United States
Environmental Protection
Agency x
Off ice of Research and
Development
Washington DC 20460
EPA/600/BP-92/001C
August 1994
External Review Draft
<&EPA Health Assessment
Document for 2,3,7,8-
Tetrachlorodibenzo-p-
Dioxin (TCDD) and
Related Compounds
Volume III 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(Do Not
Cite or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.
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Reassessment of Dioxin - External Review Draft - Contents
The following listing is provided to assist you in locating information contained in the two external
review drafts for dioxin.
Health Assessment Document for
2,3,7,8-TetracIiIorodibenzo-j?-dIoxin (TCDD)
and Related Compounds
(EPA/600/BP-92/001a, OOlb, OOlc)
Volume I of HI (EPA/600/BP-92/001a):
Chapter 1. Disposition and Pharmacokinetics
Chapter 2. Mechanisms of Toxic Actions i
Chapter 3. Acute, Subchronic, and Chronic Toxicity
Chapter 4. Immunotoxic Effects
Chapter 5. Reproductive and Developmental Toxicity
i
Chapter 6. Carcinogenicity of TCDD in Animals
I
Volume II of HI (EPA/600/BP-92/001b): !
Chapter 7. Epidemiology/Human Data
Chapter 8. - Dose-Response Relationships
Volume III of IE (EPA/600/BP-92/001c): ;
Chapter 9. Risk Characterization (Note: This third volume of the 3-volume set
integrates health and exposure information on dioxin and related
compounds; approx. 100 pages.)
Estimating Exposure to Dioxin-Like Compounds
(EPA/600/6-88/005Ca, OOSCb, OOSCc)
Volume I of HI (EPA/600/6-88/005Ca): Executive Summary (Note: This first volume of the
3-volume set summarizes the exposure information on
dioxin and related compounds; approx. 100 pages.)
Volume H of HI (EPA/600/6-88/005Cb): Properties, Sources, Occurrence, and Background
Exposure
Volume HI of ffl (EPA/600/6-88/005Cc): Site-Specific Assessment Procedures
1 i Summary Insert
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Reassessment of Dioxin - External Review Draft • Availability
To obtain a copy of other volumes or the 3-volume set of either draft document, interested parties
should contact:
CERI/ORD Publications Center
U.S. Environmental Protection 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.
Note: The two summary volumes also will be made available as a WordPerfect 5.1 file on a PC-
DOS formatted disk. Please request from CERI/ORD by EPA number as follows:
Risk Characterization Chapter (EPA/600/BP-92/001ca)
Executive Summary of the Exposure Document (EPA/600/6-88/005Caa)
Copies of both documents also will be available for purchase (individually or by set) from the:
National Technical Information Service (NTIS)
5285 Port Royal Road
Springfield, VA 22161
telephone (703) 487-4650, fax (703) 321-8547
Please provide the NTIS PB No. for each volume or set ordered.
Health Assessment Document: Vol.1 PB94-205465
Vol. II PB94-205473
Vol. HI PB94-205481
3-volume set PB94-205457
Exposure Assessment:
Vol. I
Vol. II
Vol.m
PB94-205507
PB94-205515
PB94-205523
3-volume set PB94-205499
Both draft documents will be made available for review at the EPA's Headquarters Library (ORD's
Public Information Shelf), 401M 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 EPA regional and laboratory libraries.
The documents also will be made available to the U.S. Government Depository Libraries. Through
the Depository Library program, government publications are provided free of charge to over 50 regional
depositories throughout the United States and its territories. An additional 1,350 depositories in the system
choose to receive select publications of interest to meet local needs.
Summary Insert
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DRAFT i EPA/600/BP-92/001c
DO NOT QUOTE OR CITE August 1994
1 External Review Draft
Health Assessment Document for
2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
and Related Compounds
Volume III of III
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.
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Health Assessment Document for
2,3,7,8-TetrachIorodibenzo-/>-dioxin (TCDD)
and Related Compounds
TABLE OF CONTENTS - OVERVIEW
Volume I !
I
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 HI
9. RISK CHARACTERIZATION OF 2,3,7,8-TETRACHLORODIBENZO-/7-DIOXIN
(TCDD) AND RELATED COMPOUNDS
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Health Assessment Document for
2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
and Related Compounds
CONTENTS - VOLUME III
List of Tables ,, . . Ill-vi
List of Figures Ill-vi
List of Abbreviations and Acronyms III-vii
Preface III-xv
Authors, Contributors, and Reviewers III-xix
9. RISK CHARACTERIZATION OF 2,3,7,8-TETRACHLORODIBENZO-p-DIOXIN
(TCDD) AND RELATED COMPOUNDS 9-1
9.1. INTRODUCTION . 9-1
9.2. CHEMICAL STRUCTURE AND PROPERTIES 9-3
9.3. ENVIRONMENTAL FATE 9-7
9.4. SOURCES 9-8
9.4.1. Levels in the Environment and in Food .". 9-13
9.4.2. Background Exposure Levels 9-14
9.4.3. Highly Exposed Populations 9-19
9.5. DISPOSITION AND PHARMACOKINETICS 9-23
9.6. MECHANISMS OF DIOXIN ACTION 9-30
9.7. TOXIC EFFECTS OF DIOXIN 9-36
9.7.1. General Comments . . 9-36
9.7.2. Chloracne 9-38
9.7.3. Carcinogenicity 9-39
9.7.4. Reproductive and Developmental Effects . 9-44
9.7.5. Immunotoxicity 9-48
9.7.6. Other Effects , 9-50
9.7.6.1. Circulating Reproductive Hormones 9-51
9.7.6.2. Diabetes and Fasting Serum Glucose Levels 9-51
9.7.6.3. Enzyme Induction 9-51
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CONTENTS (continued)
9.7.6.4. Gamma Glutamyl Transferase (GOT) Activity 9-52
9.7.6.5. Endometriosis i .............. 9-52
9.8. DOSE-RESPONSE CONSIDERATIONS ; 9.53
9.9. USE OF TOXICITY EQUIVALENCE 9.59
9.10. KEY ASSUMPTIONS AND INFERENCES 9.70
9.11. OVERALL CONCLUSIONS REGARDING THE IMPACT OF DIOXIN
AND RELATED COMPOUNDS ON HUMAN HEALTH 9-74
REFERENCES FOR CHAPTER 9 9_89
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LIST OF TABLES
9-1. Toxicity Equivalency Factors (TEF) for CDDs and CDFs 9-5
9-2. Effects of TCDD and Related Compounds in Different ,
Animal Species • 9-37
9-3. Estimated Body Burdens of Experimental Animals and Humans Exposed to
Dioxins: Responses in Humans Causally Associated With Exposure to
Dioxins and Comparable Effects in Experimental Animals 9-55
9-4. Estimated Body Burdens of Experimental Animals and Humans Exposed to
Dioxins: Responses in Humans Associated With Dioxin Exposure and
Comparable Effects in Experimental Animals 9-59
9-5. Estimated Body Burdens of Experimental Animals and Humans Exposed to
Dioxins: Low-Dose Effects in Animals Exposed to Dioxins and Their
Relationship to Background Human Exposure 9-62
9-6. Comparison of the Effects of TCDD Exposure on Human and Animal Tissue
In Vitro - - 9-64
LIST OF FIGURES
9-1. Dioxin and similar compounds—chemical structure 9-4
9-2. Schematic representation of the complex sequence of molecular and biological
events Involved in dioxin-mediated toxicants • 9-33
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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-p-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-p-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 Diemylnitrosamine
DHT 5ce-Dihydrotestosterone
DIS Diagnostic Interview Schedule
DMBA Dimethylbenzanthracene
DMSO Dimethyl sulfoxide
DNA Deoxyribonucleic acid
DRE Dioxin-responsive enhancers
DTH 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-Ethoxycoumarin-O-deethylase
EEG Electroencephalogram
EGF 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
HPAH
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)
j
Gamma glutamyl transpeptidase
i
Gonadotropin-releasing hormone ,
Glutathione-S-transferase
Graft versus host
Halogenated aromatic hydrocarbons
Hexachlorobenzene
Hexachlorodibenzo-j>dioxin j
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 L-lactate dehydrogenase
LH Luteinizing hormone
LDL Low density liproprotein
LMS Linearized multistage
LPL Lipoprotein lipase activity
LOAEL Lowest-observable-adverse-effect level
LOEL Lowest-observed-effect level
LPS Lipopolysaccharide
MACDP Metropolitan Atlanta Congenital Defects Program
3-MC 3-Methylcholanthrene
MCDF 6-Methyl-l,3,8-trichlorodibenzofuran
MCF-7 (breast cancer cell)
MCM Millon Clinical Multiaxial Inventory
MCPA (4-Chloro-2-methylphenoxy)acetic acid
MCPB 2-Methyl-4-chlorophenoxybutyric acid
MCPP 2-(4-Chloro-2-methylphenoxy)-propanoic acid
MFO Mixed function oxidase
MMPI Minnesota Multiphase Personality Inventory
MLE Maximum likelihood estimate
mRNA Messenger RNA
MNNG //-methyl-A'-nitrosoguanidine
NADP Nicotinamide adenine dinucleotide phosphate
NADPH Nicotinamide adenine dinucleotide phosphate (reduced form)
NaTCP Sodium 2,4,5-trichlorophenate
NHL Non-Hodgkin's lymphoma
NIEHS National Institute of Environmental Health Sciences
III-x
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
NIOSH National Institute for Occupational Safety and Health
NK Natural killer ' ', '
NOAEL No-observable-adverse-effect level :
NOEL No-observed-effect level
NTP National Toxicology Program
OCDD Octachlorodibenzo-p-dioxin <
OCDF Octachlorodibenzofuran ' <
OR Odds ratio
OVX Ovariectomized !
