8-9, some idea
of the degree of nonlinearity in the dose response for these cancers can be derived. For the
Fingerhut et al. (1991) cohort, the ratio of 14.3 for high to low LADDs (63.0/4.4)
corresponds to a ratio of increased risk of 14 (0.42/0.03) for respiratory cancer and 23
(0.46/0.02) for all cancer mortality combined. For the Manz et al. (1991) cohort, the
comparisons are also consistent with linearity; the LADD ratio of 2.4 (60/25) corresponds to
an increased risk ratio of 2.6 (1.11/0.43). The Zober et al. (1990) chloracne versus no-
chloracne cohort with a high-to-low dose ratio of 12 suggests some low-dose sublinearity, but
the low-dose estimate is very close to background and no quantitative comparisons can be
derived, since the relative risk estimates for this low-dose group are < 1. For the Saracci et
al. (1991) cohort, no direct comparison can be made either, except to note that the relative
risk for lung cancer for the low-exposure group was actually higher than that for the high-
8-98 06/30/94
-------
Table 8-12. Estimates of U.S. EPA Unit Cancer Risk for TCDD Oral Intake, Based on Animal and Human Studies and U.S. EPA Current and Proposed Estimates
Source
Animal (Female
Sprague-Dawley rat)
All Based on Kociba
et al. (1978)
Human (Males)
Fingerhut et al.
(1991)
Zober et al. (1990)
Manz et al. (1991)
U.S. EPA Proposed
(1988)
U.S. EPA Currently
Proposed (1994)
Cancers
Liver
All Giver, lung,
hard palate/nasal
turbinate)
Lung
All
Oral Dose
Range
1-100 ng/kg-day
1-60 pg/kg-day"
Model
2-Stagea
LMS
LMS
LMS
Multistage Weibull
(Incidental Tumor
Analysis)
Additive Risk
Multiplicative Risk
Additive Risk
Multiplicative Risk
Based on reciprocal of
risk specific dose for
10"* incremental risk
Estimates of Unit Risk
(pg/kg-day)-1
MLE 95% Upper
Limit
0.9x10^ -
- 0.5x10"
- 0.8x10"
1.2x10" 1.6x10""
2.1x10" 3.1x10"
4.8x10" •
3.0x10"
27x10"
17x10" •
0.1x10"
1x10"
Comments
Liver Pathology
Readings by Sauer
and Goodman (1992)
Liver Pathology by
Kociba (1978) and
Squire (1980)
Calculations based on
combined cohorts
External Review
Drafts
Ref . for
Calculation
Chapter 8, Section
8.2.2
U.S. EPA (1992)"
U.S. EPA (1992)"
U.S. EPA (1985)
U.S. EPA (1988)
Chapter 8, Section
8.5.3.3
U.S. EPA (1988)
Chapter 9, Risk
Characterization
oo
'Animal estimate of 0.24x10" (pg/kg-day)'1 times rat-to-human default conversion of (70/0.350)0-25.
'Unpublished.
'Estimates based on total concentration x time equivalence.
"Estimate currently used by U.S. EPA.
"EPA practice is to use MLEs for estimates based on human data.
-------
DRAFT-DO NOT QUOTE OR CITE
exposure group. Thus, for two of the cohorts there is some suggestion of sublinearity in
dose response, but for the other two linearity is appropriate.
Estimates of LADDs of TCDD intake in the general U.S. population range from 0.3
to 1.0 pg/kg-day (U.S. EPA, 1994). The LADDs estimated in three epidemiology studies
analyzed range from 0.4 to 63 pg/kg-day above background, with increased risks suggested
in the 2-5 pg/kg-day. The LADD estimates themselves are just too imprecise for more
definitive statements (see discussion in Section 8.5.3.5 below).
8.5.3.5. Uncertainties in Estimates From Human Epidemiology
There are many uncertainties associated with the unit risk estimates just derived from
the epidemiology studies, the two largest being hazard identification and dose estimation.
The evidence for a dioxin lung cancer hazard in humans is suggestive but not conclusive (see
Chapter 7), while that for all cancers combined has less certainty. The estimates of dose,
while based on actual body measurements, may lack both representativeness and precision.
Although 253 subjects were sampled in the Fingerhut study, they were all taken decades after
last exposure and were from two plants. Subjects from the larger plant, plant 1, had the
higher dioxin levels but a lung cancer SMR=72 based on seven deaths, while the smaller
plant had only one death from lung cancer (SMR=155). Analysis by plant in the Fingerhut
study would have been possible if body measurements at these other 10 plants had been
available.
Two choices of parameters, both of which affect LADD estimates by approximately a
factor of two, provide some estimate of uncertainty. First, for back calculation for estimates
of total body burden, a one-compartment first-order elimination model with a human half-life
of 7.1 years has been assumed. Recent data, however, suggest a longer half-life of 11.3
years (Wolf et al., 1994). Use of this longer half-life would increase the unit risk estimate
by about 40%. On the other hand, the body levels measured were quite variable and not
symmetrically distributed within each study. This led to the selection of a median rather than
a mean fat concentration for back extrapolation of the dioxin levels. If the mean had been
used, the unit risk estimates based on these studies would have been approximately 50% to
70% less and much closer to those from the animal studies. Also, the estimated exposure in
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DRAFT-DO NOT QUOTE OR CITE
the Zober chloracne subcohort is lower than levels thought to cause chloracne. This could
explain why the unit risk estimates based on the Zober study are higher than those of the
other two.
Another uncertainty is that of a possible interaction between dioxin and tobacco
smoking. In mice, dioxin and 3-methylcholanthrene (3-MC, one of the many polyaromatic
hydrocarbons in tobacco smoke) have been shown to be cocarcinogenic (Kouri et al., 1978;
U.S. EPA, 1985). Other studies of mouse skin tumors have shown that dioxin can have
anticarcinogenic properties when administered before initiation with either 3-MC or
benzo(a)pyrene (U.S. EPA, 1985). Furthermore, dioxin's tumor-promoting ability suggests
that two-stage models would be more appropriate if individual smoking histories were
known. Individual smoking histories are presented only for the 37 cancer cases and deaths in
the Zober cohort; only 2 were stated as being nonsmokers. All seven men with lung or
larynx cancer were smokers. While characteristics of the two subcohorts in Fingerhut et al.
(1991) suggest similar smoking prevalence, the effects with higher levels of dioxin could be
synergistic.
A synergism of dioxin with 4-aminobiphenyl, a known human bladder carcinogen, has
been suggested (Collins et al., 1993) for one of the plants in the Fingerhut study, but tobacco
smoking would be present in all the plants and would seem to be a more likely universal
effect modifier.
8.5.3.6. Conclusions
Epidemiology studies suggest that the lung in the human male is a much more
sensitive target organ for TCDD than is the liver and that the human is a sensitive species for
cancer response, probably more sensitive than is the rat. Although smoking may be a
modifier for the lung cancer response, the studies also show increases for all cancers
combined. Estimates derived from the human data suggest a unit risk for lung cancer of 3 to
SxlO"4 (pg/kg-day)"1; for all cancers combined the unit risk estimate is 2 to 3xlO"3 (pg/kg-
day)"1. While unit risk estimates based on rat tumors are somewhat less, they are within the
range of uncertainty of those based on human data. Both animal and human responses are
consistent with low-dose linearity.
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8.6. KNOWLEDGE GAPS
Considerable information is now available on the mechanisms of action responsible
for TCDD's effects in experimental animals and humans, and important new information is
now being generated. These data are, of course, essential to the development of reliable
biologically based models for the estimation of human risks as a consequence of exposure to
TCDD and its structural analogs such as the CDFs and coplanar PCBs. Uncertainty in such
models reflects incomplete knowledge of mechanisms and inadequacies in exposure/tissue
dose relationships. In the process of developing and evaluating biologically based models,
we can identify those knowledge gaps that create uncertainty. The idea that interaction of
TCDD with the Ah receptor is an essential first step in most, if not all, of dioxin's effects
has been considered as a reasonable assumption for over a decade. The recent Banbury
Conference on dioxin formalized this as a general consensus among dioxin researchers. The
development of models that accurately predict risks also requires tissue and cell dosimetry
data in experimental animals and humans. This kind of dosimetry information is available
for blood, liver, and adipose tissue, but dosimetry data in other target tissues such as the
lung, skin, pituitary, and reproductive tract are not available or incomplete. It would be
especially relevant to the development of biologically based dose-response models to have
dosimetry data (relationship between exposure, dose, and cell-specific dose) in target cells
when the target cell is known. For example, the lung is composed of numerous cell types,
but the identity of the target cell(s) for TCDD-mediated lung cancer is not known nor are
there many data on dose-response relationships for concentrations of TCDD in whole lung or
discrete cell types. Since the vast majority of dioxin is found in the liver and adipose tissue
under chronic exposure steady-state conditions and the lung is clearly a target organ for
biochemical and toxic effects, it would seem that the lung and perhaps other organs require
far less tissue/cell levels of dioxin to exhibit toxic effects than the liver.
One of the most confounding yet important knowledge gaps in the development of
mechanistic models is the evaluation of the adverse health consequences, if any, of current
background exposure to the CDDs and CDFs, which is estimated at 1 to 3 pg TCDD
equivalents/kg/day. More accurate information on the potency of dioxin-like PCBs is also an
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DRAFT-DO NOT QUOTE OR CITE
essential component in evaluating the health impact of background exposure to chemicals that
bind the Ah receptor.
Many of the molecular events that follow binding of TCDD to the Ah receptor are
now known for transcriptional activation of the CYP1A1 gene. However, there is little
information on the characterization of analogous events for dioxin's many other effects on
gene expression such as Ah receptor-mediated alterations in the EOF or estrogen receptor.
Most of the mechanistic or dose-response information on dioxin's effects has been generated
on changes in gene expression of single genes such as CYP1A1 induction. There is only
limited information on the complex interaction of biochemical, molecular, and biological
events that are necessary to produce a frank toxic effect such as cancer, developmental
defects, reproductive effect, or neurological effects. Figure 8-12 summarizes the series of
interconnected steps within the three major components of receptor-mediated events
(recognition, transduction, and response). Although this scheme is simplified (i.e., each step
may comprise several events), it does provide a framework for identifying knowledge gaps
that create uncertainty. Clearly, interactions with other endocrine systems are involved in
some effects, and our ability to construct accurate dose-response models for noncancer end
points would be enhanced if we had a better understanding of TCDD/endocrine interactions.
One of the more active areas of research on hormone action is directed at identifying
the cell-specific factors that produce diversity of responses for receptor-mediated responses,
that is, how do a single receptor and a single ligand produce the wide spectrum of cell-
specific responses characteristic of exposure to a given hormone. Since TCDD is acting like
a potent and persistent hormone agonist/antagonist, the mechanisms responsible for
qualitative and quantitative differences in dose-response relationships for Ah receptor-
mediated events might be similar to those mechanisms identified for steroid hormones.
Fuller (1991) has summarized some of the mechanisms responsible for generating diversity,
and these are listed below:
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Oioxin Exposure
Free Oioxin in Tissues
Dioxin Binding to the
Ah Receptor in Tissue
Ah Receptor - Dioxin Complex
Binding with DNA
I
Gene Regulation
I
m-RNA Regulation
Protein Synthesis
Biochemical Alterations
Early Cellular Responses
(cell growth stimulation)
Late (irreversible) Tissue
Response (cancer, terata)
Interactions
of Multiple
Target Genes
Figure 8-12. Biologically based risk assessment approaches for dioxin: Filling the gaps
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ligand < agonist, antagonist
target tissue < receptor gene expression
activating or inactivating enzymes
binding proteins (extra- or intracellular)
receptor < cytoplasmic versus nuclear
isoforms-differential
splicing
gene duplication
dimers < hetero- or homodimers
DNA binding factors
nuclear factors < antagonist isoforms
squelching
response elements < consensus versus nonconsensus
number of copies
position
proximity of other response elements
transactivation < gene-specific factors
cell-specific factors
In addition to the above considerations, there is considerable speculation regarding the
normal cellular functions of the Ah receptor and the identity of any endogenous ligands for
the Ah receptor. If sound scientific information were available on the normal functions of
the receptor, especially if those functions involve regulation of cell proliferation and
differentiation, it would greatly enhance our ability to predict the health consequences of
low-level dioxin exposure. It would also help considerably in the selection of appropriate
animal models for estimating dioxin risks.
Interindividual variation in human responses to TCDD and its structural analogs is one
of the most difficult issues to accommodate in the development of biologically based dose-
response models. We know from epidemiology studies that some individuals develop
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DRAFT-DO NOT QUOTE OR CITE
chloracne from a given exposure to dioxin, whereas other individuals exposed to the same
amount of dioxin do not develop chloracne. The mechanisms responsible for sensitivity or
resistance to the chloracnegenic actions of dioxin are not known, nor is there any information
on the relationship of chloracne to other toxic effects. For example, are individuals who are
susceptible to chloracne also susceptible to the carcinogenic actions of dioxin? Likewise,
there are considerable differences in the magnitude of enzyme induction when human cells
are cultured with dioxin. We need to understand the molecular mechanisms responsible for
these differences and whether high inducers are more or less susceptible to the toxic effects
of dioxin and its structural analogs. These kinds of data would allow the development of
epidemiologic and laboratory approaches for evaluating health consequences in both sensitive
or resistant populations.
In summary, we have gained considerable and valuable insights regarding mechanisms
of dioxin and dose-response relationships for dioxin effects. These data are not yet complete
but are appropriate for the development of preliminary biologically based models that may
eventually be useful for estimating dioxin's risks to humans. When sufficiently developed,
these models should provide increased confidence and decreased uncertainty than are present
with the current default approaches (LMS or safety factor). They should also accommodate
new scientific information from research directed at filling knowledge gaps to further reduce
uncertainty. Based on the model structures presented in this chapter, it should be possible to
design specific experiments to fill key knowledge gaps.
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REFERENCES FOR CHAPTER 8 AND APPENDICES C AND D
Abbott, B.D.; Bimbaum, L.S. (1989) TCDD alters medial epithelial cell differentiation during palatogenesis.
Toxicol. Appl. Pharmacol. 99: 276-288.
Abbott, B.D.; Bimbaum, L.S. (1990a) Rat embryonic palatal shelves respond to TCDD in organ culture.
Toxicol. Appl. Pharmacol. 103: 441-451.
Abbott, B.D.; Bimbaum, L.S. (1990b) Effects of TCDD on embryonic ureteric epithelial EOF receptor
expression and cell proliferation. Teratology 41: 71-84.
Abbott, B.D.; Bimbaum, L.S. (1991) TCDD exposure of human embryonic palatal shelves in organ culture
alters the differentiation of medial epithelial cells. Teratology 43: 119-132.
Abbott, B.D.; Diliberto, J.J.; Bimbaum, L.S. (1989) 2,3,7,8-tetrachlordibenzo-p-dioxin alters embryonic palatal
medial epithelial cell differentiation in vitro. Toxicol. Appl. Pharmacol. 100: 119-131.
Abbott, B.D.; Harris, M.W.; Bimbaum, L.S. (1992) Comparisons of the effects of TCDD and hydrocortisone
on growth factor expression provide insight into their interaction in the embryonic mouse palate.
Teratology 45: 35-53.
Abraham, K.; Krowke, R.; Neubert, D. (1988) Pharmacokinetics and biological activity of
2,3,7,8-tetrachlordibenzo-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.
Andersen, M.E.; Mills, J.J.; Birnbaum, L.S.; Connolly, R.B. (1993a) Stochastic dose-response modeling of
hepatic promotion by dioxin [abstract]. Toxicologist 13: 000.
Andersen, M.E.; Mills, J.J.; Gargas, M.L.; Kedderis, L.; Birnbaum, L.S.; Neubert, D.; Greenlee, W.F.
(1993b) Modeling receptor-mediated processes with dioxin: Implications for pharmacokinetics and risk
assessment. Risk Analysis 13(1): 25-36. (Appendix B of this document).
Arley, N.; Iverson, S. (1952) On the mechanism of experimental carcinogenesis, III. Further development of
the hit theory of carcinogenesis. Acta. Path. Microbiol. Scan. 30: 21-53.
Armitage, P.; Doll, R. (1954) The age distribution of cancer and a multistage theory of cancer. Br. J. Cancer
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TOXICOLOGY AND APPLIED PHARMACOLOGY 120, 138-154 (1993)
APPENDIX A
A Mechanistic Model of Effects of Dioxin on Gene Expression
in the Rat Liver
MICHAEL C. KOHN, GEORGE W. LUCIER, GEORGE C. CLARK, CHARLES SEWALL,
ANGELIKA M. TRITSCHER, AND CHRISTOPHER J. PORTIER
National Institute of Environmental Health Sciences, P.O. Box 12233,
Research Triangle Park, North Carolina 27709
Received June 15, 1992; accepted January 5, 1993
A Mechanistic Model of Effects of Dioxin on Gene Expres-
sion in the Rat Liver. KOHN, M. C., LUCIER, G. W., CLARK,
G. C., SEWALL, C., TRITSCHER, A. M., AND PORTIER, C. J.
(1993). Toxicol. Appl. Pharmacol. 120,138-154.
Improved methods for estimating the shape of the response
curve for effects of exposure to 2,3,7,8-tetrachlorodibenzo-p-
dioxin (TCDD) are needed in order to evaluate possible adverse
health effects of TCDD. A mathematical model has been con-
structed to describe TCDD-mediated alterations in hepatic pro-
teins in the rat. In this model it was assumed that TCDD medi-
ates increases in the liver concentration of transforming growth
factor-a (TGF-a) by a mechanism which requires the aryl hy-
drocarbon (Ah) receptor. TGF-a subsequently binds to the epi-
dermal growth factor (EGF) receptor, a process which is known
to cause internalization of this receptor in hepatocytes. This
action is thought to be an early event in the generation of a
mitogenic signal. Because TCDD decreases binding of EGF in
the livers of intact female rats but not in ovariectomized rats,
this effect was further assumed to be dependent on estrogen
action. The model postulates Ah receptor-dependent effects on
the concentration of cytochrome P450 1A2 (CYP1A2), which is
involved in the metabolism of estradiol, and on the concentra-
tion of the estrogen receptor. The model also incorporates in-
formation on induction of cytochrome P450 1A1 (CYP1A1) by
TCDD. The biochemical response curves for all these proteins
were hyperbolic (Hill exponents in the equations for their ex-
pression were found to be 1), indicating a proportional relation-
ship between target tissue dose and protein concentration at low
administered doses of TCDD. The model successfully repro-
duced the observed tissue distribution of TCDD, the concentra-
tions of CYP1 Al and CYP1A2, and the effects of TCDD on the
Ah, estrogen, and EGF receptors over a wide dose range. © 1993
Academic Press, Inc.
2,3,7,8-Tetrachlorodibenzo-/»-dioxin (TCDD) is a potent
carcinogen which is associated with increased incidence of
liver tumors in female rats but not in male rats (Kociba et
ai, 1978; National Toxicology Program, 1982; Clark et al,
1991). Similarly, TCDD enhances hepatocyte proliferation
and stimulates development of enzyme-altered hyperplas-
tic foci in intact female rats but not in ovariectomized rats
(Lucier et al., 1991), suggesting that estrogens play a major
role in TCDD-mediated hepatocarcinogenesis.
Biochemical signals regulating cellular proliferation are
mediated in many tissues, including the liver, by the epider-
mal growth factor (EGF) receptor (Schlessinger et al.,
1983). The EGF receptor is a member of a family of plasma
membrane receptors that, on binding of ligand, transduce
signals by tyrosine kinase activity and by internalization of
the liganded receptor (Hunter and Cooper, 1985). Phos-
phorylation of various proteins by this tyrosine kinase leads
to alterations in cellular regulation and mitotic activity.
Livers of TCDD-treated rats show a dose-dependent de-
cline in the maximal binding of EGF (Lucier, 1991) al-
though TCDD has no effect on the amount of mRNA for
the EGF receptor in mouse liver (Lin et al, 199la). The
effect of TCDD on mRNA for the EGF receptor is a tissue-
specific response. TCDD affects the amount of mRNA for
the EGF receptor in the uterus (Astroff et al, 1990) but not
in keratinocytes (Osborne et al, 1988). Antibodies raised
against the EGF receptor stain the plasma membranes of
hepatocytes in control rats, but in TCDD-treated rats there
is an apparent redistribution of the receptor into the cytosol
(C. Sewall and A. M. Tritscher, unpublished results). This is
consistent with the notion that the loss of EGF binding
capacity in liver plasma membranes is due to internaliza-
tion of the liganded receptor.
The liver does not produce EGF, but there is evidence
that it does produce several other EGF-like peptides such as
transforming growth factor-a (TGF-a), another ligand of
the EGF receptor (Mead and Fausto, 1989). TCDD induces
expression of TGF-a in keratinocytes (Choi et al, 1991),
suggesting that TCDD may also induce TGF-a in the liver.
Increased production of this peptide would subsequently
stimulate EGF receptor-mediated events in that organ.
In order to obtain a quantitative relationship between
exposure to TCDD and consequent alterations in the prop-
erties of the EGF receptor, we have constructed a mathe-
matical model of the tissue distribution of TCDD in the rat
004I-008X/93S5.00
Copyright © 1993 by Academic Press, Inc.
All rights of reproduction in any form reserved.
A-l
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MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
and its effect on the concentrations of several important
liver proteins. The model includes equations for the aryl
hydrocarbon (Ah) receptor-dependent induction of cy-
tochrome P450 isozymes 1A1 (CYP1A1) and 1A2
(CYP1A2) and of the Ah receptor itself. The complex be-
tween this receptor and TCDD (Ah-TCDD) is treated as
inducing expression of TGF-a although the induced ligand
of the EGF receptor actually may be another EGF-like pep-
tide. It is also assumed that estrogen action is required for
TCDD-mediated induction of TGF-a. TGF-a is modeled
as released into the liver interstitium. From there it binds to
the EGF receptor, causing its internalization. The model
calculates the distribution of the receptor between the
plasma membrane and cytosol.
Because estrogens appear to be required for TCDD-me-
diated effects on the EGF receptor, production of the estro-
gen receptor, CYP1 A2-catalyzed formation of catechol es-
trogens, and deactivation of estrogens by glucuronidation
were included in the model. In this effort, data on the stimu-
lation of estrogen receptor synthesis by estradiol and the
inhibition of estrogen receptor synthesis by TCDD were
incorporated.
The model's predictions were compared to the data of
Tritscher et al. (1992) and Sewall et al. (1993). Their experi-
ments were a two-stage protocol in which female Sprague-
Dawley rats were injected with an initiating dose of diethyl-
nitrosamine (DEN). After 14 days the rats were subjected to
biweekly gavage with TCDD in corn oil at doses equivalent
to 3.5-125 ng/kg/day for 30 weeks. The rats were killed 1
week following the last dose, and their livers were assayed
for TCDD, CYP1A1, CYP1A2, and plasma membrane
binding of EGF. All such measurements were performed
using samples from the same livers.
This model was used to predict tissue concentrations of
TCDD and concentrations of induced proteins following
administration of TCDD. A model which reproduces the
dose-response relationships of experimental data and is
consistent with the biochemical and physiological processes
known to occur in rats exposed to TCDD might permit
extrapolation of responses beyond the range obtained from
experimental data and lead to scientifically sound ap-
proaches for estimating risks of adverse health effects of
exposure to TCDD.
METHODS
The present model, hereafter referred to as the NIEHS model (see Ap-
pendices 1 and 2 for a complete description), was adapted from the physio-
logically based pharmacokinetic model of Leung et al. (1990), hereafter
referred to as the LPMA model. To fit the data of Tritscher et al. (1992)
and Sewall et al (1993), the LPMA model has been modified as described
below. A flowchart of the resulting model is given in Fig. 1.
TCDD in the gut compartment is periodically increased by orally ad-
ministered TCDD according to the experimental schedule (Tritscher ft al.,
\ 992). TCDD is extracted from the gut into the blood compartment, where
some of it binds to unspecified serum protein. From the blood it is distrib-
uted to the tissue compartments.
TCDD in the liver is distributed between "metabolically available" and
protein-bound pools. The latter category includes binding to CYP1A2 and
to the Ah receptor, which is presumed to mediate all of the tissue's re-
sponses to TCDD. Unbound TCDD is converted to metabolites which are
either transferred to the gut via the bile or released into the blood. The
metabolites are treated as partitioning between the liver and blood in the
same manner as TCDD. Gut metabolites ultimately appear in the feces,
and blood metabolites appear in the urine. TCDD is also cleared from the
liver by lysis of dead cells consequent to prolonged exposure.
Equations for the pharmacodynamic effects of TCDD described above
were also included in the NIEHS model. Proteins whose synthesis is repre-
sented in the model are removed by proteolysis or, in the case of the EGF
receptor, by endocytosis.
In many cases quantities that were reported in a variety of units had to be
converted to those units (nanomoles of material, volume in liters, time in
days) used in the model. These calculations were performed using conver-
sion factors of 100 mg cytosolic protein/g liver (Poland and Knutson,
1982), 18 mg microsomal protein/g liver (based on average yield of mem-
brane protein; Tritscher et al., 1992), and 4.2 mg plasma membrane pro-
tein/g liver (based on average yield of membrane protein; Sewall et al.,
1993). Rates obtained from in vitro measurements conducted at tempera-
tures other than body temperature were adjusted to 37°C using a doubling
of the rate for every 10°C increase in temperature, a factor which is typical
of enzymatic reactions.
The animals used in these experiments (Tritscher et al., 1992) grew, on
average, from 220 to 390 g during the 31 weeks of the investigation. To
account for the dilution of material with growth the parameters in the
model were expressed per liter in the appropriate compartment. Because of
the increase in compartment volume, a steady-state concentration of a
protein whose synthesis is represented in the model actually corresponds to
an increase in the tissue content of the protein with time.
TCDD distribution and clearance. The LPMA model is said to be
flow-limited; the rate of uptake by tissues is limited by the rate of delivery
via the blood. In such a model, material is transferred from the arterial
blood to various tissues at a rate given by
Blood Content x Flow
Blood Volume '
where Flow represents the rate of blood flow through the particular tissue.
Transfer of material from the tissue to the venous blood is given by
Tissue Content x Flow
Tissue Volume X Partition CoefT"
where the partition coefficient is the ratio of tissue to blood concentrations
at equilibrium. The LPMA model includes compartments for gut. blood,
fat, muscle (plus skin), liver, and other viscera. The cell membranes of
these tissues constitute a barrier to transport of TCDD. In the NIEHS
model the rates of transport given by the above expressions were multi-
plied by a "diffusion coefficient" to account for the membrane permeabil-
ity. Different compartments were assigned different coefficient values as
required to match the time course data of Abraham et al. (1988). For these
calculations the dose was absorbed directly into the blood from the site of
injection with a scaled rate constant of 0.37 kg°'73/day, bypassing the gut
compartment.
The observed blood levels of TCDD (A. M. Tritscher, unpublished re-
sults) were an order of magnitude smaller than those predicted by the
LPMA model. The LPMA model includes binding of TCDD to unspeci-
fied blood lipoproteins and uses a factor of 2.5 for the ratio of protein-
bound to free TCDD in the blood. This constant was replaced in the
A-2
-------
KOHN ET AL.
Fat
TCDD
Blood
Bound Metabolite
TCDD
UrineJ
Metabolite
FIG. 1. Flowchart of the NIEHS model of TCDD distribution and consequent effects on gene expression in the rat liver.
NIEHS model by a reverse hyperbolic function (Appendix 1) for the maxi-
mal amount of TCDD-binding blood protein. Although the form of this
equation suggests that TCDD decreases the hepatic production of blood
lipoprotein, the equation is intended as just an empirical relationship.
TCDD has been found to bind to CYP1A2 in rat liver (Voorman and
Aust, 1989) and P450d in the mouse liver (Poland et ai. 1989) with an
apparent Kd of 30 nM. The estradiol 2-hydroxylase activity of this cy-
tochrome is inhibited as a consequence of this binding (Voorman and
Aust, 1987). As Poland el al. (1989) concluded that TCDD congeners bind
at or near the active site on this enzyme, such binding is treated as competi-
tive against estradiol in the NIEHS model. It was further assumed that
TCDD binds only to the CYP1A2 that is not complexed with NADPH:cy-
tochrome P450 reductase.
The appearance of unabsorbed TCDD in feces and TCDD metabolite(s)
in both feces and urine is included in the NIEHS model. Birnbaum el al.
(1980) administered a single oral dose of HC-labeled 2,3,7,8-tetrachlorodi-
benzofuran (TCDF) to rats. They calculated rate constants of 0.381 day"1
for the appearance of metabolites in the feces and 0.536 day"1 for the
appearance of metabolites in the urine. However, the calculated rate con-
stant for clearance of radiolabel from the blood during the first 3 hr in their
experiments (32 day"1) was an order of magnitude higher than after 3 hr»
(0.6 day"1). Because excretion of metabolites of TCDF is limited by their
rates of formation, the rate constants for clearance at late time most likely
reflect this limitation. This notion is supported by nearly identical calcu-
lated rate constants for clearance of radiolabel from liver and from blood
after 3 hr (Birnbaum et al, 1980). The higher rate constant for clearance at
early time more likely reflects the rate of uptake of labeled material from
the blood by the various tissues, which therefore cannot be rate limiting for
clearance. The rates of transport of TCDF's and TCDD's metabolites
across cell membranes should be similar. This leads to the conclusion that
the specific rates of transport of TCDD's metabolites from liver to bile and
from blood to urine must be larger than the rate constants for net clearance
for times greater than 3 hr. Therefore, the above rate constants were in-
creased by a factor of 10 for this model to ensure that clearance of metabo-
lites is not limited by transport.
Rose et al. (1976) found as much TCDD in the livers of rats after 7 weeks
of chronic dosing as Tritscher el al. (1992) found after 31 weeks at compara-
ble doses. A preliminary version of the NIEHS model indicated that param-
eter values which enabled the model to reproduce the results of Tritscher et
al (1992) would uniformly underestimate the results of Rose et al. (1976)
by about 50%. The preliminary model included a first-order rate constant
for metabolism of TCDD whose value could not be altered to fit the data of
Rose et al (1976) without destroying the fit to the data of Tritscher et al
(1992). Therefore, an additional mechanism for clearance from the liver
was necessary to permit fitting both data sets simultaneously.
The cellular proliferation rate in the livers of rats exposed to biweekly
doses of 1.4 Mg/kg of TCDD for 30 weeks is 7.3% of the liver cells/week
(Luciere/ ai, 1991). At this rate, one would expect a 226% increase in the
A-3
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MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
weight of the liver over the course of the experiment. Only a 90% increase
was observed (Lucier el al., 1991; A. M. Tritscher, unpublished results),
suggesting that toxic effects of cumulative exposure to TCDD result in cell
death. The NIEHS model includes loss of TCDD from the liver by lysis of
dead cells. The specific rate of clearance by cell lysis was assumed to in-
crease as a hyperbolic function of the cumulative exposure to unbound
liver TCDD (Appendix 1). No information regarding the fate of TCDD
from lysed cells is available. Therefore, this feature of the model merely
represents the net contribution to clearance by cell turnover.
TCDD-induced changes in gene expression. The complex formed by
binding of TCDD to the Ah receptor has been clearly implicated in induc-
tion of CYP1A1 (Fisher et al., 1990). The liganded Ah receptor modulates
expression of CYP1A1 by forming a ternary complex with at least one
other transcription factor (Hoffman et al., 1991), and this complex then
binds to enhancer sequences for the CYP1A1 gene. This mechanism was
used as a prototype for all Ah receptor-dependent protein synthesis in the
model. There are no data for the concentrations of the additional tran-
scription factors required for expression nor for the competition among
various receptors for binding to them. Therefore, the NIEHS model com-
bines binding of additional transcription factors to the Ah-TCDD com-
plex into the equations for the rate of Ah-dependent induction by treating
induction of this protein as proceeding with saturation kinetics subsequent
to binding of the Ah-TCDD complex itself to the appropriate enhancer
sequences. That is, the Ah-TCDD complex was treated as though it were
the "substrate" of the rate-limiting reaction in expression.
TCDD increases the liver concentration of the Ah receptor itself (Poland
and Knutson, 1982; Sloop and Lucier, 1987). In the NIEHS model, in-
creased synthesis of the Ah receptor is postulated to occur by an Ah-depen-
dent mechanism similar to that outlined above. Bradfield et al. (1988)
observed the Ah receptor's Kd for 2-iodo-7,8-dibromodibenzo-/>-dioxin to
decline from 0.16 to 0.012 nM on 32-fold dilution of the Ah receptor
protein. A dissociation equilibrium constant depends on ligand solubility,
pH, ionic strength, etc. The hydrophobicity of TCDD apparently makes
the apparent Kd very sensitive to the above conditions. Therefore, the
model uses the value of 0.27 nM reported by Poland and Knutson (1982)
for the A^ of TCDD under conditions more representative of the intracel-
lular milieu.
A rate law as described above treats the synthesis of gene product as
occurring by a single event. The appearance of gene product is actually the
result of many steps and requires some time before measurable amounts of
protein are synthesized. The data of Sloop and Lucier (1987) show that 6 hr
elapses between administration of TCDD and increased production of Ah
receptor protein. To account for such a delay, the concentration of the
Ah-TCDD complex at time I - 6 hr is substituted into the rate law to
compute the rate of synthesis of protein at time I. The same approach was
used for binding of all receptor complexes in all gene expression processes
in the model. The model's predictions were observed to be fairly insensi-
tive to reasonable numerical values of this delay.
Fisher el al (1990) found four Ah-responsive enhancer sites associated
with the CYP1A1 structural gene. This finding raises the possibility of
cooperative binding of Ah-TCDD complexes to these binding sites and, by
analogy, to the other Ah-TCDD binding sites postulated in the model. To
test for such cooperativity, Hill exponents were included in the rate equa-
tions for protein synthesis and those values which would enable the model
to fit experimental observations were identified.
Proteins in the model were treated as being removed from the system
with first-order kinetics. The rate constant for the proteolytic degradation
of all proteins whose synthesis is represented in this model was determined
from the data of Lucier et al. (1972).
Effects of TCDD on epidermal growth factor receptor. Choi ct at.
(1991) found TCDD to induce expression ofTGF-« by an Ah-dependent
mechanism in keratmocyte cultures. The TGF-« produced was exported
to the extracellular fluid. A similar mechanism has been included in the
NIEHS model, representing the product peptide as appearing in the inter-
stitial space of the liver.
Intact female rats subjected to 30 weeks of oral dosing with 125 ng
TCDD/kg body weight/day show a 65% reduction in the maximal binding
capacity (Smax) of the EOF receptor in liver plasma membranes (Sewall et
al, 1993) at a liver TCDD content of 0.063 nmol/g (Tritscher et al., 1992).
Ovariectomized rats exhibit only a 19% reduction of EOF Bmax at a liver
TCDD content of 0.106 nmol/g after 30 weeks of oral dosing at 100 ng
TCDD/kg/day (Clark et al., 1991). Sunahara et al. (1989) observed a 56Tt
reduction in the Bma>l in female rats 10 days following gavage with 10 ^g
TCDD/kg. Madhukar et al. (1984) observed a 45% decline in the Bmax of
the EOF receptor of male rats 10 days after a single intraperitoneal injec-
tion of 10 Mg TCDD/kg. The NIEHS model predicts a liver TCDD content
of 0.217 nmol/g for females and 0.223 nmol/g for males at the end of the
10-day period. The responses of males, Ovariectomized females, and intact
females may indicate different sensitivities for transcriptional activation of
the TGF-a gene by the Ah-TCDD complex as a consequence of differ-
ences in the amount of hepatic estrogen (or some other ovarian hormone).
Intact females, having much more circulating estrogen than either males or
Ovariectomized females, should be the most sensitive.
Both the Ah-TCDD and estrogen receptor-estrogen (ER-E) complexes
may be acting as transcriptional activators of the TGF-a gene, but a syner-
gistic interaction between them appears to be necessary to fully activate
transcription. In the NIEHS model, binding of the ER-E complex is
treated as activating the TGF-a gene for subsequent binding of the Ah-
TCDD complex. However, binding of this complex to the TGF-a gene has
not been proven, and other mechanisms are possible.
Estrogen metabolism and the production of the estrogen receptor are
included in the NIEHS model to account for the effects of estrogen on
alterations in the EOF receptor. The kinetic constants for estradiol 2-hy-
droxylase activity (Appendix 2) would predict an increase in enzymatic
activity comparable to the ninefold rise in CYP1A2 resulting from expo-
sure to TCDD. Only a threefold increase in activity was actually observed
(Graham et al., 1988) in female rats. The smaller increase in activity was
attributed to rate limitation by the availability of NADPHicytochrome
P450 reductase.
This reductase forms a complex with cytochrome P450 (Coon et al,
1977). The complex binds molecular oxygen and is reduced b\ NADPH to
eliminate water, leaving a single oxygen atom to combine with substrate
(Estabrook and Werringloer, 1977). As all P450 isozymes compete for
binding to reductase. there might not be sufficient reductase to bind to all
the increased CYP1A2. Indeed, there is evidence (Ullrich and Duppel,
1975) that P450 reductase is limiting for other mono-oxygenase reactions
that are not induced. Therefore, binding of the reductase to both CYP1 Al
and CYP1A2 was included in the NIEHS model. Estradiol and reductase
were allowed to bind to CYP1A2 in random order, and oxygen and
NADPH were treated as saturating.
The catechol estrogens produced by the activity of CYP1A2 bind to the
estrogen receptor an order of magnitude more weakly than does estradiol
(Li et al, 1985). Such binding is included in the NIEHS model, and the
resulting complex was allowed to function identically to the ER-estradiol
complex. This means that the ER-estradiol and ER-catechol estrogen
complexes produce the same effect on gene expression, but that 10 times as
much catechol estrogen as estradiol is required to achieve the same quanti-
tative response. Therefore, the symbol ER-E is used in this report to repre-
sent complexes between the estrogen receptor and either estradiol or cate-
chol estrogen.
The estrogen-binding capacity of the estrogen receptor in the livers of
Ovariectomized or immature female rats is increased by administration of
estradiol (Romkes ct al., 1987). One interpretation of this observation is
that there is synthesis of additional receptor protein consequent to binding
of the ER-E complex to the estrogen receptor gene. Female rats treated
with 100 ng TCDD/kg/day exhibit a 55% reduction in the estrogen recep-
tor (Clark ct al. 1991). A 40% reduction in estrogen receptor levels was
A-4
-------
KOHN ET AL.
0.10-
O)
a>
o
c
Q
Q
O
0.05-
0.00-
i | i i
100
Time, days
200
FIG. 2. Computed time courses of the accumulation of TCDD in liver
and fat of rats treated with 125 ng TCDD/kg body weight/day.
observed in the livers of mice exposed to a single 100 ng/kg dose of TCDD
(Lin el al., 1991b). "Ah-unresponsive" mice having defective or insuffi-
cient Ah receptors require higher TCDD exposures than do normal mice
to achieve a given reduction in estradiol binding capacity (Lin el al,
1991b). These observations may be interpreted as indicating that induc-
tion of estrogen receptor is inhibited by binding of the Ah-TCDD complex
at another DNA binding site. This mechanism is represented in the NIEHS
model as noncompetitive inhibition of receptor synthesis by the Ah-
TCDD complex (Appendix 1).
The NIEHS model was implemented in the SCoP simulation program
(Kootsey el al, \ 986; Kohn el al, 1993). Values for parameters that could
not be obtained from the literature were adjusted to make the model repro-
duce the observations of Tritscher el al. (1992) and Sewall el al. (1993).
Where experimental data were available, parameters were estimated using
the "praxis" algorithm (Brent, 1973) in the SCoPfit program (part of the
SCoP package).
The fully assembled model was composed of 33 first-order ordinary
differential equations. There were 77 constants in these equations, of
which 15 were freely adjustable parameters. The remaining 62 constants
were obtained from the literature, constrained by experimental data, or
mathematically determined by the law of conservation of mass. The
model's equations are listed in Appendix 1, and all parameter values (esti-
mated or from the literature) are given in Appendix 2. The experiments of
Tritscher el al. (1992) and Sewall el al. (1993) were simulated by integra-
tion of the differential equations to 217 days of biweekly exposure to
TCDD using a C language translation of the LSODA Gear integration
package from Lawrence Livermore Laboratory.
RESULTS
The NIEHS model predicts that 92.5% of the ingested
TCDD is absorbed into the bloodstream and 7.5% appears
unchanged in the feces. As some of the TCDD cleared from
the liver by cell lysis may be returned to the gut (the model
does not treat the fate of this material), this result may be an
underestimation of the amount of TCDD in the feces. Rose
el al. (1976) report an average value of 90.3% of 14C-TCDD
extracted from the guts of female rats. The model predicts
that 92.2% of the metabolite appears in the feces and 7.8%
appears in the urine at a dose of 125 ng/kg/day. The dose of
TCDD did not have a significant effect on these predic-
tions. Urinary radioactivity at doses comparable to those of
Tritscher et al. (1992) was often undetectable in the chronic
dosing experiments of Rose et al. (1976).
The computed time courses for TCDD in the liver and
fat for biweekly oral doses of 125 ng/kg body weight/day
(Tritscher et al., 1992) and for a single subcutaneous injec-
tion of 300 ng/kg (Abraham et al., 1988) are given in Figs. 2
and 3, respectively. From the fit to the data of Abraham et
al. (1988) the NIEHS model predicts an initial half time of
11.8 days for clearance of TCDD from the liver and an
overall liver half time of 13.5 days. Abraham et al. (1988)
report an initial half time of 11.5 days and an overall half
time of 13.6 days in the liver. The model predicts a half time
of 22.3 days in the fat; Abraham et al. (1988) report a half
time of 24.5 days in fat.
The predicted dose dependency of TCDD concentration
in the blood, liver, and fat after 217 days of exposure (bi-
weekly oral doses of TCDD for 31 weeks) is given in Table
1. The NIEHS model predicts a linear relationship between
administered dose and the concentration in the liver at
doses between 3.5 and 125 ng/kg/day (Fig. 4) in agreement
with the experimental observations (Tritscher et al., 1992).
Table 2 gives the model's predictions for repeated doses,
and Table 3 gives the model's predictions for single doses.
The Hill exponents in the equations for the rates of in-
duction of CYP1A1 and CYP1A2 were estimated from the
80
Time, days
FIG. 3. Fit to the time courses in rats given a subcutaneous injection
of 300 ng/kg body weight of TCDD (Abraham et al., 1988). Circles are for
observations in liver, and squares are for observations in fat.
A-5
-------
MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
TABLE 1
Computed TCDD Distribution
TABLE 2
Fit of Model to Observed TCDD Concentrations Following
Dose,
ng/kg/day
0.01
0]
.1
0.3
1.0
3.5
7.0
10.7
15.0
20.0
25.0
30.0
35.7
45.0
55.0
65.0
75.0
85.0
100.0
115.0
125.0
Ql/-krtJ-l
DlOOQ,
nMa
0.0000967
f\ AAAQTfi
U.WVy&j
0.00254
0.00672
0.0150
(0.0127-0.0174)
0.0218
0.0269
(0.0143-0.0251)
0.0315
0.0361
0.0400
0.0435
0.0471
(0.0323-0.103)
0.0523
0.0573
0.0618
0.0661
0.0700
0.0754
0.0806
0.0839
(0.0401-0.134)
Liver,
nmol/g*
0.00000394
fi n/WI?Q<
u.uuuujyj
0.000118
0.000389
0.00142
(0.000652-0.00267)
0.00313
0.00515
(0.00435-0.00683)
0.00754
0.0105
0.0134
0.0165
0.0199
(0.0155-0.0311)
0.0252
0.0307
0.0360
0.0417
0.0468
0.0543
0.0618
0.0673
(0.0385-0.134)
CVj*
rat,
nmol/g
0.00000432
f\ f\f\ClC\A ") 1
U.Ui/UUHZ I
0.000120
0.000350
0.000954
0.00162
0.00223
0.00289
0.00361
0.00431
0.00496
0.00568
0.00683
0.00804
0.00925
0.0104
0.0115
0.0133
0.0149
0.0160
Repeated Doses
Dose, ng/kg
10 100
Dosing 5 days/week for 1 week
Liver 0.000898 0.0141
(0.0) (0.00870-0.0155)
Fat 0.0004063 0.00302
(0.0) (0.000932-0.00466)
Dosing 5 days/week for 3 weeks
Liver 0.00235 0.0373
(0.002 1 7-0.00280) (0.0298-0.0435)
Fat 0.00108 0.00726
(0.000932) (0.0059-0.0109)
Dosing 5 days/week for 7 weeks
Liver 0.00408 0.0505
(0.00342-0.00652) (0.05 1 9-0.07 1 1 )
Fat 0.00170 0.0107
(0.000620-0.00 1 24) (0.0 118-0.0161)
" Concentrations in nmol/g tissue; observed values
are in parentheses.
1000
0.147
(0.143-0.165)
0.0249
(0.0217-0.0404)
0.302
(0.227-0.457)
0.0586
(0.0497-0.0963)
0.329
(0.471-0.796)
0.0766
(0.0767-0.305)
(Rose el al., 1976)
" Range of observed values (A. M. Tritscher, unpublished results) is
given in parentheses.
b Range of observed values (Tritscher et al, 1992) is given in parenthe-
ses.
data of Tritscher et al. (1992). Hill exponents of 1.0 for
induction of CYP1A1 (Fig. 5) and CYP1A2 (Fig. 6) repro-
duced the observed responses. Because binding of TCDD to
0.15-
D)
JD
0.10-
a"
a
o
0.05-
0.00-
0 50 100
Dose TCDD, ng/kg/day
FIG. 4. Computed relationship between liver TCDD concentration
and administered dose. Experimental data points (filled circles) are the
average values of Tritscher et al. (1992). Error bars denote the range of
observed values.
the Ah receptor is the initial step in the induction of cy-
tochrome P450 isozymes, the fractional occupancy of the
receptor will have a strong effect on the predictions of the
model. As fractional occupancy is dramatically affected by
cooperativity in ligand binding, estimation of the Hill expo-
nent for this binding was included in the course of estimat-
ing the gene induction parameters for CYP1A1 and
CYP1A2. The optimal value of the exponent was 0.96, in-
dicating a lack of cooperativity (i.e., hyperbolic binding).