PAA Phenoxyacetic acid
PAH Polyaromatic hydrocarbon
PBA , Phenoxybutyric acid
PBB Polybrominated biphenyl
PBF Percent body fat
PBL Peripheral blood lymphocytes
PB-PK Physiologically based pharmacokinetic .
PCB Polychlorinated biphenyl
i
PCBA Phenoxybutyric acid
PCDD Polychlorinated dibenzodioxin
PCDF Polychlorinated dibenzofuran I
PCP Pentachlorophenol
I
PCPA Parachlorophenoxyacetic acid
PCQ Quaterphenyl
PCT Porphyria cutanea tarda
PeCDD Pentachlorinated dibenzo-/?-dioxin
PeCDF Pentachlorinated dibenzo-j7-dioxin
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LIST OF ABBREVIATIONS AND ACRONYMS (continued)
PEPCK Phosphoenol pyruvate carboxykinase
PFC Plaque-forming cell
PGEa Prostaglandin E2
PGF2tt 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
SGOT Serum glutamic oxaloacetic transaminase
SGPT Serum glutamic pyruvic transaminase
IH-xii
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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
tt/t Half-time [
TBB Tetrabromobiphenyl
TBDD Tetrabrominated dibenzo-p-dioxin
TBDF Tetrabrominated dibenzo-p-furan
TBG Thyroxine-binding globulin
TBP Thyroxine-binding protein
TCAOB tetrachloroazoxybenzene
TCB Tetrachlorobiphenyl
TCDD Tetrachlorodibenzo-p-dioxin
TCDF Tetracblorodibenzofuran
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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i
PREFACE
i ,
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-jp-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. i
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-tefrachlorodibenzo-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
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PREFACE (continued)
Assessment Document for 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 noncsmcer, 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 arid 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.
i
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
Ulf GJAhlborg
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
Melviri 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
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
*Involved with an early draft, but no longer working on the reassessment project.
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
Peter W. Preuss
Dwain Winters
Executive Director, Risk Assessment Forum, Office of Research and
Development, U.S. Environmental Protection Agency, Washington,
DC
Director, Office of Science, Planning, and Regulatory Evaluation, U.S.
Environmental Protection Agency, Washington, DC
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
M. Judith Charles
Michael Denison
Department of Pharmacology and Toxicology, Dartmouth
Medical School, Hanover, NH
Department of Environmental Sciences and Engineering,
University of North Carolina, Chapel Hill, NC
Department of Biochemistry, Michigan State University, East
Lansing, MI
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Phillip Enterline
Mark Feeley
Thomas A. Gasiewicz
James Gillette
Claude Hughes
Curtis D. Klaassen
Daniel Krewski
Suresh Moolgavkar
Jay Silkworm
Thomas Webster
Emeritus Professor of Biostatistics, University of Pittsburgh,
Pittsburgh, PA
Toxicity Evaluation Division, Bureau of Chemical Safety,
Health, and Welfare, Ottawa, Ontario, Canada
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
Associate Administrator for Science, Agency for Toxic
Substances and Disease Registry, Atlanta, GA
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Clinical Assistant Professor, Department of Social and
Preventive Medicine, State University of New York,
Buffalo, NY
Professor, Department of Environmental Health, University of
Washington, Seattle, WA
Emeritus Professor of Biostatistics, University of Pittsburgh,
Pittsburgh, PA
Director of Epidemiology, BASF Corporation, Parsippany, NJ
Professor of Epidemiology, University of California,
Berkeley, CA
Assistant Professor of Epidemiology, School of Public Health,
University of Texas, Houston, TX
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*
Germaine Buck
Harvey Checkoway
Phillip Enterline
M. Gerald Ott
Allan H. Smith
Anne Sweeney
Karen Webb
'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.
III-xxiv
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AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued)
Department of Labor (Occupational Safety and Health Administration)
Department of Veterans Affairs I
Executive Office of the President
Office of Science and Technology Policy
Council of Economic Advisors ;
Domestic Policy Council
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9. RISK CHARACTERIZATION OF DIOXIN AND RELATED COMPOUNDS
9.1. INTRODUCTION
Chlorinated dibenzo-p-dioxins and related compounds (commonly known simply as
dioxins) are contaminants present in a variety of environmental media. This class of
compounds has caused great concern in the general public as well as intense interest in the
scientific community. Much of the public concern revolves around the characterization of
these compounds as among the most potent "man-made" toxicants ever studied. Indeed,
these compounds are extremely potent in producing a variety of effects in experimental
animals based on traditional toxicology studies at levels hundreds or thousands of times lower
than most chemicals of environmental interest. In addition, human studies demonstrate that
exposure to dioxin and related compounds is associated with subtle biochemical and
biological changes whose clinical significance is as yet unknown and with chloracne, a
serious skin condition associated with these and similar organic chemicals. Laboratory
studies suggest the probability that exposure to dioxin-like compounds may be associated with
other serious health effects including cancer. Human data, while often limited in their ability
to answer questions of hazard and risk, are generally consistent with the observations in
animals. Whether the adverse effects noted above are expressed in humans, or are detectable
in human population studies, is dependent on the dose absorbed and the intrinsic sensitivity
of humans to these compounds. Recent laboratory studies have provided new insights into
the mechanisms involved in the impact of dioxins on various cells and tissues and,
ultimately, on toxicity. Dioxins have been demonstrated to be potent modulators of cellular
growth and differentiation, particularly in epithelial tissues. These data, together with the
collective body of information from animal and human studies, when coupled with
assumptions and inferences regarding extrapolation from experimental animals to humans and
from high doses to low doses, allow a characterization of dioxin hazards.
This chapter presents a risk characterization for dioxin and related compounds. In the
risk characterization, key findings pertinent to understanding the hazards and risks of dioxin
and related compounds are described and integrated. All of the available information is
considered in proposing hypotheses or in reaching conclusions. The risk characterization is
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not meant to be an executive summary of the extensive data base that has been analyzed in
detail in preceding chapters and in the Exposure Document. Risk characterization requires a
discussion of likely routes, patterns, and levels of exposure as well as aspects of hazard and
dose response. Information contained in the document titled Estimating Exposure to Dioxin-
like Compounds (U.S. EPA, 1994), hereafter referred to as the Exposure Document, will be
integrated with the health effects information on this class of compounds found in previous
chapters of this assessment. The risk characterization articulates the strengths and
weaknesses of the available evidence and clearly presents assumptions made and inferences
used. Risk is characterized in both qualitative and quantitative terms, as appropriate.
Finally, overall conclusions regarding the health risks of dioxin and related compounds are
presented.
The process for developing this risk characterization of dioxin and related compounds
has been an open and participatory one. The Health Assessment and Exposure Documents
that provide the basis for this characterization have been developed in collaboration with
scientists from within and from outside of the Federal Government. Each of these has
undergone extensive internal and external review, including review at a meeting of experts
after a first draft was completed. Additional input to this characterization comes from
comments on those draft chapters as well as from the panel of experts that met in September
1992. Panel members were asked to provide their perspective on themes to be carried into
the characterization and their contributions are reflected here. Finally, the characterization,
as presented here, reflects review and comment by both those Federal scientists involved in
developing the health assessment and exposure chapters as well as representatives of other
Federal agencies. However, the views expressed in this characterization are those of the
collective authors and, as a draft undergoing public comment and further external review, no
Agency-level endorsement should be inferred at this time.
Once fully peer reviewed and revised accordingly, this risk characterization is meant
to provide a balanced picture of the scientific findings of the health and exposure assessments
for use by risk managers in selecting risk management options, based on this and other
information. As an integrated analysis of a complex data base, it is meant to answer key
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questions concerning the science behind concerns for dioxins and! should be useful in
developing strategies for risk communication.
i
9.2. CHEMICAL STRUCTURE AND PROPERTIES
Polychlorinated dibenzodioxins (PCDDs), polychlorinated dibenzofurans (PCDFs),
and polychlorinated biphenyls (PCBs) are chemically classified as halogenated aromatic
hydrocarbons. The chlorinated and brominated dibenzodioxins and dibenzofurans are
tricyclic aromatic compounds with similar physical and chemical properties, and both classes
are similar structurally. Certain of the PCBs (the so-called coplanar or mono-ortho coplanar
congeners) are also structurally and conformationally similar. The most widely studied of
these compounds is 2,3,7,8-tetrachlorodibenzo-/>-dioxin (TCDD). This compound, often
called simply dioxin, represents the reference compound for this class of compounds. The
structure of TCDD and several related compounds is shown in Figure 9-1.