This value agrees with the estimate of 1.0 obtained by Ga-
siewicz (1984) and by Bradfield et al. (1988).
Table 4 compares the NIEHS model's predictions to the
observed terminal concentrations of the Ah receptor and of
the induced P450 isozymes. The model predicts that the
fractional occupancy of the Ah receptor by TCDD (not in-
cluded in Table 4) rises from 13.4% at a dose of 3.5 ng
TCDD/kg/day to 69.3% at 125 ng/kg/day. The liver con-
centration of CYP1A1 is 0.02 nmole/g in the absence of
administered TCDD (Tritscher et al., 1992), consistent
with (but not proof of) the notion of an endogenous ligand
or a ligand of dietary origin for the Ah receptor. A liver
concentration of 1.5 pM (in TCDD equivalent units) of this
ligand reproduces the observed basal level of CYP1A1.
Needham et al. (1991) detected a background level of
TCDD of 2 pM in human livers.
Table 5 gives the predicted concentration of interstitial
TGF-a after 217 days of exposure to TCDD and the conse-
A-6
-------
KOHN ET AL.
Fit of
Dose,
ng/kg
1000
0.25
f\ C
0.5
1
10
100
1
3
10
30
100
300
1000
3000
"Ohse
TABLE 3
Model to Observed TCDD Concentrations Following
Single Doses'
Liver,
nmol/g
22 days following gavage*
0.0114
(0.00869-0.0145)
4 days following gavagec
0.00000366
(0.0000035)
f\ nn/vwn IA
U.UUUUU/J4
(0.0000035)
0.0000147
(0.0000078)
0.000155
(0.000066)
0.00198
(0.0011)
7 days following subcutaneous injection
0.0000119
(0.0000096)
0.0000365
(0.000032)
0.000129
(0.000126)
0.000456
(0.000503)
0.00218
(0.00217)
0.00928
(0.0105)
0.0377
(0.0332)
0.116
(0.087)
rved values are in narentheses.
Fat,
nmol/g
0.00389
(0.0010-0.0222)
d
0.0000101
(not detected)
0.0000302
(0.000043)
0.0000995
(0.000153)
0.000290
(0.000432)
0.000885
(0.00104)
0.00234
(0.0025)
0.00699
(0.0063)
0.0201
(0.0114)
, , , , !,,,,!,,
6- -r -
-
_O> - ji -
1 4- L-^~~"^ 1 -
C " s*
T-" " T s
< T/
£ 2~ / L
° '. / '.
' T
0 7 1 1 1 1 | I I 1 1 | 1 1
0 50 100
Dose TCDD, ng/kg/day
FIG. 5. Induced CYP1A1 concentration vs administered dose of
TCDD. Experimental data points (filled circles) are the average values of
Tritscher et al. (1992). Error bars denote the range of observed values.
obtained with a model which required prior binding of the
ER-E complex to activate binding of the Ah-TCDD com-
plex to the TGF-a gene. An alternative model which postu-
lates random-order binding of the ER-E and Ah-TCDD
complexes produced quantitatively similar results.
The NIEHS model predicts that 10 days after administra-
tion of a single dose of 1 jug TCDD/kg there should be a
20.1% decrease in the fimax of the plasma membrane EGF
receptor in female rat livers (14% observed; Sunahara et al.,
1989), but only a 9.6% decrease in male rat livers (10%
observed on average; Madhukar et al., 1984). This com-
puted twofold greater responsiveness of the female over the
male persisted over all doses in the range of 0. 1 to 10 Mg/kg.
As estrogen plays a major role in regulating TCDD's ef-
fects on the EGF receptor, estrogen action was examined in
* Rose et al. (1976).
c J. Vanden Heuvel, NIEHS, personal communication.
d Abraham et al. (1988).
quent distribution of the EGF receptor between the plasma
membrane and cytosol. Figure 7 gives the decline in EGF
binding capacity of the plasma membranes with the dose of
TCDD and corresponds to that fraction of the receptor
which had not been internalized consequent to binding of
TGF-a. Figure 8 relates the predicted degree of internaliza-
tion of the EGF receptor to the TGF-a concentration. The
points on this curve corresponding to the predicted re-
sponses for the four dose groups of Sewall et al. (1993) are
also shown. Note that the TGF-a concentration is given in
pmol/g liver; multiplying by 10 gives the concentration in
the interstitial fluid in nanomolar. These interstitial con-
centrations are comparable to the concentrations produced
by TCDD in the extracellular fluid of keratinocyte cultures
exposed to TCDD (Choi et al., 1991). These results were
10-
o
c
sf
5-
50 100
Dose TCDD, ng/kg/day
FIG. 6. Induced CYP1A2 concentration vs administered dose of
TCDD. Experimental data points (filled circles) are the average values of
Tritscher et al. (1992). Error bars denote the range of observed values
A-7
-------
MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
TABLE 4
Ah Receptor and Induced Cytochrome P450 Isozymes
TABLE 5
Effects of TCDD on the EGF Receptor
Dose,
ng/kg/day
0.0
0.01
0.1
0.3
1.0
3.5
7.0
10.7
15.0
20.0
25.0
30.0
35.7
45.0
55.0
65.0
75.0
85.0
100.0
1 1 ? f\
1 15.U
125.0
Ah receptor,
pmol/g
2.09
(2.1)*
(2.1-3.2F
2.10
2.14
2.23
2.45
3.21
4.01
4.68
(3.0-5.7)'
5.25
5.77
6.16
6.50
(5.5-7.8)c
6.80
7.16
7.43
7.63
7.82
7.95
8.11
(4.9-10.5)c
8-) A
.24
8.32
(8.5)'
CYP1A1,
nmol/g"
0.0234
(0.0081-0.0351)
0.0271
0.0591
0.126
0.328
0.881
(0.269-0.953)
1.501
2.011
(1.89-3.13)
2.443
2.841
3.135
3.377
3.589
(2.91-4.47)
3.845
4.038
4.183
4.314
4.409
4.519
A £. 1 ")
4.612
4.672
(3.72-5.99)
CYP1A2,
nmol/g°
0.557
(0.352-0.714)
0.561
0.583
0.628
0.771
1.190
(0.840-2.31)
1.699
2.145
(1.39-3.56)
2.544
2.927
3.224
3.476
3.704
(2.39-4.21)
3.990
4.211
4.380
4.537
4.652
4.785
4.899
4.975
(2.86-10.3)
Dose, EGF receptor,
ng/kg/day pmol/g°
0.0 2.553
(2.06-2.61)
0.01 2.553
0.1 2.545
0.3 2.53
1.0 2.48
3.5 2.34
(1.45-2.95)
7.0 2.16
10.7 2.01
(1.18-2.45)
15.0
20.0
25.0
30.0
35.7
(0.8S
45.0
55.0
65.0
75.0
85.0
100.0
115.0
.88
.75
.65
.55
.46
9-1.64)
.36
.28
.22
.15
.11
.07
.02
125.0 0.990
(0.702-1.13)
" Experimental data from Sewall
TGF-a, Internalized EGF
pmol/g receptor, pmol/g
0.0
0.00000995
0.0000999
0.000291
0.000901
0.00280
0.00540
0.00806
0.0107
0.0138
0.0165
0.0194
0.0222
0.0263
0.0299
0.0326
0.0361
0.0386
0.0429
0.0446
0.0470
etal. (1993) are given
0.0
0.000849
0.00854
0.0246
0.0747
0.219
0.391
0.541
0.674
0.807
1.00
1.09
.20
.28
.34
.40
.44
.49
.53
.56
in parentheses.
1 Experimental data from Tritscher el al (1992) are given in parenthe-
ses.
* Poland and Knutson (1982).
c Sloop and Lucier (1987).
d Maximal induction from Poland and Knutson (1982).
is induced; at a dose of 125 ng/kg/day, 39.8% of CYP1A2 is
calculated to be complexed with reductase. The lack of stoi-
chiometric amounts of bound reductase is consistent with
the model. Chronic exposure to TCDD was found to reduce
the concentration of the hepatic estrogen receptor in DEN-
initiated female rats from 5.1 to 2.3 pmol/g (Clark et al.,
1991). The model reproduces this decrease in the estrogen
receptor level (Table 6). It also predicts that the reduction is
dependent on the dose of TCDD with a curve shape similar
to that obtained for mouse liver by Lin et al. (1991b).
The calculated liver concentrations of estradiol and cate-
chol estrogen and the calculated rate of the estradiol 2-hy-
droxylase activity of CYP1A2 at various doses are given in
Table 6. At a dose of 125 ng TCDD/kg/day the calculated
rate of the estradiol 2-hydroxylase is 3.1-fold greater than in
the absence of administered TCDD whereas CYP1A2 is
computed to increase 8.9-fold over its basal level. The
NIEHS model predicts that only 44.9% of the CYP1A2 is
complexed with reductase in the absence of administered
TCDD. This percentage decreases slightly as more enzyme
,o> 3-
"o
CL
O
"5.
o
O
LU
CO
2-
1-
I
0
T I | I I I I [
50 100
Dose TCDD, ng/kg/day
FIG. 7. Reduction of Smax of EGF receptor vs administered dose of
TCDD. Experimental data points (filled circles) are the average values of
Sewall el al (1993). Error bars denote the range of observed values.
A-8
-------
KOHN ET AL.
0.6-
TJ
0>
=
"co
I
0.4-
CL
8
8>
LL
a
LU
g
"o
ro
0.2-
0.0-
0.00
0.02
TGF-a, pmole/g
0.04
FIG. 8. Calculated fraction of EOF receptor redistributed from the
plasma membrane into the cytosol as a function of the concentration of
TGF-a. Vertical bars indicate model's predictions for the four dose groups
ofSewallera/. (1993).
cies (Cerutti, 1978). The proliferative effect of EGF recep-
tor-mediated events could also provide genetically altered
cells with a selective growth advantage. This mechanism is
consistent with the general agreement that TCDD is a tu-
mor promoter, but not an initiator, in two-stage models for
hepatocarcinogenesis (Kociba et ai, 1978; Pitot and Sirica,
1980; Lucier et al, 1991) and a complete carcinogen in the
2-year chronic animal bioassay (Huff el al., 1991).
EGF Receptor and Cellular Proliferation
Figure 7 shows the predicted and observed responses of
the maximal binding of EGF to liver plasma membranes vs
dose of TCDD. Although the liver TCDD concentration is
linear in the administered dose over the range 3.5-125 ng/
kg/day, the response of the EGF receptor resembles a hy-
perbolic curve. As this response may be involved in the
mechanism of tumorigenesis in TCDD-treated rats, it
would be expected that it would correlate with tumor inci-
dence better than does tissue dose. If this is true, linear
extrapolation of effects at high doses to low doses by a line
that passes through the background incidence would un-
derestimate low-dose effects. However, such hyperbolic
curves are approximately linear in the low-dose range, indi-
the conclusion of Graham et al. (1988) that P450 reductase
is limiting for the hydroxylation of estradiol. However, at a
dose of 125 ng/kg/day, the estradiol concentration in the
liver is computed to fall to 39% of its level in the absence of
administered TCDD. As the computed estradiol concentra-
tion is below its Km for this enzyme, the enzymatic rate is
probably limited more by depletion of substrate than by
availability of reductase.
DISCUSSION
The purpose of this modeling effort was to suggest plausi-
ble biochemical mechanisms which may explain the ob-
served tissue effects of TCDD in female rats. Internaliza-
tion of the EGF receptor in response to induction of TGF-a
is a possible origin for the mitogenic signal important to
carcinogenesis. The NIEHS model's prediction of TGF-a
concentrations that are comparable to those observed in
cell cultures is consistent with this mechanism, and the
model's success in reproducing observed responses to
TCDD supports the proposed mechanism. However, other
EGF-like peptides may be involved although we have
treated TGF-a as the nominal ligand for our modeling. In
addition, the production of a considerable quantity of cate-
chol estrogens may play a role in carcinogenesis (Li et al.,
1985) either by co valent binding to protein or DNA (Tsibris
and McGuire, 1977) or by generation of active oxygen spe-
TABLE 6
Effects of TCDD on Estrogen Metabolism in Liver
Dose,
ng/kg/day
0.0
0.01
0.1
0.3
1.0
3.5
7.0
10.7
15.0
20.0
25.0
30.0
35.7
45.0
55.0
65.0
75.0
85.0
100.0
115.0
125.0
Estrogen
receptor,
pmol/g"
5.19
(5.1)
5.19
5.17
5.12
4.98
4.59
4.18
3.85
3.58
3.33
3.16
3.01
2.88
2.73
2.62
2.54
2.46
2.41
2.35
2.30
2.26
(2.3)
Catecho!
Estradiol, estrogen,
pmol/g pmol/g
0.154 0.610
0.154 0.613
0.153 0.633
0.151 0.672
0.145 0.785
0.130
0.114
0.103
0.0962
0.0878
0.0828
0.0792
0.0760
0.0722
0.0692
0.0670
0.0653
0.0638
0.0642
0.0610
0.0601
.067
.336
.517
.641
.734
.792
.828
.854
.877
.888
.893
.896
.895
.891
.889
.888
Estradiol 2-
hydroxylase.
nmol/g/day
0.0347
0.0348
0.0359
0.0381
0.0446
0.0605
0.0757
0.0860
0.0931
0.0984
0.102
0.104
0.105
0.106
0.107
0.107
0.107
0.107
0.107
0.107
0.107
' Experimental data of Clark el al (1991) are given in parentheses.
A-9
-------
MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
eating that linear extrapolation from low doses to extremely
low doses should still be valid.
Because the interstitial space occupies about 10% of the
liver volume, the accumulated TGF-a is 10 times more
concentrated in the interstitium than it would be in the
cytosol. This concentration effect has significant conse-
quences for the predicted reduction in the 5max of the EGF
receptor. If the TGF-a were actually produced in some
other tissue, it would have to be transported to the liver
interstitium via the blood. This would greatly dilute the
concentration of the peptide and require synthesis of 15
times as much material to result in the observed decrease in
EGF-binding capacity of the plasma membrane.
Although production of TGF-a by the liver was modeled
as being induced by the Ah-TCDD complex, there is evi-
dence (Gaido el al, 1992) for post-transcriptional regula-
tion as well. TCDD may act to "stabilize" the mRNA for
that peptide. If information regarding the mechanism of
such an effect were available, it could readily be incorpo-
rated into this model. As it is, the model merely states that
TCDD stimulates the rate-limiting step in the net synthesis
of hepatic TGF-a.
Under normal conditions, one of the functions of TGF-a
is as a signal for regenerative hyperplasia of injured tissue
(Burgess, 1989; Mead and Fausto, 1989). The substrates for
the internalized EGF receptor's tyrosine kinase may be in-
volved in modulating this process, and their enzymatic ac-
tivities may be regulated by phosphorylation. According to
the NIEHS model, TCDD causes inappropriate stimula-
tion of this process. Cells which had previously sustained
nonlethal DNA damage could be stimulated to reproduce
before the damage could be repaired. With the damage
fixed in the genome, clonal expansion of the transformed
cells could provide a necessary step in tumor development.
Different populations of cells may differ in the sensitivity
of their proliferative responses to TCDD. The small popula-
tion of polyploid cells in the livers of rats treated with DEN
and ethinylestradiol are especially rich in EGF receptors
(Vickers and Lucier, 1991). It is possible that these cells
would show a correspondingly larger concentration of inter-
nalized EGF receptors in response to induction of TGF-a
and, hence, a stronger mitogenic signal. If the population of
polyploid cells includes preneoplastic cells, proliferation of
these previously transformed cells may represent the most
important effect of TCDD in hepatocarcinogenesis.
The NIEHS model predicts that after 31 weeks of treat-
ment, as much EGF receptor is internalized in the male at a
dose of 125 ng TCDD/kg/day as is internalized in the fe-
male at a dose of 15 ng/kg/day. This behavior is consistent
with the observed lower sensitivity of the male (Madhukar
et al., 1984). It arises because serum estradiol in the male is
about 10% of that in the female (D. Schomberg, Duke Uni-
versity, personal communication), resulting in concentra-
tions of the ER-E complex which only slightly stimulates
production of TGF-a. Because the male rat is predicted tc
produce far less TGF-a than the female, the rate of cellula
proliferation in the male at low doses of TCDD should b(
indistinguishable from that at zero dose. This prediction o
the model could explain the absence of excess liver tumor:
in male rats exposed to TCDD (Kociba et al., 1978; Na-
tional Toxicology Program, 1982).
Cytochrotne P450 and Estrogen Metabolism
As estrogen is necessary for TCDD-mediated internaliza-
tion of the EGF receptor, induction of CYP1A2 may also
be an important index of cancer risk because this cy-
tochrome catalyzes oxidation of estradiol to catechol estro-
gen. The predicted response curve for this indicator (Fig. 6)
is hyperbolic. This behavior is attributable to the Hill coeffi-
cient of 1 determined for induction of this protein.
The finding of at least four enhancer sequences for induc-
tion of CYP1A1 (Fisher etal., 1990) suggests the possibility
that several enhancer sites must be occupied to enable tran-
scriptional activation of the CYP1A1 gene. Such a mecha-
nism could lead to sigmoidal response curves for the induc-
tion of this cytochrome. The Hill exponent of 1, which gave
the best fit to the data of Tritscher el al. (1992), yields a
hyperbolic response curve (Fig. 5). Because many factors
other than cooperative interactions among binding sites
(e.g., heterogeneous uptake of TCDD in the liver and recy-
cling of TCDD among Ah receptor molecules) could influ-
ence the shape of the response curve, the value of the Hill
exponent is not necessarily related to the number of bind-
ing sites for the Ah-TCDD complex.
Using steady-state kinetics and a simpler model, Portier
el al. (1993) estimated Hill exponents of 1.86 for CYP1A1
induction and 0.534 for CYP1A2 induction. A possible ori-
gin of the difference in Hill exponents is the steady state
model's neglect of the increase in the concentration of the
Ah receptor with increasing dose of TCDD. The maximal
slope of the curve for induction rate vs inducer concentra-
tion is steeper for cooperative kinetics (Hill exponents not
equal to 1) than for hyperbolic kinetics. The increase in the
concentration of the Ah receptor with dose can mimic the
steeper rise of rate of synthesis with dose exhibited by coop-
erative kinetics.
Portier el al. (1993) showed the importance of the mecha-
nism by which putative alternative ligands bind to the Ah
receptor and/or modulate protein synthesis. In the NIEHS
model, constitutive expression of all proteins except
CYP1A1 was assumed to be independent of the Ah recep-
tor. Background CYP1 Al levels were treated as being pro-
duced by an additional Ah receptor-dependent mechanism.
An "independent" mechanism accentuates any sigmoidic-
ity in the response curve. An "additive" mechanism mini-
mizes such sigmoidicity and leads to linearity in the low-
dose response. Such an effect may be involved in producing
A-10
-------
KOHN ET AL.
the hyperbolic response of CYP1 Al (Fig. 5). It is also possi-
ble that CYP1A1 is constitutively expressed at a low rate by
a mechanism that does not involve the Ah receptor. Addi-
tional experiments on the mechanism of expression of
CYP1A1 in the absence of administered TCDD could re-
solve this issue.
As catechol estrogens bind more weakly to the estrogen
receptor than does estradiol, it is conceivable that their pro-
duction from estradiol by the estradiol 2-hydroxylase activ-
ity of CYP1A2 could result in reduced ligation of the estro-
gen receptor. Because it was assumed that the ER-E com-
plex induces synthesis of the estrogen receptor protein, a
reduced calculated rate of synthesis of the receptor could
result from the enzymatic activity of CYP1A2. An alterna-
tive model, in which the ER-E complex was a transcrip-
tional activator of the estrogen receptor gene, but the Ah-
TCDD complex was not a transcriptional inhibitor of the
gene, predicted only an 18.9% decrease in the estrogen re-
ceptor concentration. As a 55% reduction in the estrogen
receptor concentration was observed (Clark et al., 1991), an
Ah receptor-mediated effect of TCDD on expression of the
estrogen receptor such as that incorporated into the NIEHS
model is likely. This result is consistent with the observa-
tion (Lin et al., 1991b) of only a 10% decrease in receptor
level in Ah-deficient mice.
Serum estradiol in the female rat oscillates from 15 to
120 pg/ml over a 5-day estrus cycle. Including such oscilla-
tions in the NIEHS model could make as much as a 12%
difference in the model's predictions, depending on where
in the cycle the simulation was begun. The serum levels,
50-60 pg/ml, reported by Shiverick and Muther (1983) are
close to the average level over the entire cycle, and the exper-
imental data are the average results for nine individuals.
Therefore, use in the model of Shiverick and Muther's
value for the circulating estradiol concentration is justified.
TCDD Tissue Dose
The LPMA model predicts a liver TCDD content 2.4-
fold higher than observed at a dose of 125 ng/kg/day
(Tritscher et al., 1992). The LPMA model assumes that
TCDD binds to total CYP1A2 with a Kd of 7 nM, and it
predicts a liver CYP1A2 concentration 60% higher than
observed at a dose of 125 ng/kg/day (Tritscher et al., 1992).
The resulting large calculated amount of bound TCDD is
most likely responsible for the overprediction of liver
TCDD.
Although the NIEHS model accurately reproduces the
observed CYP1A2 concentration, much less binding had to
be assumed in order to reproduce the experimental data.
An alternative model in which TCDD is bound nonspecifi-
cally to an unspecified protein with Kd between 5 and 8 nM
required only about 0.2 nmol/g liver of that protein to re-
produce the observed liver TCDD content. Setting the bind-
ing protein concentration to values comparable to that of
CYP1A2 necessitated a higher rate constant for clearance
of TCDD from the liver in order to match the experimental
data and resulted in a half time in the liver of only 7 days,
half the experimentally observed value.
Livers of TCDD-treated rats often show an increase in
lipid content. From the data of Tritscher et al. (1992) the
relationship (fraction lipid) = 0.01865 X (nmol total liver
TCDD) + 0.1443 X (kg body weight) was obtained. Allow-
ing the TCDD to partition between the aqueous phase of
the liver cytosol and lipid droplets with partition coeffi-
cients between 140 and 350 did not sequester enough
TCDD to match the data without also requiring induction
of TCDD-binding protein.
The dose-response relationships illustrated by the
NIEHS model are important for understanding the risks of
adverse health effects following exposure to TCDD. Sig-
moidicity in the response requires a higher concentration to
produce a given response at low dose than does hyperbolic
response exhibiting the same concentration for half-maxi-
mal effect. Such behavior can have dramatic consequences
for the estimated minimum exposure to TCDD that should
produce an unacceptable risk of adverse health effects (Por-
tier et al., 1993). All of the amounts of protein calculated in
this model show hyperbolic dependence on dose. The fact
that the model presented here does not predict sigmoidicity
in the observed responses indicates that the response is ap-
proximately linear at very low doses. The model's success in
reproducing these responses in a biochemically realistic
way suggests that it should also be useful in the analysis of
more complex dose-response relationships, such as for cel-
lular proliferation rates and tumor incidence.
APPENDIX 1: MODEL EQUATIONS
Compartment and Flow Characteristics
CardiacOutput
StartWeight
BodyWeight
BloodVolume
= 404 X BodyWeight075 (Delp et al., 1991)
= 0.237 kg (Tritscher et al., 1992)
= StartWeight + 0.219674 X e'0-002859 x d°se x time/( 116.345 + time) (BodyWeight was estimated
from the data of Tritscher et al. (1992) as a function of dose and time by nonlinear regres-
sion)
= 0.054 X BodyWeight (Delp et al., 1991)
A-ll
-------
MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
APPENDIX 1: Continued
FatVolume
MuscleVolume
VisceraVolume
LiverVolume
LiverlnterstitialVolume
LiverFlow
FatFlow
MuscleFlow
VisceraFlow
Distribution ofTCDD
GutTCDD
GutTCDD'
FecesTCDD'
cum _ exposure'
BloodProtein
BloodTCDD'
BloodBoundTCDD'
Muscle TCDD'
FatTCDD'
VisceraTCDD'
LiverTCDD'
LiverCYPl A2 _ TCDD' =
: 0.08 X Body Weight (average of literature values: Leung et al., 1990; Delp et al., 1991)
: 0.59 X BodyWeight (Delp et al., 1991) (this compartment also includes skin and other slowly
perfused tissues)
: 0.083 X BodyWeight (Delp et al., 1991) (this compartment includes all richly perfused tissues)
•• BodyWeight X (0.03324+ 0.01927 Xdose/(88.91 + dose)) (LiverVolume was estimated from
the data of Tritscher et al. (1992) as a function of dose by nonlinear regression)
: frac _ interstitium X LiverVolume
•• 0.15 X CardiacOutput (Delp et al, 1991)
•• 0.07 X CardiacOutput (Delp et al., 1991)
= 0.36 X CardiacOutput (Delp et al., 1991)
= 0.40 X CardiacOutput (Delp et al., 1991)
GutTCDD + dose(0 X BodyWeight/322 (doses at time t—in ng/kg—were introduced into the
gut compartment according to the experimental schedule; dose(/) = 0 if no dose is scheduled
at time t)
-k _ absorption X GutTCDD/Body Weight0 7S - k _ feces X GutTCDD
A:_fecesX GutTCDD
LiverTCDD
MaxBloodProtein X LiverVolume X K, _ BloodProtein/(A;, _ BloodProtein + LiverTCDD/
LiverVolume)
k - absorption X GutTCCD/Body Weight0 75 - BloodTCDD X MuscleDiffusion X Muscle-
Flow/BloodVolume - BloodTCDD X FatDiffusion X FatFlow/BloodVolume -
BloodTCDD X VisceraDiffusion X VisceraFlow/BloodVolume - BloodTCDD X Liver-
Diffusion X LiverFlow/BloodVolume + Muscle TCDD X MuscleDiffusion X MuscleFlow/
(Muscle Volume X MusclePartition) + FatTCDD X FatDiffusion X FatFlow/(FatVolume
X FatPartition) + VisceraTCDD X VisceraDiffusion X VisceraFlow/(VisceraVolume x
VisceraPartition) + LiverTCDD X LiverDiffusion X LiverFlow/(LiverVolume X LiverPar-
tition) - k _ binding X BloodTCDD X BloodProtein + k _ binding X K _ BloodProtein X
BloodVolume X BloodBound TCDD
A:-binding X BloodTCDD X BloodProtein - k-binding X K_ BloodProtein X BloodVol-
ume X BloodBoundTCDD
BloodTCDD X MuscleDiffusion X MuscleFlow/BloodVolume - MuscleTCDD X Muscle-
Diffusion X MuscleFlow/(Muscle Volume X MusclePartition)
BloodTCDD X FatDiffusion X FatFlow/BloodVolume - FatTCDD X FatDiffusion X Fat-
Flow/(FatVolume X FatPartition)
BloodTCDD X VisceraDiffusion X VisceraFlow/BloodVolume - VisceraTCDD X Viscera-
Diffusion X VisceraFlow/(VisceraVolume X VisceraPartition)
BloodTCDD X LiverDiffusion X LiverFlow/BloodVolume - LiverTCDD X LiverDiffusion
X LiverFlow/(LiverVolume X LiverPartition) - fc_ metabolism X LiverTCDD/Body-
Weight03 - k- binding X LiverTCDD X LiverCYPl A2 + A:-binding X AT-TCDD X
LiverVolume X LiverCYPlA2 _ TCDD - k- binding X LiverTCDD X LiverAhReceptor
+ k - binding X K- AhR X LiverVolume X LiverAhR _ TCDD + k _ proteolysis X (Liver-
AhR _ TCDD + LiverCYPl A2 _ TCDD) - k _ lysis X cum _ exposure/(crit _ exposure +
cum _ exposure) X LiverTCDD
k- binding X LiverTCDD X LiverCYPlA2 - k- binding X K_ TCDD X LiverVolume X
LiverCYPlA2_TCDD
TCDD Metabolite Clearance
LiverMetabolite'
= k - metabolism X LiverTCDD/Body Weight0 3 - k - bile X LiverMetabolite + BloodMetabo-
lite X LiverDiffusion X LiverFlow/BloodVolume - LiverMetabolite X LiverDiffusion X
LiverFlow/(LiverVolume X LiverPartition)
A-12
-------
KOHN ET AL.
APPENDIX 1: Continued
BloodMetabolite'
UrineMetabolite'
GutMetabolite'
FecesMetabolite'
= —BloodMetabolite X LIverDiffusion X LiverFlow/BloodVolume + LiverMetabolite X Liver-
Diffusion X LiverFlow/(LiverVolume X LiverPartition) — k _ urine X BloodMetabolite
= k - urine X BloodMetabolite
= k - bile X LiverMetabolite - k _ feces X GutMetabolite
= k _ feces X GutMetabolite
Ah Receptor and P450 Induction
LiverAhR _ TCDD'
LiverAhR _ Induced
LiverAhReceptor'
LiverCYPlAl'
LiverP450Reductase'
LiverCYPlAl _R'
= k - binding X LiverTCDD X LiverAhReceptor - k _ binding X K- AhR X LiverVolume X
LiverAhR _ TCDD - k _ proteolysis X LiverAhR _ TCDD
= k-binding X LiverVolume X Natural_ Inducer X LiverAhReceptor - A:-binding X K-
AhR X LiverVolume X LiverAhR _ Inducer - k _ proteolysis X LiverAhR _ Inducer
= F_ AhRinduction X LiverVolume/( 1 + (K _ AhRinduction X LiverVolume)/LiverAhR _
TCDD(r - It)) + AhRexpression X LiverVolume - k _ binding X LiverTCDD X Liver-
AhReceptor + k - binding X K- AhR X LiverVolume X LiverAhR _ TCDD - k _ binding
X LiverVolume X Natural _ Inducer X LiverAhReceptor + A;-binding X K _ AhR X
LiverVolume X LiverAhR _ Inducer - k _ proteolysis X LiverAhReceptor
= V _ CYP1A1 induction X LiverVolume/( 1 + (K _ CYP1A1 induction X LiverVolume)/(Li ver-
AhR _ TCDD(/ - //) + LiverAhR _ Inducer(/ - /;))) - k _ proteolysis X LiverCYPlAl -
A:-binding X LiverCYPlAl X LiverP450Reductase + k_binding X AT-Reductase X
LiverVolume X LiverCYPlAl _ R
= P450Red _ expression X LiverVolume — k _ proteolysis X LiverP450Reductase - k _ binding
X LiverCYPlAl X LiverP450Reductase + k _ binding X K- Reductase X LiverVolume X
LiverCYP 1A1 _ R - k _ binding X LiverCYP 1A2 X LiverP450Reductase + k _ binding X
K- Reductase X LiverVolume X LiverCYPlAl _ R - k_ binding X LiverCYPlA2 _ E2 X
LiverP450Reductase + A:-binding X A:-Reductase X LiverVolume X LiverCYP 1A2 _
R_E2
= A:_ binding X LiverCYPlAl X LiverP450Reductase - k_ binding X AT_ Reductase X Liver-
Volume X LiverCYPlAl _ R
Estradiol Metabolism
LiverE2'
Li verCYP 1A2'
Li verCYP 1A2
LiverCYPl A2
Li verCYP 1A2_
LiverE2OH'
= ConcBloodE2 X LiverFlow - LiverE2 X LiverFlow/LiverVolume - k _ binding X Liver-
CYP1A2 X LiverE2 + A: -binding X AT_E2 X LiverVolume X LiverCYPl A2 _ E2 -
A; -binding X LiverCYP 1 A2 _ R X LiverE2 + A: -binding X A"_E2 X LiverVolume X
LiverCYPl A2 _ R _ E2 + k _ proteolysis X LiverER _ E2 - k _ binding X LiverEReceptor
X LiverE2 + k _ binding X K- ER _ E2 X Liver Volume X LiverER _ E2
= K-CYP1 A 1 induction X LiverVolume/(l + (K- CYPlA2induction X LiverVolume)/Liver-
AhR_TCDD(; - It)) + CYPlA2expression X LiverVolume - k _ proteolysis X (Liver-
CYP 1A2 + LiverCYP 1A2 _ E2) - k _ binding X LiverCYP 1A2 X LiverP450Reductase +
k - binding X K- Reductase X LiverVolume X LiverCYPl A2 _ R - k - binding X Liver-
CYP 1A2 X LiverE2 + k _ binding X A"_ E2 X LiverVolume X LiverCYP 1A2 _ E2
binding X LiverCYPl A2 X LiverP450Reductase - k _ binding X A"_ Reductase X Liver-
Volume X LiverCYPlA2 _ R + F_ E2H X LiverCYPl A2 _ R _ E2
= A: -binding X LiverCYPl A2 X LiverE2 - A; -binding X K-E2 X LiverVolume X Liver-
CYPl A2 _ E2 - k - binding X LiverCYPl A2 _ E2 X LiverP450Reductase + k _ binding X
A"_ Reductase X LiverVolume X LiverCYP 1A2 _ R _ E2
R _ E2' = k - binding X LiverCYPl A2 _ E2 X LiverP450Reductase - k _ binding X K- Reductase X
LiverVolume X LiverCYPl A2 _ R _ E2 + k _ binding X LiverCYPl A2 _ R X LiverE2 -
Ac -binding X K-E2 X LiverVolume X Li verCYP 1 A2 _ R _ E2 - F_E2H X Liver-
CYPl A2_R_E2
= V _ E2H X LiverCYPl A2 _ R _ E2 - k _ binding X LiverEReceptor X LiverE2OH + A- _ bind-
ing X K - ER _ E2OH X LiverVolume X LiverER _ E2OH - k _ conjugation X LiverE2OH
_E2'
= k
A-13
-------
MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
APPENDIX 1: Continued
LiverEReceptor'
LiverER _ E2'
LiverER _ E2OH'
V _ ERinduction X LiverVolume/(( 1 + K- ERinduction X LiverVolume/(LiverER _ E2(t -
It) + LiverER _ E2OH(f - It))) X (1 + LiverAhR _ TCDD(/ - lt)/(Kt•_ ERinduction X
LiverVolume))) + ERexpression X LiverVolume — k _ proteolysis X LiverEReceptor —
k - binding X LiverEReceptor X LiverE2 + k _ binding X K- ER _ E2 X LiverVolume X
LiverER_E2 - A:-binding X LiverEReceptor X LiverE2OH + A:-binding X AT_ER_
E2OH X LiverVolume X LiverER _ E2OH
—k _ binding X LiverEReceptor X LiverE2 + k _ binding X K _ ER _ E2 X LiverVolume X
LiverER _ E2 - k _ proteolysis X LiverER _ E2
-k - binding X LiverEReceptor X LiverE2OH + k _ binding X K- ER _ E2OH X LiverVol-
ume X LiverER _ E2OH - k _ proteolysis X LiverER _ E2OH
EGF Receptor and TGF-a
LiverEGFReceptor1
LiverTGF
LiverlnternalEGFR'
= EGFRexpression x LiverVolume - k _ endocytosis X LiverEGFReceptor
= F_ TGFinduction X LiverVolume/( 1 + K _ TGFinduction X LiverVolume/(LiverAhR _
TCDD(? - It) X (1 + Ka - TGFinduction X LiverVolume/(LiverER _ E2(t - It) + LiverER
_E2OH(f - //))))) - A:-proteolysis X LiverTGF - A:-binding X LiverEGFReceptor X
LiverTGF + A:_ binding X K- TGF X LiverVolume X LiverlnternalEGFR
= k - binding X LiverEGFReceptor X LiverTGF - k _ binding X K _ TGF X LiverVolume X
LiverlnternalEGFR - k _ endocytosis X LiverlnternalEGFR
Initial Conditions
LiverAhReceptor
LiverAhR _ Inducer
LiverCYPlAl
LiverCYPlAl_R
LiverCYPlA2
LiverCYPlA2_R
LiverP450Reductase
LiverE2
LiverEReceptor
LiverER _ E2
LiverER _ E2OH
LiverEGFReceptor
(All other state variables
= 0.017806 nmol (Sloop and Lucier, 1987)
= 0.000674 nmol (determined from conservation of mass)
= 0.063174 nmol (Tritscher et al, 1992)
= 0.094902 nmol (determined from conservation of mass)
= 2.4199 nmol (Tritscher el al, 1992)
= 1.9799 nmol (determined from conservation of mass)
= 0.602536 nmol (Miwa el al, 1978)
= 0.00123 nmol (assumed at equilibrium with blood)
= 0.01495 nmol (Clark et al., 1991)
= 0.02325 nmol (determined from conservation of mass)
= 0.0066792 nmol (determined from conservation of mass)
= 0.022563 nmol (Sewall et al., 1993)
start at zero)
APPENDIX 2
Physiological Parameters
frac _ interstitium
It
ConcBloodE2
crit _ exposure
Production of Blood Proteins
MaxBloodProtein
K, - BloodProtein
AT_BloodProtein
= 0.1
= 0.25 day (Sloop and Lucier, 1987) (delay between binding of liganded receptor to DNA
and appearance of induced proteins in the liver cell)
= 0.185 nmol/liter (Shiverick and Muther, 1983)
= 0.3 nmol
= 5 nmol/liter (determined from the data of Tritscher et al., 1992)
= 0.1 nM (determined from the data of Tritscher et al., 1992)
= 0.4 nM (determined from the data of Tritscher et al., 1992)
A-14
-------
KOHN ET AL.
APPENDIX 2: Continued
Receptor Binding Constants
A'-AhR
A:_ER_E2
,K_ER_E2OH
K-TGF
K-TCDD
= 0.27 nM (apparent Kdin biological materials; Poland and Knutson, 1982)
= 0.13 nM (Vickers et al, 1987)
= \3nM(Uetal., 1985)
= 0.28 nM (Hudson et al., 1985)
= 30 nM (Poland et al., 1989; Voorman and Aust, 1989)
Constitutive Expression Rates
AhRexpression = 1.4553 nmol/liter/day (determined from conservation of mass)
ERexpression = 0.67273 nmol/liter/day (determined from conservation of mass)
CYPlA2expression = 213.4 nmol/liter/day (determined from conservation of mass)
P450Reductase _ expression = 47.45 nmol/liter/day (determined from conservation of mass)
EGFRexpression = 0.69376 nmol/liter/day (determined from conservation of mass)
Endogenous Ligand of Ah Receptor
Natural _ Inducer
Gene Induction Parameters
V _ AhRinduction
K _ AhRinduction
V_CYPlAlinduction
K _CYP1 A1 induction
F_CYPlA2induction
A:_CYPlA2induction
V _ ERinduction
K _ ERinduction
K, _ ERinduction
V _ TGFinduction
K _ TGFinduction
Ka _ TGFinduction
Metabolic Parameters
F_E2H
/C_E2
K- Reductase
Rate Constants
k _ absorption
k — binding
k _ proteolysis
k - endocytosis
k _ metabolism
k - urine
k - bile
k - feces
k - conjugation
k - lysis
= 0.0015 nmol/liter (determined from the data of Tritscher et al., 1992; assumed same Kd as
for TCDD)
7.2 nmol/liter/day (Lower bound of 4 nmol/liter/day; Sloop and Lucier, 1987)
4 nM (Sloop and Lucier, 1987)
3319 nmol/liter/day (determined from the data of Tritscher et al., 1992)
4.279 nM (determined from the data of Tritscher et al., 1992)
4197 nmol/liter/day (determined from the data of Tritscher et al., 1992)
7.458 nM (determined from the data of Tritscher et al., 1992)
3.2488 nmol/liter/day (Romkes et al., 1987)
0.35 nM (determined from the data of Clark et al, 1991)
3.1 nM (determined from the data of Clark et al., 1991)
1.5 nmol/liter/day (determined from the data of Sewall et al., 1993)
9 nM (determined from the data of Sewall et al., 1993)
0.75 nM (determined from the data of Sewall et al., 1993)
8496 day"1 (Graham et al., 1988)
9400 nM (Voorman and Aust, 1987)
83.5 nM (Miwa et al., 1987)
4.8 kg°-75/day (Leung el al., 1990)
105 nmol"1 day"1 (value selected to ensure that binding reactions are always close to
equilibrium)
0.693 day"1 (Lucier et al., 1972)
0.271 day"1 (determined from the data of Sewall et al., 1993)
2.75 day"1
5.36 day"1 (Birnbaum et al., 1980)
3.81 day"1 (Birnbaum et al., 1980)
1.152day-'(Lutz#a/., 1977)
56.693 day"1 (Lucier and McDaniel, 1977)
9.5 day"1
A-15
-------
MODEL OF EFFECTS OF DIOXIN ON HEPATIC GENE EXPRESSION
APPENDIX 2: Continued
Partition and Diffusion Coefficients
FatPartition
MusclePartition
VisceraPartition
LiverPartition
FatDiffusion
MuscleDiffusion
VisceraDiffusion
LiverDiffusion
= 350 (Leung et al, 1990)
= 40 (Leung et al, 1990)
= 20 (Leung et al, 1990)
= 20 (Leung et al, 1990)
= 0.1 (selected to fit the data of Abraham et al,
= 0.5 (selected to fit the data of Abraham et al,
= 0.5 (selected to fit the data of Abraham et al,
1988)
1988)
1988)
= 0.4 (selected to fit the data of Abraham et al, 1988)
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Ullrich, V., and Duppel, W. (1975). Iron- and copper-containing mono-
oxygenases. In The Enzymes (P. D. Boyer, Ed.). 3rd ed., Vol. 12, pp.
253-297. Academic Press, New York.
Vickers, A. E. M., and Lucier, G. W. (1991). Estrogen receptor, epidermal
growth factor receptor and cellular ploidy in elutriated subpopulations
of hepatocytes during liver tumor promotion by 17a-ethinylestradiol in
rats. Carcinogenesis 12, 391-399.
Vickers, A. E. M., Nelson, K., McCoy, Z., and Lucier, G. W. (1989).
Changes in estrogen receptor, DNA ploidy, and estrogen metabolism in
rat hepatocytes during a two-stage model for hepatocarcinogenesis using
17«-ethinylestradiol as the promoting agent. Cancer Res. 49, 6512-
6520.
Voorman, R., and Aust, S. D. (1987). Specific binding of polyhalogenated
aromatic hydrocarbon inducers of cytochrome P450d to the cy-
tochrome and inhibition of its estradiol 2-hydroxylase activity. Toxicol.
Appl. Pharmacol. 90, 69-78.
Voorman, R., and Aust, S. D. (1989). TCDD(2,3,7,8-tetrachlorodibenzo-
p-dioxin) is a tight binding inhibitor of cytochrome P-450d. J. Biochem.
Pharmacol 4, 105-109.
A-17
-------
Risk Analysis, Vol. 13, No. 1, 1993
APPENDIX B
Modeling Receptor-Mediated Processes with Dioxin:
Implications for Pharmacokinetics and Risk Assessment
Melvin E. Andersen,1'6 Jeremy J. Mills,1 Michael L. Gargas,1 Lorrene Kedderis,2
Linda S. Birnbaum,3 Diether Neubert,4 and William F. Greenlee5
Received April 20, 1992; revised July 24, 1992
Dioxin (2,3,7,8-tetrachlorodibenzo-/>-dioxin; TCDD), a widespread polychlorinated aromatic hy-
drocarbon, caused tumors in the liver and other sites when administered chronically to rats at doses
as low as 0.01 (xg/kg/day. It functions in combination with a cellular protein, the Ah receptor, to
alter gene regulation, and this resulting modulation of gene expression is believed to be obligatory
for both dioxin toxicity and carcinogenicity. The U.S. EPA is reevaluating its dioxin risk assess-
ment and, as part of this process, will be developing risk assessment approaches for chemicals,
such as dioxin, whose toxicity is receptor-mediated. This paper describes a receptor-mediated
physiologically based pharmacokinetic (PB-PK) model for the tissue distribution and enzyme-
inducing properties of dioxin and discusses the potential role of these models in a biologically
motivated risk assessment. In this model, ternary interactions among the Ah receptor, dioxin, and
DNA binding sites lead to enhanced production of specific hepatic proteins. The model was used
to examine the tissue disposition of dioxin and the induction of both a dioxin-binding protein
(presumably, cytochrome P4501A2), and cytochrome P4501A1. Tumor promotion correlated more
closely with predicted induction of P4501A1 than with induction of hepatic binding proteins.
Although increased induction of these proteins is not expected to be causally related to tumor
formation, these physiological dosimetry and gene-induction response models will be important
for biologically motivated dioxin risk assessments in determining both target tissue dose of dioxin
and gene products and in examining the relationship between these gene products and the cellular
events more directly involved in tumor promotion.
KEY WORDS: PB-PK modeling; dioxin; gene regulation; cytochrome P450; risk assessment; pharmacoki-
netics; pharmacodynamics.
1. INTRODUCTION
Dioxin (2,3,7,8-tetrachlorodibenzo-/>-dioxin;
TCDD), a carcinogen'1-2* and tumor promoter'3' in ro-
1 Chemical Industry Institute of Toxicology, P.O. Box 12137, Re-
search Triangle Park, North Carolina 27709.
2 Curriculum in Toxicology, Center for Environmental Medicine, Uni-
versity of North Carolina, Chapel Hill, North Carolina 27599.
3 Health Effects Research Laboratory, U.S. Environmental Protection
Agency, Research Triangle Park, North Carolina 27711.
4 Institute of Toxicology and Embryopharmacology, Free University
Berlin, Garystr 5, D-1000 Berlin 33, Germany.
5 Department of Pharmacology and Toxicology, School of Pharmacy
and Pharmacal Sciences, Purdue University, West Lafayette, Indiana
47907-7880.
6 To whom all correspondence should be addressed.
dents, also causes immunotoxic,(4) reproductive/5' and
teratogenic(6) effects. The present U.S. EPA cancer risk
assessment, based on the incidence of liver tumors in
female rats dosed with dioxin for 2 years,(1) utilizes the
standard linearized multistage (LMS) approach, with a
one in a million risk level, to derive a virtually safe dose
of 6 femtograms dioxin/kg body weight per day.(7) Other
Western countries treat dioxin as a tumor promoter, with
a threshold for its effects, and arrive at acceptable or
tolerable daily intake values up to 1000-fold higher than
that used by the United States.(8)
All of the toxic effects of dioxin noted above are
believed to be dependent on the interaction of dioxin
with a specific cellular protein, the Ah (oryl /tydrocar-
bon) receptor. The Ah receptor-dioxin complex binds to
0272-4332/93/0200-0025S07.00/I O 1993 Society for Risk Analysis
B-l
-------
Andersen et al.
specific sites on DNA modifying regulation of various
genes, some of which can alter cell growth and differ-
entiation, while others affect metabolic processes (Fig.
1). Due to the obligate role of the/l/i receptor in its toxic
effects, dioxin has been referred to as a "receptor-me-
diated" carcinogen. Dioxin is only one of many poly-
halogenated dibenzo-p-dioxins, dibenzofurans, and
biphenyls that alter cell growth characteristics via inter-
actions with the Ah receptor. The EPA is reviewing the
current dioxin risk assessment with the stated intention
of developing a generic, biologically realistic approach
to risk assessment for these "receptor-mediated" agents.(9)
With improved knowledge of the biological basis of the
action of dioxin at the molecular level, it may well prove
possible to reconcile the present disparity among differ-
ent countries.
Dioxin demonstrates dose-dependent kinetics: as the
administered dose increases, a larger proportion of the
total dose is found in the liver.(10) Among the hepatic
proteins induced by dioxin is a cytochrome, P4501A2,
which has a high affinity for dioxin.(U-12) It appears likely
that dose-dependent hepatic sequestration is related to
induction of P4501A2, but other proteins may also be
involved. Any comprehensive model of dioxin phar-
macokinetics must include the induction of these dioxin-
binding proteins, mediated by the interaction of a dioxin-
,4/z-receptor complex with specific binding sites on DNA.
Leung and co-workers developed a physiologically based
pharmacokinetic (PB-PK) model for dioxin and conge-
ners in mice(13> and rats.(14) Their model included in-
duction of a single hepatic binding species occurring in
direct proportion to the fractional occupancy of the avail-
able Ah receptor by dioxin, but did not include direct
interactions of the Ah-TCDD complex with DNA. In
mice, dioxin pretreatment increased metabolic clearance
of dioxin-like analogs.(15)
This paper extends the earlier PB-PK models by
including induction of binding proteins/enzymes and of
dioxin metabolism in response to ternary interactions of
dioxin, the Ah receptor, and DNA binding sites and
analysis of repeated dose exposure situations. This phys-
iological model for dioxin disposition and enzyme in-
duction is discussed in relation to its potential role for
estimating dosimetry and gene regulation in receptor-
based risk assessment strategies for dioxin.
2. PB-PK MODEL
2.1. Model Structure
The dioxin PB-PK model (Fig. 2) consists of five
compartments—liver, fat, slowly perfused tissues (e.g.,
Ah + TCDD
tl
TCE
ti
Ah -TCDD
Ah -TCDD
CYPIA2 - TCDD
II
TCDD
+
HEPATIC BINDING PROTEIN
(CYPIA2)
DNA
Transcription
UNA
Translation
PROTEIN
Fig. 1. Schematic of the molecular mechanisms of action of dioxin. Dioxin binds to a cellular protein, the Ah receptor, and the dioxin-receptor
complex interacts with DNA sites in regulatory regions of certain genes. Interaction with these sites, called dioxin-responsive elements (DREs),
leads to changes in rates of gene transcription (i.e., the rate at which mRNA is produced from these genes). The mRNAs serve as templates for
protein synthesis. Changes in cellular mRNA levels lead to altered protein concentrations for cytochrome P4501A1, hepatic, dioxin-binding proteins,
metabolizing enzymes, and growth factors. The change in concentration of certain of these proteins is believed to be associated with the various
toxic effects of dioxin.
B-2
-------
PB-PK Modeling with Dioxin
sr
^-~.
i
V.
v^
V
BLOOD
FAT
„ „ , TCDD
Pp=375 I 1
rr
TCDD
SLOW
TCDD
TCDD
RICH
TCDD
TCDD
LIVER
Ah-TCDD Pr-TCDD
TCDD
^x
^
4^
^
J
Metabolism
Table I. Parameters in the Physiological Dosimetry Model for
Dioxin1
Fig. 2. The PB-PK model for dioxin. A five-compartment diffusion-
limited model was developed with metabolism and protein binding
included in the liver. Uptake after subcutaneous administration was
described as a first-order process with the chemical appearing in mixed
venous blood. Symbol definitions are in Table I.
muscle, skin, etc.), richly perfused tissues (e.g., kidney,
brain, etc.), and blood. The relevant mass-balance dif-
ferential equations appear in the Appendix and the terms
are defined in Table I. Each of the four tissue compart-
ments (denoted by subscript,) has both a specified blood
flow (Q,), tissue compartment volume (V,), and a tissue
blood volume (Vlb). The tissue blood volumes were es-
timated from Bischoff and Brown.<16) Movement of
chemical from the tissue blood into the tissue is modeled
to be proportional to a permeation coefficient-surface
area cross-product (PA) for the tissue (,). Tissue uptake
is diffusion-limited when PA,
-------
Andersen et al.
partment which effectively decreased the rate of tissue
uptake. Our revised model, because of the diffusion-
limited tissue compartments, does not require blood
binding to match tissue uptake time-course behavior.
Initial estimates of partition coefficients were ob-
tained from Leung et al.(14) and were adjusted in fitting
the data sets here. As expected, the fat has the highest
partition coefficient due to the highly lipophilic nature
of dioxin. In addition, metabolism and protein binding
are included in the liver, where both ih&Ah receptor and
the inducible binding protein act to sequester dioxin via
capacity-limited binding processes. The binding protein
was assumed to be cytochrome P4501A2. The concen-
trations of P4501A2 in naive and induced rats differ by
about 10-fold.(17) The differential equation for the liver
(Eq. 5) is solved for the total amount of dioxin in liver,
which is related to partitioned dioxin, dioxin bound to
the Ah receptor, and dioxin bound to P4501A2 (Eq. 6).
This conservation equation is used to solve for the free
dioxin in the liver tissue, C,f, which is then utilized in
calculating the concentration of the Ah receptor-dioxin
complex available for inducing cytochrome P4501A2 and
cytochrome P4501A1 activity. This latter protein, an ox-
idative microsomal enzyme induced by dioxin, is fre-
quently used as a marker of Ah receptor-mediated protein
induction. All these binding interactions are described
by simple, reversible equilibrium relationships. This ap-
proach is valid as long as the rate constants for associ-
ation/dissociation are large (i.e., the processes are rapid).
This description, like all models, is really a simplifica-
tion of the more complex series of biological events.
The model can readily be extended to include new in-
formation on these biological associations as it becomes
available.
induction of P4501A1 activity (Figs. 3 and 4). Basal
levels of P4501A1 are described by a zero-order pro-
duction rate (K0) and a first-order elimination rate con-
stant (fcj). The induction process increases the production
rate as specified by the Hill relationship and the maxi-
mum observable enhancement (K0m!a) (Eq. 7). Data on
the time course of cytochrome P4501A1 from Abraham
et al.(10) were used to estimate the first-order degradation
rate constant (&j) for P4501A1 degradation.
Induction of hepatic binding capacity was also mod-
6.0 -i
10'
1000
Dose TCDD - (ig/kg bw
Fig. 3. Dose-dependence of dioxin tissue disposition in female Wistar
rats 7 days following single subcutaneous doses. Tissue concentrations
from Abraham et a/.(10) are expressed normalized to dose, as %dose/
tissue. Model parameters are given in Table I.
2.2. Enzyme/Protein Induction
The induction process is described by assuming that
the /4/i-TCDD complex is formed based on Ah receptor-
binding parameters (BMl and KB^ and free dioxin in
the liver (C,f). This complex then binds to unspecified
sites on DNA with an affinity, Kd. Since the binding to
DNA does not directly alter the^/i-TCDD concentration
in this present model configuration, it is tacitly assumed
that the DNA sites are present at much lower concentra-
tions than the>l/j-TCDD complex. The induction is mod-
eled with a Hill plot binding relationship (Eq. 6), where
n provides a measure of interaction for multiple Ah-TCDD
complex binding sites. The value of n is estimated by
comparison of model simulations to data on the dose-
dependence of liver sequestration of dioxin or on the
n
o
500 -
250 -
Dose TCDD - ng/kg bw
Fig. 4. Dose-dependence of cytochrome P4501A1 induction in female
rats 7 days following single subcutaneous doses. Data from Abraham
et a/."0> for activity of EROD (Ethoxyresorufm-0-deethylase) in pmoles
resorufin formed/min/mg protein.
B-4
-------
PB-PK Modeling with Dioxin
eled as if it were instantaneously altered by the Ah re-
ceptor-dioxin-DNA interactions (Eq. 8). As data becomes
available for the time course of P4501A2 induction, syn-
thesis and degradation rate constants for this protein can
also be directly included in this description. For both
cytochrome P4501A1 and 1A2 induction, it was as-
sumed that/1/z-TCDD complex formation was equivalent
(similar Kbl and BM± values), but that the n value and
Kd for the two responses differed (see Table I). Our
model also allows an induction of dioxin metabolism
following dioxin treatment. However, in contrast to the
mouse, induction of metabolism in the rat, if present at
all, was negligible. The constants employed assume at
most only a doubling of the metabolic rate.
2.3. Data Analysis
The data used for this analysis were from two pre-
viously published studies with Wistar rats. The first
study,'10' with female rats, provides both dose-response
characterization of liver concentrations and of liver
P4501A1 activity and time-course characterization of
dioxin tissue concentrations and enzyme activities. The
second study(18) examined liver and fat concentrations in
male Wistar rats dosed weekly for periods up to 6 months.
Model simulations were performed using ACSL
(Advanced Continuous Simulation Language, Mitchell
and Gauthier Assoc., Concord, Massachusetts).
to 1.0 suggests little interaction among dioxin responsive
DNA binding sites involved in expression of this partic-
ular gene. The complex behavior above 10 jig/kg bw is
commented on more fully in the Discussion.
The dose-response curve for cytochrome P4501A1
induction (Fig. 4) was described in a similar manner,
but required a larger value for nl (2.3) to fit the data,
indicating possible interactions among DNA binding sites
for the Ah receptor-TCDD complex with this gene. The
half-maximal induction response for P4501A1 occurred
at about a 10-fold higher dose than the half-maximal
response of the binding protein (Kdj = 180 pM).
The PB-PK model configured for these dose-re-
sponse curves was also used to examine time-course
elimination/induction after a single dose of 300 ng/kg
(Figs. 5 and 6). Analysis of these results were especially
useful for setting the rate constant for metabolism of
dioxin and the PA, cross-products (i.e., the diffusion
limitations) for liver and fat. The simulation of these
curves requires inclusion of time-dependent growth pa-
rameters over the 100 days of the experiment. Growth
rates and volumes of fat were estimated from growth
curves for these rats available from suppliers.
3.2. Repeated Dosing
The primary health concerns with dioxin are asso-
ciated with repeated low or chronic exposures. We ana-
lyzed data from a study in which liver and fat
3. RESULTS
3.1. Dose-Response
In this study(10) rats received a single subcutaneous
dose of dioxin and were killed 7 days later. The dispo-
sition of dioxin in liver and fat was highly dose-depen-
dent in the concentration range between 1 ng/kg and
10,000 ng/kg (Fig. 3). Normalized concentrations (%
dose/g tissue) would be horizontal lines if disposition
were dose-independent. The curvature appears to be due
to the induction of a dioxin-binding protein, presumably
cytochrome P4501A2. The smooth curves were obtained
with the PB-PK model based on the parameters in Table
I. For induction of the binding protein, n and Kd were
estimated by fitting the data from Wistar rats(10) and
were, respectively, 1.0 and 50 pM. Using measured con-
centration estimates of basal and induced P4501A2,(17)
the affinity of the binding protein (KB2) was estimated
from the curve fitting to be 6.5 nM. A value of n close
Q
n
36 54 72
Time - days
Fig. 5. Time course of liver and fat tissue dioxin concentrations fol-
lowing a single subcutaneous dose of 300 ng/kg in female Wistar rats.
Concentrations are expressed as ng/g tissue, and the data are from
Abraham et a/.(10) Parameters used were the same as in Figs. 3 and 4;
however, growth of the rat and fat compartments were also included.
B-5
-------
Andersen et al.
to
I
300-
200-
36 54 72
Time - days
Fig. 6. Time course of cytochrome P4501A1 activity following a sin-
gle dose of 300 ng dioxin/kg in female Wistar rats. Data are from
Abraham et a/."01 and parameters as in Fig. 5.
g
1000
100:
Liver
Fat
32
64 96
Time - days
128
160
Fig. 7. Time course liver and fat concentrations in rats dosed weekly
with 5jig dioxin/kg starting 7 days after a loading dose of 25 |J.g/kg.
These data, expressed as ng/g tissue, are from Krowke et al. (18) The
model parameters were as specified in other figures with several changes:
Ps, P,, Pafc, and kfc were, respectively 250, 50, 0.1 and 1.5. These
contrast to values of 375, 30, 0.20, and 1.65.
concentrations were determined in male rats for up to 6
months during weekly dosing with dioxin.(18) This study
(Fig. 7) used a loading dose (25 ng/kg) followed by a
fivefold smaller dose (5 ng/kg) every week. The simu-
lations for this dosing scenario were conducted with model
parameters very similar to those used in the single-dose
exposures (Figs. 3-6). Small differences in fat and slowly
perfused tissue compartment parameters are noted in the
figure legends.
3.3. Correlation with Responses
The present EPA dioxin risk assessment is based
on liver tumors in female rats.(1) Liver tumors are also
increased in female and male mice.(2) Liver initiation-
promotion studies'3-19-20) demonstrate that dioxin is a
promoter and that the dose-response curve for its pro-
moting action on altered hepatic foci closely follows the
dose-response curve of its hepatocarcinogenic activity
(Table II). Both of these dose-effect relationships are
very steep.
Based on our preliminary repeated-dose PB-PK
model for these Wistar rats, several measures of dose
can be calculated for comparison with the promotional
efficacy and carcinogenicity of dioxin in Sprague-Daw-
ley rats (Table II). These include integrated total liver
dioxin concentration during the treatment period, or in-
tegrated free liver dioxin concentration. In addition,
measures of tissue dose related to enhanced expression
of cytochrome P4501A1 and hepatic binding proteins
can be calculated for the duration of the subchronic ex-
posure. The measures of tissue dose associated with en-
Table II. Relationship Between Hepatic Exposure to Dioxin,
Hepatic Protein Induction, and Hepatic Toxicity During Subchronic
Exposures
Altered
foci" Tumors' AUCL-free"AUCL-total'AU-lA2^AU-lAl*
Dose" volume (liver)
C, (BM2,) (IND)
Control
0.0001
0.001
0.
0.
1
10
,01
,1
0.
,8
1/86
0.2 -
0.
0.
2.
,3
,7
,8
0/50
2/50
11/49
—
—
6
6
5
4
3
3
0
.7x10-'
.4x10°
.3x10'
.2xl02
.8x10'
.8xl04
3
5
1
1
1
5
0
.5 x 10'
.9xl02
.1x10"
.2x10*
.OxlO6
.4x10*
7
5
2
3
3
3
0
.OxlO3
.7xl04
.IxlO3
.2 xlO3
.SxlO5
.5x10*
0
2.3x10'
3.9 xlO3
2.7 xlO5
2.0 xlO"
3.0 x 10s
3.1x10"
• Biweekly doses used; dose expressed as average dose per day in (xg/
kg.
* Volume as percent of liver occupied by altered cells, from Pilot et
a/.<18>
c Liver tumors observed in female rats in the bioassay study of Kociba
era/.")
d Total area under the curve for free dioxin in the liver, units are
nmoles-hr/liters.
' Total area under the curve for total dioxin in the liver, units are
nmoles-hr/liters.
1 Integrated exposure to induced levels of binding protein; units are
nmoles-hr/liter; and the calculation integrates the third term on the
right side of Eq. (6).
* Integrated exposure to induced levels of cytochrome P4501A1; units
are enzyme unit-hr; and the calculation only considers the enhanced
level of activity due to induction.
B-6
-------
PB-PK Modeling with Dioxin
hanced gene expression (assumed to be related to P4501A1
activity or the sequestration of dioxin in liver by cyto-
chrome P4501A2) are the integrated level of these gene
products (as protein concentration of 1A2, or activity of
1A1) over time. Enhanced expression refers to the in-
crease over any basal levels of expression of these gene
products. This correlative approach does not assume that
there should be a direct relationship between induction
of these cytochromes and the promotional action of dioxin
at all dose levels.
The tumor promotional response of dioxin in the
rat liver is most closely correlated with the integrated
expression of the P4501A1 gene (Table II) under these
exposure conditions. For instance, the responses (col-
umns 2 and 3) increase rapidly with increasing dose be-
tween 0.01 and 0.1 pig/kg/day. The integrated level of
cytochrome P4501A1 increases by a factor of about 10
in this range, while the integrated amounts of binding
protein (P4501A2), whose induction is already nearly
saturated by 0.01 |xg/kg/day, only increases by 50% in
this region of dose. Dioxin concentrations increase in
this dose range, but dioxin itself is not believed to be
responsible for toxic effects.
4. DISCUSSION
4.1. Pharmacokinetics
The half-life of dioxin in male rats was initially
reported to be 20-30 days.(21) These early studies were
conducted at high, inducing doses; elimination curves
were obtained for only about two half-lives; and analysis
relied on chemical detection of dioxin in tissues. The
half-lives in hamsters (10-15 days) and guinea pigs (30-
40 days) were also estimated from time course data with
a simple one-compartment model for dioxin kinetics but
used radiochemical detection of labeled dioxin.(22>23) These
various studies were relatively insensitive to the dose-
dependent effects apparent only when disposition is ex-
amined at much lower doses (Fig. 3) and most apparent
when both fat and liver concentrations are measured.
The alteration in the ratio of liver-to-fat concentration
ratio is the most sensitive marker for this dose-depend-
ence (compare Figs. 3 and 8b). These low dose effects
can now be readily evaluated by use of high specific
activity 3H-labeIed dioxin. Single doses of 1 ng/kg caused
minimal induction of P4501A1 (Fig. 4) or of the dioxin
binding protein (Fig. 3). Protein induction, however,
becomes significant at a dose of 30-50 ng/kg dose and
10°
'i io-4^
E
e
a.
10'6|
i
-------
Andersen et al.
and, in a preliminary form, for dioxin itself26'. With the
exception of the work by Leung and colleagues,'13-14'
none of these descriptions have included specific hepatic
binding. In all these other models, chemicals were dis-
tributed simply by nonspecific, dose-invariant partition-
ing and focused on behavior at higher doses with
substantial induction of binding proteins. Because of the
limited dioxin-binding capacity of the liver, even in fully
induced animals, the liver/fat concentration ratio reaches
a maximum and is predicted to decrease at higher body
burdens (Fig. 3), where hepatic binding to induced pro-
teins approaches saturation. While this region of dose
was not examined by Abraham et al. ,(10) the time course
of elimination has been examined after a single dose of
600 u-g/kg in hamsters, a species which is more resistant
to the acute toxic effects of dioxin than are rats. In this
study, the liver/fat ratio increased from 1.0-2.7 as the
body burden fell from 600 to 100 u,g/kg(22), consistent
with the predictions in Fig. 3 for the dose range above
10 ^g/kg bw.
Several people ingested relatively large amounts of
dioxin-like polychlorinated dibenzofurans in the Yusho
and Yu Cheng poisoning incidents in Japan and Taiwan,
respectively. In several instances, liver and fat samples
were obtained at autopsy, and the proportion of dose in
the liver was found to be dependent on the total body
burden. The % dose/liver increased from about 5 at am-
bient exposure levels to over 60 at a body burden of the
furans lexicologically equivalent to a burden of 3-5 |ig
dioxin/kg body weight. The body burden for half max-
imal sequestration in the liver in people was estimated
by Carrier et a/.(27) to be 0.5-1.5 u,g dioxin/kg body
weight. Thus, this dose-dependent hepatic distribution
is common both to rats and people with very similar
body burdens required for half-maximal sequestration.
The liver/fat concentration ratio, however, varies for dif-
ferent isomers. Higher chlorinated isomers tend to have
ratios higher than for TCDD itself. The reasons for these
differences are unknown; however, one suggestion de-
rived from modeling efforts is that KB2, the dioxin-bind-
ing protein dissociation constant, is lower for the higher-
chlorinated analogs (i.e., they have higher affinities for
the binding proteins than does dioxin itself).
4.2. Protein Binding/Induction
The challenge in providing a biologically realistic
PB-PK model for dioxin is the need not only to account
for the determinants of disposition (i.e., tissue partition-
ing, biotransformation rates, and protein binding con-
stants), but also to describe the pharmacodynamic events
related to the induction of specific dioxin-binding pro-
teins in the liver. Knowledge of the molecular mecha-
nisms of the interactions of the dioxin-/lA receptor complex
with regulatory regions of specific genes is growing rap-
j(jly(28,29)) ancj shows clearly a role for the ternary dioxin-
Ah receptor-DNA complexes in regulating gene tran-
scription. Our present model of disposition is based on
a ternary complex being obligatory for the pharmaco-
dynamic activity of dioxin at the genomic level. The
description of these interactions requires estimates of
binding constants between the Ah receptor and dioxin
and between the Ah receptor-dioxin complex and sites
on DNA. The ternary interactions with DNA are be-
lieved to enhance transcription of mRNA leading even-
tually to increased amounts of specific dioxin-binding
proteins, particularly cytochrome P4501A2.
The binding maximum (BMl) and binding affinity
(KB1) for TCDD binding to the Ah receptor have been
previously reported. The values used in our model differ
significantly: our estimate of BMl is only about 10% of
the value determined by Gasiewicz and Rucci(30) and our
KBl is also much lower than the value reported by Brad-
field and Poland.(31) The discrepancy in KBl is due sim-
ply to our decision to reference binding to the free dioxin
in the liver. If we had used total partitioned dioxin, our
number would increase by a factor of 20 and be more
similar to the in vitro value.(31) However, the lower BMl
value represents a more fundamental discrepancy be-
tween in vitro measurements and in vivo pharmacoki-
netics.
In the present study, the binding parameters of the
Ah receptor were largely estimated from the tissue con-
centrations of TCDD at noninducing levels, shown in
Fig. 3. The liver concentration at doses below 0.01 jig/
kg is especially sensitive to BM^ the Ah-binding maxi-
mum. In addition, the dose at which hepatic sequestra-
tion is half-maximal is extremely sensitive to KB1, the
Ah-b'mding affinity. If BMl were increased to the liter-
ature value, the liver concentration of TCDD in the low
dose region (/. e., noninduced animals) would be greatly
overestimated. One possibility to explain this discrep-
ancy is that only a relatively small proportion of the
available receptors participate in binding dioxin and
maintaining induction at physiologically realistic con-
centrations of TCDD in vivo. In mouse hepatoma cell
lines, the total receptor concentration decreases after ex-
posure to dioxin.(32)
The other important binding constant Kd, for the
DNA interactions, also affects the placement of the in-
duction curve, whether for 1A2 or 1A1, along the dose
axis. The Hill-type coefficients control the steepness of
the induction response. With the two responses exam-
B-8
-------
PB-PK Modeling with Dioxin
ined, induction of liver sequestration (i.e., 1A2) shows
higher affinity (lower Kd value) but lower cooperativity
(smaller n value) than does induction of 1A1. This ap-
parent cooperativity with 1A1 («j = 2.3) is consistent with
the observation that there are at least four dioxin re-
sponsive elements in the regulatory region of this gene.(28)
The PB-PK model structure developed here as-
sumes that a single type of Ah receptor-dioxin complex
interacts with DNA binding sites of variable affinity to
regulate different genes. More complex models with dif-
ferent Ah receptor-dioxin complexes (due to other pro-
tein interactions, for instance) might lead to different
conclusions about cooperativity in these two responses,
but such models do not seem warranted by available
biological data at this time.
progress in physiological modeling of dioxin pharma-
cokinetics now relies on the progress in uncovering the
biology of the Ah receptor,(37) especially regarding its
ability to affect the expression of dioxin-binding pro-
teins.
The current PB-PK model greatly simplifies the se-
quence of processes involved in activation of the recep-
tor by dioxin. Normally, the receptor is complexed with
heat shock protein(s) 90, which probably dissociate after
the receptor binds dioxin.(38>39) Another protein species
is then required for the interaction of the complex with
DNA sites.(40) In the future, it will be necessary to model
these events if they are found to be critical control ele-
ments for gene regulation; however, they are not under-
stood in sufficient detail to justify their inclusion in the
dosimetry model at this time.
4.3. Pharmacokinetics of the Ah Receptor
The biological activity of dioxin is a consequence
of both delivery of dioxin to target cells and the dynam-
ics of the processes that regulate Ah receptor concentra-
tion and Ah receptor binding characteristics. The present
model extends our understanding of the biological de-
terminants of dioxin action in vivo; however, it does not
account for time variant changes in the receptor. In com-
pletely induced animals, only a relatively small fraction
of total basal receptor binds to nuclear sites, even with
complete induction of P4501A1 activity.(33) Further-
more, in mouse hepatoma cells in vitro, the total cellular
receptor concentration decreases after treatment of the
cells with dioxin.(32) In contrast, there is an apparent
increase in Ah receptor in the rat liver following TCDD
treatment .(34> These complexities in/4/j-receptor kinetics
should have minimal effects on our analysis, which fo-
cuses on longer-term behaviors, after 7 days or over the
100-day excretion period. New experimental techniques
with appropriate antibodies*35* should soon be available
for measuring the total amount of cellular Ah receptor.
With these techniques, Ah receptor kinetics can be more
fully investigated by examining total amount of receptor,
receptor distribution, tissue dioxin, mRNA, and protein
for P4501A2 over the first several hours to days after
intravenous administration of dioxin. Such studies will
improve both our appreciation of the pharmacokinetics/
pharmacodynamics of the Ah receptor, and of the dosi-
metry of dioxin itself. It bears repetition that the deter-
minants of dioxin pharmacokinetics cannot be examined
in the absence of additional information. The pharma-
cokinetics of dioxin, of the Ah receptor, and of hepatic
binding proteins are inextricably coupled.(36) Continued
4.4. Risk Assessment Implications
Greenlee et al. ,(29) in a recent perspective on bio-
logically based risk assessment for dioxin, outlined the
biological steps involved in receptor-mediated growth
modulatory effects. They are recognition, transduction,
and response (Fig. 1). Our PB-PK model includes ele-
ments of recognition (binding of dioxin to the Ah recep-
tor), transduction (binding of the complex to DNA), and
response (the changes in rate of transcription/translation
of P4501A1 and the binding protein). The responses of
these two proteins are not thought to be causally linked
to the adverse effects of dioxin. Specific mitogenic and
mito-inhibitory cellular processes altered by dioxin treat-
ment are more likely causally related to toxicity.'41'42'
Even if we had modeled regulation of growth factor pro-
teins, it is unclear what specific role they play, singly
or in combination, in cell replication, cell differentia-
tion, or toxicity. It will be of interest to examine the
kinetic characteristics of induction (or repression) of these
other proteins to see if they are controlled in a manner
similar to P4501A1 or the sequestering protein. Since
data on modulation of growth regulatory genes is not yet
available, we have correlated various tissue dose mea-
sures from the present model with the promotional ef-
ficacy of dioxin. Of the several measures of tissue dose
examined (Table II), the integrated exposure to P4501A1
is most closely correlated with promotion and with he-
patic tumors. However, there is no expectation of caus-
ality between tumor responses and these induced proteins,
and this correlation should be regarded cautiously.
In using the integrated P4501A1 exposure, the low
dose extrapolation slope approaches the value of the Hill
B-9
-------
Andersen et al.
coefficient (nl = 23), reflecting the cooperativity which
appears to exist for induction of this gene by the ternary
TCDD-Ah receptor-DNA interactions (Fig. 8a). The dif-
ferences in the low dose extrapolation with measures of
tissue dose related to protein induction (Fig. 8b) is re-
lated both to their respective slopes (« values) and the
ease of induction by dioxin (the values of Kd and Kd^),
The mathematical formulation of these cooperativity in-
teractions is based on a (n +1) order interaction. Such
interactions are highly unlikely since the rate of these
association processes fall sharply with decreasing ligand
concentration. Normally, cooperative relationships arise
due to enhanced affinity for a second ligand molecule
after the binding of the first ligand. In this way, all
events involved are bimolecular. The Hill equation is
simply a convenient way to treat this more complex se-
ries of events. With these multiple bimolecular steps,
the low dose induction behavior should have a slope of
1 and be characterized by the binding parameters of the
first ligand. Dose measures related to free or total dioxin
in liver are also complexly related to administered dose;
neither curve is strictly linear with dose and the depar-
tures from linearity are in opposite directions.
Appropriate measures of tissue dose for risk as-
sessment must focus on specific cellular events, such as
cell proliferation rates with developing preneoplastic foci
in the livers of dioxin-treated rodents. Among the im-
portant ongoing efforts to improve the biological basis
of dioxin risk assessment are studies analyzing regula-
tion of dioxin-responsive growth regulatory genes in
liver(43) and the effect of these genes on replication and
focal growth. Eventually, biological response models will
have to predict the relationship between replication or
differentiation and the tissue exposure to these growth
factors. Only when this level of resolution is achieved
will it be possible to more fully justify any particular
approach to low-dose extrapolation with dioxin.
The basic behavior of dose-dependent hepatic se-
questration is a characteristic of many congeners of
dioxin(44) and is observed in several species including
people.(45) This basic PB-PK model structure appears to
be applicable to many chemicals that act via interactions
with the Ah receptor and should prove, in time, useful
for risk assessment applications with this broad group of
important contaminants with common receptor-mediated
mechanisms of toxicity. These dosimetry models will
have to be combined with emerging quantitative descrip-
tions of cell and tissue responses to develop a complete
biologically motivated risk assessment model. Some of
the challenges of this approach have been outlined in
this paper.
ACKNOWLEDGMENTS
The work on dioxin pharmacokinetics at CUT has
been generously sponsored by a grant from the American
Paper Industry. We also thank Drs. R. Conolly and R.J.
Preston for helpful comments and discussions.
Although the research described in this article has
been supported by the United States Environmental Pro-
tection Agency, it has not been subjected to agency re-
view and therefore does not necessarily reflect the views
of the Agency and no official endorsement should be
inferred. Mention of trade names or commercial prod-
ucts does not constitute an endorsement or recommen-
dation for use.
APPENDIX
There are two mass balance differential equations
for each tissue, one for tissue blood (tb) and another for
tissue (f). Respectively,
dAJdt = Q,(Ca - Cvl) + PA,(C,/Pt - Cvl) (1)
dAJdt = PA,(CVI - C,/P,) (2)
For fat, slowly perfused, and richly perfused tissues,
these equations are integrated for amount and tissue con-
centrations calculated by dividing amount (A,) by vol-
ume (V,).
A,=
C, = AJVt
(3)
(4)
The tissue free (diffusible) concentration is calculated by
dividing the tissue concentration (Ct) by the tissue par-
tition coefficient (P,)— Eq. (1) and (2). The liver equa-
tion [Eq. (5)] also contains a term for loss due to
metabolism:
dAJdt = PA,(Cvl - Clf)
- V,kf(\ + fold(C///(C//
(5)
In the development of the mathematical description used
here, there is the possibility for inducing metabolism in
direct proportion to the fractional occupancy of the Ah
receptor by dioxin. The extent of maximal induction as
increase over the basal rate is controlled by the fold term
in the above equation. For rat, this value was set to 1.0;
it may vary in other species, such as the mouse. (13)
The total mass in the liver is then apportioned be-
tween free dioxin (C,f) and bound forms of dioxin.
R-10
-------
PB-PK Modeling with Dioxin
Al =
[BM1Cvl(KBl
+ [BM2£yf(KB2 + C,,)} (6)
The second term on the right of the equality is the con-
centration of the Ah-TCDD receptor complex (4/i-TCDD)
and the third term is the concentration of the complex
of TCDD and the induced binding protein.
The activity of P4501A1 at time, /, is calculated
from the basal synthesis rate (K0), the maximum increase
in synthesis rate (K0max), the Ah receptor-TCDD com-
plex concentration, the complex-DNA dissociation con-
stant (RAJ, the appropriate Hill term (nj, and the
P4501A1 degradation rate constant (k^.
d(P45QlAlt)/dt =
+ Kd?))
(7)
The concentration of P4501A2 at any time, BM2,, was
calculated assuming an instantaneous adjustment of
binding protein concentrations and Ah-TCDD levels,
BM2, = BM20
+ BM2,(/M-TCDD)"/((/l/i-TCDD)'' + Kd")) (8)
where BM2: is the maximum induction of the binding
protein. This formulation assumes very rapid induction
of the binding proteins in these studies.
For the longer duration studies, the changes in total
body weight and proportion of weight as fat compart-
ment volume were included via table functions available
in the ACSL software package.
For the single-dose studies in female rats,
Simulated time (hr)
Body weight (kg)
Fat (%bw)
0
0.215
0.07
840
0.3
0.09
1344
0.34
0.13
0
0.35
0.10
840
0.39
0.12
1344
0.41
0.13
For the repeated dose studies in male rats,
Simulated time (hr)
Body weight (kg)
Fat (%bw)
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B-12
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DRAFT-DO NOT QUOTE OR CITE
APPENDIX C
PARAMETERS FOR ANALYZING PRENEOPLASTIC LESIONS AND TUMOR
INCIDENCE IN RAT HEPATOCYTES
Under the assumptions of 8.2, when fitting the two-stage model to the Kociba et al.
(1976) data, there is the potential to estimate as many as 24 parameters (four dose groups,
each with its own two-stage model having six parameters). As discussed in this Appendix,
the data given in this study make it impractical to estimate this many parameters with any
degree of accuracy (Kopp-Schneider and Portier, 1991). To reduce the number of
parameters in the model, (8N and 5N were assumed to be known without error. Assuming
that, on average and over finite time, the population of normal cells is effectively constant, it
is reasonable to assume that j8N = 5N. This assumption may not hold true for TCDD
because the labeling index for normal cells seems to increase with increasing exposure to
TCDD (Lucier et al., 1991). However, to illustrate the use of the two-stage model, this
assumption will be employed. Labeling data suggest that normal hepatocytes undergo mitosis
at an average rate of one mitotic event per 300 days (a 2% labeling index for a 6-day
labeling experiment), suggesting that 0N = 3.333xlO'3. This value was used in the analysis
that follows. Other values for /3N = 5N were tried and had no effect on the resulting model
parameters as long as it was assumed that changes in /3N do not imply changes in /tN_r (Portier
and Kopp-Schneider, 1991). The number of normal hepatocytes was assumed to be 6xl08
for the female rat.
There are 16 parameters to be estimated from the tumor incidence data. Table C-l
shows one set of parameter estimates resulting from this fitting exercise (labeled
"Unrestricted" under the model heading). Multiple attempts at fitting this model to the tumor
incidence data indicated that the model was not identifiable; this means that there are an
infinite number of parameter estimates that give effectively the same answer for these data.
The main problem seems to be estimating /tN.r and /ir.M simultaneously (for example, in the
high-dose group, any model with p^ • /ir.M equal to 3.4 • 10"15 seemed to yield equivalent
C-l 06/30/94
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fits). The main reason for including the "Unrestricted" model in Table C-l is that this model
represents the best fit we can possibly achieve, and the likelihood (column 6) represents a
best possible measure of goodness-of-fit. We will compare the likelihood for other models
with this target likelihood.
The first restricted model to be considered is a model in which the effect of TCDD is
only on the mutation rate from normal cells to initiated cells in the two-stage model. This
model is fit by forcing fa, 6,, and jt,_M to be constant across all dose groups and allowing ^NJ
to vary freely as dose changes. The parameter estimates for this model are given in Table
C-l as model (2). There are nine fewer parameters in this model than in the unrestricted
model. Comparing the two likelihoods, it is evident that this model fits the data significantly
worse than does the unrestricted model (X29 = 12.71, p<0.01, as a technical note, since
there is a problem with identifiability, the degrees-of-freedom, 9, for this X2 random variable
is inflated, which would inflate the p-value and the significance of the result would remain).
The problem with identifiability does not abate by restricting the parameters in the
model. This is illustrated by the next model (3) in Table C-l. In this model, it is assumed
the effect of TCDD is restricted to the rate of mutation from the initiated state to the
malignant state. In this model, fa, 5,, and /iN.j are held constant over all groups and /*!.M
changes as dose changes. Three points indicate the problem of nonidentifiability with this
model. The first is that the likelihoods for this model (3) and the model changing ^N.j
(2) are identical. This indicates that the two models explain the same amount of noise in the
data. The second point is that, even though the magnitude of the birth rate fa has changed,
the difference fa - 5j has not (~0.7xlO"2). Further modeling with these data indicates this to
be true over a wide range of fa values. The third point is that, for each dose group, /tNJ •
/*!_M is the same in the two models. Again, use of different fixed values for ftN.r in model (3)
and /itM in model (2) supports this result.
The net result is that, unlike the results seen by Moolgavkar and Luebeck (1992)
when fitting colon cancer data to the two-stage model, the magnitude of the mutation rate
/*N_! relative to ^_M cannot be addressed for the TCDD data. However, the relative change
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in the product of these two mutation rates as a function of dose can be studied. This will be
done later in this document.
A second restricted model is to consider that dioxin's only effect is on the birth rate
of initiated cells (/Si). This is in fact the modeling approach used by Thorslund (1987). This
model assumes that /XN.J, fy, and /xtM do not change as a function of exposure to dioxin. This
model is given as model (4) in Table C-l. This model is also inconsistent with the data
(likelihood = 67.57) and provides a fit that is worse than the mutational effect model given
by (2). [Technical note: It is difficult to directly compare these two models since they
constitute nonnested models.] This implies that the effect of TCDD on tumor incidence in
these rodents is more likely due to a mutational effect than a mitogenic effect on initiated
cells. Caution must be taken when interpreting this result. First, no statistical confidence
can be placed on this statement, so the observed difference in the two models may be due
solely to random change. Second, this statement is only justified within the restricted context
of this two-stage model of carcinogenesis. If any assumptions of the model are incorrect
(independent cell action, constant rates, two stages, etc.), the interpretation based on this
model could be biased.
The problem with identifiable parameters also remains with this formulation of the
model. This is illustrated by the fifth model in Table C-l. In this model, the birth rate in
the lowest dose group was fixed to be zero. The resulting likelihood and estimates of /%_!
and /itM remained the same. The estimated birth rates 03r and death rate 6j) changed, but the
difference between the two in the various dose groups did not change. Repeated applications
of this formulation of the model confirmed this problem of identifiability in this case.
With this problem of identifiability, there are basically only two parameters to be
estimated for the two-stage model for each dose group from these data. These are the
mutation parameter, given by /* = /t^ • jti_M, and the proliferation parameter, given by
n = ft - Sj. Table C-2 gives the estimates of these parameters for the simple models
considered here. The first model corresponds to the case where we allow both the mutation
parameter and the proliferation parameter to change as a function of dose (model (1) in Table
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Table C-2. Net Parameter Estimates From the Two-Stage Model of Carcinogenesis
Two-stage parameters changed
Both
Both (proliferation rate <0.02)
Mutation rates
Proliferation rates
Doses
Control
1.37X10'22
0.0376
2.80X10-20
0.02
3.34x10-"
-0.0055
1 ng/kg/day
O.OTTxlO-22
0.0317
1.91xlO-M
0.0196
1.79x10-"
-0.0121
10 ng/kg/day
8.86X1022
0.0283
18.79xlO-M
0.0196
17.8x10-"
0.00362
100 ng/kg/day
3.43xlO-15
0.0040
3.58x10"
0.00198
76.0x10-"
0.00669
Other parameters
Proliferation rate = 6.90x10 3
Mutation rate = 7.73X10"1'
0
§
u>
o
VO
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C-l). For this model, as dose increases, the mutation parameter drops to one-half the
control value for the 1 ng/kg/day dose group and increases substantially in the remaining two
dose groups (6.5-fold for 10 ng/kg/day and 25xl06-fold for 100 ng/kg/day). The hepatocyte
replication parameter drops as a function of dose. In the 100 ng/kg/day dose group, the
replication parameter is small relative to control (tenfold smaller) to adjust for the much
larger mutation parameter. The reason for this particular pattern is clear if one studies the
tumor incidence data from the Kociba et al. (1978) study. Over time, the tumor incidence in
the highest dose group is larger than the others (the mutation parameter and the replication
parameter combine to control the magnitude of the response) but climb less steeply with time
(this is controlled by the replication parameter).
The replication rates in the control group and the two lowest dose groups are very
high. Assuming the death rate is zero, this would correspond to a labeling index of 40.9%
for a 7-day labeling experiment in the control animals. If the death rate is >0, the labeling
index would be even larger. Also, if one cell entered the initiated state on day 1, by 730
days (2 years), one would expect a clone of size SxlO11, ~ 140 times the size of a normal rat
liver. To control for this problem, the same model was fit to the data with the replication
parameter constrained to be less than 0.02 (a first-day clone would be expected to have a size
of 2xl06 by study end). This resulted in the same pattern of mutation parameters but with a
mutation parameter two orders of magnitude larger in the control and two lowest dose
groups. For all three groups, the replication parameter was estimated to be at the boundary
(0.02). The parameters for the high-dose group did not change.
The model in which the replication parameter is constant across dose groups and the
mutation parameter changes for each new dose is given in row 3 ("Mutation rates") of Table
C-2. The pattern in this case is very clear and matches the observed cumulative tumor
incidence well, with the mutation rate dropping in the 1 ng/kg/day dose group then rising in
the remaining groups. This observed drop is not statistically significant from the control
mutation parameter (likelihood ratio test, p>0.05). The constant replication rate (6.9xlO'3)
is reasonable and is not likely to produce unrealistically large clones of initiated cells. As
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mentioned before, this model provided a significantly worse fit to the data than did the
unrestricted model.
Finally, the estimated parameters for the model in which the replication parameter
changed with dose and the mutation parameter was constant over all doses are given by the
last row ("Proliferation Rates") in Table C-2. The change in replication rates mirrors the
tumor response in the same manner as was seen for the mutation rates. The replication rates
are of reasonable size and should not produce impossibly large clones. As mentioned earlier,
this model does not fit the data as well as the unrestricted model or as well as the model with
fixed replication rate over dose and varying mutation rate over dose.
There is other evidence that can be used to examine the adequacy of this model for
tumor incidence from exposure to TCDD. Lucier and colleagues recently conducted an
initiation/promotion study in female Sprague-Dawley rats (Tritscher et al., 1992; Sewall et
al., 1993; Maronpot et al., 1993). In this study, they measured number and size of
preneoplastic foci in liver sections. It has been suggested that the cells in these lesions
correspond to the initiated cells in the two-stage model of carcinogenesis. If this is true, it is
possible to apply the methods of Dewanji et al. (1989) to these data to analyze the growth
characteristics of these cells (Moolgavkar et al., 1990). However, as before, we run into the
problem of nonidentifiability. Since the data in these studies were collected at only one time
point, it may not be possible to estimate all of the parameters in the first half of the two-
stage model (i.e., /*N_r, ft, and 5j) and get a unique solution. For each dose group, one of
these parameters should come from outside information or continuous dose-response models
should be assumed.
Maronpot et al. (1993) measured the rate of cell proliferation in these rodents
(Tritscher et al., 1992) using immunocytochemical detection of cells that had undergone
replicative DNA synthesis. The average labeling index by dose group is given in Table C-3
for the uninitiated animals (saline controls) in this study. It is seen there is a slight drop in
mean labeling index (p>0.05) from control to the group given 3.5 ng/kg/day. The labeling
index then increases with increasing dose. Under the assumption of a linear birth death
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42
a
4>
a
-o
o
V
i
3
6
•s
H
S
-§
Para
•3
I
tiv
£
S!
ON
15
IS
s
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process, it is possible to convert these labeling indices into estimated birth rates for these
cells using the formula:
2t n [ 1 - LIJ
where 0 is the birth rate, t is the number of days over which labeling was done, and LI is
the labeling index (Moolgavkar and Luebeck, 1992). This conversion is given in row 2 of
Table C-3. These rates are in nonfocal hepatocytes (normal cells) and correspond to an
average of one to two births per year per hepatocyte. TCDD seems to double the control
birth rate for a dose of 125 ng/kg/day (a relative change of almost 100%-row 3 of Table
C-3).
Maronpot et al. (1993) also calculated the labeling index in focal hepatocytes in the
high-dose group from this study. The mean LI for animals in the initiated group exposed to
125 ng/kg/day of TCDD was 42.5%, which corresponds to a birth rate of 0.0395. The ratio
of birth rate in nonfocal cells to birth rate in focal cells is 0.0395/0.00526 = 7.5. Assuming
this ratio is constant over all dose groups, we can rescale the birth rates in the remaining
dose groups to correspond to birth rates for focal cells resulting in the rates shown as row 4
in Table C-3.
Using the same method as that used by Moolgavkar et al. (1990), given these birth
rates, it is possible to estimate /XN.J and 5r for the focal lesion data of Lucier et al. (1991).
The resulting model did not fit these focal lesion data.
To get a handle on how large the labeling index would have to be to result in a
reasonably good fit to the data, the model was again fit to the liver foci data, allowing ft to
roam freely and the ratio Vft to be <0. The resulting parameter estimates are given in
Table C-4.