For purposes of this document, dioxin-like compounds are defined to include the
subset of this class of compounds, which are generally agreed to produce dioxin-like toxicity.
These compounds are assigned individual toxicity equivalence factor (TEF) values as defined
by international convention (U.S. EPA, 1989). Results of in vitro and in vivo laboratory
studies contribute to the assignment of a relative toxicity value. TEFs are estimates of the
toxicity of dioxin-like compounds relative to the toxicity of TCDD, which is assigned a TEF
of 1.0. All chlorinated dibenzodioxins (CDDs) and chlorinated dibenzofurans (CDFs) with
chlorines substituted in the 2,3,7, and 8 positions are assigned TEF values. Additionally, the
analogous brominated dioxins and furans (BDDs and BDFs) and .certain polychlorinated
biphenyls have recently been identified as having dioxin-like toxicity and thus are also
included in the definition of dioxin-like compounds. Generally accepted TEF values for
chlorinated dibenzodioxins and dibenzofurans are shown in Table 9-1. A recent World
Health Organization/International Program on Chemical Safety meeting held in The
Netherlands in December 1993 considered the need to derive internationally acceptable
interim TEFs for the dioxin-like PCBs. Recommendations arising from that meeting of
experts (Ahlborg et al., 1994) suggest that in general only a few of the dioxin-like PCBs are
likely to be significant contributors to general population exposures to dioxin-like
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Table 9-1. Toxicity Equivalency Factors (TEF) for CDDs and CDFs
Compound
TEF
Mono-, Di-, and Tri-CDDs
2,3,7,8-TCDD
Other TCDDs
2,3,7,8-PeCDD
Other PeCDDs
2,3,7,8-HxCDD
Other HxCDDs
2,3,7,8-HpCDD
Other HpCDDs
OCDD
Mono-, Di-, and Tri-CDFs
2,3,7,8-TCDF
Other TCDFs
1,2,3,7,8-PeCDF
2,3,4,7,8-PeCDF
Other PeCDFs
2,3,7,8-HxCDF
Other HxCDFs
2,3,7,8-HpCDF
Other HpCDFs
OCDF
0
1
0
0.5
0
0.1
0
0.01
0
0.001
0
0.1
0
0.05
0.5
0
0.1
0
0.01
0
0.001
Source: EPA, 1989.
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compounds. Dioxin-like PCBs may be responsible for approximately one-fourth to one-half
of the total toxicity equivalence associated with general population environmental exposures
to this class of related compounds. Both the refinement of the toxicity equivalence factors
for dioxin-like PCB congeners (DeVito et al., 1993) as well as a compilation and analysis of
all available data on relative toxicities of dioxin-like PCBs with respect to a number of end
points (Ahlborg et al., 1994) support these findings. Although these findings have been
published recently, additional review and data collection will be needed. In addition, this
panel urged investigation of companion TEFs for ecotoxicoiogical use, based on data from
ecotoxicological studies.
Throughout this document, concentrations of dioxin and related compounds will be
presented as TEQs. TEQs are determined by summing the products of multiplying
concentrations of individual dioxin-like compounds times the corresponding TEF for that
compound. At times, levels will be presented as concentrations of TCDD because many past
studies monitored this congener alone. At most times, TEQs for CDDs and CDFs will be
discussed. When TEQ values include the dioxin-like PCBs as well, this will be specifically
mentioned. Readers of this chapter are encouraged to review previous chapters in the Health
Assessment Document and the Exposure Document for more details on estimates of TEQ
presented in this chapter. The strengths and weaknesses as well as the uncertainties
associated with the TEF/TEQ approach are discussed later in this chapter.
There are 75 individual compounds comprising the CDDs, depending on the
positioning of the chlorine(s), and 135 different CDFs. These are called individual
congeners. Likewise, there are 75 different positional congeners of BDDs and 135 different.
congeners of BDFs (see Exposure Document, Table 2-1). Only 7 of the 75 congeners of
CDDs or of BDDs are thought to have dioxin-like toxicity; these are ones with
chlorine/bromine substitutions in, at least, the 2, 3, 7, and 8 positions. Only 10 of the 135
possible congeners of CDFs or of BDFs are thought to have dioxin-like toxicity; these also
are ones with substitutions in the 2, 3, 7, and 8 positions. While this suggests 34 individual
CDDs, CDFs, BDDs, or BDFs with dioxin-like toxicity, inclusion of the mixed
chloro/bromo congeners substantially increases the number of possible congeners with
dioxin-like activity. There are 209 PCB congeners. Only 13 of the 209 congeners are
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thought to have dioxin-like toxicity; these are PCBs with 4 or more chlorines with just 1 or
no substitution in the ortho position. These eompounds are sometimes referred to as
coplanar, meaning that they can assume a flat configuration with rings in the same plane.
Similarly configured polybromirtated biphenyls are, likely to have similar properties;
however, the data base on these compounds with regard to dioxin-like activity has been less
extensively evaluated. Mixed chlorinated and brominated congeners also exist, increasing the
number of compounds considered dioxin-like. The physical/chemical properties of each
congener vary according to the degree and position of chlorine and/or bromine substitution.
Very little is known about occurrence and toxicity of the mixed (chlorinated and brominated)
dioxin, furan, and biphenyl congeners.
In general, these compounds have very low water solubility, high octanol-water
partition coefficients, and low vapor pressure and tend to bioaccumulate. Volume II of the
Exposure Document presents congener-specific values for water solubility, vapor pressure,
partition coefficients, and photo quantum yields and discusses other physicochemical
characteristics of the chlorinated dioxins and dibenzofurans. These physicochemical
properties result in the environmental fate and transport discussed below. Expanded
discussions will be required in future documents to account for dioxin-like PCBs and for
brominated or mixed halogenated congeners.
I
9.3. ENVIRONMENTAL FATE
Despite a growing body of literature from laboratory, field, and monitoring studies
examining the environmental fate and environmental distribution of CDDs and CDFs, the
fate of these environmentally ubiquitous compounds is not yet fully understood, and the
following represents our best understanding, based on available data. In soil, sediment, the
water column, and probably air, CDDs/CDFs are primarily associated! with particulate and
organic matter because of their high lipophilicity and low water solubility. They exhibit little
potential for significant leaching or volatilization once sorbed to particulate matter. The
available evidence indicates that CDDs and CDFs, particularly the tetra- and higher
chlorinated congeners, are extremely stable compounds under most environmental conditions,
with environmental persistence measured in decades. The only environmentally significant
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transformation process for these congeners is believed to be photodegradation of chemicals
not bound to particles in the gaseous phase or at the soil- or water-air interface. Brominated
congeners are significantly more readily transformed by photodegradation. CDDs/CDFs
entering the atmosphere are removed either by photodegradation or by dry or wet deposition.
Although some volatilization of dioxin-like compounds on soil does occur, the predominant
fate of CDDs/CDFs sorbed to soil is to remain in place near the surface of undisturbed soil
or to move to water bodies with erosion of soil. CDDs/CDFs entering the water column
primarily undergo sedimentation and burial. The ultimate environmental sink of these
CDDs/CDFs is believed to be aquatic sediments.
Little specific information exists on the environmental transport and fate of the dioxin-
like PCBs. However, the available information on the physical/chemical properties of
dioxin-like PCBs, coupled with the body of information available on the widespread
occurrence and persistence of PCBs in the environment, indicates that these PCBs are likely
to be associated primarily with soils and sediments and to be thermally and chemically stable.
Soil erosion and sediment transport in water bodies and emissions to the air (via
volatilization, dust resuspension, or point source emissions) followed by atmospheric
transport and deposition are believed to be the dominant transport mechanisms responsible
for the widespread environmental occurrence of PCBs. Photodegradation to less chlorinated
congeners followed by slow anaerobic and/or aerobic biodegradation is believed to be the
principal path for destruction of PCBs. Similar situations exist for the polybrominated
biphenyls (PBBs). Little information is available on the occurrence and fate of biphenyl
congeners containing both chlorine and bromine, but their contribution to dioxin-like activity
in the environment is thought to be small.
9.4. SOURCES
The chlorinated and brominated dioxins and furans have never been intentionally
produced other than on a laboratory-scale basis for use in chemical analyses. Rather, they
are generated as by-products from various combustion and chemical processes. PCBs were
produced in relatively large quantities for use in such commercial products as dielectrics,
hydraulic fluids, plastics, and paints. They are no longer produced in the United States but
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continue to be released to the environment through the use and disposal of these products. A
similar situation exists for the commercially produced PBBs, which were produced for a
number of uses like flame retardants. .