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Table C-4. Estimated Model Parameters in the Two-Stage Mathematical Model of Carcinogenesis (Figure 1) for the Noninitiated
(Saline)
Dose1
XoM,"
(mutation rate)
A. Fully nonparametric in dose
Control (0.0)
3.5
10.7
35.7
125.0
3.851 10"
4.346 10"
3.860 10-*
9.964 10"
7.0435 10"
ff
(birth rate)
f
(death/birth)
Log-likelihood
3.041 102
3.548 10*
3.749 10-2
3.576 10^
3.664 10-2
6.077 10-2
6.062 10-2
1.267 10 '
3.354 lO'2
6.210 10-2
282.79
B. Constant mutation rate, other parameters nonparametric in dose (pure promoter)
Control (0.0)
3.5
10.7
35.7
125.0
8.423 10"
6.778 10-2
6.045 ID"2
9.510 10^
3.509 104
3.372 10^
6.385 10"'
4.998 101
6.988 10-'
0.000 10-'
o.ooo 10-'
281.62(p=0.673)
C. Constant ratio (p) and birth rate (ff), nonparametric for Xgft, (pure initiator)
Control (0.0)
3.5
10.7
35.7
125.0
0.241 10-3
0.415 lO'3
0.339 10-3
1.066 10-3
0.731 10J
3.578 10-2
5.386 10 2
280.78 (p=0.855)
D. Constant ratio (p) and hyperbolic' birth rate (ff), XD/J,
Intercept
v™
kd
3.070 10"
5.909 10"
18.17
3.003 10-2
4.948 103
2.20
3.825 10-3
276.73 (p=0.146)
"ng/kg body weight/day.
'Mutations per day per mm3 of liver.
'Births per cell per day.
•"Deaths per birth.
"Parametric form is rate (dose) = intercept + £„,„ dose/(D0 5+dose).
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There is considerable instability in the estimated model parameters for the
nonparametric model (A) in Table C-4, possibly due to the fact that there may not be
sufficient data in each dose group to allow for the estimation of individual parameters. This
is especially true for the low-dose groups in which there are few focal lesions and they are
generally of small size. However, the model does fit the data rather well.
The first task is to test if the effect of dose, within the context of this model, is
significant or not. This can be restated as the hypothesis that one set of model parameters
will fit all five treatment groups. This is accomplished by fitting the same two-stage model
to all five data sets, treating them as one large experimental group (this model is not shown
in the table). Using a likelihood ratio test, there is a significant change in the two-stage
model parameters with p< 0.001. The next step is to attempt to simplify the final model to
include fewer parameters and provide a continuous description of the effect of dose on the
various model parameters.
Based on the existing understanding of how TCDD induces PGST+ foci in this
biological two-stage model, it would be natural to assume that Xl^, the mutation rate from
normal cells to initiated cells, does not change as a function of dose. Model B in Table C-4
addresses this basic model. In fitting a new model to these data, the first task is to assess,
using a proper statistical test, if this model fits the data as well as the original model. Using
the likelihoods, it is possible to test this hypothesis using a likelihood ratio test. For these
data, this yields p=0.673, indicating that a model with a constant mutation rate fits the data
as well as the fully nonparametric model and suggesting that TCDD at most causes a
marginal increase in either XQ, nlt or both. Maronpot et al. (1992) noted some evidence of
hepatotoxicity via histopathological changes, small increases in relative liver weight and
alterations in serum chemistry. The increases in relative liver weight would suggest the
effect may be on XQ, but the hepatocytes were also enlarged, making this determination
difficult. As discussed in Kohn et al. (1993) and Portier et al. (1993), it is possible for
TCDD to have a small effect on mutations in these animals via a secondary pathway. This
finding is consistent with that conjecture.
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Over the entire dose range, the ratio of death rate to birth rate follows an erratic
pattern that varies over several orders of magnitude. Sensitivity analysis shows that the
likelihood is insensitive to rather large changes in the ratio, suggesting it may be possible to
get an adequate fit to these data with a constant value for p across all dose groups. Analyses
done by Moolgavkar and Luebeck (1992) and Luebeck et al. (1991) relied on this
assumption. A model with p constant across all dose groups indicates that there is no effect
of TCDD on the ratio of the death rate to the birth rate of PGST+ foci in uninitiated
animals.
Model C, in contrast to model B, tests the hypothesis that there is no effect of TCDD
on growth rates of initiated cells. The results are not significant, indicating a pure initiation
model fits these data as well as a pure promotion model. Model C also provides a much
clearer pattern to the effect of dose on the mutation rate and birth rate for the two-stage
model. The mutation rate (X^) basically increases until the highest dose group where there
is a slight decrease. This decrease may be due to a single animal in the 35.7 ng/kg/day dose
group. This would suggest that we might model the mutation rate (and the birth rate) by the
hyperbolic function; for the mutation rate, this would be given as
= intercept + Vmaxdose/(kd+ dose) [1]
A similar model is used for the birth rate. The parameter estimates for this model are
presented in Table C-4 as model D. A likelihood ratio test indicates that this model provides
an adequate fit to these data (p =0.146).
Several other models were fit to these data. These models were various combinations
of constant, nonparametric, linear, and hyperbolic dose-response functions for the three rate
functions (X^, fi, and p) for the two-stage model. The model given in Table C-4 as model
D seems to be the best parameterization that agrees with the data and the biological
understanding of the carcinogenicity of TCDD.
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The DEN-initiated animals must be handled differently from the noninitiated group.
It is believed that initiation with DEN results in a number of the normal cells undergoing
mutation. This process is assumed to occur in a very short period. To accommodate
initiation, Luebeck et al. (1991) introduced a second mutation rate process at the start of the
experiment. These data were fit using the methods of Luebeck et al (1991). The resulting
model agrees very closely with the model for the saline-treated animals, the main difference
being a larger mutation rate.
The calculations have hidden assumptions that could have a bearing on the
interpretation of the results. First, we have ignored the changes in labeling index in the
normal hepatocytes. This could change the size of the pool of normal hepatocytes, which is
assumed constant for these calculations. This might alter the parameter estimates. In
addition, these changes in the rate of replication in nonfocal tissue could alter the mutation
rate /tN_r (Portier and Kopp-Schneider, 1991). As for tumor incidence, the model assumes
constant rates and independent cell kinetics. Finally, the large estimated percentage of cells
that are initiated can only imply most of these initiated cells are single cells and cannot be
observed on histopathological examination. This has not been verified. The analysis is very
sensitive to the choice of values for the radius of a cell (Moolgavkar et al., 1990) and to the
minimum size of a detectable focus. These issues cannot be discussed fully in this brief
document.
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APPENDIX D
CONSIDERATIONS FOR USING DOSE SURROGATES
IN ESTIMATING TUMOR INCIDENCE
It is possible to compare the parameters estimated for the two-stage model to
predictions from the PB-PK models to try to locate a reasonable mechanistic link between the
two classes of models, to aid in species extrapolation, and to help guide us in choosing the
most appropriate curvature for low-dose extrapolation. The two-stage model parameters that
were estimated from the tumor incidence data are compared to 14 dose surrogates from the
three main PB-PK models reviewed earlier. Leung et al. (1990a) suggested using occupancy
of the cytosolic (Ah) receptor, averaged over the length of the study. They also suggested
using average binding to the microsomal proteins. These predictions are shown in Table D-l
for the study of Kociba et al. (1978), along with the predicted final concentration of TCDD
in the liver of these animals. The correlation coefficients of these parameters with the two-
stage model parameters are shown in the last three columns of Table D-l under "Both,"
"Mut. Rates," and "Prol. Rates." It is clear that these three surrogates correlate well with
all but the two-stage model in which the effects of TCDD were treated as pure promotional
effects. The high correlations and low p-values for these correlations je uiiven by the order
of magnitude differences in the parameters and the dose surrogates. (Caution should be used
in judging these correlations to carry any weight of scientific evidence.)
Andersen et al. (1993b) suggested four dose surrogates to be used in any risk
assessment for TCDD. These are based on free TCDD in the liver, total TCDD in the liver,
amount of induced CYPIA1, and amount of induced CYPIA2. As for Leung et al. (1990a),
these are integrated over the lifespan of the animal (in this case, total integrated rather than
average, but this will not affect the correlations). As for Leung et al. (1990a), the
correlation with all two-stage model formulations/parameters is significantly different from
zero for three of the surrogate doses (excluding CYPIA2) and all but the promotion-only
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Table D-l. A Comparison of Dose Surrogates With Parameters Estimated From the
Two-Stage Model of Carcmogenesis
Dose surrogate
Leung et al. (1990s)
Concentration
Ali-receptor occupancy
Microsomal binding
Andersen et al. (1993b)
Free TCDD in liver
Total TCDD in liver
Induced CYPIA1
Induced CYPIA2
Dote
Control
0.0
0
25
0
0
0
0
Kohn et al. (1993)
Free TCDD in liver
Ah-receptor/TCDD
complex
CYPIA2/TCDD complex
TGF-«
Internalized EGF receptor
Total CYPIA1
Total CYPIA2
0
0
0
0
0
190
4,517
1 ng/kg/day
0.3
2
29
6.4
590
3,900
57,000
0.1237
1.0202
1.6857
0.005899
0.4829
2,198
5,999
10 ng/kg/day
100 ng/kg/day
4.4
17
132
53
11,000
270,000
210,000
69.4
61
132
420
120,000
200,000
320,000
Correlation*
Both
1.00++
-1.00++
0.96+
-0.96+
0.96+
-0.96+
0.99++
-0.99++
1.00++
-1.00++
0.99++
-0.99+
0.79
-0.79
0.7663
8.8617
27.3142
0.0540
8
3.7213
14,213
15,747
5.0522
49.352
464.85
0.3788
13.668
42,196
44,579
0.99+
-0.99*
0.99+
-0.98+
1.00++
-1.00++
0.99++
-0.98+
0.97+
-0.96
0.95
-0.95
0.96+
-0.96+
Mut. Rates
0.99+
1.00++
1.00++
1.00++
1.00++
1.00++
0.89
Prol.
Rates
0.70
0.81
0.81
0.73
0.72
0.74
0.87
1.00++
1.00++
0.99+
1.00++
1.00++
0.99++
1.00++
0.74
0.76
0.69
0.74
0.83
0.83
0.81
•Pearson correlation coefficient -+ indicates statistical significance at the 0.05 level, + + indicates significance at the 0.01 level (caution should be
used in applying the p values too literally due to the small sample sizes involved).
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TCDD model. The correlation of induced CYPIA2 with the two-stage model parameters is
not significantly different from zero for all two-stage models. Note that the highly nonlinear
CYPIA1 curve correlates well with the first two two-stage models. If this were used as a
dose surrogate for TCDD toxicity (in the independent framework chosen by Andersen et al.
[1993b]), the resulting low-dose risks would be substantially smaller than those of any other
dose surrogate. Induced CYPIA2 fails to correlate due to the strong saturation seen at the
two highest doses; an effect not observed in the two-stage model fits.
Kohn et al. (1993; Appendix A) did not suggest specific dose surrogates but did
suggest mechanisms that naturally lead to the choice of certain surrogates. Seven such
surrogates are given in Table D-l: free liver TCDD, Ah-receptor bound TCDD, CYPIA2-
bound TCDD, integrated TGF-a, internalized EOF receptor, integrated CYPIA1, and
integrated CYPIA2. With the exception of integrated CYPIA1, all dose surrogates correlated
as well with the two-stage parameters as the dose surrogates from the other two PB-PK
models. None of the surrogates resulted in correlations with the promotion model, which
were significantly >0. The internalized EGF receptor correlates well with the promotion
model but not significantly better than 0 and not better than induced CYPIA2 in the Andersen
et al. (1993b; Appendix B) model. Total CYPIA2, which Kohn et al. (1993) suggest could
be tied to secondary mutagenic effects of TCDD, correlates well with the mutation model.
All of the dose surrogates predicted by the Kohn et al. (1993) model have positive slope at
dose 0 and would behave, in the low-dose region, as a linear model. Risk estimates utilizing
these dose surrogates and the two-stage model would likely result in a less than tenfold
change in risk over what would be estimated by applying a one-stage model to the Kociba et
al. (1978) data.
It should be possible to correlate these same dose surrogates with the model
parameters arising from the liver foci analysis. However, none of the authors calculated
dose surrogates for this design situation. This can be done at a later time.
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One final point on dose surrogates: The choice of which measure of exposure one
decides to correlate with which measure of effect is a somewhat arbitrary decision. The
choices above were mostly chosen because they have traditionally been used in this context.
However, there is no pressing mechanistic reason why these particular choices are best. The
major two variables in this decision are which end points(s) and how to consider time
factors. For example, we could also have considered induced amount of Ah receptor as a
dose surrogate; there is no a priori reason to exclude it. It is also possible to integrate over
shorter periods of time. This is not really mechanistically justifiable for TCDD, but for
other compounds, which show a short mitogenic effect or get rapidly metabolized into toxic
compounds of brief duration, shorter integration periods would be more appropriate. Thus,
some thought should be given to a choice of dose surrogate based on mechanistic
considerations. The paper by Kohn et al. (1993) discussing their mechanistic model can be
used to provide considerable direction on this topic (see Appendix A).
D-4 06/30/94
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APPENDIX E
TETRACHLORODIBENZO-P-DIOXIN
EMPIRICAL RELATIONSHIPS FOR NON-CANCER ENDPOINTS
Lynne F. McGrath, K.R.Cooper, P.Georgopoulos, and M.A. Gallo
Environmental and Occupational Health Sciences Institute
UMDNJ-Robert Wood Johnson Medical School, Piscataway, NJ
L INTRODUCTION
Dose-response assessment is an essential element in determining health risks associated with
environmental contaminants. The evaluation of dose-response requires information on a range of
quantifiable doses or exposures and corresponding measured responses. The study of adverse
effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and the use of dioxins as tools to elucidate
underlying biochemical and physiological mechanisms of cellular action are pivotal in the overall
evaluation of dioxin exposure. The determination of the shape, slope and of the dose-response
functions of TCDD for several endpoints, over multiple time points and in multiple species, is
critical to the understanding of the potential human health risks and exposure to dioxin.
Dose-response relationships for several endpoints can be established using human data
(Poole, this document) however quantification of exposure can be difficult. The assessment of
TCDD presented here primarily employs controlled animal studies in which administered dose is
known and delivered dose is either measured or can be estimated. The endpoints described were
those found to have some consistency across species and strains. This extensive review of dose-
response relationships has been completed to better elucidate key biomarkers or surrogates for
cancer and other life threatening diseases.
TCDD and members of this family of compounds are exogenous ligands for the Ah
receptor(AhR). This receptor had originally been hypothesized to be a member of the nuclear
hormone receptor family. The overall hypothesis of TCDD action, put forth by several groups, (see
DRAFT September 15, 1993 DO NOT CITE OR QUOTE
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Whitlock chapter this document) can be simplified as follows: TCDD and the other members of
the family (ligands) enter the cell by mass action, bind to the AhR (receptor), the receptor-ligand
complex translocates to the nucleus via specific transport proteins and subsequently binds to
specific sequences on the gene. This in turn evokes the production (or perhaps the suppression) of
several mRNA species. Recent results have identified a specific consensus sequence on the gene
which has been named the xenobiotic response element (XRE)(Whitlock,1991, Denison, 1991).
In addition to the DNA sequence identification, the structure and amino acid sequence of the AhR
protein was reported by Burbach et al., (1992). Both the XRE(s) and the structure of the AhR are
analogous to the steroid receptors and their respective genomic response elements. This similarity
is important in regard to biological models of TCDD action and risk assessment.
Dose-response relationships have been established for several endpoints in intact and
surgically altered animals. In vitro experiments have been used to determine critical concentration
and structural relationships for TCDD effects at the cellular and molecular level. In the vast
majority of these studies the role of the AhR-ligand interaction has been essential but not
necessarily sufficient to evoke a detectable biological response.
Although all toxic endpoints for TCDD may not be Ah locus mediated the evaluation of
dose response allows us to conclude that the critically sensitive events are Ah receptor mediated.
The multiple steps in receptor mediated processes suggest that the dose-response relationships are
not linear. This series of cellular events can be best modeled as a simple series of log-differential
equations which basically take the form of Michaelis-Menten hypotheses, which include Michaelis-
Menten and Hill functions. This format allows for slope estimates, as well as inhibition and
saturation phenomena. Hence, if curvilinearity of the dose-response exists, it should be possible
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to model using kinetic equations. Indeed, this may well be the case for some end effects of
TCDD, however, recent rodent experiments suggest the induction of cytochrome P-450IA1
(CypIAl) is linear through zero (curve fitting with a Hill function) using long term exposure and
very low doses (Lucier et al., 1991a).
Low dose phenomenon like thymic atrophy and immunotoxicity have also been described
and are the result of multistep processes in which many of the steps are unknown, and others are
non-specific. Non-cancer endpoints have generally not been subject to dose-response modeling
such as that used for assessment of cancer endpoints although many have been consistently reported
after TCDD administration. The focus of this dose-response assessment is on toxic endpoints for
which a mechanism has been hypothesized and on the description of low dose effects. In addition,
there is the practical criteria of using a satisfactory number of dose points.
To improve our understanding of non-cancer endpoints and to elucidate possible mechanistic
relationships between endpoints, several simple functional forms were used in the empirical
analysis of relationships between TCDD dose and non-cancer endpoints. Throughout the document
datasets have been selected for a more rigorous dose-response assessment. The criteria used for
inclusion is endpoint and dataset specific, but generally it depends on the number and range of dose
points. Specific rational for selection is described with the specific endpoint. Because of the
exploratory nature of this analysis the number of assumptions regarding dose-response relationships
were kept to a minimum. Each functional form used will be briefly discussed with comments on
the nature of these assumptions. For a more extensive mathematically detailed explanation of these
functions see Appendix.
Initially a x-y plot of the data was made to provide a visual examination of dose-response.
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Attempts to use the Linear Multistage Model for fitting the dose response did not give satisfactory
correlation coefficients, in addition the underlying biological assumptions for this model would not
be appropriately applied to the non-cancer endpoints described in this document. Although these
graphs are not included in this document they were used as a first step in the analysis. The first
and simplest model used to describe selected data sets was the linear function.
The Linear Model
y = mx + b
(x = dose; y = response; m = slope; b = Background response)
Many dose-response relationships in pharmacology and toxicology are known to fit the
linear model. For example, the elimination of some chemicals (ie., chloroform) from plasma obeys
1st order kinetics and as such is modeled as a linear function. In many cases the actual dose-
response relationships are nonlinear but appropriate transformations of the independent and
dependent variables are used to obtain a linear relationship. For example, Scatchard and
Lineweaver-Burke plots represents ligand-receptor binding using equations in a linear form
however, the binding phenomenon is non-linear (Tallarida and Jacob, 1979). The success of the
linear fit is typically evaluated via least squares criteria resulting in an R2 value which quantifies
the "goodness of fit". However, in the case of a nonlinear phenomenon that has been transformed
into a linear relationship, the transformation introduces a bias in the R2 estimate. (These R2 values
are not directly compared to those derived form linear functions).
The second functional form used to evaluate the TCDD dose-response relationships was the
Michaelis-Menten equation. This is a hyperbolic function often used to fit the relationship between
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enzyme-substrate or receptor ligand interactions.
The Michaelis-Menten Model
y = ax / (b+x)
(x = dose; y = response; a = maximum response; b = concentration of x at 1/2 maximum response)
The assumption inherent when plotting the ligand-receptor binding relationship is that there is a
single rate function that represents the relationship between the free receptor and the bound receptor
over specified doses, and that a finite number of binding sites must be occupied to initiate a
response and that saturation of the receptors occurs, and that this maximum interaction is reached
at some finite dose (Goldstein et al., 1974).
The third function used was the Power Law formalism. The relevant equation is nonlinear
and includes a power (not necessarily an integer) as a term in the function.
Power Law Function
y = axs
(x = dose; y = response; g = kinetic order; a = limiting response)
This functional form has been historically used to represent a variety of nonlinear biological
phenomenon including the saturation of, or Synergy in, enzyme systems; thus the name "S-
Systems" (Savageau, 1991). This formalism has been used to describe processes obeying power
laws which provides adequate descriptions of many nonlinear interactions that occur in nature. It
is not a surrogate for Michaelis-Menten function, but provides an approximation of individual
biological interactions or associations. The exponent in the Power Law function ("g") is related
to the pharmacological kinetic order (rate of response) of the observed process. If the "g" value is
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greater that 1 the response is superlinear and positive cooperatively is assumed, whereas is the
response is less that 1 and sublinear, then negative cooperativity is assumed. The Power Law has
fewer degrees of freedom than the Hill function (see below) and it can fit subsets of doses allowing
for different reaction rates to be determined. Although it has not been specifically applied to
toxicological results for TCDD, the generalized approach of the Power Law formalism is
compatible with the pleiotropic responses seen after TCDD administration (Savageau, 1991). The
Hill and Michaelis-Menten functions describe a response over an entire dose range. However, the
Power Law, when applied to subsets of a data, may provide a mechanism for describing changes
in underlying biology from low to high doses.
The fourth function used in this analysis was the Hill function. This nonlinear function is
a generalization of the Michaelis-Menten equation as it incorporates in it the flexibility of the
Power Law approach.
The Hill Function
y = (ax8) / (b + x8)
(x = dose;y = response;g = kinetic order;b = concentration of x at 1/2 maximum response;a = maximum response)
It has been used to model receptor-ligand interactions, and because of its wide acceptance and
flexibility an attempt to fit this formalism was made for selected datasets (Boeynaems and Dumont,
1975). As in the Power Law formalism, the value of the "g" exponent in the Hill plot determines
the behavior of dose response; if greater than 1 the response is superlinear, if the exponent is less
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than 1 it is sublinear1. The Hill plot is not as simple as the Power Law function because of the
greater number of variables. The Hill plot assumes saturation at the high doses, as well as an
"inertial"2 type of response at low doses. If the underlying biology of a system indicates that
saturation occurs at high doses, then the Power Law function, because it is exponential, does not
accurately describe the high dose phenomenon.
In the analysis of empirical relationships selected data sets were tested using various
functional fits. The Michaelis-Menten and Hill functions were adopted for use because it is the
most widely used model that assumed receptor-ligand or substrate-enzyme interactions. The
nonlinear model used to fit all data sets was the Power Law function because of its simplicity.
Initially these functions were used to fit the percent-response form of data which was used as a
relative normalization procedure. Subsequently, these functions were used to fit raw data. The
Power Law function, because it was not constrained by underlying assumptions, was then used to
fit ranges of doses to look for changes in the exponent possibly detecting rate changes. Analysis
of the relationships for specific endpoints are discussed in detail throughout this chapter with the
descriptions of the specific endpoints.
Because many of the data sets have few values much of the available statistical analysis
(mean square error of fit, analysis of co-linearity) were not performed. Hypothesis testing to
evaluate curve fits was also limited by the number of data points. These data could be made
1 A superlinear dose-response could generally suggest positive cooperativity or synergy through
the measured dose-response range, whereas, a sublinear dose response could imply negative
cooperativity or an inhibitory effect.
2 The definition of inertia is " a property of matter by which it remains at rest or in some uniform
motion in the same direction unless acted upon by a net external force" Merkin, 1985
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more robust by the analysis of raw data. Since rigorous statistical analysis was not applied to the
data three criteria were used to conclude a "poor fit" existed. First a low R2 or a high chi square
value, second, visual observation of the curve fit, and third the inability to fit the data to a
particular function.
Table I: Summary of Dose Response Functions
LINEAR MODEL
Assumes biological response is linear.
R2 > 0.8 typically assumes good
fit
MCHAELIS-MENTEN
FUNCTION
1. Assumes a non-linear biological phenomenon.
2. Assumes minimal or no response at low doses.
3. Assumes maximum activity is reached in dose
range.
Chi Square "goodness of fit"
dependent on degrees of freedom
HILL FUNCTION
1. Assumes a non-linear biological phenomenon.
2. Assumes low dose response is negligible or
"inertial".
3. Assumes saturation or maximal activity is reached
at high doses.
Chi Square "goodness of fit"
dependent on degrees of freedom
POWER LAW
FUNCTION
1. Assumes a non-linear biological phenomenon.
2. Assumes synergy or cooperativity.
R2 > 0.8 typically suggests a good
fit; however it should be taken
into account that the R2 estimate
is logarithmically biased.
In assessing potential health risks, an analysis of dose-response provides an estimate of the
severity of the effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin. The focus of this dose-response
assessment is on studies using multiple doses and on evaluating multiple endpoints in several
species. The hope, in evaluating responses in a variety of species, is that consistency will add
confidence to the estimates of human health risks subsequently made. This report emphasizes
revaluation of effects that occur at low doses; these being the most relevant to environmental
health. Responses at higher doses are also included to demonstrate species differences and
estimations of what might occur under incidental high exposures (such as industrial accidents).
Experimental evidence indicates that TCDD is extremely toxic in some systems at low doses.
However, although qualitatively similar, the magnitude of response is species and strain specific.
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IL ACUTE TOXICTTY
A large body of data exists describing the acute toxicity of TCDD in a variety of species.
Acute data on lethality is briefly described here and will be referred to throughout the document.
Acute responses are generally considered Ah receptor mediated segregating with the Ah locus. In
this analysis of non-cancer endpoints this acute toxicity information is used as a basis of comparing
tissue, species and strain sensitivity. Sensitivities that differ from those listed in Table n will be
probed for endpoint specific factors that may modify the response.
The acute responses listed in Table II are LDJO's which vary almost 10,000 fold. The
accumulated levels of TCDD causing lethality after continuous administration are consistent with
the levels causing effects after single exposure which strongly suggests that over short periods of
time the total body burden is the rate limiting factor in morbidity. It is generally recognized (Poland
and Knutson, 1982) that the guinea pig is most sensitive to the lethal effects of TCDD followed
by the rat (sensitive strain), the mouse (sensitive strains), rat and mice (non-sensitive strains) and
by the hamster. However, the rhesus monkey (see below) may be more sensitive than the guinea
pig. Because of the variety of protocols used to assess acute toxicity, several studies are reviewed
here. In a continuous feeding study, rhesus monkeys were administered 500 ppt TCDD for 9
months (Allen et. al, 1977). Four months after this treatment period 4 of the 8 animals died after
receiving an estimated total dose of 18.6 ug/animal (approximately 3.5 ug TCDD/kg or 0.013 ug
TCDD/kg/day). Based on these data the LD50 may be considerably less than the 70 ug TCDD/kg
(lowest dose tested) reported by McConnell et al., (1978) (Table II). In a continuous feeding study
in Hartley guinea pigs, 5 of 10 males died after a total dose administration of 0.6 ug TCDD/kg to
1.17 ug TCDD/kg. Four of 10 females died after a total dose administration
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Table II
Comparative Acute Toxicity of TCDD
IJ>50 Values
Species
Hartley guinea pig
Hartley guinea pig
Hartley guinea pig
Golden Syrian Hamster
Golden syrian Hamster
Golden Syrian Hamster
Sherman (Spartan) rat
Sherman (Spartan) rat
Han/Wistar rat
Long Evans
Fischer rat Charles River
Fischer rat Harlan
Fischer rat Fredrick Cancer
Center
CD rat Charles River
New Zealand albino rabbit
New Zealand albino rabbit
Rhesus monkey
Rhesus monkey
DBA/2J mice
B6D2,/J mice
C57BL/6J mice
C57BL/6J (Ahb'b)
C57BL/6J (Ah""1)
C57BL mice
Sex
Male
female
male
male
male
male
female
male
male
male and
female
male and
female
female
female
male
male
male
male
male
male
oral
oral
oral
i.p.
oral
oral
oral
oral
i.p.
oral
oral
dermal
oral
oral
oral
oral
oral
oral
oral
oral
LD50 ug
TCDD/kg
0.6
2.5
2.1
>3000
1157
5051
22
45
>3000
10
164
340
303
297
115
275
<70 ug
TCDD/kg
3.5 ug TCDD/kg
0.013 ug
TCDD/kg/day
2570
296
182
159
3351
114
Time to Death
5-34 days
42 days
19-42 days
50 days
50 days
26-43 days
9-27 days
13-43 days
24.8 ± 0.6
28.3 ± 0.5
25.9 ± 0.8
24.5 ± 1.0
6-39 days
12-22 days
approx 4 weeks
28-44 weeks
30 days
30 days
30 days
22 days
22 days
15-24 days
Reference
Schwetz,1973
Silkworth et al, 1982
Schwetz et al., 1973
Olson, et al., 1980
Olson et al., 1980
Henck et al., 1981
Schwetz et al., 1973
Schwetz et al., 1973
Pohjanvirta et al., 1990
Walden, 1985
Schwetz et al., 1973
Schwetz et al., 1973
McConnell, et al., 1978
Allen, et al., 1977
Chapman and Schiller,
1985
Chapman and Schiller,
1985
Chapman and Schiller,
1985
Birnbaum, et al., 1990
Bimbaum, et al., 1990
Vos et al., 1973
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of 1.3 to 1.8 ug TCDD/kg (DeCaprio et al., 1986). Although this phenomenon has not been fully
evaluated, female animals in general appear to be less sensitive to the acute toxicity of TCDD than
do the males. This is exemplified by female guinea pigs and Sherman rats which were found to be
less sensitive to the lethal effects of TCDD than the male animals (Table II). The magnitude of
the response may vary across or within species, but there are a number of consistent lesions. At
the measured LD50, in all animals examined, there is a species specific delay in onset of the
appearance of toxicity, generally about three weeks. Animals begin to lose weight immediately and
continue to lose up to 50% of their body weight, depleting their adipose tissue until their death.
This is characterized as the wasting syndrome. Body weight loss was also observed after a single
TCDD treatment in male Fischer F-344 rats (po 100 ug TCDD/kg), male albino guinea pigs (ip 2
ug TCDD/kg) and male C57BL/6 mice (po 360 ug TCDD/kg). Based on observations of food
consumption of these animals and pair fed animals, Kelling et al. (1985) concluded that
hypophagia appeared to cause some loss of adipose tissue accounting for weight loss. The precise
mechanism underlying the wasting syndrome is not known however, body weight loss appears to
contribute more to lethality in the some species and strains (ie., guinea pig and the Sprague Dawley
rats than in others (ie., Fischer rats and C57B1/6 mice) (Kelling et al., 1985; Peterson, et al.,1984).
BO. SYSTEMIC EFFECTS
A. INTRODUCTION
Administration of TCDD has consistently been shown to cause a variety of toxic effects.
Most pathological changes were found to affect epithelial tissue; however, specific dermal lesions
are observed in some species. In man, exposure to TCDD results in chloracne, a skin lesion, which
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is characterized as "follicular hyperkeratosis in the absence of inflammatory comedones"3 hi
monkeys there is hair loss along with thickening and keratinization of glandular tissue. Oral
administration in monkeys, rabbits, and hairless mice results in dermal lesions. Gastrointestinal
lesions characterized as hyperplasia of the gastric mucous or intestinal epithelium have been found
in monkeys, cows and hamsters but not in rats, and guinea pigs. Urinary tract hyperplasia has been
reported in the guinea pig, monkey and cow. Liver toxicity, immunotoxicity, and thymic atrophy
occur in several species at very low doses. In addition, tissue specific fetal development is
affected by TCDD in several species (Poland and Knutson, 1982). These effects occur at low doses
and some dose-response relationships have been established for a few of these endpoints. The
comparison of dose response of several endpoints is the focus of this report, hi all mammalian
species tested thymic atrophy is one of the most sensitive clinical indicators of toxicity.
B. THYMIC ATROPHY
Guinea pigs are most sensitive to the acute lethal effects and to may of the systemic effects
of TCDD. Harris et al. (1973) administered eight weekly doses of 0.0, 0.008, 0.04, or 0.2 ug
TCDD/kg to female guinea pigs which resulted in a statistically significant4 decrease in thymus
weight in the 0.04 ug TCDD/kg dose group (LOAEL = 0.0057 ug TCDD/kg/day equivalent dose,
NOAEL = 0.0011 ug TCDD/kg/day). The Michaelis-Menten functional fit of this data (Graph
A I)5 is poor primarily because of the minimal number of data points. Additional dose-response
3Zugerman, C. (1990) Chloracne Clinical Manifestation and Etiology()ermatologic Clinics 8(1),
209.
4A11 statistical significance described in this report was determined by the study investigators.
5 All graphs discussed in text are located at the end of the document.
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analysis was not performed on this data set. In A 90 day study, Decaprio et al. (1986) administered
2, 10, 76, and 430 ppt TCDD in the feed. Daily intake was calculated as 0.00012, 0.00061, 0.0049,
and 0.026 ug TCDD/kg/day in males and 0.00012, 0.00068, 0.0049, and 0.031 ug TCDD/kg/day
in females. In males, the more sensitive species in this study, the absolute and relative thymus
weight was significantly different from controls at the 0.0049 ug TCDD/kg/day (See Table HI).
This level is comparable to that seen in the Harris et al. (1973) study.
In female rats, 31 daily doses of 0.0, 0.1, 1.0, and 10 ug TCDD/kg resulted in a decrease
thymus weight after 24 days in the 0.1 ug TCDD/kg/day dose group (cumulative dose 0.24
ug/kg)(Harris, et al., 1973). In a 13 week study Kociba et. al., (1976) demonstrated a dose related
decrease in thymus weight in male and female Sprague-Dawley (S-D) rats administered 0.1 ug
TCDD/kg/day orally 5 days/week. Graphic depiction of this percent decrease in thymic weight in
male and females (Graph A2) demonstrates a "good fit" to the Michaelis-Menten type response.
The female Sprague Dawley rats appeared slightly more sensitive to this response, although the
doses where a decrease in thymic weight was observed was the same. Microscopic examination
revealed complete involution of the cortical region of the thymus of rats administered 1.0
ug/kg/day and slight involution in those administered 0.1 ug TCDD/kg/day. The statistically
significant decrease in thymus weight appeared to be a sensitive indicator of microscopic changes
in the thymus. A dose-dependent decrease in absolute thymus weight was reported in Rhesus
monkeys administered acute oral doses of 70 ug TCDD/kg (lowest dose administered) or greater
(McConnell, 1978). However, this dose may not accurately reflect the sensitivity of this endpoint
in the monkey, but does provide information on the consistency of this response.
Thymic atrophy has been reported in rats, guinea pigs, monkeys, hamsters and there are
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suggestions of thymic effects in humans. Effects on thymic hormones have been observed in
humans exposed (not quantified) to TCDD. A statistically significant lower mean serum thymosin
aipha-iwas observed in a study that compared 94 people allegedly exposed to TCDD in contaminated
residential areas to a nominally unexposed control population. The study found thymosin levels
were inversely associated with years of living in the contaminated area (up to 7 years) when
controlling for age and socioeconomic factors. Thymosin is produced in the epithelioid cells of the
thymus and functions to modulate the maturation of prothrombocytes to mature thrombocytes.
Decreased serum thy mo sin avlM-i suggests an effect on the thymus in this human population exposed
to environmental levels of TCDD. Unfortunately concentrations of TCDD in these areas not serum
levels (a more accurate estimate of TCDD exposure) were reported (Stehr-Green et al., 1989).
In six studies a range of single doses of TCDD were administered orally to C57BL/6J mice.
Four of these studies (McConnell et al., 1978a, Bimbaum, et al., 1990, Chapman and Schiller,
1985, and Bombick et al., 1988) tested doses that were greater than 30 ug TCDD/kg and found a
statistically significant decrease in thymic weight or thymic atrophy at all doses tested. Vos, et al.,
(1974) administered single oral doses of 0, 0.2, 1.0, 5.0, and 25 ug TCDD/kg/day. At 1.0 there was
a statistically significant decrease in thymus/bwt, with a NOAEL at 0.2 ug TCDD/kg. Kerkvliet
and Brauner (1990) tested similar doses; 0, 0.2, 1, 2.0, and 5.0 ug TCDD/kg, in female C57BL/6J
mice reporting a dose-dependent decrease in relative thymic weight. Graphic comparison of these
two studies (Graph Bl) demonstrates similar responses in the 1 to 10 ug/kg dose range, although
the male mice of a different strain appear more sensitive. Since the linear range of the dose
response curve lies between 0.2-10 ug TCDD/kg a linear model was also fit to this range therefore
data sets could be directly compared. [Graph B2 demonstrates that the slope of both these dose-
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response curves are similar (Vos, male mice = 0.57 and Kerkvliet, female mice = 0.46)]. It
becomes apparent when comparing this Graph B2-(A) to Graph B2-(B) that the range of doses used
for comparison is important. Although many studies describe thymic atrophy, few have enough
data to plot comparative dose-response curves, therefore removing data points for this type of
comparison will generally be problematic.
Poland and Glover (1980) found a dose-dependent decrease (graphically depicted) in thymic
atrophy in C57BL/6J and DBA mice after a single i.p. injection. The ED50 was 10'8 mol/kg(3.22
ug TCDD/kg) in the C57BL/6J mice and the DBA/2J mice were 10 X less sensitive to this effect.
The ED30 for induction of the Aryl Hydrocarbon Hydroxylase (AHH) in these animals was one-
tenth the ED30 for thymic atrophy in both DBA and C57BL/6J mice.
In a review of the dose-response literature it is critical to note that the intra- and inter-
species sensitivities are linked to the Ah locus (Whitlock, 1987). Different responses for TCDD
in a variety of strains of mice were investigated to determine if the toxicity segregated with the Ah
responsive mice. Thymic atrophy was observed after a single i.p. injection of 20 ug TCDD/kg
TCDD was administered to sensitive species C57BL, C3H, and BALE mice. The same dose
produced no thymic atrophy in ODD, AKR, and DBA mice (less sensitive species). These effects
corresponded well with the induction of aryl hydrocarbon hydroxylase (AHH) activity in the livers
of these strains of mice. The toxicity of TCDD to the thymus segregates with the Ah locus.
(Nagayama, et al., 1989) Intraspecies differences do not appear to be related to different levels of
cytosolic Ah receptor, but to a different cellular response to the ligand-receptor complex. Bombick
et. al., (1988) administered multiple doses of TCDD in a single i.p. injection to DBA and C57BL/6J
mice. The LOAEL for thymic atrophy in the DBA (115 ug TCDD/kg) was four times greater than
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the LOAEL (30 ug TCDD/kg lowest dose tested) in the C57BL/6J mice. A comparison of thymic
atrophy in wild type (sensitive) and congenic strains of C57BL/6J by Bimbaum, et al., (1990)
indicated that the sensitivity of these strains is associated with the Ah locus.
Four strains of mice with differing susceptibility to Ah receptor mediated toxicity were used
to investigate the role of this receptor in thymic atrophy and related immunotoxicity. C57BL/6,
DBA, C3H/HEN, B6D2F1 male mice were administered a single i.p. injection of 0, 1.2, 6.0, and
30 ug TCDD/kg TCDD. All but the DBA mice showed a statistically significant decrease in
thymic weight at the two highest doses. The dose-response relationships using a "Michaelis-
Menten" type fit (See table HI and Graph B3) demonstrate a good correlation between receptor
induction and thymic atrophy (Vecchi et al., 1983). The fit to some of the data sets was not
satisfactory for the C57BL/6 and DBA mice, therefore the use of other functions was explored.
These strain sensitivities were are also demonstrated with the Power law function (Graph B4) with
good correlation coefficients in the responsive strains, 0.84 to 0.97. Linear fits to these data can
also be used to demonstrate relative sensitivity, however correlation coefficients are generally lower
(Graph B5) resulting in a poor fit.
The comparison of Long Evans (LE) and Han/Wistar (H/W) rats provide a unique
opportunity to observe intraspecies differences. Long Evans rats are more sensitive to the acute
toxic effects of TCDD than Han/Wistar rats, differences which not are believed to correspond with
the Ah locus. L-E and H/W rats were administered a single i.p. dose of 5, 50, and H/W rats were
additionally administered 500 ug TCDD/kg. Tissues were weighed at day 1, 4, 8, 16, and 32. A
decrease in thymus to body weight ratio was observed 8 days post dosing in H/W rats at a dose
of 5 ug TCDD/kg. No effect was seen at this dose on day 1 or 4 post dosing. In L-E rats this
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same dose produced a decreased thymus to body weight ratio, but was observed earlier at 4 days
post dosing. (Pohjanvirta, et al., 1989, 1990)
The distribution of TCDD in these two strains of rats was also examined by Pohjanvirta,
et al., (1990). Both strains were administrated a single i.p. injection of 5 ug/kg 14C-TCDD.
Although many tissues were examined, liver, thyroid, kidney, spleen, and lung had very similar
distribution patterns across tissues, but the distribution of TCDD to the thymus was very different.
hi Long Evans rats there is a continuous increase in TCDD concentration from 4 hrs to 16 days
to a maximum of 0.18% of total dose concentrated in the thymus, followed by a decrease at the
32 day endpoint. In the Han/Wistar rat the concentration of TCDD was about half the level in the
L-E rat (0.07%), and remained constant from day 1 to day 8, followed by a decrease (Pohjanvirta
et al., 1990). Further studies are needed to explain the differences in tissue distribution and TCDD
sensitivity in these two strains of rats.
Olson et al., (1980) evaluated the toxicity of TCDD in the Golden Syrian Hamster, a species
not highly sensitive to the effects of TCDD having a 50 day LD50 of greater than 3000 ug
TCDD/kg, ip ; and an oral LD50 of 1157 ug TCDD/kg. Olson et al. (1980) also evaluated the
thymic response to TCDD after a single ip injection of 0, 5.0, 25, 100, 500, 1000, 2000, and 3000
ug TCDD/kg. Because of the number and range of data points, these data were evaluated more
extensively using all the functional forms described in the introduction. Graph B6 depicts a
"Michaelis-Menten" dose-response relationship, and a dose-dependent decrease in thymic weight
was noted at 100 ug TCDD/kg which becomes significant (see Table HI) at 500 ug TCDD/kg/day
and greater. Reduced thymus weight was the most sensitive alteration noted in gross and
histopathological examination. Olson et. al.(1980a) measured tissue distribution of radioactive
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TCDD in the hamster after a single oral dose of 650 ug TCDD/kg. The distribution to the thymus
reached a maximum at day 1 post injection and by day 3 TCDD content (measured as percent dose
TCDD/g tissue) had begun to decrease from the thymus. The peak concentration in other tissues
varied from day 1 to day 10 (liver was day 3) and had decreased substantially in all tissues by day
20. An important factor in tissue specificity may be excretion and distribution rates.
The Olson et al. (1980) data set (Table HI) was used for additional dose-response analysis
because of the number of data points in this study. In this data set, hamsters were administered
single oral doses from 5 to 3000 ug TCDD/kg and thymus weight was measured at 50 days post-
dosing. The percent response data were evaluated using the Linear, Michaelis-Menten and Hill
functional forms described in the introduction. These data normalized as percent response provided
a better fit to the functions than raw data. This could be due to the imprecision of changes in
tissue weight as a toxic endpoint, or that normalized tissue weight is a more biologically relevant
indicator of toxicity. Because negative values of percent response are not compatible with the
Power Law function, raw data was used when evaluating this function. The linear model of
normalized response (Graph B6-(A) did not fit well (R2 = 0.65) presumably because it not reflect
the biology of receptor-ligand interactions. The graphic representation of these data using the
Michaelis-Menten function does begin to show a maximum response through the top dose of 3000
ug TCDD/kg (Graph B6-(B) and appears to reflect biological phenomenon. The third functional
form used was the Hill plot (graph B6-(C) which also incorporated the data at the high doses. The
log-linear plot of the Hill function depicts an "inertial" type of response at low doses with a "g"
value over the entire dose range of 1.09. a linear/superlinear dose response. The Power Law
function (Graph B-7(A) graphed over the entire dose range demonstrated a sublinear (g = -0.39.
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R2 = 0.76) decrease in absolute thymus weight. The Power Law function was subsequently applied
systematically to subsets of doses to obtain optimal R2 values. The Michaelis-Menten and Hill
plots could not be used for this additional analysis because of the restrictions within these
functions. The result of mis exercise was to graphically observe a change in kinetic order from the
low dose, 5.0 to 500 ug TCDD/kg, (Graph B-7(B) with a "g" value of-0.17 (R2 = .85) to the high
dose data, 500 to 2000 ug TCDD/kg (Graph B7-(C) with a "g"value of -1.57 (R2 = .99). This
increase in kinetic order with increasing doses suggests an inertial response at low doses and a
synergistic response occurring at high doses suggesting that an inhibitory protein may be saturated
at higher doses or induction of additional factors may occur at high doses to enhance the response.
However, this is very speculative due to the uncertainties inherent in the use of clinical endpoints
and reported means values, the constraints of using % response normalized data, and the
limitations of evaluating a single species.
Several factors including tissue distribution, nuclear uptake, receptor concentration, and
receptor affinity, could also be responsible for the differences in toxicity observed in hamsters, rats
guinea pigs and mice. Pohjanvirta et al., (1990) identified strain differences in tissue distribution,
Lund et al., (1982) found that nuclear uptake of TCDD in the thymus is only 6% of the uptake in
the liver, and Carstedt-Duke (1979) studied the distribution of the TCDD receptor in various
tissues in the Sprague-Dawley rat. The highest concentration of receptor, in tissues examined, was
found in the thymus 25.2 fmol/mg protein. Cultures of several tissues from the Sprague Dawley
rat were incubated with 10 nM TCDD to determine cytosolic Ah receptor concentrations. In the
thymus, receptor concentrations remained relatively constant between day 2 and day 70 (Gasiewicz,
1985). Although species differences in these factors are not well characterized the differences seen
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in tissues provide evidence that these factors are also involved in species differences.
Studies indicate that thymus cells have a high concentration of Ah receptors, and that TCDD
induces receptor mediated enzyme induction in this tissue (Cook et. al., 1987). Dose related enzyme
induction (7-ethoxycoumarin O-deethylase activity) has been observed in human thymic epithelial
(HTE) cultures with a NOAEL of 0.1 nM and LOAEL of 1 nM. In HTE cultures it was observed
that individual cultures differ in their sensitivity and that the Ah receptor concentration is not an
accurate predictor of enzyme activity. In addition, in HTE cultures there is an Ah receptor
mediated decrease in thymocyte maturation. This was determined by co-culturing thymocytes with
HTE and examining the responsiveness of these cultures to mitogens (concanavlin A and
phytohemagglutinin). Concentrations of TCDD suppressing thymocyte maturation ranged from 0.1
nM to 1.0 nM. (Cook et. al., 1987)
The pathology of TCDD induced thymic atrophy has been characterized as a depletion of
small cortical cells in the thymus, thinning cortex and an increase in macrophage in the cortex
characteristic of cell damage (McConnell, et al., 1978a). Kociba found in a 13-week repeated dose
study in rats, a pronounced decrease in thymocyte number in the cortical region of the thymus
(Kociba, et al., 1976)
Thymocytes develop into mature T-lymphocytes, and apoptosis occurs as a natural part of
the selection process in this maturation. TCDD and glucocorticoids both increase thymocyte
apoptosis. This could have a dramatic effect on the immunological status of the animal (Nebert
et al., 1990). McConkey and Orrenius (1989) proposed that the decreased number of thymocytes
seen in vivo experiments results from an increase in Ca2+-mediated endonucleases, normally seen
as part of the thymocyte maturation process. It has been suggested that this effect is similar to the
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effect seen with exogenous glucocorticoid treatment. Analogies in the mechanisms of action have
been noted between glucocorticoid and TCDD nuclear receptor mediated mechanisms of action.