Dioxin-like compounds are released to the environment in a variety of ways and in
varying quantities, depending on the source. Studies of sediment cores in lakes near
industrial centers of the United States have shown that historical environmental deposition of
dioxins and furans was quite low until about 1920, peaked around 1980, and has declined
thereafter. This trend suggests that the presence of dioxin-like compounds in the
environment has occurred primarily as a result of industrial practices and is likely to reflect
changes in release over time. Further work to confirm declining trends in environmental
samples and to relate these data to human exposures will be required.
Although these compounds are released from a variety of sources, the congener
profiles of CDDs and CDFs found in sediments have been linked to combustion sources
(Kites, 1991). Three theories have been suggested to explain formation of CDDs and CDFs
during combustion: (1) CDDs and CDFs are present in the fuels or feed materials and pass
through the combustor intact; (2) precursor chemicals are present in the fuels or feed
material and undergo reactions catalyzed by particulates and other chemicals to form CDDs
and CDFs; and (3) the CDDs and CDFs are formed de novo from organic and inorganic
F '
substrates bearing little resemblance in molecular structure. ;
The principal identified sources of environmental release of CDDs and CDFs may be
grouped into four major types:
• Combustion and Incineration Sources: Dioxin-like compounds can be
generated and released to the environment from various combustion processes
when chlorine donor compounds are present. These sources can include
incineration of wastes such as municipal solid waste, sewage sludge, hospital
and hazardous wastes; metallurgical processes such as high-temperature steel
production, smelting operations, and scrap metal recovery furnaces; and the
burning of coal, wood, petroleum products, and used tires for power/energy
generation.
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• Chemical Manufacturing/Processing Sources: Dioxin-like compounds can be
formed as by-products from the manufacture of chlorine and such chlorinated
compounds as chlorinated phenols (e.g., pentachlorophenol), PCBs, phenoxy
herbicides (e.g., 2,4,5-T), chlorinated benzenes, chlorinated aliphatic
compounds, chlorinated catalysts, and halogenated diphenyl ethers. Although
the manufacture of many chlorinated phenolic intermediates and products, as
well as PCBs, was terminated in the late 1970s in the United States,
production continued elsewhere around the world until 1990, and continued,
limited use and disposal of these compounds can result in releases of CDDs>
CDFs, and PCBs to the environment.
• Industrial/Municipal Processes: Dioxin-like compounds can be formed through
the chlorination of naturally occurring phenolic compounds such as those
present in wood pulp. The formation of CDDs and CDFs resulting from the
use of chlorine bleaching processes in the manufacture of bleached pulp and
paper has resulted in the presence of CDDs and CDFs in paper products as
well as in liquid and solid wastes from this industry. Municipal sewage sludge
has been found to occasionally contain CDDs and CDFs.
• Reservoir Sources: The persistent and hydrophobic nature of these compounds
causes them to accumulate in soils, sediments, and organic matter and to
persist in waste disposal sites. The dioxin-like compounds in these
"reservoirs" can be redistributed by various processes such as dust or sediment
resuspension and transport. Such releases are not original sources in a global
sense, but can be on a local scale. For example, releases may occur naturally
from sediments via volatilization or via operations that disturb them, such as
dredging. Aerial deposition and accumulation on leaves can lead to releases
during forest fires or leaf composting operations.
As awareness of these possible sources has grown in recent years, a number of
changes have occurred in the United States, which should reduce the release rates. For
example, releases of dioxin-like compounds have been reduced due to the switch to unleaded
automobile fuels (and associated use of catalytic converters and reduction in halogenated
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scavenger fuel additives), process changes at pulp and paper mills, new emission standards
and upgraded emission controls for incinerators, and reductions in the manufacture of
chlorinated phenolic intermediates and products and the use of pesticides such as 2,4,5,-T
and pentachlorophenol.
Although dioxins in the environment may arise from a number of sources as discussed
above, the Exposure Document presents recent analyses of only air emissions of CDDs and
CDFs for several European countries in terms of total toxic equivalents based on
international TEFs for CDDs and CDFs. These studies assume that emissions to air make up
the major portion of dioxins released to the environment. Estimates of total release in these
countries range from approximately 100-1,000 g TEQ/year in West Germany and 100-200 g
TEQ/year in Sweden to approximately 1,000 and 4,000 g TEQ/year maximum emissions in
The Netherlands and United Kingdom, respectively. Similar nationwide estimates for the
United States have not been compiled prior to this reassessment effort. The Exposure
Document estimates the U.S. emissions to be in the range of 3,300-26,000 g TEQ/year, with
a central estimate of 9,300 g TEQ/year. These estimates were derived from data from
emission tests at relatively few facilities in each combustor class. These data were used to
develop emission factors and then extrapolated to a nationwide basis using the total amount
of waste feed material in each class. Variability of measured emissions from facilities within
a class and the uncertainty in estimating the total amount of waste feed material in each class
lead to the wide range presented above. Qualitatively speaking, major contributors to this
total include medical waste incinerators, municipal waste incinerators, cement kilns, and
industrial wood burning. Because of the limited number of measurements and the large
number of potential sources for each of these emissions, total estimated emissions from these
sources are considered highly uncertain. Municipal waste incineration has more
measurement data than other air sources, but emissions are highly variable among facilities
so that the overall estimate remains uncertain. Diesel-fueled vehicles; hazardous waste
burning, forest fires, and metal smelting are more moderate contributors of dioxin-like
compounds, but the magnitude of the contribution is also highly uncertain. Sewage waste
incineration and residential wood burning as well as a few minor processes round out the
current analysis and provide lower range estimates of medium to low certainty. Although
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still other sources are recognized and releases to land and water in addition to air are
discussed in the Exposure Document, it is clear from this exercise that additional
measurement data will be needed to gain an adequate appreciation for the nature and
magnitude of major U.S. sources of CDD and CDF emissions.
Several investigators have attempted to conduct "mass balance" checks on the
estimates of national dioxin releases to the environment. Basically, this procedure involves
comparing estimates of the emissions to estimates of aerial deposition. Such studies in
Sweden (Rappe, 1991) and Great Britain (Harrad and Jones, 1992) have suggested that the
deposition exceeds the emissions by about tenfold. These studies are acknowledged to be
quite speculative due to the strong potential for inaccuracies in emission and deposition
estimates. In addition, the apparent discrepancies could be explained by long-range transport
from outside the country, resuspension, and deposition of reservoir sources or unidentified
sources. Bearing these limitations in mind, this procedure has been used in this reassessment
to compare the estimated emissions and deposition in the United States.
Deposition measurements have been made at a number of locations in Europe and two
places in the United States (see discussion of these studies in Volume II of the Exposure
Document). These limited data suggest that a deposition rate of 1 ng TEQ/m2-yr is typical
of remote areas and that 2-6 ng TEQ/m2-yr is more typical of populated areas. Applying
these values, the total U.S. deposition can be estimated as 20,000 to 50,000 g TEQ/yr. This
range can be compared to the range of emissions for the United States (3,300-26,000 g
TEQ/yr) as presented in the Exposure Document. As noted above, interpreting such
comparisons is highly speculative and supports the need to conduct further emissions testing
into all media and deposition measurement, if we are to understand the total mass balance for
these compounds.
While all of the above emission and deposition values are given in the form of TEQs,
it should be noted that neither emission nor deposition is equivalent to exposure or intake.
Significant changes in composition can occur to complex mixtures of CDDs, CDFs, and
PCBs as they move through the environment. Measurements at or near the point of human
contact provide the best estimates of human exposure. TEQs are most relevant to potential
for hazard and risk when they represent intake values.
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9.4.1. Levels in the Environment and in Food
CDDs, CDFs, and PCBs have been found throughout the world in practically all
media, including air, soil, water, sediment, fish and shellfish, and agricultural food products
such as meat and dairy products. The highest levels of these compounds are found in soils,
sediments, and biota; very low levels are found in water and air. The widespread occurrence
observed, particularly in industrialized countries, is not unexpected, considering the
numerous sources that emit these compounds into the atmosphere and the overall resistance
of these compounds to biotic and abiotic transformation.
The average levels of these compounds found in the various media in North America
have been compiled in the Exposure Document. The levels shown for environmental media
and for food in North America are based on few samples and must be considered uncertain.
However, they seem reasonably consistent with, levels measured in a number of studies in
Western Europe and Canada. The consistency of these levels across industrialized countries
adds some confidence to the limited data from the United States and provides some
reassurance that the U.S. estimates are reasonable. A major concern raised regarding all of
these data is that few if any of these studies had a statistical design that was satisfactory for
generalization to national food supplies. This adds to the uncertainty of extrapolations using
these findings and argues for additional data collection to evaluate national and regional
differences of levels of dioxin-like compounds in the environment and in food.
This assessment proposes the hypothesis that the primary mechanism by which dioxin-
like compounds enter the terrestrial food chain is via atmospheric deposition. Dioxin and
related compounds enter the atmosphere directly through air emissions or indirectly, for
example, through volatilization from land or water or from resuspensipn of particles.
Deposition can occur directly onto soil or onto plant surfaces. Soil deposits can enter the
food chain via direct ingestion (e.g., grazing animals, earth worms, fur preening by
burrowing animals). Dioxin-like compounds in soil can become available to plants by
volatilization and vapor absorption or particle resuspension and adherence to plant surfaces.
In addition, dioxin-like compounds in soil can adsorb directly to underground portions of
plants. Uptake from soil via the roots into above-ground portions of plants is thought to be
insignificant.
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Support for this air-to-food hypothesis is provided by Hites (1991) who concluded that
"background environmental levels of dioxin-like compounds are caused by dioxin-like
compounds entering the environment through the atmospheric pathway." His conclusion was
based on demonstrations that the congener profiles in lake sediments could be linked to
congener profiles of combustion sources. Further arguments supporting this hypothesis .
include: (1) numerous measurements show that emissions occur from multiple sources and
deposition occurs in most areas, including remote locations; (2) atmospheric transport and
deposition are the only mechanisms that could explain the widespread distribution of these
compounds in soil; and (3) other mechanisms of uptake into food, for instance, from direct
contamination or through packaging, are much less plausible. Direct uptake into food from
soil or sediments is possible and could be important for "local" exposures. These routes are
less likely to explain the general background level of dioxin and related compounds found in
the diet of the general population.
At present, it is unclear whether atmospheric deposition represents primarily "new"
contributions of dioxin and related compounds from all media reaching the atmosphere or
whether it is "old" dioxin and related compounds that persist and recycle in the environment.
Understanding the relationship between these two scenarios will be particularly important in
understanding the relative contributions of individual point sources of these compounds to the
food chain and assessing the effectiveness of control strategies focused on either "new" or
"old" dioxins in attempting to reduce the levels in food.
9.4.2. Background Exposure Levels
The term "background" exposure has been used throughout this reassessment to
describe exposure of the general population, who are not exposed to readily identifiable point
sources of dioxin-like compounds, that results in widespread, low-level circulation of dioxin-
like compounds in the environment. The primary route of this exposure is thought to be the
food supply, and most of the dioxin-like compounds are thought to come from non-natural
sources. For the purposes of estimating background exposures to dioxin-like compounds via
dietary intake the upper-range background toxicity equivalent values (i.e., those calculated
using one-half the detection limit for the nondetects) were used in the Exposure Document.
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Uncertainties associated with the use of TEQs have been described throughout this chapter.
The estimates are based on intake of dioxin-like CDDs and CDFs and do not include
estimates for dioxin-like PCBs or other dioxin-like compounds. Inclusion of dioxin-like
PCBs could raise these estimates by 35-50%. The net effect of these calculations is that we
may be overestimating background levels based on the use of one-half of the detection limit
and underestimating background levels by not including the dioxin-like PCBs or other dioxin-
like compounds.
A background exposure level of 120 pg TEQ/day for the United States was
estimated. These estimates are comparable to analogous estimates for European countries.
These include estimates for Germany, which range from 79 pg TEQ/day based on Furst et
al. (1990) to 158 pg TEQ/day based on Furst etal. (1991), 118-126 pg TEQ/day exposure
via numerous routes in The Netherlands (Theelen, 1991), and 140-290 pg TEQ/day for the
typical Canadian exposed mainly through food ingestion (Oilman and Newhook, 1991). It is
generally concluded by these researchers that dietary intake is the primary pathway of human
exposure to CDDs and CDFs. These investigators among others suggest that greater than 90
percent of human exposure occurs through the diet, with foods from animal origins being the
predominant pathway.
This conclusion, that food is the predominant pathway of exposure, remains to be
validated in the United States. Although data are derived from multiple studies from around
the world, the data represent limited numbers of samples. Use of one-half of the detection
level for nondetects is a reasonable but conservative approach to estimating low levels in
samples. For some data sets, use of zero values for nondetects would result in significantly
lower estimates. Setting nondetects equal to zero, however, does not significantly change the
average TEQ levels estimated for most categories of U.S. food. In the current assessment,
similar estimates of TEQs derived from different data sets, developed by different
investigators in several countries, strengthen the probability that this inference represents the
exposure of the general population in industrialized countries to dioxin and related
compounds.
Data on human tissue levels suggest that body-burden levels among industrialized
nations are reasonably similar (Schecter, 1991). These data can also be used to estimate
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background exposure through the use of pharmacokinetic models. Using this approach,
exposure levels to 2,3,7,8-TCDD in industrialized nations are estimated to be about 20 to 40
pg TCDD/day (0.3-0.6 pg TCDD/kg/day). This is generally consistent with the estimates
derived using diet-based approaches to estimate total TCDD intake. Pharmacokinetic
approaches have not been applied to estimate exposures to CDDs or CDFs other than TCDD,
which contribute substantially to the body burden of dioxin-like compounds. Estimates of
exposure to dioxin-like CDDs and CDFs based on dietary intake are in the range of 1-3 pg
TEQ/kg/day. Estimates based on the contribution of dioxin-like PCBs to toxicity equivalents
raise the total to 3-6 pg TEQ/kg/day. This range is used throughout this characterization as
an estimate of average background exposure to dioxin-like CDDs, CDFs, and PCBs.
The U.S. study of CDD/F body burdens contained in the National Human Adipose
Tissue Survey (NHATS) (U.S. EPA, 1991) analyzed for CDD/Fs in 48 human tissue
samples which were composited from 865 samples. These samples were collected during
1987 from autopsied cadavers and surgical patients. While this was an important study of
chemical residues occurring in human fat, numerous technical shortcomings of this study
have been described. For instance, the sample compositing prevents use of these data to
examine the distribution of CDD/F levels in tissue among individuals. However, it did allow
conclusions in the following areas:
• National Averages: The national averages for all TEQ congeners (but
excluding dioxin-like PCBs) were estimated and totaled to 28 pg TEQ/g lipid
adjusted value (28 ppt).
• Age Effects: Tissue concentrations of CDD/Fs were found to increase with
age.
• Geographic Effects: In general, the average CDD/F tissue concentrations
appeared fairly uniform geographically.
» Race Effects: No significant differences in CDD/F tissue concentrations were
found on the basis of race.
• Sex Effects: No significant differences in CDD/F tissue concentrations were
found between males and females. -
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• Temporal Trends: The 1987 survey showed decreases in tissue
concentrations relative to the 1982 survey for all congeners. However, it is
not known whether these declines were due to improvements in the analytical
methods or actual reductions in body-burden levels. The fact that areas
surveyed in this study changed over time (due to drop-out of areas) also makes
interpretation of time trends difficult. The percent reductions among
individual congeners varied from 9% to 96%. The relationship of these data
to an apparent declining trend of dioxin-like compounds in environmental
samples, especially sediments, is currently unclear.
More recent data (Patterson et al., 1994) show similar decreasing trends with regard
to levels of dioxin-like PCBs in blood and fat. In addition, these data showed a wide
variability of PCS congeners in human adipose tissue samples as compared to concentrations
of CDDs and CDFs, which were less variable.
Inclusion of dioxin-like PCBs in TEQ calculations raises the average body burden to
40-60 pg TEQ/g (40-60 ppt). Because available data from the two studies discussed above
do not provide a representative population sample, these conclusions must be regarded as
uncertain. Additional measurements will be necessary to confirm this hypothesis. Use of a
protocol for sampling that allows an evaluation of age-adjusted population averages will be
critical for understanding the current body-burden situation and evaluating impacts of future
efforts to further reduce exposures to this class of compounds.
Levels of dioxin-like compounds found in human tissue/blood appear similar in
Europe and North America. Schecter (1991) compared levels of dioxin-like compounds
found in blood among people from U.S. pooled samples (100 subjects) and Germany (85
subjects). Although mean levels of individual congeners differed by as much as a factor of
two between the two populations, the total TEQ averaged 42 pg TEQ/g (42 ppt) in the
German subjects and was 41 pg TEQ/g (41 ppt) in the pooled U.S. samples. These values
do not include TEQs for PCBs.
New information on levels of dioxin-like compounds in human adipose tissue and
blood has recently been published (Patterson et al., 1994). This study reports measurements
of dioxin-like PCB congeners as well as CDD and CDF levels in samples from 28 Atlanta
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residents. These measurements show that concentrations of dioxin-like PCBs can be more
than an order of magnitude higher than concentrations of TCDD. Comparison with other
published information suggests much higher levels of nondioxin-like congeners of PCBs and
the possibility that concentrations of both types of congeners will depend heavily on previous
human activities such as fish consumption. These data are consistent with the previous
statement that dioxin-like PCBs may account for approximately one-third of the total TEQ in
the general population. Values in Patterson's study calculated TEQs for PCBs using the data
of Safe (1990), which were acknowledged by the author as being conservative and, based on
more recent data, overestimate the contribution of dioxin-like PCBs.