Similarity of their action on thymocyte Ca+2 endonucleases as measured by %DNA fragmentation
led researchers to propose a receptor mediated mode of action on thymocytes. Additional evidence
that TCDD and glucocorticoids act in a similar manner to increase thymic atrophy through
depletion of thymocytes includes the findings (McConkey and Orrenius 1989) that inhibitors of
glucocorticoid induced thymic suicide, reduced the effects of TCDD, and reduced the glucocorticoid
sensitivity in thymocytes. This does not preclude the possibility of different mechanisms of action
with shared modulators of receptor-ligand interaction.
An alternative mechanism to induce thymic atrophy was proposed by Bombick, et al., 1988.
After administration of TCDD to sensitive mice an increase in protein-tyrosine kinase was observed.
Previous reports (Bombick et al., 1988) found that protein tyrosine kinases are associated with
lymphocytes, and that they may be involved in lymphocyte maturation. This seems very plausible
especially since the protein kinase p60 activates oncogenes and may play a role in mitogen signal
transduction. Hence, any inhibitor of mitogenesis and signal transduction initiated by TCDD on
the DNA would stop maturation and produce thymic atrophy. A dose-dependent increase in protein
kinase was observed following in vitro exposure of 2-10 nM TCDD to B lymphocytes. The
increase occurred rapidly (increases observed 5 minutes after exposure) and was significant at
approximately 2.5 nM (Clark et al. , 1991). The significance of this observation to TCDD
mediated immune alterations in vivo has not been determined.
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Table III Species Comparison of Thymus Weight Changes After TCDD Administration
S lecies
DHA/21
C57BU6J
female mice
C57BL/6J
C57BL/6 Sell
male
C57BL/6
DBA
C3H/HeN
B6DF1
male mice
Long E\ ans
Sprague
Sprague
Dawley rats
Golden
Syrian
Hamster male
Guinea pigs
Hartley
Uuitiea pigs
Hartley
em a e
Dosing Regimen
rtinglu i {> (9. 1 corn
oil/acetone)
single oral
single i.p. (9.1 corn
oil/acetone)
single oral
single i.p. injection
measurements 12 days
post dosing
single i p
oral 5 X week 13
weeks
oral 5 X week
13 weeks
single ip injection
measure 50 days poet-
dosing,
continuous dose in
feed 90 days
8 X weekly oral
Effect
dotryuso tliymuu wuiglit (day 2)
decrease tbymic/bwt 7 days post
dosing and absolute thymuB weight
decrease thy m us weight
(day 2)
decrease thymus weight (week 2 and
week 6)
% decrease thymus weight
decrease thymus weight
decreased thymua weight
decreased thymus weight
decrease thymue weight
decrease relative thymus weight
decrease thymus weight (day 36)
Dose ug TCDD/kg
0
30
60
115
0
0.2
1.0
20
5.0
0
30
60
115
0
0.2
1.0
50
25.0
0 ug/kg
1.2
6.0
30.0
0
5.0
50.0
0 (0) (tot dose
0.001 (0.065)
0.01 (065)
0.1 (6.5)
1.0 (65)
0 (0) (tot dose)
0.001 (0.065)
0.01 (0.65)
0.1 (6.5)
1.0 (65)
0
5
25
100
500
1000
2000
3000
Male Female
0 ng/kg/day 0
0.12 012
0.61 0.68
4 90 4.90
26 31
ug/kg/week (total dose)
0 0
0 008 (0 064 total)
0.04 (0.32 total)
02 (16 total)
Response % change Thymus Weight
0
4.7%
106%
40 1%*
0 tbymus/bwt
+3%
-6.6 %
-16.7 %
-23.4%
0
+45 %
-12.9 %
-19.9 %
0%
27.2%*
35.0%*
53.1%*
0%
9.3%
12.5%*
33.1%*
79.9%*
C57BL/6 DBA/2
0 % 0 %
1 1 % 19 %
44 %* 16 %
54 %* 27 %
0%
3 5%
18.4%*
456%*
79.8%*
C3H/HeN B6D2F1
0 % 0 %
20 % 8 %
40 %* 24 %*
27 %* 44 %*
0 Day 4 0 Day 8 Day 16
20%* 40%*
20%* 60%*
0
1.9%
3.8%
21.0%
750%
0
7.5%
150%
42.5%
77.5%
00
1.9%
2.9% increase
32.7%
53.5%*
84.2%*
94.0%*
83.2%
Males
0 %
15%
13%
24%*
0 101 g (raw data)
0.099 g
0.104
0.068
0047*
0.016
0006
0017
Females
0%
6%
16%
12%
0
17.8%
25 4%*
47 2%*
NOAEL
1.2 ug/kg
C57BL,
C3H, B6
LOAEL
115 ug /kg
30 ug/kg » 0.92
nM/kg
1.0 ug/kg « 3.1
nM/kg
6.0 ug/fcg
C57BL
C3H
B6D2F1
500 ug/kg »
155.0 nM/kg
0.04 ug/kg «
0 12 nM/kg (0.32
total * 0 96
tlM/kg)
Reference
Bombick, et al ,
Kerkvltet and
Brauner, 1990
Bombick, et al.,
1988
Voa, et al , 1974
Vecchi et al., 1983
Pojhanvitta, et al.,
1989
Kociba, et al., 1976
Kociba, et al., 1976
Olson, et al , 1980
DeCaprio, et al.,
1986
Harris, et al., 1973
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The development of mature thymocytes involves the passage of cells through a variety of
subpopulations with unique phenotypes. Antigenic determinants can be used to identify these
subpopulations. Kerkvliet and Brauner (1990) used flow cytometric analysis to quantify these
alteration in thymocyte subpopulations after p.o. administration of 2 ug TCDD/kg in C57BL/6J
female mice. There was a significant TCDD dependent increase in the percentage of CD4"CD8"
double negative (DN) thymocyte phenotypes and a decrease in the percentage of double positive
(DP) CD4+CD8+ thymocytes. These changes were not observed in the lymphocyte population of
the spleen indicating that the T-cell population of the thymus is uniquely sensitive to TCDD
toxicity. Although dose-response trends were not reported in this study, the alteration in the percent
of thymocyte subpopulations was a sensitive indicator of TCDD immunotoxicity. The alteration
of these subpopulations has also been demonstrated in vitro further supporting the hypothesis that
an imbalance in percent of cell types may be responsible for TCDD induced alterations in
immunocompetence (Lundberg et al. 1990).
Fine, et al.(1990) investigated the effects of TCDD on the intrathymic development of T-
cells by quantifying the effects of a single i.p. injection TCDD on subpopulations of cells in BALB
male mice. All populations of thymocytes were reduced substantially, 50-76% decrease in total
number of cells of each type, in addition the total cell population as measured by thymic cellularity
was decreased approximately 90%. Although the total number and type of cells decreased the
percent of cells in each population changed. This suggests that TCDD is working through two
different mechanisms. There was an increase in the percent of CD4-CD8- double negative (DN)
CD4+CD8-, and CD8+CD4- single positive (SP) thymocytes cell types, and a decrease in
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CD4+CD8+ double positive (DP) cell types. These data are supported by Baylock et al. (1992)
who found that fetal thymic cell population percentages varied in a similar manner. They suggested
that TCDD altered the maturation of fetal thymocytes affecting the transition from DN to DP.
Holladay, et al. 1991) found the effect of TCDD on fetal thymocyte maturation was more
significant at gestation day (GD)18 than at day 6 or day 14 after birth, indicating a sensitivity in
fetal thymocytes that decreases with age (Blaylock et al., 1992).
Decreased thymus weight in the perinatal period is related to a decrease in immune system
function, and as such may be of critical importance in the understanding of the toxicity of TCDD.
Pregnant Wistar rats were administered a single oral dose of 0.0, 0.005, 0.5, 0.125, 0.25, and 0.35
ug TCDD/kg on gestation day (GD) 16. A dose related decrease in pup thymus weight was
reported. In addition, there was a dose related increase in 7-ethoxycoumarin deethylase activity,
and on biphenyl 2-hydroxylase activity at the above administrated doses, with a NOAEL for
enzyme induction of 0.005 ug TCDD/kg (Madsen, et al., 1989).
TCDD produces a consistent response of thymic atrophy in the guinea pig, rat, hamster,
monkey and in the sensitive mice. This effect is receptor mediated and although the mechanism
is not completely understood, several plausible mechanisms have been proposed. Because the
thymus is a key organ in immunological responses it is not surprising that TCDD affects the
immune system.
C. Immunotoxicity
TCDD-mediated immunotoxicity and the implications of enhanced and suppressed immune
function are described in detail in the Immunotoxicity chapter of this document (Kerkvleit and
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Burleson 1992). In addition, detailed evidence that the immunotoxicity of TCDD and related
halogenated aromatic hydrocarbons is, at least in part, Ah-receptor mediated is presented. Studies
with multiple doses and protocols that most resembled other studies (similar species, strains and
dosing regimens) were evaluated for comparison of immunological endpoints. Because of .the
variety of protocols tested to evaluate TCDD immunotoxicity some endpoints have not shown clear
dose-response relationships, however there are some that do. The premise that immunotoxicity is
at some level Ah receptor mediated leads to the assumption that the response of some
immunotoxicity endpoints may be related to the response of other Ah receptor mediated events.
Exposure to TCDD results in thymic atrophy, alterations in bone marrow and general
immunosuppression in almost every species examined. (Luster et al., 1979,1980) Suppression of
cell-mediated immunity, (CMI), measured as a response to phytohemagglutin (PHA) and
concanavalin A (Con A), may be observed in the neonate during thymic organogenesis; the most
sensitive time for this type of immunotoxicity (Luster, et al., 1987). Vos (1974a) investigated the
role of maternal treatment of (0, 1.0, 5.0 ug TCDD/kg) administered prenatally GD 11 and GD 18,
and/or postnatally day 7 or 14 on cell-mediated immunity (CMI) of Fischer-344 rat pups. The
response of spleen cells to PHA was significantly decreased in 25 day old pups after maternal
postnatal exposure only at the 5 ug TCDD/kg level. However, after pre and postnatal maternal
exposure there was a significant decrease in the PHA response at 1.0 ug TCDD/kg indicating that
prenatal exposure is a more sensitive period for immunotoxicity. There was significant fetal
toxicity associated with immunosuppression at the 5 ug TCDD/kg level. The susceptibility of pups
is believed to be greater than in adult rats because of the role the thymus plays in CMI. Although
the relevance of the thymic atrophy seen in animals as it relates to the adult human population may
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be questioned, the thymus is an important organ in the development and function of the human
immune system. Similar results were reported by Faith and Moore (1977) in thymus cells from 25
day old female Fischer rat pups exposed postnatally to TCDD. There was a significant decrease in
responsiveness to PHA at the 5 ug TCDD/kg level, but not to Con A. The impact of pup age on
CMI is assessed by comparing the responsiveness to PHA in 1 and 4 month old mice. The
responsiveness to PHA was significantly reduced in 1 month old mice that were administered 4
weekly doses of 25 ug TCDD/kg TCDD. Four month old mice administered the same doses (0,
1.0, 5.0, and 25 ug TCDD/kg) weekly for 6 weeks had no differences in PHA responsiveness when
compared to controls although all doses showed significant thymic atrophy. The suppression of cell
mediated immunity in sexually mature (young adult) rats and mice was demonstrated to be a
relatively insensitive endpoint which differed significantly from controls at the 25 ug TCDD/kg
dosage level only. The difference in sensitivity based on age is consistent with what is known
about the development of the thymus (Vos and Moore 1974a).
Thomas and Hindsdill (1979) described a dose-related decrease in relative thymus weight
in offspring of adult female Swiss mice exposed to TCDD in feed at concentrations of 1, 2.5, 5 and
10 ppb for 4 weeks prior to mating. A statistically significant decrease in relative thymus weight
and number of SRBC (sheep red blood cells) plaque forming cells per gram of spleen (both
indicators of immunotoxicity) were observed in the offspring of animals exposed to 2.5 or 5 ppb
TCDD. Although food consumption data was not reported in this study and daily intake and fetal
dose can not be determined, the spectrum of toxic effects presented are relevant. Both endpoints
are indicators of immunotoxicity. No overt signs of maternal toxicity were observed at doses
producing immunotoxicity.
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Another effect observed after prenatal exposure was that allograft rejection times were
prolonged in a dose-dependent manner in 23 day old C57BL/6J mice exposed pre and postnatally
(GD 14 and GD 17) to 2.0 and 5.0 ug TCDD/kg levels of TCDD (see Table IV) and in rats
exposed postnatally on day 1,8, and 15 to the 5.0 ug TCDD/kg level (Vos and Moore 1974a).
Vecchi et al. (1980) examined the response of C57B1/6J adult mice splenocytes to Con A
and lipopolysaccharide (LPS) after TCDD treatment to assess cell-mediated immunity. Male mice
were administered 0, 1.2, 6 and 30 ug TCDD/kg. At 7 and 14 days post dosing, splenocytes from
these animals were tested for their response to Con A and LPS. At 30 ug TCDD/kg a significant
increase in mitogenic response was observed at 1.6 and 3.2 ug/sample Con A, NOAEL=6.0 ug
TCDD/kg. Their was no splenocyte response to LPS in mice administered 30 ug TCDD/kg. This
study indicated that relatively high doses (30 ug TCDD/kg) are required to affect cell-mediated
immunity in adults.
Vos et al. (1973) also examined cell-mediated immunity in female Hartley guinea pigs
receiving 8 weekly doses of 0, 0.008, 0.04, 0.2, and 1.0 ug TCDD/kg and in female CD rats
administered 6 weekly doses of 0, 0.2, 1.0 and 5.0 ug TCDD/kg (Table IV). Vos et al. (1973)
selected delayed hypersensitivity to tuberculin toxin in guinea pigs and rats to assess CMI. The
results are reported as diameter of skin reaction. In guinea pigs administered TCDD, sensitized
with tuberculin on day 35 post dosing and challenged on day 47 and 54, there was a dose
dependent decrease in diameter of skin reaction which was significant at the 0.04 ug
TCDD/kg/week level (0.32 ug TCDD total dose). TCDD inhibited the delayed hypersensitivity
response in these animals. There was no significant change in diameter of skin reaction in rats
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sensitized on day 28 and challenged on day 42 (see Table IV). This response is consistent with
results of acute toxicity data that found guinea pigs are more sensitive to the toxic effects of TCDD
than rats. However, rats are generally not used for delayed hypersensitivity testing. Hence, the
guinea pig results stand on their own.
Another method Vos et al. (1973)usedto assess cell-mediated immunity is to examine local
graft vs. host reactions (GVH). B6D2F1 recipient mice administered 0.0, 0.02, 1.0, and 5.0 ug
TCDD/kg were injected with C57BL/6J mice spleen cells. To assess the GVH response the ratio
of the weight of the right and left popliteal lymph nodes was compared. A dose dependent
decrease in ratio was observed and was significant at 5.0 ug TCDD/kg (Table IV)
Rosenthal et al. (1989) assessed the response to endotoxin (E. coli LPS, 0.25, 5 and 25
mg/kg) administered 7 days post TCDD treatment to female B6C3F1 at single oral doses 0, 50,
100, and 200 ug TCDD/kg. Effects observed included an increase in mortality from 5 mg/kg
endotoxin at the 200 ug TCDD/kg level and a decrease in the ability of mice administered 50 ug
TCDD/kg to clear 2.5 mg/kg endotoxin from their blood. The sensitivity of this endpoint was not
tested, as all administered doses resulted in an some effect (Table IV). The increased
responsiveness to endotoxin segregated with the Ahb/b locus (sensitive strain) and was not seen in
the Ahd/d (less sensitive strain) congenic mice. Furthermore, the response to endotoxin and TCDD
was reversed by administration of methyl prednisolone leading the authors to conclude that an
inflammatory component is involved in the response.
The most sensitive indicator of cell-mediated immunity is the reduction in cytotoxic T-
lymphocytes (CTL). Nagarkatti et al.(1984) demonstrated a decrease in CTL response at a
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cumulative dose of 4 ng/kg. Kerkvliet et al. (1990a)noted a dose-related decrease in CTL at 10 and
20 ug TCDD/kg (single oral dose TCDD). Graphic representation of there data indicated a dramatic
decrease in the CTL response in C57BL/6J mice and a minimal decrease at these doses in non-
responsive congenic mice. A controversial study by Clark et al. (1981) demonstrated a dose-related
suppression of CTL in C57BL/6J mice. Doses as low as 0.004 ug TCDD/kg to 40 ug TCDD/kg
resulted in a decrease in CTL activity. It has also been reported by Clark et al. (1981) that TCDD
does not deplete the number of CTL precursors, but induces a population of suppressor T-cells.
Clark's data demonstrate suppression at extremely low levels, however, the protocol used is very
difficult to precisely replicate (Holsapple et al. 1991, 199la).
Cell mediated immunity is only one type of response the body employs as a defense
mechanism. Humoral immunity may be assessed by measuring the serum level of induced anti-
toxin in response to toxoid challenge. Vos et al. (1973) measured the response to the tetanus toxoid
in guinea pigs administered 8 weekly doses of 0, 0.008, 0.04, and 0.2 ug TCDD/kg. Guinea pigs
were injected with toxoid on day 28 post TCDD administration and anti-toxin levels were measured
7,21, and 28 days after initial injection. There was a significant increase in serum anti-toxin levels
at 0.008 and 0.04 ug TCDD/kg levels 7 days after toxoid injection, however, not at 21 and 28 day.
There was a significant decrease in anti-toxin levels in the 0.2 ug TCDD/kg dose at these two time
points suggesting a dose-dependent response (Table V).
Another method of assessing humoral immunity is to quantify the production of specific
antibodies in response to an antigen characterized as thymus-dependent and thymus-independent
antigens. The sheep red blood cell (SRBC) antigen is the most common antigen used to assess
immune competence after TCDD treatment. This assay has produced consistent and reproducible
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results while using multiple doses. Therefore this endpoint will be evaluated more empirically.
Vecchi (1980) assessed the immune response in C57B1/6J male mice after i.p. treatment with 0,1.2,
6, or 30 ug TCDD/kg. 4 X 108 SRBC, a T-dependent antigen, and 0.5 ug of type HI
pneumonococcal polysaccharide, a T-independent antigen, were used to assess humoral immunity.
At all doses, with both antigens, and at all time periods assessed, TCDD significantly decreased the
plaque forming cells/106 splenocytes (a measure of immune response) in a dose-dependent manner
(Table V). The antibody-mediated immunity was suppressed in C57BL/6J mice at the lowest dose
tested, 1.2 ug TCDD/kg, a dose lower than that found to suppress CMI.
Davis and Safe (1988) measured plaque forming cells (PFC)/106 splenocytes) in C57BL/6J
mice after administration of 0.322, 0.422, 1.2, and 3.1 ug TCDD/kg followed by an i.p. injection
of 4 X 108 SRBC. Although statistical significance was not assessed in this study, there was a
dose-dependent decrease (Table VI) in PFC/106 viable cells seen in all dose levels. These authors
also evaluated the response of other poly chlorinated dioxin compounds to examine structure activity
for Ah receptor affinity and concluded that the humoral response to T-dependent antigens is Ah
receptor mediated. In addition the authors suggest that TCDD-receptor complex may modulate
early B-cell differentiation.
Narasimhan et al. (1993) assessed the same endpoint in B6C3F1 mice and found a
statistically significant decrease in PFC/106 viable cells at 0.05 ug TCDD/kg (Table VI). Analysis
of dose-response for immunotoxicity is complicated by the variety of endpoints and dose ranges
reported in the literature. However, the use of PFC/spleen was a consistent measure of humoral
immunity in three studies. The Kerkvliet and Brauner (1990); Narasimhan et al.(1993)
unpublished); Silkworth et al. (1989); and the Davis and Safe (1988) data sets were selected for
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additional dose-response analysis. These data sets normalized to percent response, were fitted to
the Michaelis-Menten function (graph Cl) using doses in the same range6. All but the Kerkvliet
and Brauner (1990) data set fit this function poorly. The fit to the Power Law function for all four
data sets (graph C2) shows a consistent response between the Silkworth (DBA2) and the Kerkvliet
data, ie, both fit the function well (R2 = 0.99 and 0.98 respectively), and had comparable exponents
in the function ("g" = 0.32. 0.30). The Davis and Safe (1988), Silkworth et al. (1989) and
Narasimhan et al. (1993) data sets all had higher "g" exponents and poorer fits. A plausible
explanation for these variation could be that test methodologies differed(ie. use of different control
solutions) or there are genuine species and sex differences that modulate this response. It is
interesting to note that the shape of the dose-response curves for the DBA and BALB mice used
in the Silkworth et al. (1989) experiment are very different (Graph C2) and this suggests that the
dose values used are capturing different segments of the dose-response curves for these two strains
of mice.
The Narasimhan et al. (1993) data set was used for additional analysis because of the
number and range of doses reported. The functions that best fit the raw data points were the
Michaelis-Menten and the Hill functions (Graph C3-A, C3-B). The Power Law Function (Graph
C4-A) fit all raw data values for PFC's/spleen with an R2 value of 0.89 and a "g" value of -0.43.
A Power Law functional fit of the entire data set using normalized percent response data could
performed because at very low doses there was a negative value which is not mathematically
6The entire data set from Safe, 1992 was not used for this analysis. Direct comparison to
Kerkvliet and Brauner (1990) and Davis and Safe (1988) used data that from a similar dose range.
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consistent with this function. Systematic selection and analysis of data subsets using % response
normalized data (excluding extremely low values) resulted in values of "g" = 2.2; and R2 = 0.99,
in the low dose range (0.025 to 0.1 ug TCDD/kg) (Graph C4-B) suggesting superlinearity. In
contrast, the high dose range (0.5-2.5 ug TCDD/kg) resulted in "g" = 0.23 and R2 = 0.87,
suggesting sublinearity or a lower rate of response (Graph C4-B). The higher exponent at the low
dose may represent synergy or positive cooperativity. However it is also consistent with the
existence of non-nuclear receptor-mediated interactions at low doses, or may suggest induction of
additional inhibitory factor at higher doses. When an analysis of subsets of this data was applied
to raw data a different pattern of dose-response was observed. At the low administered doses the
absolute response resulted in a sub-linear fit with a "g" value of-0.57, R2 = 0.89, whereas at higher
doses a steeper linear fit was observed, "g" = -.99, R2 = 0.98 (depicted in graph C4-C). Whether
raw data of percent response normalized data is a better representation of low dose phenomenon
can not be concluded from this analysis.
Kerkvliet et al. (1990) demonstrated differences in humoral immunity related to the Ah
responsiveness of C57BL/6J congenic mice after acute exposure to TCDD. The dose-dependent
decrease in splenic antibody response to the T-cell independent antigen TNP-LPS revealed
differences in sensitivity which were consistent with strain Ah responsiveness. The response of
these strains to T-cell dependent antigen was more complex and therefore could not be directly
related to Ah receptor responsiveness (See Table V). Graph C5-A depicts the different sensitivities
of the two strains using a Michaelis-Menten function with a reasonably good fit. The Power Law
function also depicts strain sensitivity (Graph C5-B) although the fit is poor.
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Vecchi et al. (1983) compared multiple doses (3) in three strains of mice to demonstrate that
the decrease in antibody production as measured by PFC's/106 spleen cells segregated with the Ah
receptor (See Table IV, Graph Dl-A, Dl-B). The Michaelis-Menten and the Power Law function
are both good representations of the species differences in sensitivity. Because of the differences
in the correlation and the "g" value in the Power Law, it appears that the best fit to this functional
form is at the doses where a change in response is measured.
In an acute study (Morris et al. 1992) a dose-dependent decrease in anti-SRBC IgM AFC
(antibody forming cells)/spleen was observed in all doses (4.2, 14, and 42 ug TCDD/kg)
administered orally to the B6C3F1 mouse (Ah responsive strain) however, only at the 42 ug
TCDD/kg dose in DBA/2J mice (less responsive). Subchronic administration of the same
cumulative dose revealed statistically significant decrease (not clearly dose dependent) at all doses
in both strains. This study clearly shows (Table V) that exposure conditions can alter response.
The acute exposure study demonstrates that B6C3F1 mice are approximately 10 times more
sensitive than the DBA mice, however the results of the subchronic study suggest that there may
be a modification of the Ah mediated strain sensitivity. These results support what is known about
Ah receptor mediated toxicity that cumulative doses will establish an equilibrium between receptor-
ligand formation, receptor biosynthesis, and receptor-mediated effect (EROD induction, or
suppressed immune response). A lower dose can maintain this equilibrium whereas a higher dose
would be required to induce the same effects with a single dose.7 Clearly Morris et al. (1992) has
demonstrated this phenomenon with humoral immunity.
7In comparing Bimbaum and DeVito's data (Table X) and Safe's data (Table IX) the dose to
induce a statistically significant increase in EROD after acute administration was 0.5 ug/kg whereas
administration of 0.0975 cumulative dose in a subchronic protocol induced tharne level of EROD
induction.
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Kramer et al. (1987) using highly purified murine B-cells suggested early events may be
the target of TCDD in these cells since TCDD increased basal kinase activity in the presence of
added phospholipid, but did not increase protein kinase C (PKC)activity. The authors proposed that
TCDD affects basal kinase activity at the level of protein phosphorylation. Hence, protein kinase
involved in phosphorylation and early activation may be the target of TCDD and result in altered
B-cell differentiation. Changes in basal protein kinase activity, although Ah receptor mediated, may
not require translocation to the nucleus and therefore this endpoint may respond very differently
than other events mediated through XRE activation.
It has been hypothesized that B-cells are a direct target of TCDD toxicity (Dooley and Holsapple
1988). The effect of TCDD on T- and B- cell interactions in the antibody response was
investigated using a T-dependent antigen SRBC's and two T-independent antigens DNP-Ficoll
(dinitrophenol-Ficoll)andTNP-LPS(trinitrophenol-lipopolysaccharide). The antibody response was
assessed by enumeration of antibody forming cells (AFC's) in spleens of female B6C3F1 mice
administered 0, 0.1, 1.0 and 10 ug TCDD/kg daily for 5 days (Table V). Five ug TCDD/kg total
dose produced a decrease in AFC induced by all three antigens. Their findings that the response
to T-dependent and T-independent antigens was decreased at the same dose level of TCDD led the
authors to conclude that B-cells were the primary target of TCDD and that this imnrunosuppression
is at the level of B-cell differentiation. These doses are high doses for immunomodulating effects
(100 X the LOAEL in the Narasimhan et al. (1993) study) and therefore may not represent the
lowest dose effect.
To assess the suppression of humoral immunity by TCDD and the role of the Ah receptor
in that suppression, Holsapple et al. (1986), and House et al. (1990) administered a single p.o. dose
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of 0, 0.05, 0.1, 0.5, 1.0, and 2.0 ug TCDD/kg to female B6C3F1 mice and subsequently
administered SRBC a T-dependent antigen. The response was measured by the enumeration of
IgM antibody forming cells (AFC) in the spleen. Significant dose-dependent reduction in IgM AFC
was observed at 1.0 ug TCDD/kg (Table V).
The immune system is very complex with many interactions across its components.
Susceptibility to infection may be affected by TCDD. Thigpen et al. (1975) found that C57BL/6J
male mice treated with weekly doses of 0, 0.5, 1.0 and 5.0 ug TCDD/kg for 4 weeks (equivalent
to 0.07, 0.14, and 0.7 ug TCDD/kg/day) and subsequently administered a challenge of Salmonella
bern responded with a dose dependent increase in mortality at the 1.0 and 5.0 dosage levels. The
increase in mortality was not seen after administration of a challenge dose of the virus Herpesvirus
suis. Thigpen et al.(1975) attributes this to differences in pathogenesis of the infections. In this
study, the administered doses of TCDD resulted in no clinical signs of toxicity; yet resulted in
immune suppression. House et al.(1990) attempted to elucidate the effects of TCDD on a variety
of immune system factors. Using the influenza virus to assess responses. Survival after
administration of 0.1, 1.0, and 10 ug TCDD/kg was significantly decreased in a dose-dependent
manner in all doses tested. In addition it was reported that TCDD did not effect interferon
production; a system generally initiated by a viral infection.However, TCDD may directly interfere
with the effector cells in this system by modulating the production of cytotoxic T-cells and natural
killer cells (House et al. 1990). It was also reported that a dose-dependent increase in mortality
in TCDD treated mice following an influenza challenge in which the LOAEL (lowest dose tested)
was a single dose of 0.1 ug TCDD/kg (House et al. 1990)
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Table IV Cell-Mediated Immunity
Species
C57BL/6J
23 day old
mice pups
F-344 28
day old rat
pups
Female
albino rats
CD
Male
B6D2F1
mice
Female
B6C3F1
mice
C57BL/6
male mice
Hartley
guinea pigs
Dosing regimen
0. 2.0, 5.0 ug/kg pre and
postnatally QD 17 and
17 and day 1,8,15
0 and 5 ug/kg maternal
postnatally on days 0,
7,14
0, 0.2, 1.0, 5.0 ug/kg.
weekly for 6 weeks.Day
28 rats were injected
with M. tuberculosis and
day 42 injected with 5 ug
tuberculin PPD
0, 0.2,1.0 and 5.0 ug/kg
weekly for 4 weeks
C57BL mice. Recipient
mice received graft of
spleen cells in right hind
foot pad.
Single oral dose of 0, 50,
1OO or 200 ng/kg
followed by 7 day iv
treatment with E.coli LPS
endotoxin 5 or 25 mg/kg
weekly Lp. injection X 4
wks At week 5
sensitized punting ears
with 3% oxazalone
challenged with 1%
oxazalone
8 X weekly oral
administration
Effect
increase rejection time of tail-
skin graft
increase rejection time of tail-
skin grafts
Thickness of skin reaction on 24
and 48 hrs after final injection
Response measured as difference
between right and left popliteal
lymph node.
Response mortality at 48 hrs
ear thickness 10' mM
measured
day 35 0.05 ml injection of
tuberculin bacillus
hyperseositivity tested day 47
and 54 with tuberculin injections
(1.25 ng) skin reactions
measured 24 and 48 hrs after
tuberculin injections skin
diameter measurements %
difference from control
Dose (ug TCDD/kg) - Response (%change)
0
2.0
50
00
5.0
0
0.2
10
5.0
ug/kg (total dose ug/kg)
0 0
02 0.8
1.0 4.0
5 0 20.0
ug/kg
0
50
100
200
Dose
0
0.4 ug/kg
4 ug/kg
40 ug/kg
Dose ug/kg/wk (total dose)
0
0.008(.064)
0.04 (0.32)
0.2 (1.6)
0
19%*
25%*
0
+ 18%*
Skin diameter Skin diameter Skio Thickness Skin Thickness
24 hrs 48 hrs 24 hrs 48 hrs
0 000
+34% -105% -9.5% -27%
+5.1% -5.9% +662% -8.2%
+0.7% -4.6% -10.5% -13%
Right/left Popliteal lymph nodes % differs from control
0
-0.2%
-38%
-66%*
Mortality 5 ug/kg endotoxin 0%
0%
0%
87.5%*
Increment in ear thickness 24 hrs after challenge
0
+54%
-38%*
-32%*
skin diameter skin skin diameter
diamet 47 d - 48 hrs 54d - 24 hrs
er
47d- 0 0
24hrs -9% -5%
-15%* -9%*
0 -42%* -22%*
-4%
-8%
-29%*
Mortality 25 ug/kg
endotoxin
0%
50%*
100%*
100%*
skin diameter
54d - 48 hrs
0
-9%
-12%*
•41%*
NOAEL
N/A
N/A
0.4 ug/kg
LOAEL
N/A
N/A
4 ug/kg
Reference
Vos and
Moore, 1974a
Vos and
Moore, 1974a
Vos et al.,
1973
Vos et al ,
1973
Rosenthal, et
al., 1989
Clark, et al.,
1981
Vos, 1973
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Table V Humoral (Antibody-Mediated) Immunity
Species
male
C57BL/6J
6-8 wks
Hartley
Guinea
pigs female
B6C3F1
mice
DBA/2J
mice
Female
B6C3F1
mice
Female
B6C3F1
mice
C57BL/6J
male mice
C57BL/6J
male mice
Ah", and
Ah"
B6C3F1
mice
Male
C57BL/6
male mice
Female
B6C3F1
mice
Dosing regimen
i.p with TCDD followed
by (SRBC) in-jection
PFC measured 7, 14 ,
2 l.and 42 d.after TCDD
8 X weekly oral
administration
oral administration
1 acute oral dose
2 14 day cumulative dose
oral aministration
1. acute oral dose
2. 14 day cumulative
dose
oral administration of
TCDD daily for i days
48 hrs after last treatment
sensitized with antigen
single p.o. TCDD, day 8
post dosing immunized
with SRBC.
i.p. injection of TCDD
weekly for 4 weeks
bacterial and viral
challenge 2 d.after dosing
single oral dose TCDD
prior to ip injection of 25
ug TNP-LPS
single i.p injection of 0,
0.1, 1.0, 10 ug/kg
weekly i.p injectionX 4
wfcs week 5 injected with
1-2 X 10, 5 days later
mj. 2x10' SRBC
14 daily doses p.o.
TCDD
Effect
measured % decrease from
control PFC/ 10' splenocytes
day 28 and day 42 0. 1 ml
tetanus toxoid injected into hind
footpad Serum antitoxin
measured day 35, 49, and 56.
decrease Formation of anti-
SRBC IgM Antibody forming
Cells
decrease Formation of anu-
SRBC IgM antibody forming
cells
Antibody forming cells
measured after sensitizing with
SRBC, DNP-Ficoll, TNP-LPS
in vivo assay
Spleen cell IgM anti-SRBC AFC
response measured as change in
IgM AFC '10' recovered cells
mortality rates reported
% decrease number of nucleated
spleen cells
Spleen cell IgM, IgO, TNP-LPS,
and TNP-FicoIl AFC
response/106 recovered cells
measured footpad 24 hrs after
challenge
1. Serum complement activity
CH50 representing the amount
of serum necessary to lyge 50%
of the target cells
2 Serum level of complement
component C3
Dose (ug/kg) - Response (%change)
0 ug/kg
1.2
6
30
0 dose ug/kg/wk
0.008 (.064)
0.04 (0.32)
0.2 (1.6)
0 ug/kg
4.2
14
42
0 ug/kg
4.2
14
42
0 ug/kg/day
O.I
1.0
10.0
0 ug/kg
0.05
01
0.5
1.0
2.0
ug/kg/week (ug/kg/day)
0 (0)
0.5 (0.07)
1.0 (0.14)
5 0 (0.7)
10 (1 4)
20 (2.8)
0 ug/kg
1
5
10
20
50
0 ug/kg
0.1
1.0
10
0
0.4 ug/kg TCDD
4 ug/kg TCDD
40 ug/kg TCDD
0 Total dose ug/kg/day
001
005
0.1
0.5
1.0
20
0 7d 0 14d
54%* 42%*
86%* 70%*
96%* 89%*
0 Day 25 0 Day 49
+ 16%* -13%
+21%* -22%
-5% -43%*
1. 0
64 %
675 %
83 %
1. 0
16 %
9.5 %
33 %
0 21d 0 42d
60%* 55%*
77%* 85%*
89%* 88%*
0 Day 56
+8%
-7%
-29%*
2. 0
78%
75%
88%
2. 0
56 %
83 %
88 %
0 Response SRBC 0 Response DNP-Ficoll 0 Response TNP-LPS
-32% -60% N/A
-72%* -86%* -64%*
-94%* -87%* ND
0 % change from control IgM AFC
+4.2%
+5%
-26%
-48%*
-66%*
Bacterial challenge mortality
25%
25%
65%
70%
65%
95%
0% (Ah ")
10.6%
17 %
2S.5 %
24%
not tested
0 IgM AFC 0 IgO AFC
-3% -14%
-57%* -22%
-77%* -42%*
0 Response footpad swelling
-3%
34%
-38%
0 CH50 %decrease
41%*
31%*
38%*
54%*
65%*
81%*
0 weight % change from control
+18%
-4%
+12%
+4%
6%
Viral Challenge
15% mortality
ND
ND
15%
0%
10%
0 % (Ah")
not tested
12.6%
22.1 %
15.1 %
25.2 %
0 TNP-Ficoll 0 TNP-LPS
-28% N/A
-66%* N/A
-91%* +6%
0 serum C3 %decrease
3%
6%
2%
8%*
19%*
26%*
NOAEL
N/A
Day 35
N/A
Day 49&
56 0.04
IgM AFC
= 0.5
Ah^*6
Ah"*'"*
IgM 0.1
IgOO.l
TNP-Ficoll
0.1
CH50=
N/A
C3 = 0 1
LOAEL
1.2 ug/kg
Day 35
0.008
Day 4902
Day 56 0.2
IgM AFC =
1.0
Ah^*8
10 ug/kg
Ah*f
IfeM, IgO,
andTNP-
Ficoll 1.0
CH50=001
C3 = 0.5
ug/kg/day
Reference
Vecchi, 1980
Vos, 1973
Morris et al.,
1992
Morris et al,
1992
Dooley and
Holsapple,
1988
Holsapple, et
al., 1986
Thigpeo, et al.,
1975
Kerkvliet et al.,
1990a
House, et al.,
1990
Clark, et al.,
1981
White, 1986
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Table VI
Dose - Response Effects of TCDD on PFC response in C57BL/2N and B6C3F1 Mice
Species
C57BL/6N mice
male
C57BL/6J
Female
Male
BALB/cByJ mice
DBA/2J male
mice
B6C3F1 Mice
female
Dosing Regimen
Single i.p injection with
TCDD followed 5 days by
a single i.p. injection of 4
X 10* SRBC
single oral dose TCDD 2
days prior to single i.p.
injection 2.5 X 10'8 SRBC
single oral dose TCDD
administered orally 2 days
prior to injection with
SRBC
single oral dose TCDD
administered orally 2 days
prior to injection with
SRBC
Effect
% decrease in
spleenic plaque
forming cells/106
spleen and per
viable cells 4
days after SRBC
treatment
% decrease in
spleenic plaque
forming cells/106
spleen cells
% decrease
PFC/106 spleen
cells
% decrease
PFC/106 spleen
cells
% decrease in
spleenic plaque
forming cells/
106spleen and per
viable cells 4
days after SRBC
treatment
Dose
0 ug/kg
0.322
0.644
1.19
3.09
0 ug/kg
0.2
1
2
5
0 ug/kg
0.25
1.0
4.0
16.0
0 ug/kg
0.25
1.0
4.0
16.0
0 ug/kg
0.005
0.010
0.025
0.050
0.1
0.5
1.0
2.5
% Response
0%
6.3%
43.6%
84.8%
91.6%
PFCs/106 spleen
777
731
438
118
65
0%
30.4 %
52.6 %
67.4 %
78.3 %
0%
6.6 %
18.2%
79.6%
98.6%
0%
23.5%
32.6%
57.1%
87.9 %
0%
-7.8%
-18.6%
2.3%
13.1%
55.7%
62.7%
84.5%
92.5%
939
1012
1114
917
816*
416*
350*
146*
70*
NOAEL
N/A
1.0 ug/kg
0.025 ug/kg
LOAEL
N/A
4.0 ug/kg (all
doses were
decrease)
at all doses
significant
decrease
observed
0.05 ug/kg
Reference
Davis and Safe,
1988
Kerkvliet and
Brauner, 1990
Silkworth et al.,
1989
Silkworth et al.,
1989
Narasimhan et al.,
1993
E - 38
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Complement is another important factor in the immune response. White et al. (1986)
investigated the effect of TCDD on complement levels and activity in female B6C3F1 mice
administrated 0, 0.01, 0.05, 0.1, 0.5, 1.0, 2.0 ug TCDD/kg total dose administered over 14 day
period. As a measure of complement activity White estimated CH50 which is the amount of
treated mouse serum required to lyse 50% of the cells. There was a significant decrease in (CH50)
activity at all TCDD doses tested (Table V). A significantly reduced level of the C3 complement
component was found at 0.5 ug TCDD/kg. However, the C3 component was the only component
measured. It is concluded that several components may be affected, hence the decrease in CH50
activity is a more sensitive indicator of toxicity to the complement system that individual
components.
In summary, immune system may be profoundly affected by TCDD administration, and may
prove to be the most sensitive indicator of TCDD toxicity. A broad spectrum of immunotoxic
effects have been reported after TCDD administration in adult animals. Significant among these
are the increases in mortality in TCDD treated mice challenged with influenza (House et al. 1990)
and the decrease in complement (CH50) activity in which the LOAELs were both less that 0.01 ug
TCDD/kg. In addition the PFC assay has been found to be an extremely sensitive indicators of
toxicity (Narasimhan et al. (1993) found LOAEL = 0.05 ug TCDD/kg). Without studies comparing
similar endpoints across species the implications of this toxicity is difficult to assess. It has been
suggested that the susceptibility of the developing immune system is even greater than that of adult
animals. This and other effects on the developing fetus are very sensitive to TCDD administration.
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D. Teratogenic/Developmental Toxicity
The effects of TCDD on proliferative cellular events make it a particular concern in the
developing fetus. Although only 0.0005% of maternally administered TCDD reaches affected
tissues in the fetus of the C57BL mouse (Abbott and Bimbaum 1989), many effects including
thymic atrophy, cleft palate, and hydronephrosis are observed (Couture et al. 1990). The responses
to maternally administered TCDD are dose dependent, and based on studies of sensitive and
resistant strains of mice, appear to segregate with the Ah receptor locus (Poland and Glover 1980;
Silkworth et al. 1989).
D.I. deft Palate
Couture et al. (1990) summarized 23 studies which described the teratogenic and
developmental toxicity of TCDD. Of the 14 studies summarizing toxicity in mice. 13 listed cleft
palate and 8 listed hydronephrosis or renal congestion as toxic endpoints. None of the remaining
9 studies, performed in other species, list cleft palate as an endpoint, and only one study in
hamsters described hydronephrosis. All species have not been studied, yet it is believed that cleft
palate is unique to mice. Although hydronephrosis is believed to occur via a mechanism related
to cleft palate it has been reported to occur in other species including rat and hamster. Various
protocols have been used to evaluate cleft palate and hydronephrosis, the two most common fetal
abnormalities, and generally the response is dose dependent.
Mice, because of their unique sensitivity to cleft palate, have been the focus of the research
on the developmental and teratogenic effects of TCDD. Coutney and Moore(1971) observed an
increase in the incidence of cleft palate and kidney anomalies in three strains of
mice DBA/2J, C57BL/6J and CD-I after dams were administered 3 ug TCDD/kg/day s.c for 10
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days (GD 6-15). In the CD-1 mice a dose dependent increase in cleft palate incidence and kidney
anomalies was observed in offspring administered 1 and 3 ug TCDD/kg/day. Courtney and Moore
(1971) also observed that CD rats administered 0.5 ug TCDD/kg/day did not have offspring with
an increase in cleft palate. The authors concluded that the increase in cleft palate was specific to
the mouse, however, there was an increased incidence of kidney anomalies (not characterized) in
these rats. In addition, no significant maternal toxicity (characterized as decreased maternal weight
gain or increase relative liver weight) was observed in these CD rat or CD-I mice. The effects of
TCDD administered orally 0,0.001, 0.01,0.1, 1.0 and 3.0 ug TCDD/kg/day on gestation days (GD)
6-15 on fetal development in Carworth CF-1 mice was investigated by Smith et al.(1976) (Table
V). At the 1.0 and 3.0 ug TCDD/kg/day dose a statistically significant increase in cleft palate was
observed. At the highest dose an increase in dilated renal pelvis was observed. Since there was a
response at only two doses it is difficult to conclude if there is a dose-response relationship using
this protocol. The NOAEL for developmental /teratologic effects was reported as 0.1 ug
TCDD/kg/day (1.0 ug/kg total dose), and LOAEL was 1.0 ug TCDD/kg/day (10 ug/kg total dose),
doses comparable to that of Courtney an Moore (1971).
Courtney (1976) also investigated the teratogenic effects of TCDD using oral and
subcutaneous routes of administration . Pregnant CD-I mice were administered 0, 25, 50, 100, and
200 ug TCDD/kg/day using both routes of administration, and additional dose of 400 ug TCDD/kg
P.O. for 10 days (GD 6-15). Maternal toxicity, evident in all dose groups, was characterized by
a decrease in maternal body weight gain, an increase in relative liver weight, and in some cases
marked edema, Parameters used to assess fetotoxitiry were number of live fetuses/litter, number
of abnormal fetus/litter, percent cleft palate, percent kidney anomalies (characterized as
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hydronephrotic kidneys), and percent pups with a club foot. A dose-dependent increase in percent
cleft palate and percent hydronephrosis was detected in all groups administered TCDD orally,
except at the highest dose where it was observed that the percent kidney anomalies decreased
probably because of the extent of maternal toxicity at this dose. (Table VII). The mice
administered TCDD s.c. had a response (increase cleft palate) close to maximal at the lowest dose
tested, although the same dose administered orally resulted in a minimal but detectable response.
Additionally, the comparison of 2,3,7,8-TCDD to 1,2,3,4 TCDD and OCDD shows 2,3,7,8 TCDD
to be the most potent fetotoxin of these congeners tested. The doses in this study were extremely
high and do not reflect the sensitivity of cleft palate and hydronephrosis as toxic endpoints
(Courtney 1976). The CD-I, CF-1, C57BL/6J and the DBA mice clearly have different sensitivity
to TCDD induced cleft palate. An incidence of 19% cleft palate was observed in CD-I mice
administered 25 ug TCDD/kg/day orally GD 7-15, while only 1.0 ug TCDD/kg/day in the CF-1
mouse (the most sensitive of the species tested) resulted in 21% cleft palate in offspring (Smith et
al. 1976). C57BL/6J administered 4 ug TCDD/kg/day and DBA administered 8 ug TCDD/kg/day
had 47% and 27% cleft palate incidence respectively (Table VII) (Silkworm et al. 1989a). After
the initial reports that TCDD was teratogenic in mice, the focus of research was to characterize
these fetal anomalies and on finding a sensitive period during gestation when these effects occurred.
Unfortunately, because of the study designs, fetal tissue dose and distribution was not obtained,
therefore, quantificative assessment of dose response relationships can not be done.
To determine if there is a sensitive period that cleft palate or hydronephrosis occurred,
Bimbaum et al. (1989) and Weber et al. (1985) administered TCDD to mice on GD 10 or 12. In
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the offspring of pregnant C57BL/6N mice administered a single oral dose on GD 10 of 0, 12, 17,
and 22 ug TCDD/kg, a significant and dose-dependent increase in percent of fetuses with cleft
palate was observed. Maternal toxicity characterized as decreased weight gain and increased
relative liver weight was significant at the 17 ug TCDD/kg dosage level.(Weber, et al., 1985).
However, kidney damage which was significant at all levels, was not clearly dose dependent
because of a near maximal response at the lowest dose. These doses are equivalent to the doses
administered for 10 days to the CF-1 (Smith, et al. 1976) mice which produced cleft palate and
hydronephrosis. Similar doses were also used by Bimbaum et al. (1989) to assess the teratogenic
effects of TCDD in C57BL/6N mice. Pregnant mice were administered a single oral dose of 0, 6,
12, 15, or 18 ug TCDD/kg on GD 10, or doses of 0, 6, 9, 12, 15 ug TCDD/kg on GD 12 (Table
VII). On GD 10 and GD 12 maternal toxicity characterized as increase relative liver weight was
observed at all doses. A statistically significant dose-related increase in cleft palate was observed
at 12 ug TCDD/kg on GD 10 and at 9 ug TCDD/kg on GD 12. Maximal cleft palate induction was
observed in C57BL/6N mice administered TCDD on gestation day 12. The observed incidence was
about 60% greater than the incidence observed at the same dose (15 ug TCDD/kg) on GD 10.
Couture et al. (1990a) reported, in an experiment using single dose administration in C57B1/6J
mice, that on GD 10 or 12 a LOAEL of 12 ug TCDD/kg for cleft palate, and a LOAEL of 3 ug
TCDD/kg on GD 6, 8, 10, 12 or 14 for hydronephrosis.
To determine when palate formation was most sensitive to TCDD, Pratt et al.(1984)
administered a single s.c. dose of 100 ug TCDD/kg to C57BL/6J mice on days 3-13 of gestation.
Greater than 95% of the fetuses from dams administered TCDD on days 8,9,and 10 had cleft palate.
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The maximum response was observed on day 10. Histological examination of palate formation in
controls revealed significant cell death of medial epithelial cells resulting in palate closure. In the
TCDD treated mice this cell death was not observed. This apoptosis (programmed cell death) is
apparently inhibited by an Ah receptor mediated mechanism (Pratt et al. 1984). The role of the Ah
receptor in the teratogenic effects of TCDD after repeat dose administration was investigated by
Silkworth et al. (1989a). By comparing C57BL/6J and DBA/2J mice Silkworth determined that
the incidence of cleft palate segregated with the Ah receptor. Pregnant C57BL/6J mice were
administered 0, 0.5, 1.0, 2.0, and 4.0 ug TCDD/kg/day during gestation day 6-15 (total dose of 0,
5, 10, 20, and 40 ug TCDD/kg). DBA/2J were administered up to 8.0 ug TCDD/kg/day.(Table V)
The most sensitive indicators of maternal effects in C57BL/6J mice was an increase in relative liver
weight and a decrease in relative thymus weight seen at all doses. A significant increase in relative
liver weight and a decrease in the relative thymus weight was seen in the DBA mice at the 2.0 ug
TCDD/kg/day and 8.0 ug TCDD/kg/day dose level, respectively, indicating a 4-8 fold decrease in
TCDD sensitivity in the DBA /2J mice consistent with other studies. In both strains of mice there
was a statistically significant increase in percent cleft palate/dam at the highest doses 4 and 8 ug
TCDD/kg/day. These doses were 4-8 times higher than the LOAEL for maternal effects. This
study indicated that the fetal abnormality (cleft palate) seen in DBA and C57BL/6J mice was a less
sensitive indicator of TCDD's effects than the measures of maternal effects, but that percent of
offspring with hydronephrosis was more sensitive being significantly increased in both strains at
0.5 ug TCDD/kg/day. This is not true with all mice strains. In the CF-1 mice studied by Smith
et al. (1976) using a similar protocol, the teratogenic effects (cleft palate) of TCDD proved to be
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a more sensitive indicator than maternal effects characterized as increase relative liver weight.
Most importantly this comparison of C57BL/6J and DBA/2J points out that in these two mouse
strains a statistically significant increase in hydronephrosis is observed at the same dose suggesting
this may be an equitoxic measure of maternal dosing of TCDD, however the magnitude of this
response is not equivalent at the LOAEL, 5 ug/kg.
Identifying the window in which maximal cleft palate occurred enabled Abbott and
Bimbaum (1989) to investigate the mechanism by which TCDD disrupted medial epithelial cell
terminal differentiation. Levels of EGF receptor were found to be a critical difference in the
development of the embryonic palate in the presence of TCDD. Using the C57BL/6N, Abbott and
Bimbaum, administered a single oral dose of TCDD on GD 10 or GD 12. Four days post dosing
measurements of thymidine incorporation, concentration of EGF receptor, and degree of palate
closure was shown to be markedly increased. It was concluded that the characteristic TCDD
mediated £l eft-pal ate formation is due to a defect in the fusion of the palatial shelves. Although the
shelves meet, the medial cells which usually die during palate closure continue to proliferate under
the influence of TCDD. The mechanism by which this occurs is shown to be related to the increase
expression of the EGF receptor in the medial epithelium on GD 10 and 12, which coincides with
the sensitive time for cleft palate formation (Abbott et al. 1989a).
The incidence of cleft palate in mice correlates well with their sensitivity to Ah receptor
mediated toxicity. Poland and Glover (1980) administered a single s.c. dose of 30 ug TCDD/kg
to pregnant dams of 10 different mice strains. Four of the five TCDD sensitive strains had a 54-
95% incidence of cleft palate, whereas all the strains with a low sensitivity to Ah receptor mediated
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toxic effects had a < 4% incidence of cleft palate. Poland and Glover also investigated the strain
differences using C57BL/6J, DBA/2J and their hybrid cross B6D2F1/J mice. Mice were
administered a single s.c. dose of 0, 3, 10, or 30 ugTCDD/kg on GD 10. A dose-dependent
increase in cleft palate/live fetus was seen in the C57BL/2J mice (Table VII). In the other two
strains cleft palate was only seen at the highest dose indicating a segregation with the Ah receptor.
Dencker and Pratt (1981) investigated the role the tissue concentration of the Ah receptor
plays in cleft palate formation using two strains of mice C57BL/6J, and AKR/J (a TCDD resistant
strain). The results of this comparison indicate that the concentration of Ah receptor is greater in
the maxillary process (the region containing palate shelves) of the C57BL/6J mice than in the
AKR/J mice. In addition,the embryonic C57BL mice maxillary process contained a higher
concentration of Ah receptor (measured on GD 13) than the embryonic brain, liver, limb buds or
skin. The Ah receptor without doubt has some role in cleft palate formation probably involving
the altered transcription of a protein involved in palate epithelial cell division or apoptosis (Dencker
and Pratt 1981).
Experiments using cultured embryonic palate tissue have enabled researchers to observe
palate development more closely and to compare very sensitive endpoints in mouse, rat and human
tissue. Abbott and Bimbaum (1989) established this methodology using C57BL/6N mice palate
shelves. An extensive database of in vivo data has made this comparison possible. GD 12 mice
palate shelves were exposed to 0, 10'13, 10'12, 10'11, 5 X 10'11, 7.5 X 10'", 10'10, and lO'9 M TCDD
(Table Vm for dose equivalents). A dose-response was observed in percent of palates not
undergoing apoptosis (an endpoint that indicates the medial epithelial cells continue to grow
resulting in cleft palate formation). The response increases until reaching the cytotoxic
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concentration of 7.5 X 10""M. The EGF receptor, measured by immunochemistry, increased in
medial epithelial cells at 10"11, 5 X 10"11, 10~10, and 10~9 M (only doses reported). An increase in
cell proliferation (thymidine incorporation) was also found to increase significantly. Table Vin
points out the relationship between in vitro culture concentrations and administered dose. These
calculations are based on the observation that 0.0003% of a 30 ug TCDD/kg dose administered to
a pregnant C57B1/6N mouse is distributed to the palate tissue (Abbott and Bimbaum 1989). The
concentrations that produce cleft palate in vitro correlate well with the doses that are found in vivo
to produce cleft palate. For example, Bimbaum et al. (1989) found an increase cleft palate after a
single dose administration of 12 ug TCDD/kg. Abbott and Bimbaum (1989) found a significant
increase in EGF receptor levels in medial epithelial cells at all dose above 10"11 (equivalent to 20
ug TCDD/kg). Decreased program cell death was observed at dose equivalents of 0.2 ug TCDD/kg
to 200 ug TCDD/kg.
Rats are not as sensitive to cleft palate formation as are mice. This appears to be because
TCDD is fetotoxic and maternally toxic. Therefore, fetal death may occur at lower doses than cleft
palate formation. The pups of Holtzman rats administered 0, 1.5, 3, 6, and 18 ug TCDD/kg on GD
10 developed cleft palate only at the 18 ugTCDD/kg dose. A dose-dependent increase in percent
fetal mortality and decrease in fetal thymus weight was observed in these rats. Although cleft
palate was not observed at the low doses, the percent of fetuses per litter affected by intestinal
hemorrhage was significantly increased in a dose-dependent manner at all doses (Olson 1989,
1990). Other endpoints were also found in other animal species, guinea pigs were treated orally
with 0, 0.15 and 1.5 ug TCDD/kg on GD 9 and a significant increase in percent fetal mortality
occurred at the 1.5 ug TCDD/kg dose (Olson 1990). The developmental toxicity of Golden Syrian
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Hamsters was investigated (Olson 1990) after administration of 0. 1.5, 3, 6, or 18 ug TCDD/kg on
GD 9. A significant dose related increase in relative liver weight was observed in dams at all
doses. In addition, a significant dose-dependent increase in neutrophilic metamyelocytes,
neutrophilic band cells, total liver P450 activity, and EROD activity was observed at all doses. A
dose-dependent decrease in liver estradiol 2-hydroxylase activity was observed in the dams, which
was significant at the 3 ug TCDD/kg dose level. Fetal liver EROD and total P450 also increased
at all doses, but statistical significance was not determined due to small sample size. Fetal thymus
area decrease and kidney congestion was observed at all dose levels. A dose-dependent increase
in the incidence of hydronephrosis was observed and was statistically significanir at the 3.0 ug
TCDD/kg (maternally administered) dose level (Olson et al. 1990). The hamster is the least
sensitive species tested to the acute toxic effects of TCDD (see Table II), but it is extremely
sensitive to the developmental abnormalities observed. Kidney congestion and hydronephrosis are
observed after a single dose of 1.5 ug TCDD/kg which is equivalent to the toxicity seen in the CF-
1 mouse (Table V). This observation raises questions regarding the ability of acute toxicity to
predict tissue specific sensitivity.
The in vitro palate culture provides a method to examine palate development in rats without
the associated maternal or fetotoxicity of TCDD. Abbott and Birnbaum (1990) culturing GD 14
palates of F344 rats with TCDD, found a dose-dependent increase in the number of palates in
which the medial peridermal cells did not go through programmed cell death (Table VIII). Since
data on distribution of TCDD to the rat fetal palate has apparently not been reported it is not
possible to determine the comparable administered dose. It is valuable to compare the tissue dose
response in several species. The mouse is about 100-1000.
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Table VH TERATOGENIC EFFECTS: DOSE RESPONSE
Carwortli CF-1 mice
CD-I mice
CD-I mice
C57BL/6J mice
C57BL/6J mice
DBA/21 mice
C57BL/6N mice
I. C57BL/6J
2. B6D2F1AI
3. DBA/2J mice
Golden Syrian
Hamsters
C57BL/6N Mice
oral dose administered daily
GD 6-15
1. daily oral dose
administration
2. daily s.c. dose
«1mim"™tii>n
GD 7-16
1. daily oral dose
administration
2. dairy s.c. doae
GD 7-16
•ingle oral dose administered
onGD 10
daily administration GD 6-15
daily administration GD 6-15
single oral admmiatration on
GD 6,8,10,12
single B.C. administration GD
10
single oral doae GD 10
•ingle oral dose administered
to dam on GD 10 or GD 12
Effect
percent liters with cleft
palate
% cleft palate /total
fetus
% kidney
anomalies/total fetus
1. percent liters with
deft palate
2. percent fetuses with
deft palate
1. % cleft palate per
2. % hydronophrOBU
1.% deft palate per
dam
2.% hydronephrosU
1. % fetuses/litter with
hy OTODOphro BIB
2. % fetuses/litter with
cleft palate
% cleft palateilive fetus
]. % kidney congestion
2. % hydronephrOBis
3. % fetal mortality
increase cleft palate
incidence
Dose
0 ug/kg/day
0.001
001
0.1
1
3
0 ug/kg/day
50
100
200
400
0 ug/kg/day
50
100
200
400
0 ue,/kg tot doae
0.01
0 1
1.0
10
30
0 ug/kg tot dose
250
500
1000
2000
4000
0 ug/kg tot dose
250
500
1000
2000
4000
Oug/kg
12ug*g
22 UK/kt
0 ug/kg/day
0.5
1.0
2.0
4.0
0 ug/kg/day
0.5
2.0
4.0
S.O
Dose
Oug/kg
6
9
12
24
0 ug/kg Total dose
5.0
10
20
40
0 ug/kg Total dose
20
40
80
1. GD 6 2. GD 6
25.7% 0.0%
96.2% 0.0%
100.0% 1.0%
100.0% 0.0%
100.0% 3.5%
100.0% 40.3%*
0 ugAg
3ug/fct
10 ng^
30w/ke
0 ug/kg
3.0
6
18
0 ug/kg GD 10
6
12
15
18
1. 0
9.7*
28.5*
83.6*
91.1*
0 ug/kg GD 12
6
9
12
15
Response
0 % palate
5%
0 %
6%
21 %*
71 %*
0 l.oral
3%
19%
66%
100%
100%
0 l.oral
34%
72%
71%
100%
50%
1. 0%
40%
82%
100%
1. 0%
0%
2.23%
2.33%
47.23%*
1. 0%
0%
3.51%
3.47%
27.24%*
1. GD 8 2. GD 8
27.0 0.3
90.0 0.0
-
.
100.0* 9.7*
100.0* 93.4*
0% kidney
0%
0%
0%
5%
28%*
0 2 s.c.
82%
79%
85%
100%
0 2. B.C.
53%
58%
95%
38%
2. 0%
8.7%*
43.9%*
77.6%*
2. 0%
39.92%*
72.79%*
74.23%*
91.83%*
2. 0%
4.95*
13.16*
36.89*
60.47*
1. GD 10 2. GD 10
18.6 0.0
89.2 0.0
-
.
100.0* 53.9*
100.0* 98.8*
1. 0% 2. 0% 3. 0%
000
300
54 13 2
2. 1.6%
11.2%
15.3%*
34.5*
0
0 GD 10 incidence
0
20.86*
49.30*
70.12*
3. 7%
10.8%
7.7%
8.7%
58.2%
0 GD 12 incidence
1.43
19.91*
51.45*
77.91*
NOAEL
0.1 ug/kg/day
N/A
N/A
if*
1. 40 ug/kg
2. Sugtg
1. GD 12 2.
6.4% 0.0
97.9%* 1.1
-
-
100.0%* 76.0%*
100.0%*
100%*
GD 10 = 6 ug/kg
GD 12 = 6 ug/kg
LOAEL
1 0 ug/kg/day
10 ug/kg
N/A
N/A
12 ug/kg
4 ug/kg/day
8 ug/kg
GD 10=12
GD 12=9 ug/kg
Reference
Smith et al.,
1976
Courtney, 1976
Courtney, 1976
Weber et al.,
1985
Silkworth et
al., 1989a
Silkworth et
al., 1989a
Couture et al.,
1990s
Poland and
Glover, 1980
Olson, 1990
Birnbaum, et
al., 1989
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times more sensitive when examining the endpoint of programmed cell death. Hence, it may be
concluded that tissue distribution of TCDD is not the rate limiting factor for species sensitivity.
Human effects of TCDD relating to pregnancy are difficult to determine, primarily because of the
confounding effects of multiple potential etiologic agents, a relatively high background of birth
defects, and a lack of robust exponential data. The cultured human palate allows direct comparisons
between similar endpoints that have been evaluated in the rat and mouse species. Abbott and
Bimbaum (1991) cultured human palatine tissues GD 52, 53, and 54 for 3-4 days. Evaluation of
palate growth and development revealed that 100% of the palate medial epithelial cells continued
to proliferate and scheduled programmed cell death did not occur at 10"8 and 10"9 M TCDD
concentrations. Since only three concentrations were reported, a dose-response relationship is
difficult to evaluate although there was an increased response observed between the 10"11 and the
10"8 M concentrations. Direct comparison between the human and mouse palate demonstrates that
the concentrations at which 25% or 100% of palates were affected (programmed cell death did not
occur) were 500 times more sensitive in the mouse than in the human palate (See Table VIII).
These in vitro observations found that the human palate is sensitive to cleft palate formation
through the same mechanism seen in mice; changes in growth factors (ie; EGF, and TGF's) that are
involved in the mechanism of altering programmed cell death in the medial epithelial cells of the
palate. The effects of TCDD on regulation of these growth factors causes epithelial cells to
continue to proliferate resulting in palatine shelves that do not fuse normally.
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Table VIII: Comparison of Effects of TCDD on In Vitro Palate Cultures
Species
C57BL/6N mice
F344 rats
Human Palate
Cultures
Gestation Days
OD 12
GD 14-15
GD 52, 53 and
54
Concentrations
Control
1 X 10'"M
1 X 10'1!M
5X 10'12M
1 X 10-" M
5 X 10-"M
7 5X 10-UM
1 X 10-10M
Control
5 X l
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D.2 Hydronephrosis
Hydronephrosis is the second well characterized fetal anomaly observed commonly in
TCDD treated mice. In contrast to cleft palate which is species and tissue specific for TCDD
treated mice, kidney toxicity occurs after pre and postnatal treatment, and occurs in several species.
The kidney lesions have been characterized as an increase cell proliferation of the epithelial cells
of the uteric lumen. Cell proliferation causes obstruction/constriction of the ureter resulting in
decreased urine outflow and "true hydronephrosis" (Abbott et al. 1987). Uteteric epithelial
hyperplasia has been observed in mice administered TCDD both pre and postnatally, in adult guinea
pigs (McConnell et al. 1978a) in Rhesus monkeys (Gupta et al. 1973) and reported as kidney
anomalies,but not further characterized in CD-rats (Courtney and Moore 1971).
TCDD administered to pregnant C57BL/6 mice (GD 10-13) at 1.0, 3.0 and 10.0 ug
TCDD/kg/day (total dose 4, 12, 40 ug/kg) resulted in a dose dependent increase in the incidence
of early stages of hydronephrosis and "renal papilla which are markedly reduced in size, or non-
existent, in a few cases resulting in an enlarged renal pelvis." These lesions were found
predominantly in the right kidney. The lesions were also found in pups exposed postnatally and
were qualitatively similar to lesions found in fetuses exposed in utero (Moore et al. 1973). Fetal
hydronephrosis is a very sensitive indicator of TCDD toxicity with an incidence of 89-98%
observed after administration of a single dose of 3 ug TCDD/kg during GD 6, 8,10, or 12 (Couture
et al. 1990a).
In mice hydronephrosis appears to be the most sensitive indicator of fetotoxicity. Silkworth
et al. (1989a) demonstrated this by administering C57BL/6J and DBA/2J mice 0. 0.5, 1, 2, and 0,
0.5, 2,4,8 ug TCDD/kg/day respectively during GD 6-15. A statistically significant increase in
hydronephrosis was observed at the lowest dose tested 0.5 ug TCDD/kg/day (total dose 5 ug
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TCDD/kg) in both strains of mice with no NOAEL observed for this endpoint. In contrast cleft
palate incidence was observed to be statistically significant (Table V) at the 4 ug TCDD/kg/day (40
ug/kg total dose) in C57B1/6J mice and at 8 ug TCDD/kg/dav (80 ug/kg total dose) in the DBA/23
mice. The sensitivity of these endpoints may be related to the window of sensitivity in fetal
development that these anomalies can occur. Studies determining the mechanism of each of these
•
lesions have helped to understand these differences.
Couture et al. (1990a) administered a single oral dose to C57BL/6N mice on GD 6, 8, 10,
12 or 14 to investigate the point in gestation that TCDD produced maximal toxicity. Unfortunately,
all administered doses (3, 6, 9, 12, 24 ug TCDD/kg produced a statistically significant increase in
the incidence of hydronephrosis from GD 6-14 so a sensitive time period was not determined
(Table VII). The incidence appeared to decrease after day 14. Couture-Haws et al. (1991, 1991a)
examined sensitivity to TCDD administered postnatally in dams administered 0, 6, 9, 12 ug
TCDD/kg on postnatal day 1, 4, 8 and 14. All doses significantly increased the incidence of
hydronephrosis on day 1 and 4. No increase was seen on PND 8 or 14. No maternal toxicity was
observed at any dose. These results indicate that the neonatal kidney is affected by the level of
TCDD available in milk.
The mechanism is similar to that for cleft palate since altered regulation of cell growth and
death are responsible for cleft palate formation in the mouse. This is mediated through a sustained
level of EGF production not found in control animals. TCDD induced hydronephrosis in the mouse
is characterized as an increase in uteric epithelial cells causing closure of the ureter. The continued
proliferation of these epithelial cells as measured by thymidine incorporation is found to correlate
with a sustained level of EGF receptors (Abbott and Birnbaum 1990a).
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Cultured ureteric epithelial cells derived from GD 10 or 12 mouse fetuses had an incidence
of 91-100% hyperplasia after exposure to 1 x 1Q-10 M (32 pg/ml) TCDD, a 4 fold increase over
control. In addition, intense EOF receptor staining (immunohistochemistry) and increase thymidine
uptake(indicating cell growth and proliferation) was observed in most of these sections (Abbott and
Birnbaum 1990a). Based on the LOAEL of hydronephrosis seen in C57BL/6N mice in vivo (3
ug TCDD/kg) the tissue dose seen in the ureter of these mice is equivalent to 0.3 pg (Abbott and
Birnbaum 1990a)
When comparing TCDD toxic endpoints, the effects on fetal development do not appear to
be the most sensitive in all of the species studied. However, without information on tissue
distribution in all species this conclusion is somewhat limited.
Hydronephrosis is a more sensitive fetotoxic endpoint in the mouse and in the hamster
occurring at doses that are not overtly toxic to dams. Cleft palate on the other hand appears to
be a TCDD-related teratogenic effect unique to the mouse, but analysis of the mechanism reveals
that the interference with programmed cell death by TCDD is not unique to the mouse. Cleft palate
is found at doses where maternal effects are observed (characterized as decrease weight gain or
increase relative liver weight) in C57BL and DBA mice, but cleft palate is not accompanied by
maternal effects in the CD-I or CF-1 mice, although these strains have not been thoroughly
examined.
D.3 Other
Giavani et al. (1982, 1982a, 1982b) attempted to characterize cleft palate and kidney
abnormalities in rabbits and rats. The most frequently observed lesion among litters of pregnant
rats administered 0. 0.125, 0.5 and 2 ug TCDD/kg/day during GD 1-3 was cystic kidneys. This
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was not statistically significant and was observed at doses that were fetotoxic and maternally toxic,
but it does demonstrate that kidney development is a target of TCDD toxicity when administered
early in gestation (Giavini et al. 1982a). Pregnant rabbits administered 0. 0.1, 0.25 0.5 and 1 ug
TCDD/kg/day GD 6-15, TCDD had a significant increase in post-implantation losses reaching 100%
at the highest dose. At maternally toxic and fetotoxic doses, a small increase in kidney
abnormalities characterized as hydronephrosis and double kidney formation was observed (Giavini
et al. 1982b). These observations indicate that the development of kidney anomalies, although
found, are not the most sensitive indicator of fetotoxicity in the rabbit or the rat.
Giavinni et al. 1982 investigated the effects of TCDD on the implantation characterized
as mean corpus lutea, mean number of implantation, and percent preimplantation losses in
Sprague-Dawley rats. Dams were administered 0, 0.125, 0.5, and 2.0 ug TCDD/kg/day for 3 days
GD 1-3 (total doses 0, 0.375, 1.5, and 6 ug/kg). Fetotoxicity measured as fetal weight on GD 21,
and maternal toxicity measured as weight gain was also evaluated. Maternal toxicity was observed
at the highest dose only, and fetotoxicity was observed in the 0.5 and the 2 ug TCDD/kg/day dose
group. The results indicate that no significant relationship between TCDD at doses up to 2 ug
TCDD/kg/day and pre or post implantation losses, mean corpus lutea, or number or implantation
sites. Fetoxicity was found to be related to maternal toxicity (Giavini et al. 1982) as suggested by
Khera (1987), who postulated that maternal toxicity may be responsible for the fetotoxicity found
after administration of certain chemicals. In Sprague-Dawley rat (Sparschu et al. 1971)
administered 0.03, 0.125, 0.5, 2.0 and 8.0 ug TCDD/kg GD 6-15 the most sensitive indicator of
fetotoxicity was decrease fetal weight and gastrointestinal hemorrhage LOAEL = 0.5 ug
TCDD/kg/day. This occurred at a dose that was not maternally toxic although decrease fetal weight
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may be associated with undetected maternal toxicity. Gastrointestinal hemorrhage was also the
most sensitive indicator of fetal toxicity observed in Holtzman rats by Olson (1990). Dams
administered 0. 1.5, 3, 6, and 15 ug TCDD/kg single p.o. dose on GD 10 had a statistically
significant increase in percent fetuses per litter affected at all doses administered.
Evaluation and comparison of reproductive endpoints are complicated by the predominance
of species specific effects and the unique problems of fetal/tissue distribution. Understanding the
underlying biochemical mechanisms responsible for adverse reproductive effects is essential to
proceeding with comparative analysis. For cleft palate these mechanisms have been thoroughly
investigated and any risk assessment will be enhanced by incorporating this information.
E. Hoimone Effects
In the classic TCDD carcinogenicity study (Kociba et al. 1978) it was observed that female
rats had a dose-dependent increase in liver cancer and a dose-dependent decrease in the incidence
of spontaneous tumors in the uterus, breast and pituitary. These observations, coupled with reports
suggesting that the mechanism of TCDD toxicity was mediated through a nuclear hormone-like
receptor mechanism (Poland and Knutsen 1982) led researchers to investigate the role of hormones
in modulating TCDD toxicity. Early work focused on thyroid hormone and the role of the thyroid
in the toxicity of TCDD (McKinney et al. 1985; Lucier, this document) The possible role of
estrogens and estrogen receptor in the toxicity of TCDD was recognized by several research teams
(Gallo and Umbreit 1988). The estrogen receptor and the Ah receptor are both ligand-specific and
response element specific (Denison 1991, Green and Chambon, 1991). Experimental evidence
(Poland and Glover, 1980; Romkes and Safe, 1988) indicates that TCDD does not bind the
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estrogen receptor(ER) nor does estradiol(E2) bind to the Ah receptor. However, there may be
direct competition for factors that modulate ligand-receptor binding, translocation, and response
element binding ie., heat-shock protein 90 (Perdew 1988, Green and Chambon, 1991). It may be
the effect of TCDD on the concentration of uterine and hepatic ER that results in modification of
estrogens sensitive tissues. In addition, subsequent alteration in estradiol levels from induction of
CYPIA1 and CypIA2 may induce organ specific changes (DeVito et al. 1992a). Altered estradiol
levels have also been implicated in a reduced immune response (Luster et al. 1985), and both an
increase and decrease in tissue specific cancers. The relationship between TCDD, estradiol and
cancer detailed in the Carcinogenicity Chapter of this document(Lucier) is still not fully understood.
The dose response relationships between TCDD and altered Estrogen Receptor(ER), and/or
changes in serum estradiol can result in a variety of responses. Evaluation of TCDD in non-human
primates revealed a decrease in ability to conceive and an increase in aborted fetus. Rhesus
monkeys administered a continuous dose of 50 ppt TCDD in the feed (total dose 1.8 ug TCDD
approx equivalent to 8.6 ng TCDD/kg/day) had increases in serum estradiol and progesterone.
Abnormal hormone levels may have been responsible for only 2/8 treated animals conceiving versus
8/8 in the control group (Allen et al. 1979). McNulty 1985 reported a dose-related increase in fetal
loss at much higher levels: 1.0 ug TCDD/kg and 5.0 ug TCDD/kg total dose (administration
divided into 9 intragastric doses GD 20-40) which was accompanied by maternal toxicity. No
hormone levels were measured in these Rhesus monkeys so its relationship to fetal loss can not be
determined.
The effects of TCDD mediated alteration in hormone levels is not restricted to the reduced
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ability to conceived or fetal loss. A decrease in uterine and body weight was observed in
prepubescent female guinea pigs, rats and hamsters administered a single 4, 50, 400 ug TCDD/kg
(an equitoxic) dose of TCDD (Hruska and Olson 1989). A statistically significant decrease in
uterine weight occurred in guinea pigs and Sprague Dawley rats. Only a slight (not significant)
decrease in uterine weight was observed in the hamster. In addition to uterine weight changes,
increases in the uterine estrogen receptor density (measured using saturation analysis, followed by
estimation of receptor density [Bmax] from Scatchard plot) were observed at above administered
concentrations. The largest change in uterine ER density (Bmax) was in the rat, followed by the
hamster, then the guinea pig. Neither the change in uterine weight nor the constitutive levels of ER
(Bmax) predicted the increase in uterine receptor density. At these doses TCDD decreased the
hepatic ER concentration significantly in rats and guinea pigs but not in hamsters. The species
differences in TCDD mediated changes in hepatic ER may be partially explained by the constitutive
levels of these receptors. In the three species studies, the constitutive level of ER in the liver
predicted the magnitude of the decrease in hepatic ER. The differences in uterine and hepatic ER
responses suggests that there may be other factors not found in the liver, but present in the uterus
that modulate this response.
Species differences in many endpoints can be directly related to their differences in hepatic
enzyme (P450) induction. Lin et al. (1991) investigated this relationship as it pertained to ER using
C57BL/6 responsive strain (Ahbb) and the non-responsive strain (Ahdd). Mice were administered
0, 0.3, 1, 3, 10, 30, 100, and 300 ug TCDD/kg single oral dose. The study found that the
magnitude of the decrease in hepatic ER in the Ahbb mice was significantly more than that of the
non-responsive mice. Although the dose (LOAEL) at which the decrease in hepatic ER was
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observed was similar in both strains, approximately 0.06 ug TCDD/kg.
In intact and ovariectomized Long Evans rats (Romkes et al. 1987) a dose dependent
decrease in hepatic and uterine ER was observed after administration of 20 ug TCDD/kg or 80 ug
TCDD/kg TCDD. This study also showed that the administration of estradiol with the TCDD
further decreased the ER in the uterus, but did not have an effect on liver ER. This again indicates
other factors (primarily hormonal) are responsible for the increase in sensitivity of the uterus to
TCDD.
The receptor mediated toxicity of TCDD has been discussed in detail in this document
(Whitlock chapter). Clearly TCDD's effective concentration and mechanism of toxicity are
comparable to that of nuclear hormones. This does not imply that there is a
direct interaction between TCDD and hormone receptors. In a study in which female Long Evans
rats were administered a single i.p injection of 0, 20, 40, and 80 ug TCDD/kg a dose-dependent
decrease in the hepatic and uterine ER and progesterone receptors (PR) was observed (Romkes and
Safe 1988). This study also demonstrated that TCDD has no binding affinity for the ER and PR
which suggests that TCDD does not exert its antiestrogenic properties through competition with the
ER or PR.
DeVito et al. (1992a) has recently demonstrated that down-regulation of the hepatic and
uterine estrogen receptor (ER) occurs at doses that increase CypIAl (Table IX) but there are no
accompanying changes of serum estradiol or increases in 2-OH estradiol (DeVito et al., 1991).
Graph E (using a Michaelis-Menten and a Power Law function) depicts the increase in hepatic Ah
activity at doses that decrease hepatic estrogen receptor. Although the Michaelis-Menten function
can be fit to this data it is the Power Law function that provides insight into this relationship. The
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similarity in the apparent kinetic order ("g" value) for the percent decrease in hepatic estrogen
receptor and percent increase in AHH activity suggests a common mechanism of action.
Supporting this is the information that this TCDD induced down-regulation of the ER segregates
with Ah receptor (Lin et al. 1991). Other investigators have suggested that, in the MCF-7 breast
cancer cell line, that TCDD induced increases in P450 enzymes increase estrogen metabolism
resulting in a decrease in the ER (Gierthy et al. 1988). This is significant because it demonstrates
that TCDD may induce decreases in estradiol which are Ah receptor mediated, and suggests that
clinical effects could result from a decrease in estradiol not directly from TCDD.
It has been shown that c-fos is a nuclear regulatory protein that plays a role in the control
of estradiol induced cell proliferation (Astroff et al 1991). At very low i.p. administered doses
(0.064, 0.128 ug TCDD/kg), reduced constitutive levels of uterine c-fos mRNA were observed in
Sprague-Dawley rats, however coadministration of exogenous estradiol increased the levels of c-fos
mRNA (Graph F-A). Although TCDD does not interact directly with ER, it may compete with
the activated estrogen receptor at the level of transcription. The power law function applied to the
TCDD induced decrease in c-fos resulted in a functional fit with R2 = 0.99, "g" = 0.66 (Graph F-B).
This is consistent with the "g" values described later in this chapter that were computed for the
TCDD induced increase in EROD activity (Abraham et al, 1988, "g" = 0.66) and benzo(a)pyrene
hydroxylase (Kitchin and Woods, 1979, "g" = 0.77), suggesting the mechanisms by which TCDD
alters c-fos mRNA may be, in part, common with those mechanisms involved in P450 induction.
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Species
Female Long
Evans Rats
Female Long
Evans Rats
Female S-D
rats
Mice CD-I
Dosing
Regimen
single ip
single ip
single ip
single ip dose
Table IX HORMONE EFFECTS: DOSE RESPONSE
Effect
Uterine
% decrease cytosolic ER
% decrease nuclear ER
% decrease cytosolic PR
% decrease nuclear PR
Hepatic
% decrease cytosolic ER
% decrease nuclear ER
% decrease cytosolic PR
% decrease nuclear PR
% decrease in c-fos
mRNA units
% decrease hepatic ER
% decrease uterine ER
% increase hepatic AHH
Dose
Oug/kg
20
50
80
Oug/kg
20
50
80
0 ug/kg
0.016
0.064
0.128
Oug/kg
1
3
10
30
Response
C-ER
0 %
16.4 %
23.6 %
46.2 %
C-ER
0%
4.0 %
7.5 %
54.3 %
N-ER
0%
12.4 %
32 0 %
68.3 %
N-ER
0%
12.3 %
45.5 %
26.9 %
C-PR
0%
14%
20%
25 %
C-PR
0%
3.5 %
18.1 %
63.4 %
N-PR
0 %
3.5 %
17.9 %
72.9 %
N-PR
0 %
21.7 %
27.5 %
64%
0 %
20 %
44 % *
82 % *
H-ER
0%
9.4%
14.1 %
23.4 %
46.9 %
U-ER
0%
4.5 %
0 %
40.9 %
36.5 %
H-Ahh
0 %
116.6 %
333 %
467 %
400 %
NOAEL
0.016 ug/kg
LOAEL
0.064
ug/kg
Reference
Romkes and
Safe 1988
Romkes and
Safe 1988
Astro ff et al.,
1991
DeVito et al
1992a
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Profound changes occur in biological systems as a consequence of minute changes in
hormone levels. Much evidence exists that TCDD acts like a hormone, primarily because of
receptor action and the low concentrations that are associated with effects, but also because of the
types of systems affected. Many qualitative changes that occur as a consequence of TCDD
administration are well documented; however, the dose-response relationships for many of these
endpoints have not been evaluated. Dose-response assessments and an increase understanding of
the interaction with other hormone systems is important and should be a focus of future
research.
F. Male Reproductive Toxicity
The male reproductive system is adversely affected by TCDD and related compounds. A
detailed review of TCDD induced male reproductive toxicity is presented in the Reproductive and
Developmental Toxicity chapter of this document (Peterson). Some of the effects are described here
to help understand their dose-response relationships and compare the specificity and sensitivity of
this response to other endpoints.
Overtly toxic doses of TCDD that decrease body weight and feed consumption decrease
plasma androgen concentrations, testis and accessory sex organ weights and alters spermatogenesis
and testicular morphology when administered to postweanling male rats (Moore et al. 1985, 1989,
1991; Mebus et al. 1987; Morrissey and Schewtz 1989; Kleeman et al. 1990; Bookstaff et al.
1990). The decreases in circulating androgens by TCDD are mediated through at least two different
pathways. TCDD inhibits testicular steroidogenesis (Kleeman et al. 1990) and alters the regulation
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of pituitary luteinizing hormone secretion (Bookstaff et al. 1990). The ED50 for the androgenic
deficiency is approximately 15 ug TCDD/kg (Moore et al. 1985). In this study Moore et al.,
administered single oral doses of 0, 6.25, 12.5 25, 50 and 100 ug TCDD/kg to male Sprague-
Dawley rats. Serum testosterone was measured 7 days post administration. Graphic representation
(Graph G) depict a dose-dependent decrease in serum testosterone. The Power Law function fit
this data (R2 = 0.92). When the power of the curve fit was compared to the power of the curve
fit of other endpoints (Benzo(a)pyrene hydroxylase induction Graph K and EROD induction Graph
M) it was found that the kinetic order (g = 0.60) was consistent with these other endpoints. There
are no studies that can directly compare biochemical endpoints quantitatively to male reproductive
endpoints because protocols similar to those designed by Moore et al. (1985) have not been used.
However, the similarity of the curve fit using the power law function to biochemical endpoints (see
biochemical endpoints discussion, section G) suggests a common mechanisms may be responsible
for the decrease in serum testosterone. These data indicate that the adult male reproductive system
toxicity may be mediated by similar mechanism but is not as sensitive to the actions of TCDD in
that androgen deficiency occurs at doses that induce overt toxicity.
While the adult male rat requires overtly toxic doses of TCDD to produce an anti-
androgenic affect, the developing male reproductive system is markedly sensitive to TCDD.
Maternal doses as low as 0.16 ug TCDD/kg decrease anogenital distance of 1 and 4 day old male
rats, delays testicular descent and decreases seminal vesicle and ventral prostate weights in rats.
The organ weights were consistently decreased from the fetal stage into adulthood (Mably et al.
1992a) Table X. Decreases in masculine sexual behavior and increases in feminization of male rats
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also occurred at maternal doses of 0.16 ug TCDD/kg (Mably et al. 1992b), in addition
spermatogenesis was altered at maternal doses as low as 0.064 ug TCDD/kg. The single
administration of 0.064 ug TCDD/kg to pregnant rats on gestational day 15 consistently decreased
epididymis and cauda epididymis weights, cauda epididymal sperm counts and daily sperm
production (Mably et al. 1992c). None of the doses of TCDD resulted in any demonstrable overt
maternal toxicity (Mably et al. 1992a,b,c). These data indicate that in utero and lactational
exposure to TCDD permanently impaired the development of the male reproductive system, as well
as possibly impairing sexual differentiation of the CNS. Hence, the developing male rat
reproductive system is one of the most sensitive to the toxic effects of TCDD.
The mechanism(s) by which in utero and lactational exposure to TCDD alters the
development of the male reproductive system is not completely understood. Decreases in
circulating androgens are mediated in part by reductions in testicular responsiveness to luteinizing
hormone (LH) stimulation (Mably et al. 1992a). Decreases in accessory sex organ weights may
be due to a combination of decreased androgen levels and decreased responsiveness of these organs
to androgens (Mably et al. 1992a). Future studies are indicated to determine the relative importance
of these effects, and of the possible role of altered testosterone metabolism in response to TCDD
induction of phase I and phase II enzyme systems.
The alterations in sexual behavior in male rats appears to be due to impairment of sexual
differentiation of the CNS which requires the presence of testosterone during perinatal development
for male differentiation (Mably et al. 1992b). Once in the CNS, testosterone is converted to
estradiol and both of these steroids must bind to their respective receptors for the proper
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Table X Male Reproductive Toxicity
Species/Strain
Holtzman Rats
Holtzman Rats
Holtzman Rats
Holtzman rats
Holtzman Rats
Sprague Dawley
rats
Dosing Regimen
single dose GD 15
single dose GD 15
single dose GD 1 5
single dose GD 15
single dose GD 1 5
single po dose
Effect
seminal vesicle wt
male offspring after
birth
ventral prostate weight
percent of control
Caudal epididymis wt
in male offspring
anogenital
distance/crown rump
length day 1 or 4 after
birth
Caudal epididymal
sperm in offspring
serum testosterone
level reported as %
decrease
Dose
0.0 ug/kg
0.064
0.16
0.4
10
0.0 ug/kg
0.064
0.16
0.4
1.0
0.0 ug/kg
0064
0.16
04
1.0
0.0 ug/kg
0.064
0.16
0.4
1.0
0.0 ug/kg
0.064
0.16
0.4
1.0
0.0 ug/kg
6.25
12.5
25
50
100
Response % decrease
0% day 32
0%
14%
27%
32%
0% day 32
21%
29%
43%
57%
0% day 32
17%
21%
42%
46%
0% day 1
6%
18%
29%
29%
0% day 63
41%
53%
65%
71%
0%
18.8%
37.5%
43.8%
93.8%
93.8%
0% day 49 0% day 63
9% 13%
13% 28%
48% 38%
65% 44%
0% day 49 0% day 63
+9% 19%
+6% 31%
25% 44
50% 47%
0% day 120
7%
10%
20%
25%
0% day 4
9%
18%
27%
23%
0% day 120
33%
37%
48%
58%
NOAEL
LOAEL
Reference
Mably,
1992a
Mably,
1992a
Mably,
1992a
Mably, et
al, 1992a
Mably,
1992c
Moore, et
al., 1985
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differentiation of the CNS. Maternal doses of 1 ug TCDD/kg decreased plasma testosterone levels in
male offspring on days 18 through 21 of gestation (Mably et al., 1992b). The reduction in plasma
testosterone may account for the demasculinization and feminization of male rats seen in adulthood.
However, it is possible that TCDD alters the sensitivity of the developing pituitary to androgens and
estrogens (Mably et al. 1992b) perhaps by altering the respective hormone receptors (Umbreit and Gallo
1988; Devito et al 1990). The mechanism of the altered sexual behavior of male rats exposed in utero
to TCDD remains undetermined. Quantitative dose response analysis of these very sensitive effects
in male offspring are complicated by the incomplete understanding of the toxicokinetics of TCDD in
the fetus. Additional research that focuses on understanding tissue distribution and metabolism in the
fetus will aid in comparative analyses and help to better understand the implications of these findings
for a human population.
G. Liver Toxicity
Hepatotoxicity is consistently found in animals across species, administered TCDD. The liver
is one of the major sites for metabolism of exogenous and endogenous chemicals. Although clinical
and morphological changes are seen in the liver, it is the induction of hepatic cytochrome P450IA1
family of metabolizing enzymes that is one of the most sensitive indicators in many species for receptor
mediated responses.
Guinea pigs are most sensitive to the acute lethal effects of TCDD. Harris et al.(1973)
administered 0, 0.008, 0.04, 0.2, and 1.0 ug TCDD/kg weekly for 8 weeks (equivalent to 0, .001 .006,
.029, and 0.143 ug TCDD/kg/day) (Table XI). Absolute liver weight was not a sensitive indicator of
toxicity, although liver to body weight ratios calculated from graphic data do show an increase at the
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highest dose. DeCaprio et al. (1986) described the effect of subchronic exposure to TCDD on guinea
pigs. Animals were administered 0, 2, 10, 76, and 430 ppt TCDD in feed for 90 days. Equivalent total
doses calculated by author were for males 0.0, 0.12, 0.61, 0.4.9 and 26 ng/kg/day and for females 0.0,
0.12, 0.68, 4.9, and 31 ng/kg/day. At the high dose 60% of the male and female animals either died
during study or were sacrificed in a moribund condition. A significant increase in liver to body weight
ratio was observed in males and females at the 4.8 and 4.86 ng/kg/day level. A dose dependent trend
of increased relative liver weight was observed. These changes were not accompanied by other signs
of hepatic toxicity. It is interesting to note that the female guinea pig is more sensitive to the chronic
effects of TCDD than to the acute toxic effects. This was also observed in S-D rats in the Kociba et
al. (1978) chronic toxicity study.
Hamsters are relatively insensitive to the acute toxic effects of TCDD, although the response
to TCDD in the liver is qualitatively consistent with other species. A dose-dependent increase in
absolute liver weight was observed in hamsters 50 days after administration of a single i.p. injection
of 0, 5,25,100, 500, 1000, 2000, and 3000 ug TCDD/kg (Table XI). Graphing the Olson et al. (1980)
data using a Michaelis-Menten function resulted in a very poor fit (Graph H). These data are extremely
variable which may be due to the timing of the measurement in relation to the T1/2 of TCDD in
hamsters. Peak liver concentrations of TCDD are reached at 3 days in hamsters, therefore liver weight
measurements taken at 50 days may not accurately reflect tissue toxicity (Olson et al. 1980a; Gasiewicz
1979).