While, as described above, evaluation of the range of background population blood
levels is difficult given existing data, the NHATS tissue data show that the maximum
measured concentrations were about two times higher than the average for most congeners
(U.S. EPA, 1991). These results are based on composite samples that each included
approximately 20 individual samples. This high level of compositing will greatly reduce the
individual variability of samples. Consequently, the range in body burdens in the entire
population is expected to be larger than that found among the samples in this study; The
Patterson et al. (1994) data show that the maximum 2,3,7,8-TCDD concentration was about
three to four times higher than the average. Similar results were seen for PCB 126. These
results are based on samples of 28 individuals. Again, the range of body burdens in the '
entire population will be greater than that found among these 28 individuals. Accordingly, it
can be concluded that body burdens of dioxin-like compounds are likely to be at least three
to four times higher than the average for some members of the population and, perhaps, even
higher. While it is difficult to know the full extent of the range of body burdens, the
Patterson data were found to fit reasonably well as a log-normal distribution. This
observation has also been made for other data sets (Sielken, 1987). With such distributions
in large populations, it is not unusual to see values that extend three standard deviations
beyond the mean. The body burdens corresponding to three standard deviations beyond the
mean (99th percentile) have been estimated (using a log-scale calculation) to be
approximately seven times higher than the arithmetic mean. Whether individuals with
background levels of dioxin-like compounds of this magnitude exist in the general population
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is unknown, but these calculations provide support for the inference; that the general
population may have a wide range of body burdens and, therefore, both average and high
end values should be considered when evaluating potential for adverse impacts of background
exposures. i
9.4.3. Highly Exposed Populations
In addition to general population exposure, some individuals or groups of individuals
may also be exposed to dioxin-like compounds from discrete sources or pathways locally
within their environment. Examples of these "special" population exposures include
occupational exposures, direct or indirect exposure to local populations from discrete local
sources, exposure to nursing infants from mother's milk, or exposures to subsistence or
recreational fishers. These exposures have been discussed previously in terms of increased
exposure due to dietary habits (see Exposure Document) or due to occupational conditions or
industrial accidents (see Chapter 7). Although exposures to these populations may be
significantly higher than to the general population, they usually represent relatively small
numbers of individuals. Inclusion of their levels of exposure in the general population
estimates would have little impact on average estimates and would obscure the potential
significance of elevated exposures for these subpopulations.
For example, consumption of breast milk by nursing infants may lead to higher levels
of exposure during the early postnatal period as compared to intake in the diet later in life.
Schecter et al. (1992) report that a study of 42 U.S. women found gin average of 16 pg
TEQ/g (16 ppt), 3,3 ppt of which was 2,3,7,8-TCDD, in the lipid portion of breast milk. A
much larger survey in Germany (n=728) found an average of 31 pg TEQ/g (31 ppt) with a
range of 6 to 87 pg TEQ/g in the lipid portion of breast milk (Beck et al., 1991). These
estimates do not include a contribution to total TEQ from dioxin-like PCBs. The level in
human breast milk can be predicted on the basis of the estimated dioxin intake by the
mother. Such procedures are presented in Volume II of the Exposure Document.
Elimination of 2,3,7,8-TCDD through mother's milk can result In higher exposure
levels to the infant than for the general population. Assuming that an infant breast feeds for
one year (a conservative assumption since, in the United States, 6 months of breast feeding is
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more typical), has an average weight during this period of 10 kg (which is on the high end
[90-98th percentile] of the average weight distribution for the first year of life), ingests 0.8.
kg/d of breast milk and that the dioxin concentration in milk fat is 20 pg/g (20 ppt) of TEQ,
the average daily dose to the infant over this period is predicted to be about 60 pg TEQ/kg/d,
not including dioxin-like PCBs. This value is 10 to 20 times higher than the estimated range
for background exposure to adults (i.e., 3-6 pg TEQ including dioxin-like PCBs/kg/d) and
would have been even higher if dioxin-like PCBs had been included in this sample analysis.
A range of alternative assumptions could be made regarding the nursing time period, infant's
body weight, and milk ingestion rate. None of these factors is likely to vary individually by
more than a factor of two and, when combined, will likely result in less than multiplicative
variability in estimates of daily intake. WHO (1988) suggested that a reasonable average
nursing scenario would be 6 months duration, 0.7 L/day ingestion rate, and a milk fat
content of 3.5%. Using a milk ingestion rate of 120 mL/kg/day (compared to 80 mL/kg/day
used above) and a milk concentration of 16.9 pg TEQ/g, WHO estimated a daily intake of 70
pg TEQ/kg/day.
If a 70-year averaging time is used to obtain an added increment of lifetime daily
dose, then the increment of lifetime average daily dose attributable to the EPA nursing
scenario is estimated to be 0.8 pg of TEQ/kg/d. On a mass basis, the cumulative dose to the
infant under this scenario is about 210 ng compared to a lifetime background intake of about
1,700 to 5,100 ng (suggesting that 4% to 12% of .the lifetime intake may occur as a result of
breast feeding for the first year of life). WHO (1988) estimated that 4% of the lifetime
intake would occur during the 6 months of nursing in their scenario. This percentage, as
well as the daily intake rate, is nearly identical to the estimates presented in the Exposure
Document although based on somewhat different assumptions. Traditionally, EPA has used
the lifetime average daily dose as the basis for evaluating incremental cancer risk and the
average daily dose (i.e., the daily exposure per unit body weight occurring during an
exposure event) as the more appropriate indicator of risk for certain noncancer end points.
The use of a lifetime average daily dose for high-level, early exposures may underestimate
cancer risk if dose rate or perinatal sensitivity is important in the ultimate carcinogenic
outcome. The average daily dose approach may be particularly important for the evaluation
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of noncancer end points if exposure is occurring during windows of sensitivity during
prenatal and postnatal development. However, data are currently insufficient to verify this
hypothesis.
In addition, consumption of unusually high levels of fish, meat, or dairy products
containing elevated levels of dioxin and related compounds can lead to elevated blood1 levels
in comparison to the general population. Most people eat fish from multiple sources, both
fresh and salt water, where levels of dioxin-like compounds are likely to be low. Even if
large quantities of fish are consumed, most people are not likely to have unusually high
exposures to dioxin-like compounds. However, individuals who fish regularly for purposes
of basic subsistence are likely to obtain their fish from a few sources and may have the
potential for elevated exposures. Such individuals may also consume large quantities of fish.
Although average consumers may eat a few fish meals a month (an average intake of
approximately 6.5 grams of fish a day), many recreational anglers near large water bodies
may consume, on average, four to five times as much (approximately 30 grams per day). Of
course, these averages include some individuals who eat no fish at all. Some individuals at
the high end of the consumption range may eat, on average, as much as 140 grams per day.
Certain members of ethnic groups who are subsistence fishers may consume two to three
times this high-end amount as an upper estimate (up to 400 grams or approximately 1 pound
per day). If high-end consumers obtain their fish from areas where content of dioxin-like
chemicals in the fish is high, they may constitute a highly exposed subpopulation. Svensson
et al. (1991) found elevated blood levels of CDDs and CDFs in high fish consumers living
near the Baltic Sea in Sweden. The highest consumers, fishermen or workers in the fish
industry, had blood level TEQs that were approximately three times that of non-fish
consumers (60 pg TEQ/g lipid versus 20 pg TEQ/g lipid). The difference in levels of
dioxin-like compounds was particularly apparent for the CDFs. Didxin-like PCBs were not
accounted for in this study. Studies are currently under way to examine fish consumption
patterns in several Native American groups. Recent results (Columbia River Intertribal Fish
Commission, 1994) suggest that Native Americans living along the Columbia River may
consume an average of 30 grams of fish a day; some individuals consume much higher
levels. Studies are currently under way to determine levels of dioxin-like compounds in fish
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from this region. No measurements of dioxin-like chemicals in the blood of these Native
American populations are currently available.
Dewailly et al. (1994) observed elevated levels of coplanar PCBs in the blood of
fishermen on the north shore of the Gulf of the St. Lawrence River who consume large
amounts of seafood. Coplanar PCB levels were 20 times higher among the 10 highly
exposed fishermen than among controls. This study also reported elevated levels of
coplanar PCBs in the breast milk of Inuit women of Arctic Quebec. The principal source of
protein for the Inuit people is fish and sea mammal consumption.
The possibility of high exposures to dioxin-like chemicals as a result of consuming
meat and dairy products is most likely to occur in situations where individuals consume large
quantities of these foods from a locality where the level of these compounds is elevated.
Most people eat meat and dairy products from multiple sources and, even if large quantities
are consumed, are not likely to have unusually high exposures. However, individuals who
raise their own livestock for basic subsistence have the potential for higher exposures if local
levels of dioxin-like compounds are high. Volume III of the Exposure Document presents
methods for evaluating this type of exposure scenario and concludes that indirect exposures
via consumption of locally produced foods represent a major pathway for human exposure
for a limited number of individuals in the population. In an example analysis contained in
Volume III of the Exposure Document based on proximity to combustor emissions, the high
end exposure estimates from food consumption were found to be about two orders of
magnitude higher than inhalation exposures at the same location. However, it should be
noted that no studies were found in the literature to demonstrate this potential increased
exposure based on measurements of dioxin-like chemicals from source to livestock to
humans.
Although the subpopulations discussed above have the potential for high exposure to
dioxin-like compounds, a careful evaluation of dietary habits and proximity to sources of
dioxin and related compounds is needed. It would generally be inappropriate to compute the
total intake of dioxin-like compounds in a subpopulation by simply adding the dioxin intake
from highly consumed food to the general population intake level. The general population
background estimate assumes a typical pattern of food ingestion, whereas a subpopulation
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that has a high consumption rate of one particular food type is likely to eat less of other food
types. Ideally, the evaluation should be based on the entire diet of the subpopulation and use
case-specific values for food ingestion rates and concentrations of dioxin-like compounds.
High blood levels of dioxin and related compounds based on high levels of exposure
have been documented for industrial exposures in segments of the chemical industry and for
industrial accidents. Health effects studies in human populations have focused on these
groups of highly exposed individuals. Results of these studies are described in detail in
Chapter 7.
9.5. DISPOSITION AND PHARMACOKINETICS
The disposition and pharmacokinetics of 2,3,7,8-TCDD and related compounds have
been investigated in several species and under various exposure conditions. These data and
models derived from them are critical in understanding the sequelae of human exposure.
Data related to disposition and pharmacokinetics of dioxin and related compounds and efforts
to develop models to further understand tissue dosimetry are described in detail in Chapter 1
of the Health Assessment Document.
The gastrointestinal, dermal, and transpulmonary absorptions, of these compounds
i
represent potential routes for human uptake. Findings of studies in experimental animals
indicate that oral exposure to 2,3,7,8-TCDD in the diet or in an oil vehicle results in the
absorption of >50%, and often closer to 90%, of the administered dose. Gastrointestinal
absorption of related compounds is variable, incomplete, and congener specific. More
soluble congeners, such as 2,3,7,8-TCDF, are almost completely absorbed, while the
extremely insoluble OCDD is very poorly absorbed. In some cases,, absorption has been
found to be dose dependent, with increased absorption occurring at lower doses (2,3,7,8-
TBDD, OCDD). The limited data base also suggests that there are no major interspecies
differences in the gastrointestinal absorption of these compounds among mammals. Limited
data (Poiger and Schlatter, 1986) from a single human volunteer suggest a high level
(>87%) of absorption of 2,3,7,8-TCDD in corn oil from the gastrointestinal tract.
Following absorption, a half-life for elimination was estimated to be 2,120 days (5.8 years).
It should be noted that this estimate of half-life is for a single individual and that longer
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median half-lives for 2,3,7,8-TCDD have been estimated (7.1 and 11.3 years) in other
studies described in this chapter and in Chapter 1.,
Additional data also indicate the importance of the formulation or vehicle containing
the toxicant(s) on the relative bioavailability of 2,3,7,8-TCDD and related compounds after
exposure. For instance, rodent feeding studies indicate that the bioavailability of 2,3,7,8-
TCDD from soil varies between sites and 2,3,7,8-TCDD content alone may not be indicative
of potential human hazard from contaminated environmental materials. Although data
indicate that substantial absorption may occur from contaminated soil, soil type and duration
of contact may substantially affect the absorption of 2,3,7,8-TCDD from soils obtained from
different contaminated sites. This uncertainty should be kept in mind as intake values and
the assumption of 50-100% absorption are often used to estimate potential risk from
environmental samples.
In experiments measuring dermal absorption for 2,3,7,8-TCDD and several CDFs,
the percentage of administered dose absorbed decreased with increasing dose while1 the
amount absorbed increased with dose. Results also suggest that the majority of the
compound remaining at the skin exposure site was associated with the outer skin layer (the
stratum corneum) and did not penetrate through to the dermis. Together, these results on
dermal absorption indicate that at <0.1 ^mol/kg, a greater percent of this administered dose
of 2,3,7,8-TCDD and three CDFs was absorbed. Nonetheless, even following a low-dose
dermal application of 200 pmol (1 nmol/kg), the rate of absorption of 2,3,7,8-TCDD is still
very slow (rate constant of 0.005 hour1). Dermal exposure of humans to 2,3,7,8-TCDD and
related compounds usually occurs as a complex mixture of these contaminants in soil, oils, or
other mixtures that would be expected to alter absorption. Available data suggest that the
dermal absorption of 2,3,7,8-TCDD depends on the formulation (vehicle or adsorbent)
containing the toxicant. Although no data are available to directly evaluate human dermal
absorption, the data available from in vitro and animal studies suggest slow dermal
absorption of these compounds, which is likely to be dependent on the vehicle or adsorbent
containing the compounds and the duration of the contact.
The use of incineration as a means of solid and hazardous waste management results
in the emission of vapors and contaminated particles that may contain TCDD and related
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compounds into the environment. Thus, exposure to TCDD and related compounds may
result from inhalation of contaminated fly ash, dust, and soil or from ingestion if air-
transported particles are deposited on fruits and vegetables. Direct exposure by the
inhalation route is usually relatively low as a percentage of overall intake. Systemic effects
occur in animals after pulmonary exposure to TCDD, suggesting that transpulmonary
absorption of TCDD does occur. Further results suggest that the transpulmonary absorption
of 2,3,7,8-TCDD and 2,3,7,8-TBDD was similar to that observed following oral exposure.
These limited data provide evidence of efficient transpulmonary absorption after intratracheal
instillation in laboratory animals. No data from humans or primates are available to address
this issue. However, these data provide support for the inference that efficient absorption
will occur when vapors and particles containing dioxin and related compounds are inhaled by
humans.
Once absorbed into blood, 2,3,7,8-TCDD and related compounds readily distribute
to all organs. Tissue distribution within the first hour after exposure reflects physiological
parameters such as blood flow to a given tissue and relative tissue size. There do not appear
to be major species or strain differences in the tissue distribution of 2,3,7,8-TCDD and
2,3,7,8-TCDF in mammals, with the liver and adipose tissue being the primary disposition
sites although human data to address this issue are quite limited. The tissue distribution of
the coplanar PCBs and PBBs also appears to be similar to that of 2,3,7,8-TCDD and 2,3,7,8-
TCDF based on evaluation in experimental animals.
Multiple studies suggest that distribution of this class of compounds to internal organs
is dose dependent. At low doses in animal studies, adipose tissue serves as the major depot;
at high doses, a major fraction is sequestered in the liver. The biochemical basis for,this
observation is under investigation. Induction of a hepatic binding protein has been
hypothesized to play a major role.
As discussed above, levels of 2,3,7,8-TCDD averaging 5-10 pg/g lipid (ppt) have
been reported for background populations. Sielken (1987) evaluated these data and
concluded that the levels of 2,3,7,8-TCDD in human adipose are log-normally distributed
and positively correlated with age. Among the observed U.S. background levels of 2,3,7,8-
TCDD in human adipose tissue, more than 10% were > 12 pg/g (ppt).
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Human body-burden measurements on dioxins were initially conducted using adipose tissue,
which required surgical samples, or, occasionally for women, using breast milk. Patterson et
al. (1988) showed that human serum was an accurate and more practical surrogate for human
adipose tissue. They found that the partitioning ratio of 2,3,7,8-TCDD between adipose
tissue and serum was approximately 1.09 when the concentrations were adjusted for lipid
content. This relationship appears to hold for at least a thousandfold concentration range in
excess of background levels. This correlation indicates that serum 2,3,7,8-TCDD, coupled
with measurement of serum lipid content, provides a valid estimate of the 2,3,7,8-TCDD
concentration in adipose tissue under steady-state, low-dose conditions.