To assess the hepatoxic effects of TCDD on C57BL/6J mice Vos et al.(1974) administered male
mice weekly doses of 0, 0.2, 1.0, 5.0 and 25.0 ug TCDD/kg. Mice were killed at 2 and 6 weeks to
assess the systemic effect of TCDD. The relative liver weights generally increased with increasing
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dose. A statistically significant increase was seen at 2 weeks and at 6 weeks at the 1.0 ug/kg dose
level (equivalent to 0.14 ug TCDD/kg/day)(Table XI). Hepatic damage was characterized as
degeneration of liver cells around the central vein and proliferation of bile duct epithelial cells (at 25
ug TCDD/kg). At the 5 ug TCDD/kg level, necrotic hepatocytes and inflammatory cells were noted in
the centrilobular region. At the lowest dose 0.2 ug TCDD/kg (equivalent to 0.28 ug TCDD/kg/day)
centrilobular fatty change was noted. Fatty change increased in severity with increasing dose.
Microscopic observation of the liver was the most sensitive indicator of frank hepatic toxicity following
TCDD exposure.
Shen et al. (1991) investigated the role of TCDD in liver injury in C57B1/6J and DBA mice
administered a single dose of 0, 3, and 150 ug TCDD/kg and 30, and 60 ug TCDD/kg respectively
(Table XI). Increased liver/bwt ratio and decreased body weight were the most sensitive indicators of
toxicity. On day 7 the C57BL/6J mice had an equivalent increase in relative liver weight at 1/1 Oth the
dose required to elicit the same response in DBA mice, however too few dose levels were used to
assess a dose response relationship. Changes in liver/bwt ratio are a good surrogate for
histopathological changes in the liver of these mice however, in guinea pigs this relationship was not
found. In the DBA mouse no fatty change was observed at 30 ug TCDD/kg, but mid-zonal fatty
change was noted at 600 ug TCDD/kg at 7 d. Focal necrosis was noted at 30 ug TCDD/kg 7 d and
at 600 ug TCDD/kg at 3 and 7 days post dosing. In the C57BL/6J mouse fatty changes were noted at
3 ug TCDD/kg at 3 and 7 days, and severe diffuse fatty changes were noted at 150 ug TCDD/kg at 3
and 7 days. Focal necrosis and focal acute and chronic inflammation was observed at 150 ug TCDD/kg
at 1 and 3 days post-dosing. (Shen et al. 1991) Hence, the sensitivity for necrosis and fatty change is
again approximatley 10-fold between the C57BL/6J and the DBA.
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The differences in TCDD sensitivity between C57BL/6J and DBA mice is believed to be
mediated in part at the level of the Ah receptor gene locus. Poland and Knutson (1982) hypothesized
TCDDs mechanism of toxicity as TCDD binding to an endogenous cytosolic protein (Ah receptor) in
a manner steroid hormones bind to their receptors. This receptor-ligand complex interacts with DNA
to regulate transcription of several proteins, including the cytochrome P-450IA1 which is assayed as
the enzyme aryl hydrocarbon hydroxylase (AHH). TCDD in not mutagenic and these DNA
interactions do not damage DNA. To further clarify the relationship between receptor binding and
toxicity, Bimbaum et al. (1990) used C57BL/6J Ah^ congenic "resistant" mice and wild type Ahb/b
"sensitive" mice. A single dose of 0, 400, 800, 1600, 2400, and 3200 ug TCDD/kg TCDD was
administered to Ah^ mice and 0, 50, 100, 200, 300, and 400 ug TCDD/kg TCDD to Ahb/b mice.
Relative liver weights were significantly increased at all doses. This is expected since doses were
much higher than the dose administered by Shen et al. 1991 in which 3 ug TCDD/kg that resulted in
increased relative liver weights. The severity of a variety of liver lesions were compared in the
congenic and wild type mice. Hepatocellular cytomegaly and hepatocellular karomegaly occurred at
8 times the dose in Ahd/d mice compared to Ahb/b wild type mice. The differences were greater for fatty
change in the liver and bile duct hyperplasia. Qualitatively the toxicity seen in these two strains of
mice were similar; but the doses that result in this toxicity differ by 8-24 times. (Bimbaum et al. 1990)
In male C57BL/6J mice administered single oral doses of 95, 145, and 190 ug TCDD/kg, a dose-
related increase liver, and liver/bwt ratio was observed. In male DBA mice administered a single dose
of 1370, 1870 and 2610 ug TCDD/kg TCDD an increase in absolute and relative liver weight was
observed at every dose, but not with a dose response trend (Chapman and Schiller 1985). These same
authors reported a dose-dependent decrease in serum glucose, total serum cholesterol,cholesterol ester,
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serum triglycerides, and an increase in liver triglycerides in both strains of mice. The dose response
data is reported graphically.
Pohjanvirta studied two strains of rats, Long Evans (L/E) a species sensitive to TCDD with a
LD50 of 10 ug TCDD/kg and the Han/Wistar (H/W) rat with a LD50 of >3000 ug
TCDD/kg.(Pohjanvirta 1990). In assessing target organ toxicity the L/E rats received a single ip
injection of 5, and 50 ug TCDD/kg. The H/W rats received an additional dose of 500 ug TCDD/kg.
Relative organ weights were assessed on day 1,4,8,16 and 39-40. The only difference observed in
relative liver weights was the initial statistically significant increase on day 1 in the H/W rats at the
two highest doses. Both species were observed to have a statistically significant increase in relative
liver weight on day 3 to day 16. There were marked strain differences observed histopathologically.
In L/E rats 50 ug TCDD/kg (5 X LD50) four days post-dosing resulted in multinucleated and swollen
hepatocytes. Only minor lesions were observed in H/W rats administered up to 500 ug TCDD/kg
(Pohjanvirta 1989). Distribution of TCDD to the liver did not account for any differences in liver
toxicity. Both strains had peak TCDD liver concentrations on day 8 and the rate of liver tissue
disappearance appeared equivalent. This is unlike the thymus distribution in which was qualitatively
very different between these two strains (Pohjanvirta 1990).
A dose-dependent increase in relative liver weight was observed in male and female Sprague-
Dawley rats (Kociba et al. 1976) administered 0.001, 0.01, 0.1, 1.0 ug TCDD/kg/day 5 days/week for
13 weeks (Equivalent doses: 0.0007, 0.007, 0.07, 0.7 ug TCDD/kg/day). Liver to body weight ratios
(Table XI) were increased in males at the 0.1 and 1.0 ug TCDD/kg dose level. Female rats were more
sensitive, having a statistically significant increase in relative liver weight additionally at the 0.01 dose
level. This is illustrated using the Power Law function (Graph I). The R2 values for males and females
(0.93, 0.94) are high for the Power Law indicating a good fit to the data.
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Table XI Liver Weight Chances
Species/Strain
Hartley Guinea Pigs female
Hartley Guinea pigs (230-340)
Males
Hartley Guinea pigs (230-340)
females
Rhesus Monkeys <70 kg
Female Syrian golden hamster
51-60g
C57BL/6J male mice
C57BL/6J male mice 20-28 g
DBA male mice 20-28g
C57BL/6J male mice
DBA male mice
Sprague-Dawley rats males
Dosing
TCDD administered
orally 1 X week for
8 weeks
oral continuous feed
90 days
oral continuous feed
90 days
single oral gavage
single i p injection
weekly doses p o
single i p in olive
oil
single i p in olive
oil
single p o
p o 5 x/wk 1 3 weeks
Effect
increase absolute liver weight
increase hver/bwt estimated from
graph
increase hver/bwt reported as %
control (total dose reported)
increase hver/bwt reported as %
control (total dose reported)
increase liver as percent bwt 4-6
wks
change liver weight 50 days
increase liver/bwt ratio reported
at (2 weeks) and (6 weeks)
increase hver/bwt ratio as %
control
increase hver/bwt ration as •/«
control 1,3,7 days post dosing
increase hver/bwt ratio 30 days
post dosing
increase hver/bwt ratio reported
as ug/kg/day and (total dose)
increase Liver/bwt ratio reported
as ug/kg//day and (total dose)
Dose ug TCDD/ kg - Response
0 ug/kg 0 ug/kg/day
0 008 001
0 04 006
02 029
1 0 143
0 ug/kg 0 ng/kg/day
0011 012 ng/kg/day
0 055 0 61 ng/kg/day
0441 49 ng/kg/day
0 ug/kg 0 ng/kg/day
0011 012 ng/kg/day
0 061 0 68 ng/kg/day
0 437 49 ng/kg/day
0
70
350
0 ug/kg
5
25
100
500
1000
2000
3000
ug/wk (2) (6)
0 (0) (0)
02 ( 4) (1 2)
1 0 (2) (6)
50 (10) (30)
25 (50)(150)
0 dose
3
150
0 ug/kg
30 ug/kg
0 ug/kg C57BL 0 DBA
95 1370
145 1870
190 2610
0 (0) ug/kg
0 001 (0 065)
0.01 (0.65)
01 (6.5)
1 0 (65)
0 (0) ug/kg
0001(0065)
0 01 (0 65)
01 (65)
1 0 (65)
0 % change liver wt 0% change hver/bwt
+0 3% +4%
-8 5% -2%
-85% +143%
N/A N/A
0
-9 7% (decrease)
+18.1%
+24.0%*
0
+4.4%
-0.6% (decrease)
+28%*
0
458% *
47.0 % *
0
-5.3%
+19%
+8.6%
+22.1%*
+42%*
+27.4%*
4.9%
0 0
1 3% 6.8%
11.2%* 17.6%*
155%* 26.4%*
17.5%* 42.6%*
0 day 1 0 day 3 0 day 7
-5% +3.6% +22%*
-2% +11%* +41%
0 day 1 0 day 3 0 day 7
-5.8% 0 +115%*
+9.8%* 0 +27%*
0 C57BL 0 DBA
13.9%* 53%
32%* 36.8%
41.7%* 46.5%
0 (measured dose) 0
N/A 3 2%
0 0037 ug/g hver 4 5%
0.0346 22%
0 2840 35%
0 (measured dose) 0
N/A 3.2%
0.0026 ug/g hver 4.5%*
0 0360 ug/g liver 22%*
0.3240 ug/g liver 35%*
NOAEL
0 055 ug/kg
0061
100 ug/kg
042 weeks
1 2 6 wks
3 - 3 days
N/A -7 days
30 ug/kg 1,3
days
N/A
0 065 ug/kg
0 065 ug/kg
LOAEL
0 44 ug/kg
044
70 ug/kg
Lowest dose
tested
500 ug/kg
2 0 -2 wk
6 0-6 wk
150 - 3d
3 -7d
30 ug/kg - 7d
95
0 65 ug/kg
0 65 ug/kg
Reference
Hams, 1973
DeCapno, 1986
DeCapno, 1986
McConnell, 1978
Olson, Holscher,
andNeal, 1980
Vos, 1974
Shen, 1991
Shen, 1991
Chapman and
Schiller, 1985
Kocibaetal., 1976
Kociba etal,, 1976
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The "g" values for these functions indicate a sublinear dose response. Attempts to fit the
Michaelis-Menten function resulted in a very poor fit suggesting that a sigmoid dose-response may
not characterize this phenomenon. The female rats were also more sensitive than male rats as
measured by histopathology, clinical chemistry and relative liver weight. Microscopic examination
revealed some hyperplasia in the bile ducts and altered hepatocytes at the 1.0 ug TCDD/kg dose
level of both sexes, characterized as multinucleated and irregularly shaped cells, lipid accumulation,
and some necrosis. Slight changes in hepatocytes "variations in size and shape" were noted in the
0.1 ug TCDD/kg level. No abnormalities were noted microscopically in the 0.001 or 0.01 dose
levels. Clinical chemistry also revealed a statistically significant increase in total bilirubin, direct
bilirubin, and alkaline phosphatase at the two highest doses in females (0.1 and 1.0 ug TCDD/kg)
and in the highest dose in males. Increase relative liver weight was the most sensitive indicator
of liver toxi city in this study. This study was terminated at 13 weeks at which time a parallel study
reported that steady state concentrations of TCDD are found in the fat and the liver (Rose 1976).
G. Biochemical Endpoints
The mechanism of TCDD toxicity involves the formation of a receptor-ligand complex
which results in an increase in the transcription of P450 liver enzymes. The increase in enzyme
production may account for the increase in liver weight, and may also be responsible for other toxic
endpoints.
Cytochrome P450 is a family of enzymes which catalyze the metabolism of a wide range
of endogenous lipophilic compounds (steroids, fatty acids, cholesterol derivatives) as well as a
number of xenobiotics. There is a large body of literature which point to major differences between
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different animal species, and humans when comparing specific cytochrome P450 forms and
substrate specificity. These differences are also important when comparing sexual differences and
age related effects in animals. The differences in species, sex and age sensitivity can be attributed
to which cytochrome P450s are constitutively expressed and the inducibility of all isozymes. If the
mechanism of toxic effects are mediated by the cytochrome P450 then the differences observed
between animals can be explained. The different responses are probably due to a number of
factors, because of the complexity of the biological response. As outlined in the Whitlock chapter
(this document), the effects caused by dioxin interaction with the receptor are a multi step process
which, with alterations at any of the steps, will alter the response. In graphing the response of
activity in a single species there is very good agreement for the levels of Ah receptor in the cytosol
and nucleus, the mRNA for P450IA1 and the induction of enzymatic activity using EROD as the
substrate in a single tissue. Tissue sensitivity may differ because of the presence of tissue specific
stuctural orthologs of P450. (Wrighton, 1990; Nebert et al. 1989). There can be differences
between species and sex as to which forms are constituitively expressed and there inducibility.
There may also be species specific forms, which make a direct extrapolation to man difficult
(Guengerich 1989). The presence or absence of a specific isozyme in a tissue or cells within a
tissue will also effect the toxicity. As stated above and discussed in this document (Witlock
chapter), the multi step process of induction of the TCDD responsive enzymes is mediated through
a cytosolic receptor. The differences in the receptor and the regulation of the genes and post-
translational events between species and sex can result in distinctly different toxic responses. The
importance of the genetics in the effects caused by TCDD is best demonstrated in the Ah
responsive and nonresponsive mice. In the homozygous mouse the animals are very sensitive to
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many TCDD mediated responses, while the heterozygous mouse strain is less sensitive.
The P450I family is present in rodents and humans and the family contains these two members
Cytochrome P450IA1 (CypIAl) and P450IA2 (CypIA2). Both of these forms are important in the
response observed following TCDD treatment. The CypIAl is an ortholog of the rat P450c (also
known as P448). In mammalian species this can be measured as aryl hydrocarbon hydroxylase
(AHH) and is expressed at very low levels in the liver and extrahepatically. In animal studies this
is the major form that is induced following poly aromatic aromatic hydrocarbon treatment. Wrighton
(1990) has reported that in man P450IA1 is rarely expressed in human liver and is predominantly
an extrahepatic form, and as such is inducible by cigarette smoke and polyaromatic hydrocarbons
(PAHs).
The second member of this family is CypIA2, which is present the untreated animals
constitutively at levels greater than CypIAl, is inducible by polyaromatic hydrocarbons(PAHs)
(Sesardic et al 1988; Quattrochi and Tukey 1989). The human CypIA2 (HLd) is expressed in all
human livers examined. This is important because this isoform is believed to be a specific (non
receptor) binding protein in hepatic microsomes, which may play an important role in the human
response to TCDD (Wrighton 1989). Unlike the CypIAl which in the human liver is not inducible,
the CypIA2 is inducible by constituents of cigarette smoke and other sources of polyaromatic
hydrocarbons. In animal studies the CypIA2 is inducible following TCDD exposure.
Hook et al. (1975) characterized the induction of microsomal mixed-function oxidases
(MFO) and cytochromes following TCDD treatment. There was a decrease in N-demethylation
enzyme activity in the liver of female C-D rats not seen in male rats; however, there was no
difference in the induction of P450° or benzo(a)pyrene hydroxylase. In addition, Hook in an
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enzyme assay using tissues from male rats administered 25 ug TCDD/kg found there was liver and
kidney induction of biphenyl 4-hydroxylase, biphenyl 2-hydroxylase, benzo(a)pyrene hydroxylase,
UDP-glucuronyl transferase and benzphetamine N-demethylase. In the lungs, intestine, brain and
testes of these animals there was no induction of biphenyl 4-hydroxylase, and biphenyl 2-
hydroxylase, but there was a significant induction of all other enzymes. In guinea pigs dosed with
0.175 ug TCDD/kg biphenyl-4-hydroxylase was reported in liver, lung, and kidney tissues 4 days
after administration. In rabbits administered 25 ug TCDD induction of 4-hydroxylase was
observed in lung tissue and not in kidney and liver. Induction of benzo(a)pyrene hydroxylase was
observed in kidney and liver, but not in the lung. These differences are difficult to interpret because
only one dose was administrated, but suggested definite species, sex and tissue differences in
response to TCDD. The clinical manifestations of these enzyme changes are not well understood,
but is a focus of current research.
To understand the tissue differences in enzyme induction, EROD; ethoxyresorufin
dealkylase activity a measure of CypIAl activity was quantified in the homogenized lung and liver
of male Swiss mice after ip injection of 5 or 50 nmol/kg TCDD. Results indicate that the induction
of P450 IA1 in the liver and lung follow a very similiar pattern. Both liver and lung have an
immediate and significant increase in EROD 15 fold and 25 fold respectively over the very low
consituitive levels (Beebe et al. 1990). Comparing the EROD activity in the lung administered
50 nmol/kg versus the liver administered 5 nmol/kg reveals that the liver has 5 times the activity
(40 nmoles resorufm/min/g liver vs. 8 nmoles/g lung). These differences may be accounted for by
differences in tissue distribution, as Gasiewicz et al. (1983) reported an approximately 7 to 10 fold
difference in TCDD distibution to the liver vs lung in the three species of mice in which
measurements have been made.
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Kitchin and Woods (1979) evaluated a dose response relationship between TCDD and the
elevation of the liver enzymes as total p-450 and benzo(a)pyrene hydroxylase in microsomes.
Additionally, they administered 2 ug TCDD/kg to rats and examined the increase enzyme activities
associated with a number of cytochrome p-448 type substrates (biphenyl 2-hydroxylase, biphenyl
4-hydroxylase, 4,4'dimethyl-biphenyl hydroxylase, benzo(a)pyrene hydroxylase, EROD, and
benzphetamine N-demethylase). All but the last enzyme was elevated significantly by an exposure
of 2 ug TCDD/ kg. In a third experiment, Kitchin and Woods (1979) administered a single po dose
of 2 ug TCDD/kg to female S-D rats and reported EROD, cytochrome P-450, and AHH levels 1,
3, and 6 months post-dosing. Results indicate that enzyme remain elevated even after 6 months
although the levels of AHH and EROD are substantially reduced after 6 months. Kitchin and
Woods (1979) do some interesting calculations based on TCDD radiotracer experiments. TCDD
levels in the liver, thymus and adipose 3 days post-dosing of 2 ug TCDD/kg resulted in tissue
concentrations of 1.08, 0.37, and 1.84% of oral dose per gram of tissue (wet weight). They
determined that as few as 65 molecules of TCDD in a hepatocyte can significantly increase the
enzyme activity of Benzo(a)pyrene hydroxylase. (Table XII). Both total P450 and BAP
hydroxylase fit the Power Law functional form with high correllation, but have different kinetic
orders (g = 0.58, 0.42) which may indicate different factors regulate each (Graphs J and K).
Abraham et al. 1988 described the response of hepatic ethoxyresorufm-O-deethylase
(EROD) and total P450 induction in female Wistar rats (See Table XII). Rats were administrated
doses ranging from 1 to 3000 ng TCDD/kg via subcutaneous injection. Subcutaneous absorption
was reported to be slow, but almost complete. TCDD induced a significant increase P450 activity
at 0.1 ug TCDD/kg and in EROD at 0.3 ug TCDD/kg. The EROD response continued to increase
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enzyme assay using tissues from male rats administered 25 ug TCDD/kg found there was liver and
kidney induction of biphenyl 4-hydroxylase, biphenyl 2-hydroxylase, benzo(a)pyrene hydroxylase,
UDP-glucuronyl transferase and benzphetamine N-demethylase. In the lungs, intestine, brain and
testes of these animals there was no induction of biphenyl 4-hydroxylase, and biphenyl 2-
hydroxylase, but there was a significant induction of all other enzymes. In guinea pigs dosed with
0.175 ug TCDD/kg biphenyl-4-hydroxylase was reported in liver, lung, and kidney tissues 4 days
after administration. In rabbits administered 25 ug TCDD induction of 4-hydroxylase was
observed in lung tissue and not in kidney and liver." Induction of benzo(a)pyrene hydroxylase was
observed in kidney and liver, but not in the lung. These differences are difficult to interpret because
only one dose was administrated, but suggested definite species, sex and tissue differences in
response to TCDD. The clinical manifestations of these enzyme changes are not well understood,
but is a focus of current research.
To understand the tissue differences in enzyme induction, EROD; ethoxyresorufin
dealkylase activity a measure of CypIAl activity was quantified in the homogenized lung and liver
of male Swiss mice after ip injection of 5 or 50 nmol/kg TCDD. Results indicate that the induction
of P450 IA1 in the liver and lung follow a very similiar pattern. Both liver and lung have an
immediate and significant increase in EROD 15 fold and 25 fold respectively over the very low
consituitive levels (Beebe et al. 1990). Comparing the EROD activity in the lung administered
50 nmol/kg versus the liver administered 5 nmol/kg reveals that the liver has 5 times the activity
(40 nmoles resorufin/min/g liver vs. 8 nmoles/g lung). These differences may be accounted for by
differences in tissue distribution, as Gasiewicz et al. (1983) reported an approximately 7 to 10 fold
difference in TCDD distibution to the liver vs lung in the three species of mice in which
measurements have been made.
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Kitchin and Woods (1979) evaluated a dose response relationship between TCDD and the
elevation of the liver enzymes as total p-450 and benzo(a)pyrene hydroxylase in microsomes.
Additionally, they administered 2 ug TCDD/kg to rats and examined the increase enzyme activities
associated with a number of cytochrome p-448 type substrates (biphenyl 2-hydroxylase, biphenyl
4-hydroxylase, 4,4'dimethyl-biphenyl hydroxylase, benzo(a)pyrene hydroxylase, EROD, and
benzphetamine N-demethylase). All but the last enzyme was elevated significantly by an exposure
of 2 ug TCDD/ kg. In a third experiment, Kitchin and Woods (1979) administered a single po dose
of 2 ug TCDD/kg to female S-D rats and reported EROD, cytochrome P-450, and AHH levels 1,
3, and 6 months post-dosing. Results indicate that enzyme remain elevated even after 6 months
although the levels of AHH and EROD are substantially reduced after 6 months. Kitchin and
Woods (1979) do some interesting calculations based on TCDD radiotracer experiments. TCDD
levels in the liver, thymus and adipose 3 days post-dosing of 2 ug TCDD/kg resulted in tissue
concentrations of 1.08, 0.37, and 1.84% of oral dose per gram of tissue (wet weight). They
determined that as few as 65 molecules of TCDD in a hepatocyte can significantly increase the
enzyme activity of Benzo(a)pyrene hydroxylase. (Table XII). Both total P450 and BAP
hydroxylase fit the Power Law functional form with high correllation, but have different kinetic
orders (g = 0.58, 0.42) which may indicate different factors regulate each (Graphs J and K).
Abraham et al. 1988 described the response of hepatic ethoxyresorufin-O-deethylase
(EROD) and total P450 induction in female Wistar rats (See Table XII). Rats were administrated
doses ranging from 1 to 3000 ng TCDD/kg via subcutaneous injection. Subcutaneous absorption
was reported to be slow, but almost complete. TCDD induced a significant increase P450 activity
at 0.1 ug TCDD/kg and in EROD at 0.3 ug TCDD/kg. The EROD response continued to increase
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even at doses 1000 times this. No maximum induction was reached in this study. The Michaelis-
Menten and Hill plots graphically depict this, phenomenon (Graph L and M). TCDD liver
concentrations continued to increase at concentrations up to 3 ug TCDD/kg. This was attributed
to steady state not being reached at day 7, but maximum-induction being reached at higher doses
should not be discounted based on this data.
Experiments by Abraham et al. (1980) and Kitchin and Woods (1979) (Table XII) provided
sufficient doses to further analyze the four functional forms described in the introduction. Kitchin
and Woods (1979) measured benzo(a)pyrene hydroxylase (aryl hydrocarbon hydroxylase - AHH)
induction and total P450 concentration in the liver of female Sprague-Dawley rats 3 days after
TCDD administration. Abraham et al. (1980) used Wistar rats and measured EROD and total
hepatic P450 at 7 days post dosing.
For the Kitchin and Wood (1979) data the linear function of dose-response was not an
appropriate fit (Graph Jl-A) for the total P450 (R2 = .67). The Michaelis-Menten fit of raw data
(Jl-B) depicts an "inertial response" in the dose range from 0.0006 to 0.06 ug TCDD/kg dose range
and appears to reach maximim at 20 ug TCDD/kg indicating that this function does describe the
data. The Hill function (Graph Jl-C) also fits the data points with a sigmoid curve and an exponent
of 1.09. A Hill exponent of 1.0 is generally interpreted as having a linear dose response at low
doses without cooperativity. However, the validity of this interpretation comes into question because
the x-y plot does not depict linearity at low doses. The Power Law fit, over the entire dose-range
(Graph J2-A), resulted in a "g" = 0.76, (R2 = 0.98) suggesting a sublinear dose-response.
Systematic application of the Power Law resulted in a sublinear response in the low dose range
(0.0006 to 0.6 ug TCDD/kg) (Graph J2-B]) with a "g" = 0.035: R2 = 0.51 which increased at high
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dose range to 0.11. R2 = 0.97. This shift of "g" values form 0.035 to 0.11 or greater defines the
"inertia! response". Graphs L depicts the Abraham et al. (1980) P450 raw data. The Hill function
applied the Abraham et al.(1980) P450 raw data has an exponent of 0.55 (sublinear indicating
negative cooperativity in the low dose range, while Kitchin and Woods (1979) had an exponent of
1.09. In addition the fit to the Michaelis-Menten function does not appear to reach a maximal
response (GraphJl-B) indicating this function does not describe the dataset.
It is the Power Law function that suggests a consistent response between these two datasets.
The change in kinetic order between the low doses (g = 0.036, R2 = .88 [Abraham et al., 1980);
(g = 0.035, R2 =0.51; [Kitchin and Woods 1979]) and high doses (g = .14, R2 = .99 [Abraham et
al. 1980]); (g = 0.11, R2 = 0.97; [Kitchin and Woods 1979]) is demonstrated with an improved
"goodness of fit".
Total P450 is a non-specific measure of response and may not be a good predictor of
clinical toxicity. The induction of enzyme activity is a measure of a specific post-transcriptional
event and as such may serve as a more consistent predictor of clinical outcome. Therefore, the
induction of benzo(a)pyrene hydroxylase was analyzed (Kitchin and Woods 1979). The Michaelis-
Menten and Hill functions both graphically depict an inertial response at low doses and a
maximum response a high doses (Graph Kl). The exponent for the Hill function is 1.07 (compared
to 1.09 for the P450 induction) suggesting that both responses may be mediated by similar
mechanisms. Systematic analysis of subsets of doses (Graph K2-B) resulted in a comparable trend
of increasing kinetic order with increased dose range. The low dose (0.0006 to 0.004 ug TCDD/kg)
kinetic order (g value) was 0.21. R2 = 0.96 which increased with doses in the 0.02 to 0.6 ug
TCDD/kg range (g = 0.77. R2 = 0.96. In the highest dose range tested (0.6 to 20 ug TCDD/kg)
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the kinetic order decreased to 0.19. R2 = 0.80.
Similar responses were observed when EROD activity was measured. Graphs M depicts the
functional forms used in this analysis. Both Michaelis-Menten(Graph Ml-B) and Hill functions
(Graph Ml-C) graphically depict a low dose "inertial" response. The Hill function has a exponent
of 1 which is consistent with the total P450 response (indicating no cooperativity and suggesting
linearity in low dose range). The Power Law functional fit on all data points resulted in a sublinear
fit, g= 0.52, R2 = 0.95 (Graph M2-A). Analysis of subsets of data resulted in a response consistent
with the Kitchin and Woods (1979) data analysis, ie., an increase in kinetic order from the low dose
range (g = 0.11) to the high dose range (g = 0.66). For EROD induction the "Goodness of fit" for
these both subsets of data was high, R2 =0.99(Graph M2-B). The analysis of these two datasets
(Kitchin and Woods, 1979; Abraham et al., 1980) suggests differences in the low and high dose
responses. The Power Law function is commonly used to describe physical phenomenon, therefore
the dose response fit of this model may be best described using laws of physics. The low dose
response can be characterized as an "inertial" in which activity is measurable, but at a reduced rate.
Not until an "external force" acts upon the system is an altered response (manifested as change in
kinetic order) measured. The Power Law function when applied to total P450, EROD, and
benzo(a)pyrene hydroxylase describes this phenomena.
In female mice (Narasimhan et al.(1993) see Table XIII) measured the individual steps of
the multistep process of enzyme induction. Graphic representation of the total nuclear Ah receptor
binding, mRNA for CypIAl, mRNA for CyplA2, and enzyme induction (EROD) plotted against
dose demonstrated parallel slopes when fitted to a Power Law function. R2 values indicate good
fit to these functions (Graph N). This data supports the conclusion that the steps of this multistep
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process are regulated by similiar mechanisms, and there is not likely to be post-translational
modification for EROD induction. However, there is a difference in Cypl Al Snd Cypl A2 mRNA
levels.
The differential dose-response in induction of CypIAl and CyplA2 has recently been
investigated (Trischer et al. 1992). Female rats (initiated with DEN or saline) were administered
TCDD biweekly p.o. for 30 weeks. Dose equivalents of 3.5,10.7, 35.7, and 125 ng/kg/day resulted
in a dose-dependent increase in hepatic CypIAl and CypIA2 (measured by RAI). The rate of
induction (different slopes) of the two enzymes differs only slightly and probably not significantly
(Graph O). The differences in the Power Law function between these two enzymes is not as great
as the difference detected by DeVito et al. (1992) (see below) for hepatic CypIAl and CypIA2,
possibly because of the different techniques of enzyme measurement used in this study; therefore
definative conclusions can not be made. Additionally, maximum induction of Cypl Al is reached
at a liver concentration of 10 ppb, wheras CyplA2 activity continues to increase after 30 weeks at
hepatic concentrations of 30 ppb. Immunolocalization of enzyme induction in this experiment
indicated that hepatocytes are not induced uniformily (Trischer et al. 1992). The mechanisms
responsible for this differential induction may help in the understanding of intraspecies variability
of response and prove useful in risk estimation.
The induction of skin, hepatic and lung EROD and hepatic 1A2 (acetanilide hydroxylase
activity) was measured in female B6C3F1 mice administered 0,1.5, 4.5,15, and 150 ng/kg/day for
13 weeks (total doses 0, 0.098, 0.29, 0.98, 9.8)(DeVito et al. 1992 see Table XIV). Graphic
representation of dose-response data indicates parallel slopes for all EROD induction, but a different
slope for 1A2 induction (Graphs P1-P4). The fit to the Power Law functional form for EROD in
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the three tissues measured had a high correllation coefficient. There was also a striking similiarity
between the Power Law for the EROD measured in the three tissues versus the power law for
hepatic CypIA2. The difference in induction between hepatic and lung EROD ranges from 5 fold
to 18 fold which may be accounted for by differences in tissue distribution, although other factors
can not be ruled until tissue dosimetry assays are run parallel to the enzyme induction assays.
Enzyme induction after a subchronic (13 week) exposure in B6C3F1 female mice was
evaluated using the Devito et al.(1992) data. The linear regression model (Graph PI) for hepatic,
skin and lung EROD induction had R2 values of 0.95, 0.91, and 0.94 respectively; however, the
R2 for hepatic IA2 induction was 0.54. The Michaelis-Menten functional forms (Graph P2)
graphically depict the "inertial" phase of the dose-response at doses less than 15 ng/kg/day (total
dose = 0.975 ug TCDD/kg) for hepatic, skin and lung EROD activity. For hepatic EROD, the Hill
function (Graph P3) exponent (1.1) was superlinear, whereas the exponent derived from application
of the Hill plot for skin and lung EROD activity was sublinear (exponent = 0.83, 0.84). The "g"
value in the Power Law function for hepatic, skin and lung EROD induction (Graph P4) depicts
a sublinear kinetic order (g = 0.68, 0.75, and 0.79; R2 = 0.92, 0.94, 0.97, respectively). These
values are remarkably consistent, and interestingly are similar to the "g1 values determined for
EROD induction by Abraham et al., 1988 at a similar range of doses( at doses 0.3 to 3.0 ug
TCDD/kg "g" = 0.66, R2 = 0.99. Systematic analysis of subsets of data does not reveal the same
change in kinetic order seen in the Abraham et al. (1988) or the Kitchin and Woods (1979) data
sets; however, the total administered doses in the Devito et al. (1992) are not in the range of doses
in which the change in kinetic order was observed. There is a notable increase in the "g" value for
hepatic EROD induction (Table XV), but the R2 values are low, and a consistent trend was not
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observed. Although no change in kinetic order was observed in the lung and skin EROD induction,
there are no acute data comparable to the Abraham et al. (1988) or Kitchin and Woods (1979) data
for lung and skin induction. Therefore, a more complete understanding of the activity in extra-
hepatic tissue is not possible with the available data.
The induction of P450IA2 was also measured (DeVito et al. 1992) using the functional
forms previously described. The initial X-Y plot and linear plots of data revealed a much higher
degree of variability in the data values. This observation is consistent with what has been found
in humans populations, that P450IA2 constitutive values vary 3000% (Grant et al, 1983). This
precludes making assumptions about the Hill or Michaelis-Menten plots.
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TABLE Xn DOSE-RESPONSE FEMALE RATS
Liver Enzyme Activity, Total P-450, Benz(a)pyrene, and EROD
Species
Female Sprague-Dawley
rats
Female Sprague-Dawley
rats
Wistar Rats Female
Wistar rats
Female
Dosing Regimen
single oral dose
single oral dose
single s.c.
injection
single s.c.
injection
Observed Effect
increase P-450 (3 days)
(Omuraand Sato, 1964)
increase
Benzo(o)pyrene
hydroxylase (nmol polar
products/mg protem/hr)
(DePierre, 1975) at 3
days
increase P-450
concentration (p-moles
X mg protein"^
measured as reduced CO
difference spectra 7-days
post dosing
increase EROD Activity
pmoles resorufin formed
X mg protein-1 X min-1
(resorufin measured
spectrofluorometrically
in liver homogenates 7
days post-dosuig)
Dose ug TCDD/kg
0
0.0006
0.002
0.004
0.02
0.06
0.2
0.6
2.0
5.0
20.0
0
0.0006
0.002
0.004
0.02
0.06
0.2
0.6
2.0
5.0
20.0
0 us/kg
0.001
0.003
0.010
0.030
0.100
0.300
1.00
3.00
8.85F
0.003
0.010
0.030
0.100
0.300
1.00
3.00
0 (measured
tissue dose)
0.0031
0.0102
0.0406
0.162
0.699
3.38
10.7
27.9
0 (measured
tissue dose)
0.0031
0.0102
0.0406
0.162
0.699
3.38
10.7
27.9
Response
0.88 0%
0.91 3.4%
0.93 5.7%
1.0 13.6%
0.87 -1.1%
.06 20.5%
.16 31.8%
.13 28.4%*
.33 51.1%*
.52 72.4%*
.68 90.9%*
4.9 0%
4.9 0%
6.7 36.7%*
7.2 46.9%*
8.3 69.4%*
14.0 185.7%*
59.0 1104%*
96.0 1859%*
155.0 3063%*
182.0 3614%*
189.0 3757%*
169 0%
182 7.1%
173 2.4%
184 8.9%
183 8.3%
199 17.8%*
230 36.1%*
269 59.2%*
319 88.8%*
14.3 0%
16.9 18.2%
19.0 32.9%*
21.9 53.1%*
40.7 184.6%*
89.5 525.9%*
204.4 1329.4%*
495 3361.5%*
759.8 5213.3%*
NOAEL
0.2
0.0006
LOAEL
0.6
0.002
0.100 ug/kg
0.003 ug/kg
Reference
Kitchin, 1979
Kitchin, 1979
Abraham, 1988
Abraham,! 988
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TABLE Xm DOSE - RESPONSE B6C3F1 Female Mice
Liver EROD, CyplAl mRNA, and CyplA2 mRNA and Total Nuclear Ah Receptor Binding
Species
B6C3F1
female mice
B6C3F1
female mice
B6C3F1
female mice
B6C3F1
female mice
Dosing
Regimen
single i.p.
injection
single i.p.
injection
single ip
injection
single ip
injection
Effect
24 hrs. post
dosing increase
EROD pmol/mg
protein/min
24 hrs post
dosing increase
relative mRNA
units CyplAl
24 hr. post dosing
relative mRNA
units CyplA2
24 hr post dosing
total nuclear AH
receptor binding
Administered
Dose
0 ug/kg
0.005
0.010
0.025
0.05
0.10
0.50
1.00
5.00
Oue/kg
0.005
0.010
0.025
0.05
0.10
0.50
1.00
5.00
Oue/kg
0.005
0.010
0.025
0.05
0.10
0.50
1.00
5.00
Oug/kg
0.005
0.010
0.025
0.05
0.10
0.50
1.00
5.00
Liver
Dose
f 1
0
0.11
0.10
0.09
0.26
0.74
2.36
11.5
53.16
0
0.11
0.10
0.09
0.26
0.74
2.36
11.5
53.16
0
0.11
0.10
0.09
0.26
0.74
2.36
11.5
53.16
0
0.11
0.10
0.09
0.26
0.74
2.36
11.5
53.16
Response
0
+51%
-8.0%
-11%
+9.0%*
+109.6%*
+252%*
+538.5%*
+1857%*
0
-0.14%
-0.14%
+171%
+100%
+571%
+900%
+1257%
+2671%
0
-0.14
-0.28
-0.28
-0.28
-157
+57
+243
+343
156 pmol/mg prot/min
168
139
170
327*
549*
996*
3053*
0.07
0.06
0.06
0.19
0.14
0.47
0.70
0.95
1.94
0.07
0.06
0.05
0.05
0.05
0.03
0.11
0.24
0.31
N/A
4.3
7.2
6.2
8.2
6.5
10.6
45.4
111.3
NOAEL
0.050
ug/kg
0.10
0.5
LOAEL
0.10 ug/kg
0.50
1.0
Reference
Narasimhan et
al., 1993
Narasimhan et
al. 1993
Narasimhan et
al., 1993 1992
Narasimhan et
al., 1993
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TABLE XIV DOSE - RESPONSE B6C3F1 Female Mice
Liver EROD, Skin EROD, Lung EROD, Hepatic Ia2 Activity
Species
B6C3F1
female mice
B6C3F1
female mice
B6C3F1
female mice
B6C3F1
female mice
Dosing Regimen
repeat dose 5 X
per week 1 3
weeks (65 doses)
oral
repeat dose 5 X
per week 1 3
weeks (65 doses)
oral
repeat dose 5 X
per week 1 3
weeks (65 doses)
oral
repeat dose 5 X
per week 1 3
weeks (65 doses)
oral
Effect
increase hepatic
EROD nm/mg
protein/min
increase skin
EROD
nm/mg/protein/
min
increase lung
EROD
nm/mg/protein/
min
increase hepatic
Ia2
acetanilide 4-
hydroxylase
Administered Dose
0 ug/kg/day (0 ug/kg total)
0.0015 (0.0975 ug/kg)
0.0045 (0.292 ug/kg)
0.015 (0.975 ug/kg)
0.045 (2.925 ug/kg)
0.150 (9.75 ug/kg)
0 ug/kg/day (0 ug/kg total)
0.0015 (0.0975 ug/kg)
0.0045 (0.292 ug/kg)
0.015 (0.975 ug/kg)
0.045 (2.925 ug/kg)
0.150 (9.75 ug/kg)
0 ug/kg/day (0 ug/kg total)
0.0015 (0.0975 ug/kg)
0.0045 (0.292 ug/kg)
0.015 (0.975 ug/kg)
0.045 (2.925 ug/kg)
0.150 (9.75 ug/kg)
0 ug/kg/day (0 ug/kg total)
0.0015 (0.0975 ug/kg)
0.0045 (0.292 ug/kg)
0.015 (0.975 ug/kg)
0.045 (2.925 ug/kg)
0.150 (9.75 ug/kg)
Response
100%
220%
260%
610%
1780%
3965%
100%
170%
280%
1040%
1760%
5280%
100%
217%
545%
1614%
2833%
8786%
100%
130%
180%
290%
360%
420%
126.25 mean EROD
271.4
323.4
763.6
2245.25 •
4996.0
0.89 mean EROD act.
1.54
2.5
9.29
15.65
47.0
4.2
9.15
22.9
67.8
119
369
277.0
357.2
618.4
798.4
920.6
1164.4
NOAEL
LOAEL
Reference
DeVito et
al.1992
DeVito et al.,
1992
DeVito et al.,
1992
DeVito et
al.,1992
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The fit to the linear model is also poor, R2 = 0.54. The Power Law for P450IA2 induction
resulted in a "g" = 0.24. R2 = 0.73 , a sublinear functional fit. In additional, the absolute
number for this exponent is lower than the "g" value for hepatic, skin or lung EROD induction.
The evaluation of data subsets reveals a completely opposite pattern than that observed with
EROD, a trend toward a decrease in the kinetic order with increasing dose. Because of the
variability of the data and low R2 the significance of these observations can not be determined.
Table XV Power Law Functional Fit
Selected Subsets of Doses from Devito et al. 1992
Dose ng/kg/day
1.5 - 15
1.5 -45
1.5 - 150
4.5 -150
15 - 150
Hepatic EROD
.47
.85
.68
.2
.82
.69
.63
.92
.96
.94
Lung EROD
.87
.78
.79
.77
.74
.92
.94
.96
.95
.91
Skin EROD
.75
.71
.75
.81
.74
.80
.85
.93
.91
.81
Hepatic 1A2
.34
.27
.24
.11
.18
.59
.65
.73
.53
.36
Human and animal P450 enzymes differ both quantitatively and qualitatively.
Understanding the role of these enzymes in the multistep process of TCDD induced
carcinogenesis is essential if enzymes are to be used as response surrogates. Additionally, the
role of these enzymes in immunotoxicity, male reproductive toxicity, teratogenicity, or other
non-cancer endpoints is not understood and is an area where additional research is needed.
The role of CypIAl in the toxicity of TCDD is not yet fully understood. CypIAl is
constituitively expressed both hepatically and extrahepatically in rodents. In man CypIAl is
present and inducible extrahepatically, and current investigations in its role in lung cancer
indicate that genetic variations of the CYPIA1 gene may increase the relative risk of lung
cancer in humans.
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Induction of CypIAl by TCDD in the liver has been used to evaluate cancer risk.
Evidence that rodents and man may induce this enzyme differently may indicate that the
estimation of human health risks based on rodent CypIAl induction is not appropriate. CypIA2
is present constitutively in human livers at widely variable levels, but is not found
extrahepatically in humans and in rodents (Wrighton and Stevens 1992). In rodents CypIA2 is
found constituitively at higher levels that rodent CypIAl (Trischer et al. 1992; Kedderis et al.
1991) The induction of CypIA2 may alter the rate of metabolism of endogenous materials
thereby altering their fuction. If enzyme induction is used as a surrogate for TCDD toxicity
then extrahepatic induction of CypIAl and hepatic induction of CypIA2 are areas of
investigation relevant to human health effects.
Sunahara et al. (1989) reported a dose-dependent decrease in hepatic EGF receptor after
administration of 1.0-25.0 ug TCDD/kg po in Sprague-Dawley female rats (Table XVI). EGF is
not produced in the liver, but extrahepatically, therefore the levels of EGF in adrenalectomized
S-D rats was also investigated. A decrease in the EGF receptor binding was found from 1.0 to
5.0 ug TCDD/kg. Graph Q depicts similiar hepatic EGF receptor binding in intact and
adrenalectomized S-D rats administered a dose range of 1 to 5 ug TCDD/kg. However, at doses
below 1.0 ug TCDD/kg in adrenalectomized animals there was a marked increase in EGF
receptor binding levels. Unfortunately, doses between 0.001 and 0.5 ug TCDD/kg were not
tested in intact rats.
An 80-90% decrease in high and low affinity hepatic EGF binding was observed in
C57BL/6J mice 7 days after treatment with 30 ug TCDD/kg. There was no decrease in EGF
receptor mRNA as a consequence of TCDD treatment indicating that the effect may ocur post-
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transcriptionally. The transcriptional alteration of an EGF receptor ligand has been a
hypothesized effect of TCDD (Lin et al. 1991). The significance of these decreases to
carcinogenesis was discussed in this document (Lucier Chapter). EGF may be involved in the
carcinogenic mechanism, but there are not clear cut dose-response relationships in the path from
decreased hepatic EGF to cancer.
Epidermal growth factor is believed to be a regulator of cell proliferation (Clarke et al.,
1991). A dose-dependent increase in EGFr detected immunohistochemically in mouse
embryonic uteric epithelial cells (GD14) of C57BL/6N mice was observed (Abbott and
Birnbaum 1990a). A consequence of this is hyperplasia of the uteric epithelial cells, and
hydronephrosis. There was not enough data points available for this endpoint to fit it to a
dose-response curve.
The level of EGR receptor in embryonic palate closure has been studied in C57BL/6n
mice (Abbott and Birnbaum 1989) and in humans (Abbott and Birnbaum 1991). Although
statistically significant changes in EGF receptor level were not found in the medial epithelial
cells of human palate after exposure to TCDD the C57BL/6N mice the EGF receptor level was
altered. In normal the normal palate, closure involves the decrease in
EGF receptor and a subsequent programmed cell death, or lack of cell growth. A sufficient
decrease in EGF receptor does not occur in the TCDD treated palate cultures preventing normal
palate closure. It has been hypothesized that other factors (TGF-a, or TGF-b play a role in its
regulation, but sufficient data is not available to determine a dose-response relationship (Abbott
et al. 1991)
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TABLE XVI DOSE-RESPONSE FEMALE SPRAGUE-DAWLEY RATS
Epidermal Growth Factor Response Intact V. Adrenalectomized tats
Species
Female
Sprague-
Dawley rats
Female
Sprague-
Dawley rats
(200-250 g)
adrenal-
ectomized.
Dosing
regimen
single p.o
single p.o
Effect
% change EOF receptor
binding levels
10 days post-dosing
% change EOF
receptor binding levels
10 days post dosing
Dose
Oug/kg
1.0
2.5
5.0
10.0
25.0
Oug/kg
0.001
0.01
0.10
0.50
1.0
2.5
5.0
Response % Change
0
-14%
-19%
-39%
-54%
-64%
-3%
+27%
+2%
+7%
-12%
-48%
-43%
NOAEL
no
significance
determined
no
significance
determined
LOAEL
no
significance
determined
no
significance
determined
Reference
Sunuhara et al.,
1989
Sunuhara et al.,
1989
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ra. CONCLUSION
Understanding the consistencies and inconsistencies of dose response relationships is
necessary and can add confidence to a risk assessment. TCDD has been extensively studies with
various species and strains of animals using different protocols resulting in many consistencies in
response. Thymic atrophy, wasting syndrome and liver toxicity are seen in most controlled animal
experiments with TCDD. TCDD is also acutely toxic in many species tested, hut the sensitivity
to acute toxicity differs with species and strain and does not appear as a constant predictor of other
endpoints. Female rats and guinea pigs are less sensitive to acute toxicity than males, but are more
sensitive to chronic toxicity to liver. Hamsters are relatively insensitive to acute toxicity but are
the very sensitive to fetal toxicity (hydronephrosis). Carcinogenic responses are less well
understood but TCDD is clearly a low-dose rodent carcinogen. Where humans respond to each of
these endpoints is a critical question in risk assessment which can not be easily answered without
a better understanding of the underlying mechanisms of toxicity.
The segregation of toxic effects with the Ah receptor is an often reported phenomenon and
it is clear that most responses to TCDD in cells in culture and in animal tissues segregate with the
Ah receptor. Generally atypical tissue responses within species are also Ah receptor mediated, for
example, the response of BALB/cByJ mice (Ahb/Ahb) to TCDD in the PFC assay for
imrmmoresponse is very sensitive, whereas the DBA mouse (Ahd/Ahd) is much less sensitive
(Silkworth et al., 1989). The kinetics of tissue distribution appear to have some role in different
sensitivities. For instance, the half-life in the thymus of the hamster is short and this species
requires a comparatively high dose of TCDD to produce thymic atrophy. Pohjanvirta et al., (1989,
1990) reported the same phenomenon in Long Evans and Han Wistar rats, finding that the half-life
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in the thymus was related to the animals sensitivity to thymic atrophy. These differences can be
accounted for by several factors; primarily pharmacokinetic and dynamic differences. However,
as with steroids, tissue and cell factors may also be involved to account for specificity.
CypIAl or IA2 induction have served as the major surrogates to TCDD activity and has led
to a better understanding of tissue susceptibility and species differences. Induction of CypIAl in
humans is difficult to quantify however, CypIA2 has been proposed as a reasonable biomarker
when coupled with tissue levels of TCDD (or congeners). The level of CypIA2 in the human liver
(which has highly variable constitutive levels) may account for differences in human responses.
The role of CypIAl and CypIA2 induction in cancer is described in this document(Lucier chapter).
The relationships between enzyme induction and the increase in toxicity at non-cancer endpoints
is not well understood, and this is an area where further research is needed.
In animal studies graphic analysis of the data revealed that the Power Law function was the
best fit for dose-response relationships of Total Ah receptor binding, CypIAl mRNA concentrations
and EROD activity. The consistency of the power function seen with the graphic analysis of these
endpoints (Graph K) indicates that there is little post-translational modification. The "goodness
of fit" to the power function was high for these responses and lines were parallel, confirming what
is known of the biological relationships between receptor, mRNA and protein. Graph M shows
excellent correlation coefficients for the EROD activity in three tissues, revealing a good fit (R2
>0.90) to the Power Law function. The consistency of the Power Law function demonstrates that
EROD activity in extrahepatic tissues is probably not modified, and from this, indicates that in mice
liver EROD activity is a good surrogate for extra hepatic tissue activity. The difference in the
power law for hepatic CypIA2 confirms what is known about CypIA2; that is, it regulated
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differently that CypIAl. Finally, the analysis of selected data sets with the Power Law function
has added to the understanding of the dose-response relationships for these endpoints. The enzyme
endpoints are easily analyzed and may prove to be a simple surrogate for some other toxic effects.
An "inertial" response in the low dose range can be characterized when applying this function to
total P450, EROD and benzo(a)pyrene hydroxylase in the liver. This suggests that the response
at low doses may be part of an adaptive response to environmental levels of TCDD ("external
forces") and that only at higher doses does the system respond with an increase in kinetic order.
The curve fitting for the CypIAl /CypIA2 regulated endpoints is contrasted to the immunotoxicity
data, where the measured endpoints are reached, presumably, through a more complex and less well
understood pathway.
Immunotoxicity and thymic atrophy are frequently reported toxic responses to TCDD
administration. The application of the Power Law function to these endpoints suggests a very
different response than that seen in enzyme induction. At low doses a high kinetic order (reaction
rate) is seen contrasted with a lower kinetic order at high doses. This suggests that additional
factors, not directly related to enzyme induction, are involved in the these clinical endpoints.
Although the immune system may be extremely sensitive to the toxic effects of TCDD there does
not seem to be an assay that has developed the sophistication to make interspecies comparisons.
Some organ weight changes were sensitive clinical endpoints. Thymic atrophy and liver
weight increases were observed after a single oral dose of 1 ug TCDD/kg in mice, and with a
continuous dose of 4.9 TCDD ng/kg/day in guinea pigs (See Table III and Table XI). Although
individual data sets fit the Power Law function with a high correlation coefficient, there wasn't
enough consistency between study methodology to adequately compare studies. Current research
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in the cellular mechanism of thymic atrophy and understanding of thymocyte maturation has
developed to the point where a mechanistic model can be proposed. Additional quantitative data
on thymocyte maturation across species is needed for this model to develop.
The development of a in vitro model for cleft palate has enabled direct species
comparison, and has been useful in understanding these species differences. Qualitatively the
measured endpoints in this model have shown consistency across species and have helped to
understand the role of the EOF receptor in programmed cell death. There were not enough data
points to analyze this endpoint using the Power Law function. Although, qualitative comparisons
indicate that rat and human palate are less sensitive than the palates of C57BL/6N mouse. Because
of the understanding of the mechanism for cleft palate formation a mechanistic model can be
developed.
From a public health perspective, the use of non-cancer endpoints may be more significant
in assessing human health risks because some of these effects occur at lower doses and have been
reported to occur under general population exposure conditions, ie. short to moderate exposure
periods, or acute accidental exposures. A comparison of tissue-dose with effect in a variety of
studies exemplifies this (Table XVII). The apparent differences between endpoints for cancer and
non-cancer risks are in the extrapolation procedure which are questioned in the modelling chapter
(Lucier and Gallo).
Kociba et al. (1978) performed a rodent bioassay to assess the carcinogenic risk of TCDD
study supplying Sprague-Dawley rats with feed containing concentrations equivalent to 0, 0.001,
0.01, and 0.1 ug TCDD/kg/day. The administered dose associated with a statistically significant
increase in hepatocellular carcinoma in female rats was 0.1 ug TCDD/kg/day. At 0.01 ug
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TCDD/kg/day female rats were found to have a statistically significant increase in hepatocellular
nodules and alveolar hyperplasia. These doses produced liver concentrations of 24,000 ppt and
5100 ppt, respectively (Kociba 1978).
Using an initiator-promoter dosing regimen, Trischer et al., (1992) found an increase in
CypIAl and Cyp IA2 in rodent liver tissue after administration of weekly doses equivalent to 3.5
ng/kg/day to 125 ng/kg/day. A dose-dependent increase in these biochemical endpoints was
significant at all doses. The measured concentration of TCDD in the liver was reported as 480 ppt
to 1990 ppt.
In contrast to these two studies, liver concentrations of TCDD (tissue dose) that statistically
increases non-cancer endpoints are as low as 0.084 ppt. Narasimhan et al., (1993) administered
a single i.p. dose of 0 to 2.5 ug TCDD/kg to B6C3F1 female mice, and found a statistically
significant decrease in splenic plaque forming cells (PFCs/spleen) at the 0.05 ug TCDD/kg dose
level. The liver concentration that resulted in this decrease was 0.084 ppt.(calculated from
Narasimhan et al. 1993)
Ma et al. 1991 administered female C57BL/6J mice single i.p. injections of TCDD and
measured the decrease in hepatic protein tyrosine phosphorylation. The EC50 for this decreased
response was 0.05 ug TCDD/kg which was approximately equivalent to a tissue dose of 0.084 ppt.
(the tissue distribution data used to calculate this was from Narasimhan et al., 1993). This data also
reports an effect that is not clearly nuclear-Ah receptor mediated, the decrease in protein tyrosine
phosphorylation.
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Table XV11 Summary LOEL Non-Cancer Endpoints vs. Cancer Endpoints
See text for description of study protocols
Endpoint
increase hepatic hyperplastic
nodules
increase CyplAl/CyplA2
increase benzo(a)pyrene
hydroxylase (Ahh)
increase EROD
reduction in spermatogenesis in
male offspring
Increase in protein tyrosine
phosphorylation
decrease splenic plaque forming
cells
LOAEL - Tissue dose - liver ppt
5100 ppt
480 ppt
5.4 ppt
10.2 ppt
maternal tissue dose 806 ppt
fetal liver dose at 0.083 ppt
0.084 ppt (estimated)
0.084 ppt
Reference
2 year Rodent Bioassay -
Kocibaet al., 1978
Initiation-Promotion
Protocol - Trischer et al.,
1992
Kitchin and Woods, 1979
Abraham et al., 1988
Mably et al. 1992
Maetal., 1992
Narasimhan et al., 1993
The increase hepatic enzyme activity is a well documented phenomenon which may
ultimately be responsible for alterations in clinical endpoints. Kitchin and Woods, 1979 found that
enzyme activity (benzo(a)pyrene hydroxylase) was significantly increased in Sprague-Dawley rats
after the single oral administration of 0.002 ug TCDD/kg (LOAEL). Based on [1,6-3H] TCDD
administration the tissue concentration in the liver that resulted in this increased enzyme activity
was 5.4 ppt. Kitchin and Woods (1979) estimated that this level is equivalent to 65 molecules of
TCDD per liver cell. Abraham et al. (1988) reported a statistically significant increase in
ethoxyresorufm O-deethylase activity (EROD) after a single s.c. administration of 0.003 ug
TCDD/kg resulting in a measured 14C-TCDD liver concentration of 10.2 ppt. These two studies
are remarkably consistent, in reporting liver concentrations which are less than 1/200 the
concentrations in the carcinogenicity bioassays.
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Other studies have also found very low TCDD concentrations that result in adverse effects.
Mably et al. (1992) reported that a single maternal dose of 0.064 ug TCDD/kg resulted in a
significant and permanent reduction in spermatogenesis in their male offspring. If, as VanBerg et
al. (1987) reported, that 0.13% of TCDD reaches the fetus, then the fetal dose that resulted in these
effects would be 0.083 ppt.
Clearly hepatocellular carcinoma is not the most sensitive effect of TCDD administration
in animals. Induction of cancer appears to occur after continuous exposure to doses that result in
high levels of enzyme induction. The most significant public health concern from TCDD exposure
may be immunotoxicity, reproductive toxicity and other effects that are seen in human and animal
after low dose acute exposures. Empirical analysis of these effects will help to determine the
biological significance of these observations and result in a more relevant risk assessment.
The evaluation of dose-response relationships has demonstrated many consistencies and
inconsistencies in the patterns of toxicity between tissues, strains, and species. The development
of new models has been proposed for cleft palate development and thymic atrophy because they
are well understood and sufficient data has been collected to analyze additional data needs and start
mechanistic model development. Future research focused on cleft palate, thymocyte maturation and
immunotoxic endpoints, and male reproductive toxicity should result in progress toward this end.
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REFERENCES
Abbott, B.D., and Bimbaum, L.S.(1989). TCDD Alters Medial Epithelial Cell Differentiation during
Palatogenesis. Toxicol. Appl. Pharmacol. 99, 276-286.
Abbott, B.D., and Bimbaum, L.S., and R.M. Pratt.(1987) TCDD-Induced Hyperplasia of the
Ureteral Epithelium Produces Hydronephrosis in Murine Fetuses. Teratology 35, 329-334.
Abbott, B.D., Diliberto, J., and Bimbaum, L.(1991). Mechanism of TCDD-mduction of cleft Palate:
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Complement Levels following Exposure to Poly chlorinated Dibenzo-p-dioxins. Toxicol. Appl.
Pharmacol. 84,209-219, 1986.
Whitlock, J.P. (1987) The Regulation of Gene Expression by 2,3,7,8-Tetrachlorodibenzo-p-dioxin.
Pharm. Rev. 147
Whitlock, J.P. (1991)Mechanism of Dioxin action: Relevance to Risk Assessment in Banbury
Report 35-.Biological Basis for Risk Assessment of Dioxins and Related Compounds Ed. Gallo,
M.A., Scheuplein, R.J. and Van Der Heijden, K.A. Cold Spring Harbor Laboratory Press, USA p.
351
Wrighton S. A. Human cytochromes P450 Responsible for Hepatic Drug Metabolism. New
Horizons in Molecular Toxicology, p 80—86. May 21-22, 1990 Indianapolis , Indiana Lilly
Research Laboratories Symposium.
Wrighton, S.A., and Stevens, J.C. (1992). The Human Hepatic Cytochromes P450 Involved in Drug
Metabolism. Crit. Rev. Toxicol.22(l):l-21.
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E - 112
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Appendix 1
The Use of Power Laws, Mkhaelis-Menten and Hill's Function in Dose
Response: The Mathematical Rationale
The use of a particular empirical function in phenomenological dose-response analysis is
guided by conceptual considerations concerning infinitesimal (differential) incremental changes in
response at a given dose level. In the following x andy denote dose and response in appropriate
units, y =f(x) the dose response function and dx, dy are the differentials of the dose and the
response. As a simple example, one may consider the case of linear response: the underlying
assumption is that the infinitesimal change in response is always proportional to, and with a
constant proportionality ratio, a, to the infinitesimal change in dose at all dose levels. So, dy = adx,
which is integrated to give y =ax+b, b being the "response at zero dose".
A nonlinear dose-response function is derived if one assumes that at any given level of dose
and infinitesimal change in response is proportional to an infinitesimal change in dose but with a
proportionality ratio that depends on the levels of dose and response. The specific
assumption that leads to a Power Law is that the incremental change in response will be
proportional to the incremental change in dose as well as to the total response at that point,
but inversely proportional to the total level of dose: dy = gdx—, where g is a
a proportionality constant. This relationship gives, y =axg, where a is an integration constant. The
differential relationship above essentially states that the local rate of change of response with
increasing dose is proportional to the total response normalized over the total dose. This concept,
DRAFT September 15,1993 DO NOT CUE OR QUOTE
E- 113
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i.e. the rate of change of the "dependent" variable with respect to the "independent" variable being
proportional to their ratio, underlies much of the S-system approach to nonlinear phenomenon and
employs a type of scaling that is common in many physiological and other biological systems.
More complicated nonlinear dose-response functions are derived if one assumes a similar but
generalized relationship between incremental response and dose locally, i.e. at a given point of the
dose-response curve, as follows:
dx=g(x)dx—
x
where the proportionality constant g is now replaced by a more general "local modifier function"
g(x). By assigning an appropriate form to g(x) one can appropriately modify the relationship
between dose and response so as to reflect dominance of different biological phenomenon over
different ranges of the dose-response curve. For example, one can choose a functional form for
g(x) such that for small doses the proportionality factor is close to unity and therefore the
relationship dy = gdx -- becomes linear or assumes another constant value, g, in which case the
power law behavior (with power g) dominates at low doses. At the same time one may define g(x)
so that e.g., at high doses the incremental change in response becomes vanishingly small. A
functional form that combines both these requirements is:
In that case
\dy_\ gb
ydx
and, using the property
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E - 114
-------
1 xfi+x8) bS b+x8
and one gets
ax8
y=
b+x*
where a is an integration constant. The above expression is the well known Hill function (which
for g=l) becomes the Michaelis-Menten equation.
DRAFT September 15, 1993 DO NOT CITE OR QUOTE
E - 115
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GRAPH Al
Dose-Response
(Graph in Log-Linear Coordinates)
for
Female Guinea Pigs
(Harris etal., 1973)
Dose: repeat oral dose 1 X weekly for 8 weeks
Response: % decrease thymic weight
0.001
0.01 0.1
Calculated Daily Dose (ug/kg/day)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "1 + Michaelis-Menten" type response)
f(x) =l+4.745175E+6*x/(2.771639E+3+x), chi*2 = 447.7731
E-116
-------
0.001
GRAPH A2
Dose-Response
(Graph in Log-Linear Coordinates)
for
Male and Female Sprague-Dawley Rats
(Kocibaetal, 1976)
Dose: repeat dose 5 X per week for 13 weeks
Response: % Decrease thymus weight
O males
• females
o.i
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "1 +Michaelis-Menten" type response)
males, f(x) = l+1.054825+2*x/(4.255322E-l+x), chi*2 = .5741008
females, f(x) = l+8.165068E+l*x/(8.180588E-2+x), chiA2= 69.15886
E-117
-------
GRAPH Bl
Dose-Response
(Graph in Log-Linear Coordinates)
for
Male C57BL/6 Sch St. mice, (Vos, 1974)
Female C57BL/6 mice (Kerkvliet and Brauner, 1990)
Dose: single oral
Response: % decrease in thymic weight 2 weeks (Vos) and 7 days
(Kerkvliet and Brauner) post dosing
Dose (ug/kg)
O Vos
Kerkvliet and
Brauner
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "1 +Michaelis-Menten" type response)
Vos, f(x) = l+6.046204E+l*x/(1.061148E+l+x), chi*2 = 50.5892
Kerkvliet and Brauner, f(x) = 1+ 1.081043E+2*x/(1.098941E+l+x), chi*2 = 49.71537
E-118
-------
GRAPH B2
Dose-Response
(Graph in Log-Linear Coordinates)
for
Male C57BL/6 Sen St. mice, (Vos, 1974)
Female C57BL/6 mice (Kerkvliet and Brauner, 1990)
Dose: single oral
Response: % decrease in thymic weight 2 weeks (Vos) and 7 days
(Kerkvliet and Brauner) post dosing
1 • • i ' • • i • ' ' i ' ' ' i • ' ' i • • ' i • ' '
25
30
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for Linear type response)
Vos,f(x) =5.767803E+0*x+4.784906E+0, R*2 = .93
Kerkvliet, f(x) = 4.613001E+0*x+-1.106652E+), RA2 = .85
10 15 20
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Linear type response)
Vos, f(x) = 2.784140E+0+8.886964E+0, RA2 = .95
Kerkvliet, F(x) = 4.613001E+0*x+-1.106652E+0, R*2 = .85
O VOS
Kerkvliet and Brauner
E-119
-------
Graph B3
Dose-Response
(Graph in Log-Linear Coordinates)
for
Male mice with differing suseptibility to he Ah induction
(Vecchietal., 1983)
Dose: single ip
Response: % decrease thymus weight 12 days post-dosing
Michaelis Menten Function
1 10
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "1+ Michaelis-Menten" type response)
B6D2F1 f(x) = y=l+5.477307E+l*x/(8.229451E+0+x),chiA2 = 0.33
C57BL/6J f(x) = y = l+6.171059E+l*x/(3.697454E+0+x), chiA2 = 48.9
C3H f(x) = y = 1+ 6.754428E+l*x/(3.767704E+0+x), chiA2 = 14.4
DBA f(x) = y = 2 + 2.20347E+l*x/(3.000336E-l+x), chiA2 =52.3
100
E-120
-------
GRAPH B4
Comparative Dose-Response
(Graph in Log-Log Coordinates)
for
Male mice with differing suseptibility to the Ah induction
(Vecchi et al., 1983)
Dose: single ip
Response: decrease thymus weight
0.1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
B6D2: f(x) = 7.884996E+0*(xA5.296098E-l),RA2=9.729347E-l
C57BL/6J: f(x) = 1.223977E+l*(xA4.942995E-l), RA2 = 8.447337E-1
C3H: f(x) = 1.957259E+l*(xA3.514898E-l),RA2 = 9.833631E-1
DBA:f(x) = 1.658810E+l*(xA1.091679E-l), RA2 = 4.339738E-1
E-121
-------
GRAPH B5
Dose-Response
(Graph in Log-Linear Coordinates)
for
Male mice with differing suseptibility to the Ah induction
(Vecchi et al., 1983)
Dose: single ip
Response: % decrease thymus weight 12 days post-dosingDose-Response
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for Linear function)
C57BL/6, f(x) = 1.145833E+0*x +2.212500E+1, RA2 = 6.173469E-1
C3H , f(x) = 1.283602E+0*x +2.475000E+1, RA2 = 8.888266E-1
DBA , f(x) = 3.360215E-1 *x +1.65000E+1, RA2 = 8.313934E-1
B6D2F1, f(x) = 1.115591E+0*x+1.150000E+l, RA2 = 9.107615E-1
E-122
-------
100
GRAPH B6
Dose-Response
for
male Syrian Hamsters
Olsen et al, 1980
Dose: single oral
Response: % response decrease thymic weight 50 days post dosing
A ... B
100
Dose (ug/kg)
1000
10000 0
500
1000 1500 2000
Dose (ug/kg)
2500 3(XX
Functional Form of Slgmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" type response
f(x) =0.101 + 1.001197E +2 *x/(3.023870E+2+x)
chi*2 = 415.4893
c
-20
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for Linear type response)
f(x) = 3.019634E-2*x +1.829978E+1
R*2 = 6.781075E-1
100 1000
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Hill plot)
f(x) = (9.709755E+1 *xA1.087476E+0)/(4.559997E+2+xM.087476E+0)
= 407.9375
E-123
10000
-------
Graph 67
Dose-Response
(Graph in Log-Linear Coordinates)
Male Syrian Hamsters
(Olsen, 1980)
Dose: single oral
Response: response thymus weight change after SO days-(A)- raw
data
0.1-
0.01-
0.001
100
Dose (ug/kg)
(Best Fit Estimate for Power Law)
f(x) = 2.931550E-l*(xA-3.945743E-l)
RA2 = 7.593673E-1
1000
10000
(B)
(C)
I
1
0.1-
o.oi.
10 100
Dose (ug/kg)
(Best Fit Estimate for Power Law Function)
(doses = 5-500 ug/kg)
f(x) =1.494819E-l*(xM.737934E-l)
RA2 = 8.487105E-1
0.001
1000 1
0.01 T
10
100
Dose (ug/kg)
1000
10000
(Best Fit Estimate for "Power Law Function)
(doses 500-2000 ug/kg)
f(x) =4.705032E+2*(xA-1.484813E+0)
RA2 = 9.992644E-1
E-124
-------
Graph Cl
Dose-Response
(Graph in Log-Linear Coordinates)
Female C57BL/6 mice (Kerkvliet and Brauner, 1990)
Male C57BL/6J mice (Davis and Safe, 1988)
B6C3F1 female mice (Narasimhan et al.,1993)
BALE and DBA male mice (Silkworth et al., 1989)
Dose: Single oral or ip administration
Narasimhan—>
* Davis and Safe
Narasimhan
• Kerkvliet
Silkworth BALE
o Silkworth DBA
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "1 +Michaelis-Menten" type response)
Davis, f(x) = 1 + 1.432263E+2*x/(1.456438E+0+x), chiA2 =849
Narasimhan et al, f(x) = l+(9.592833E+l*x/(1.627160E-l+x), chiA2= 693
Kerkvliet, f(x) = 1+ (8.143631E+l*x)/(4.432174E-l+x), chiA2 = 46.2
Silkworth BALE, f(x) = l+(1.225246E+2+x)/(3.286863E+0+x), chiA2=284
Silkworth DBA, f(x) = l+(9.344592E+l*x)/(2.176351E+0+x), chiA2=280
E-125
-------
Graph C2
Comparative Dose-Response
(Graph in Log-Log Coordinates)
Female C57BL/6 mice (Kerkvliet and Brauner, 1990)
Male C57BL/6J mice (Davis and Safe, 1988)
B6C3F1 female mice (Narasimhan et al., 1993)
BALB and DBA male mice (Silkworth et al., 1989)
Dose: Single oral or ip administration
Response: % decrease PFC/IO^ spleen cells
Narasimhan-->
* Davis and Safe
• Narasimhan
• Kerkvliet
Silkworth BALB
O Silkworth DBA
0.01
0.1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
Davis, f(x) = 4.124948E+l*xA1.125711E+0, rA2 = .7501
Narasimhan et al., f(x) = 8.380591E+l(xA6.873447E-l), ^2 = .7317
Kerkvliet, f(x) = 5.114275E+l*x(xA3.027756E-l), rA2= .9807
Silkworth BALB, f(xO =1.929342E+1 *(xA6.915990E-l), rA2=.9371
Silkworth DBA, f(x) = 3.468637E+1* (xA3.193009E-l), ^2 =.9899
E-126
-------
Chart C3
Dose-Response
(Graph in Log-Linear Coordinates)
for
Female B6C3F1 mice
(Narasimhan et al., 1993)
Response: number of PFC's/10* 6 spleen cells
1200
0
0.001 0.01 0.1 1 10
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" + background parameter type response)
f(x) =939 + -3.005346E+3 *x, chiA2 = 43,447
B
1200-
1000-
800-
600-
400-
200-
0.001
0.01
0.1
Dose (ug/kg)
10
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Hill function type response)
f(x) = 1.204213E+3*xA-7.995208E-l/6.767036E+0+ XA -7.995208E-1, chiA2 = 5834
£-127
-------
10000
I
I
o
1000^
100-
10^
0.001
100 -
10 -
a
S3
06
1 -
O.I
Chart C3
Dose-Response
(Graph in Log-Linear Coordinates)
Female B6C3F1 mice
(Safe, 1992)
Response: % decrease PFC
C
0.01
Dose (ug/kg)
Functional Form of Dose-Response
(Best Fit Estimate for Power Law)
f(x) =1.602198E+2*xA-4.269049E-l, rA2 = .8930
D
0.01
0.1 1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
PFC low dose f(x) = 7.395768E+4*(xA2.933124E+0), RA2=8.481507E-1
PFC high dosef(x) =8.103178E+1*(XA2.074615E-1), RA2= 8.233572E-1
E-128
-------
GRAPH C4
Dose-Response
(Graph in Log-Linear Coordinates)
for
Female B6C3F1 mice, Narasimhan et al., 1993)
Response: number of PFC's/spleen
or % response PFCs/lO^ spleen cells
A
10000 -3
1000-.
100-=
0.001
0.01
0.1
Dose (ug/kg)
Functional Form of Dose-Response
(Best Fit Estimate for Power Law)
f(x) =1.602198E+2 * xA-4.269049E-l, RA2 = 8.929944E-1
Dose -Response - Subsets - % response-B
Dose -Response - Subsets - raw data response-C
100-
I
•8
8
10 -
i
* %PFC low
4 %PFChigh
10 -a
0.1-
0.01
* low dose
• high dose
0.01
10 0.01
0.1 1
Dose (ug/kg) Dose (ug/kg)
Functional Forms of Dose-% Response Curves Functional Forms of Dose- Response Curves
PFC low f(x) =1.164128E44 * XA 2.298986E+0, RA2 = 9.972033E-1 PFC low f(x) = .274749E-l*(xA-5.729254E-l), RA2=9.961E-1
PFC high f(x) = 7.748043E+l*xA2.341332E-l, RA2 = 8.628811E-1 PFC high f(x)=1.474431E-l* (XA-1.078037E+0), RA2=9.594E-1
E-129
10
-------
GRAPH C5
Comparative Dose-Response
(Graph in Log-Log Coordinates)
C57BL/6J mice AhBB and Ahdd
(Kerkvleit et al., 1990a)
Dose: single oral dose TCDD 2 days prior to an ip injection of 25 ug TNP-LPS
Response: decrease spleen cellulararity
10
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
Ahbb, f(x)=l+2.7767E+l*x/(2.1853353E+0+x), chiA2=38.15
Ahdd, f(x) =l+2.385456E+l*x/(4.32568E+0+x), chiA2=56.63
B
100
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
Ahbb: f(x) = 1.089295E+1*(3.140763E-1), RA2 = 8.527404E-1
Ahdd: f(x) = 9.726441E+1*(2.284170E-1), RA2 = 4.834202E-1
E-130
AHbb
AHdd
-------
Graph Dl
Comparative Dose-Response
(Graph in Log-Log Coordinates)f
Male Mice differing in susceptiblility to Ah induction
(Vecchi, 1983)
Dose: single ip
Response: % decrease PFC/splenocytes at 5 days
A
100-r
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best fit Estimates for 1+ Michaelis-Menten)
C57BL/6:f(x) =l+9.448837E+l*x/3.957415E-l+x), chiA2=21.7
DBA: f(x) = l+7.446390E+l*x/(4.459745E+0+x). chiA2=13.7
B6D2F1: f(x) = 1+1.034210E+2*x/(l.53593 lE+O+x), chiA2=0.71
B
10-
10
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
C57BL/6: f(x) = 7.235706E+1 * (xA8.8.0771E-2), rA2 = 9.922455E-1
DBA2: f(x) = 1.477200E+1* (xA4.769771E-l), rA2 = 9.085196E-1
B6D2F1: f(xO = 4.738026E+1 * (xA2.381199E-l),rA2 =9.018871E-1
E-131
100
* C57BL/6
* DBA2
• B6D2F1
-------
1000-q
GRAPH E
Dose-Response
(Graph in Log-Linear Coordinates)
Female CD-I mice
(DeVitoetal. 1992))
Dose: single i.p.
Response: % decrease hepatic estrogen receptor increase hepatic AHH activity
A
10
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" type response)
Hepatic AHH activity : f(x)=4.808068E+2*x/(1.821770E+0+x)
Hepatic ER: f(x) = 1.7.132724E+l*x/(1.761074E+l+x)
B
100
Hepatic ER
Hepatic AHH
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
Hepatic ER: f(x) = 8.829252E+0*(xM.667998E-l),RA2 = 9.842274E-1
Hepatic: f(x) = 1.600130E+2*(xA3.533556E-l),RA2 = 6.889952E-K
E-132
-------
0.01
100-
10-
GRAPHF
Dose-Response
(Graph in Log-Linear Coordinates)
Female Sprague-Dawley Rats
(Astroffet.al, 1991)
Dose: single oral
Response: % decrease uterine c-fos mRNA units
A
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "1 + Michaelis-Menten" type response)
f(x) =1+ 6.515702E+2*x/(8.920061E-l+x)
B
0.01 0.1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
f(x) = 3.001080E+2* (\*6.6285nE-l), RA2 = 9.851087E-1
E-133
-------
Graph G
Dose-Response
(Graph in Log-Linear Coordinates)
for
Male Sprague-Dawley Rats
(Moore et al, 1985)
Dose: single oral
Response: % decrease serum testosterone
A
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for Michaelis-Menten" type response)
f(x) = y=l+1.359828E+2*x/(3.753185E+l+x)
B
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
f(x) = 7.138131E+0 *(xA5.988408E-l),RA2 = 9.252290E-1
E-134
-------
GRAPH H
Dose-Response
(Graphin Log-Linear Coordinates)
for
Female Syrian Hamsters
(Olsenetal, 1980)
Dose: single i.p.
Response: increase liver weight 50 days post dosing
45-
40-
35-
30-
8
IJ5J
I25
s
8 2o^
15-
10-
10
100
Dose (ug/kg)
1000
10000
Functional Form of Sigmoid Dose-Response
Best Fit Estimate for "1 + Michaelis-Menten" type response)
f(x) = 1 +2.156705E+1 *x/(1.447850E+l+x)
E-135
-------
GRAPHI
Comparative Dose-Response
(Graph in Log-Log Coordinates)
for
Female and Male Sprague-Dawley Rats
(Kociba, 1976)
Dose: p.o. 5x/wk 13 weeks
Response: increase liver/bwt ratio
100-
I
8
10-
Females—>
-Males
0.001
0.01
1 i
0.1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
Female: f(x) = 6.181107E+1 * (xM.611284E-l ),RA2 = 9.357175E-1
Male: f(x) = 3.820392E+1 * (xA3.805964E-l), RA2 = 9.310575E-1
* males
* females
E-136
-------
1.8-
1.6-
1.4-
t'^
I 0.6^
O
0.4-
0.2-
0-
0.0001
Graph Jl
Dose-Response
(Graph in Log-Linear Coordinates)
for
Female Sprague-Dawley Rats
(Kitchin and Woods, 1979)
Dose: single oral
Response: raw data change in P-450 at 3 days
A
0.001
0.01
1
0.1
Dose (ug/kg)
Functional form Linear Dose-Response Best Fit
f(x) =3.565270E-2*x+1.05977E+0, RA2=6.72222497E-1
B
10
100
0
0.0001 0.001 0.01 0.1 1 10
Dose (ug/kg)
"Functional Form of Sigmoid Dose-Response
Best fit for Michaelis-Menten + background parameter response)
f(x) = y=0.88 + 7.812964E-l*x/(1.022375E+0+x)
Dose (u
Functional Form of SigmoicTDose-Respoase
(Best Fit Estimate for "Hill function plot
f(x)=(6.466629E-l*xA1.08926E+0)/(3.584389El+
xA1.089267E+0)+0.88; chi square = 9.346E-2
E-137
-------
10-
GRAPH J2
Comparative Dose-Response
(Graph in Log-Linear Coordinates) for
Female Sprague-Dawley Rats (Kitchin and Woods, 1979)
Dose: single oral
Response: raw data P450 (3 days)
A
s
1-
o.i
0.0001
0.1
I
1
"T
10
10-
0.001 0.01
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
complete data set f(x) = 1.293664E+0 * (xA5.855594E-2), RA2 = 8.538984E-1
B
100
1-
0.1-
P450-low
P450-high
0.0001
0.001
0.01
1
10
0.1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
(.0006 to .6 ug/kg) f(x) = 1.152077E+0 *(xA3.457812E-l), RA2 = 6.187106E-1
(0.6 to 20 ug/kg) f(x) = 1.220259E+0 * (xA1.146001E-l), RA2=9.754329E-1
E-138
100
-------
200
Graph Kl
Dose-Response
(Graph in Log-Linear Coordinates)
Female Sprague-Dawley Rats
(Kitchin, 1979)
Dose: single oral
Response: increase Benzo(a) pyrene hydroxylase raw data
A
0.0001
TTIJ
0.01
TT1I
0.1
B
0.001 0.01 0.1 1
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Linear Model" type response)
f(x) = 8.692922E+0*x +4.796840E+1
RA2 = 4.938167E-1
10
100
0.0001 0.001 0.01 0.1 1
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Estimate for "Michaelis-Menten + background parameter")
f(x) = y=4.9 +1.939516E+2*x/(6.164672E-l+x),
chiA2 = 197.2
0.0001 0.001 0.01 0.1 1
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best for" Hill Function" + background parameter)
f(x) =1.0909172E+2 *xA1.066342E+0/ 5.697987E-1 + xA1.066342E+0)
chiA2 = 163
E-139
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K2
Comparative Dose-Response
(Graph in Log-Log Coordinates)
Female Sprague-Dawley Rats
(Kitchin and Woods. 1979)
Dose: single oral
Response: increase Benzo(a) pyrene bydroxylase
A
1000-
100-
8,
-8
10-
0.0001
O.OO1
O.O1
O.I 1
Dose (ug/kg)
Functional Form of Dose-Response
(Best Fit Estimate for Power Law)
f(x) =8.044850E+1 * (xA4.195383E-l), RA2 = 9.310005E-1
1O
1OO
1000 n
100-
10 -
1-
0.0001
0.001
0.01
0.1 1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
(x = 0.0006 to 0.004) f(x) = 2.351223E+l*(xA2.091992E-l) ,RA2 = 9.633734E-1)
(x = 0.02 to 0.6)f(x) • 1.572746E+2 * (xA7.726666E-l), RA2 = 9.638735E-1
(x m 0.6 to 20)f(x) = 1..2000001E+2 * xA1.887329E-l),RA2 = 8.048891E-1)
E-140
10
100
* BaP hydroxylase -Low
• BaP hydroxylase Mid
• BaP hydroxylase -High
-------
350-
300-^
250-
I
200
150-^
100-|
50-
0-
0.001
350
Graph LI
Dose-Response
(Graph in Log-Linear Coordinates)
Female Wislar Rats
(Abraham, 1980)
Dose: single sc
Response raw data P450 at 7 days
A
0.01
1
0.1
Dose (ug/kg)
Functional Form of Dose-Response
(Best Fit Estimate for Linear Model)
f(x) =4.676353E=1 * X+1.913979E+2, RA2 * 8.781851E-1
10
0.001 0.01 0.1 1 10 0.001 0.01 0.1
Dose (ug/kg) Dose (ug/kg)
Functional Form of Sigmoid Dose-Response Functional Form of Dose-Response
(Best Fit Estimate for" Michaelis-Menten + background parameter" (Best Fit Estimate for Hill funcction + background parameter]
f(x) = y= 169 + 3.876606E+2*x/(3.886855E+0+x) f(x) = 3.994308E+2*xA5.478666E-l/(3.021723E+0+
ChiA2 =2769 xA5.478666E-l) + 169, ChiA2 =169
E-141
-------
Graph L2
Dose-Response
(Graph in Log-LogCoordinates)
Female Wistar Rats
(Abraham et al., 1980)
Dose: single sc
Response: raw data: P450
A
1000-
100-
10-
0.001
0.01
f(x)
0.1 1
Dose (ug/kg)
Functional Form of Dose-Response
(Best Fit Estimate for Power Law)
= 2.615298E+2* (xA7.141240E-2), RA2 = 8.300624E-1
10
1000-
JL
100-
10-
* P450-Low
4 P450-High
0.001
0.01 0.1 1
Dose (ug/kg)
Functional Form of Dose-Response
(Best Estimate for Power Law)
P450-Low f(x) = 2.134553E+2*(xA3.591655E-2), RA2 = 8.762258E-1
P450-High f(x) =2.719892E+2*xA1.377786E-l, RA2 = 9.982764E-1
E-142
10
-------
800-
a
I
•o
u
700-
600^
500^
400^
300-^
200^
100-
0-
0.001
Graph Ml
Dose-Response
(Graph in Log-Linear Coordinates)
Female Wistar Rats
(Abraham et al., 1980)
Dose: single sc
Response: raw data: EROD - 7 days
A
^ 1
0.01
1 r
0.1
10
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Linear "type response)
f(x) = 2.527667E+2*x + 6.538808E+1, RA2 = 0.911
1 10 0.001 0.01 0.1
Dose (ug/kg) Dose (ug/kg)
Functional Form of Sigmoid Dose-Response Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" + background parameter" (Best Fit Estimate for "Hill plot with background parameter1)
type response) f(x)=(9.877489E+2*xA1.088899E+0)/1.072719E+0+
f(x) =14.3 + 1.066009E+3*x/(1.273480E+0+x), ChiA2 = 343 xA1.088899E+0)+ 14.3, ChiA2 = 176
E-143
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Graph M2
Comparative Dose-Response
Dose-Response
(Graph in Log-Log Coordinates)
Female Sprague-Dawley Rats
Abraham et al., 1988
Dose: single sc
Response: raw data EROD (7 days)
A
1000-,
i
o.ooi
o.oi
i
o.i
Dose (ug/kg)
Functional Form of Dose-Response
(Best Fit Estimate for Power Law)
f(x) = 3.724271E+2 *(xA5.226669E-l), RA2 =9.528637E-1
1000 -a
I
"8
* EROD - low
EROD - high
0.001
0.01 0.1 1
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
EROD - low f(x) = 3.671388E+1 *(xA1.126439E-l), RA2 = 9.991464E-1
EROD - high f(x) =4.236494E+2 * (xA6.572701E-l), RA2 = 9.912442E-1
E-144
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GRAPH N.
Comparativ Dose-Response
(Graph in Log-Log Coordinates)
for
Female B6C3F1 Mice
(Safe 1993)
Dose: single i.p. injection
Response: 3 days post dosing increase EROD, Cypl Al,CyplA2, Total nuclear Ah Receptor Binding
10000^
1000-
100 -.
0.01
1000
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
EROD f(x) = 5.810675E+1 * ( xM.071389E-l), RA2 = 8.276920E-1
CyplAl f(x) 2.529055E-2 * (xA5.273765E-l), RA2 = 9.472254E-1
Total Binding f(x) = 1.694760E+0 * (xA4.264534E-l), RA2 = 8.042876E-1
CyplA2 f(x) =2.169244E-2*(xA2.824955E-l), RA2 = 6.557616E-1
10000
E-145
* EROD pmol/mg/min
4 CyplAl mRNA
• total receptor binding
O CyplA2
-------
GRAPH O
Comparative Dose-Response
(Graph in Log-Log Coordinates)
Female Sprague-Dawley Rats
(Trischer et al., 1992)
Dose: p.o. biweekly for 30 weeks
Response: increase enzyme activity
1000-
100-
•s
S
10-
* CYPIA2
CYPIA1
10
100
1000
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
CypIAl f(x) = 6.47089E+1 * (xA3.193647E-l), RA2 = 9.546921E-1
CypIA2 f(x) = 3.453639E+1 * (xM.540268E-l), RA2 = 9.869648E-1
E-146
-------
GRAPH PI
Comparative Dose-Response
(Graph in Log-Linear Coordinates)
for
Female B6C3F1 Mice
(DeVito et al., 1992a)
Response: 13 week dosing 5 x per week % increase Hepatic EROD, Hepatic Ia2,
Skin and Lung EROD
7000
6000-
* Hepatic Ia2
4 Hepatic EROD
• lung EROD
D skin EROD
100
1000
Dose (ng/kg/day)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Linear Function)
Hepatic EROD: f(x) = 4.188324E+0*x+5.909044E+2, RA2 = 5.364690E-1
Hepatic Ia2: f(x) = 3.2.205153E+l*x+3.213650E=2, RA2 = 9.458424E-1
skin EROD: f(x) =2.996198E-l*x+2.254425E+0, RA2 = 9.113437E-1
lung EROD: f(x)= 2.360102E+0*x+1.561678E+l, RA2 = 9.399031E+1
E-147
-------
Graph P2
Dose-Response
(Graph in Log-Log Coordinates)
Female B6C3F1 mice
(DeVito et al, 1992a)
Dose: 5 X weekly 13 weeks
Response:% increase CyplA2 level, Hepatic EROD, Lung EROD and Skin EROD
B
1600
7000-
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" type response
Hepatic 1A2 f(x)=l+1.080720E+3*x/(3.669929E+0+x)
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" type response
Hepatic EROD f(x)= 1.08857 lE+4*x/(1.766552E+2+x)
70
1
1000
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "1 Michaelis-Menten" type response)
Skin EROD f(x) = 1.734444E+2*x/(4.054377E+2+x)
1 10 100
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" type response)
Lung EROD f(x) = l+-2.521625E+9*x(-1.0431179+9+x)
1000
E-148
-------
1600-
Graph P3
Dose-Response
(Graph in Log-Log Coordinates)
Female B6C3F1 mice
(DeVito et al, 1992a)
Dose: 5 X weekly
Response:% increase CyplA2 level, Hepatic EROD, Lung EROD and Skin EROD
A B
1400-
1200-
g 1000 H
I
800-j
600-
400-^
200-
0-
7000-
6000-
5000^
I
4000-
3000-
JO
O
2000-
1000-
1
1000
10 100
Dose(ng/kg/day)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for" Hill function + background parameter)
1A2 levels f(x) = 1.242448E+3*xA5.924782E-l) /
4.586583E+0+xA5.924782E-l)+ 135
1
10 100
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Hill function" type response)
Hepatic EROD f(x) =9.115251E+3*xA1.115666E+0/
2.283763E+2+xAl.l 15666E+0 +80)
D
1000
1 10 100 1000
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Hill Function" type response)
Skin E ROD f(x) =2.397953E+2*xA8.615018E-l
/3.332969E+2+xA8.615018E-l+0.83
1 10 100 1000
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Hill Function" type response)
Lung EROD f(x) =2.070050E+3*xA8.436144E-l/
(3.480857E+2+xA8.436144E-l) +3.19
E-149
-------
Response:
Graph P4
Comparative Dose-Response
(Graph in Log-Log Coordinates)
Female B6C3F1 Mice
(DeVito et al., 1992a)
Dose: p.o. 5 X per week 13 weeks
> inaease Hepatic EROD.Hepatic Ia2, Skin and Lung EROD
10000 -q
1000-
I
T3
OJ
i
S
* Hepatic Ia2
© Hepatic EROD
lung EROD
skin EROD
0.1
1000
Dose (ug/kg)
Functional Forms of Dose-Response Curves
(Best Fit Estimates for Power Laws)
Hepatic EROD: f(x) = 1.482129E+2 * (xA6.769077E-l), RA2 = 9.229244E-1
Hepatic Ia2: f(x) = 3.650628E+2 * (xA2.423515E-l), RA2 = 7.277352E-1
Skin EROD: f(x) = 9.676016E-1 * (xA7.538292E-l), RA2 = 9.363354E-1
Lung EROD: f(x) = 6.679660E+0 * (xA7.910685E-l ), RA2 = 9.658391E-1
E-150
-------
0.001
Graph Q
Dose-Response
(Graph in Log-Linear Coordinates)
for
Sprague-Dawley Rat
Dose: single oral
Response: % change in liver EGR receptor binding levels
Dose (ug/kg)
Functional Form of Sigmoid Dose-Response
(Best Fit Estimate for "Michaelis-Menten" + initial response parameter)
Intact f(x) = -14+-8.004886E+l/(1.319891E+l+x)
Adrenaalectomized f(x) = -3-1.072898E=2*x/(6.766685E+0+x)
E-151
U.S. GOVERNMENT PRINTING OFFICE: 1994 — 550-001 /00162
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