In a study of potentially heavily exposed Vietnam veterans, the Centers for Disease
Control and Prevention (MMWR, 1988) reported an Air Force study of Ranch Hand veterans '
who were either herbicide loaders or herbicide specialists in Vietnam. The herbicide 2,4,5-T
(Agent Orange) that was used in Vietnam was contaminated with a low percentage of
2,3,7,8-TCDD. The mean serum 2,3,7,8-TCDD level of 147 Ranch Hand personnel was 49 •
pg/g (ppt) in 1987, based on total lipid-weight, while the mean serum level of the 49 controls •
was 5 pg/g (ppt). In addition, 79% of the Ranch Hand personnel and 2% of the controls had
2,3,7,8-TCDD levels > 10 pg/g (ppt). The distribution of 2,3,7,8-TCDD levels in this
phase of the Air Force health study indicates that Ranch Hand veterans have had higher
lifetime exposures than controls and that a small number of Ranch Hand personnel had
unusually heavy 2,3,7,8-TCDD exposure. Pirkle et al. (1989) estimated the median half-life
of 2,3,7,8-TCDD in humans to be approximately 7 years on the basis of 2,3,7,8-TCDD
levels in serum samples taken in 1982 and 1987 from 36 of the Ranch Hand personnel who
had 2,3,7,8-TCDD levels > 10 pg/g (ppt) in 1987. Similar tissue concentrations were
obtained by Kahn et al. (1988) in a report comparing 2,3,7,8-TCDD levels in blood and
adipose tissue of moderately exposed Vietnam veterans who handled herbicides regularly
while in Vietnam and matched controls. Although this study can distinguish moderately
exposed men from others, the data do not address the question of the difficulty of
characterizing the exposures of persons whose exposures are relatively low and who
constitute the bulk of the population, both military and civilian, who may have been exposed
to greater than background levels of 2,3,7,8-TCDD. Despite the fact that their exposures
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may result in slightly elevated levels of 2,3,7,8-TCDD, these individuals are
indistinguishable from the general population with similar blood levels spanning a range of
nondetect to > 10 ppt. Recently, a follow-up analysis to the Ranch Hand study described
above has been published. This study (Wolfe et al., 1994) describes half-life measurements
based on 337 Ranch Hand veterans. The estimate of the median half-life of TCDD is
predicted to be 11.3 years. The implications of this longer half-life on our understanding of
TCDD kinetics and on the back-calculations of historic intake values arid body burdens will
need to be fully described in future versions of this report.
The metabolism of 2,3,7,8-TCDD and related compounds is required for urinary and
biliary elimination and therefore plays a major role in regulating the; rate of excretion of
these compounds and determining their half-life. Although early in vivo and in vitro
investigations were unable to detect the metabolism of 2,3,7,8-TCDD, there is now evidence
that a wide range of mammalian and aquatic species are capable of slowly biotransfofming
2,3,7,8-TCDD to polar metabolites. Although metabolites of 2,3,7,8-TCDD have not been
directly identified in humans, recent analytic data from feces samples from an individual in a
self-dosing experiment suggest that humans can slowly metabolize 2,,3,7,8-TCDD (Wendling
and Orth, 1990). Direct intestinal excretion of the parent compound is another route for
excretion of 2,3,7,8-TCDD and related compounds that is not regulated by metabolism.
Some investigators have questioned whether the parent compound or metabolites are
responsible for dioxin toxicity. Structure-activity studies of 2,3,7,8-TCDD and related
compounds support the widely accepted principle that the parent compound is the active
species, arid the relative lack of biological activity of readily excreted monohydroxylated
metabolites of 2,3,7,8-TCDD and 3,3',4,4'-TCB suggests that metabolism is a detoxification
process necessary for the biliary and urinary excretion of these compounds. This concept
has also been generally applied to 2,3,7,8-TCDD-related compounds, although data are
lacking on the structure and toxicity of metabolites of other CDDs, BDDs, CDFs, BDFs,
PCBs, and PBBs. It is still possible, however quite unlikely, that low levels of unextractable
and/or unidentified metabolites may contribute to one or more of the toxic responses of
2,3,7,8-TCDD and related compounds.
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Physiologically based pharmacokinetic (PB-PK) models have been developed for
2,3,7,8-TCDD in mice, rats, and humans. PB-PK models incorporate known or estimated
anatomical, physiological, and physicochemical parameters to describe quantitatively the
disposition of a chemical in a given species. PB-PK models can assist in the extrapolation of
high-to-low dose kinetics within a species, estimating exposures by different routes of
administration, calculating effective doses, and extrapolating these values across species.
These models are particularly important given the limited empirical data on individual dioxin-
like congeners.
Chapter 8 contains a review of biologically based models of dioxin pharmacokinetics.
The early studies in rodents have recently been extended to describe protein induction and
tissue distribution data in the mouse (Leung et al., 1990b) and rat (Leung et al., 1990a).
Andersen et al. (1993) refined the model to include induction of CYP1A1 and diffusion-
limited tissue distribution. CYP1A1 is one of a family of proteins involved in the activation
and detoxification of both endogenous and exogenous chemicals. The model described by
Kedderis et al. (1993) for 2,3,7,8-tetrabromodibenzo-p-dioxin extended the use of PBPK
models to the brominated congener of TCDD. Portier et al. (1993) modeled the steady-state
induction of CYP1A1 and CYP1A2 using Hill equations. Their analysis stressed the
importance of the mechanism of endogenous protein expression on the shape of the dose-
response curve in the low-dose region. Kohn and Portier (1993) extended this result to a
general class of models and discussed implications of these models for risk assessment.
Kohn et al. (1993) used approaches to describe tissue dosimetry of TCDD and additionally
incorporated dioxin-mediated effects on growth factors, induction of the Ah-receptor, and
several models for endogenous induction of CYP1A1, CYP1A2, and the EOF receptor.
Other models have been proposed recently to describe effects of TCDD on lipid metabolism
(Roth et al., 1993).
An empirical dose-dependent model by Carrier (1991) relates the varying fraction of
the body burden of TCDD associated with the liver in humans to the total body burden of
TCDD. This model is consistent with the animal results described by the PB-PK models of
Andersen et al. (1993) and Kohn et al. (1993).
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Our uncertainty in the validity of predictions from PB-PK models is primarily driven
by the limited availability of congener and species-specific data that accurately describe the
dose- and time-dependent disposition of 2,3,7,8-TCDD and related compounds. As
additional data become available, particularly on the dose-dependent disposition of these
compounds, more accurate models can be developed. In developing a suitable model in the
human, it is also important to consider that the half-life estimate of 7.1 years for 2,3,7,8-
TCDD was based on two serum values taken 5 years apart, with the assumption of a single
compartment, and assuming a first-order elimination process (Pirkle et al., 1989). It is likely
that the excretion of 2,3,7,8-TCDD in humans is more complex, involving several
compartments, tissue-specific binding proteins, and a continuous daily background exposure.
Furthermore, changes in body weight and body composition should also be considered in
developing PB-PK models for 2,3,7,8-TCDD and related compounds in humans. Data
contained in the recently reported, expanded study of half-life of 2,3,7,8-TCDD in Ranch
Hand Veterans (Wolfe et al., 1994) and additional follow-up studies using blood level
information from the 1992-1993 physical examination should allow for better estimates of
TCDD half-life, provide important additional data to evaluate whether TCDD follows single-
compartment, first-order kinetics, and provide additional information with which to study the
influence of percent body fat on TCDD elimination in these veterans. '
It is known that exposure occurs to the developing fetus through placental transfer of
dioxin-like compounds in maternal blood via the placenta. In addition, exposure is likely to
increase in the early postnatal period through intake of mother's milk containing dioxin-like
compounds. Redistribution of body burdens is likely to occur with growth and development,
depending on relative intakes and changes in body fat content. Fasting, aging, and disease
are all thought to alter steady-state levels of dioxin during life. These changes complicate
standard pharmacokinetic models and present the possibility for temporary but potentially
important increases in blood or tissue levels of dioxin-like compounds during critical periods
of development, growth, and aging. Additional data on both distribution and dose to target
organs and response to the tissue-specific dose in relation to development and growth will be
required to refine our perspectives on the importance of these issues in evaluating dioxin
hazards and risks.
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An understanding of the relationship between exposure and dose is an important
aspect of an adequate characterization of risk. The data base relating to this issue is
extensive for 2,3,7,8-TCDD but is lacking for many of the related compounds. Nonetheless,
evaluation of available data and the development of physiologically based models has led to a
better understanding of the disposition and pharmacokinetics of dioxin and related compounds
than for most other environmental chemicals. This is particularly important because this
characterization relies extensively on estimates of body burden, which is a function of the
uptake, distribution, metabolism, and excretion of this complex mixture of structurally
related compounds. Estimates of half-life in the body facilitate the understanding of
bioaccumulation as a function of intake over a lifetime and of the impact of incremental
exposures on blood or tissue levels both over the short and long term. In addition, these
estimates allow some estimation of historical body burdens to complement effects analysis in
human populations presumed to have high exposures in earlier decades.
9.6. MECHANISMS OF DIOXIN ACTION
Knowledge of the mechanisms of dioxin action may facilitate the risk assessment
process by imposing bounds on the assumptions and models used to describe possible
responses to exposure to dioxin. In this document, the relatively extensive data base on
dioxin action has been reviewed, with emphasis on the contribution of the specific cellular
receptor for dioxin and related compounds, the Ah receptor, to the mechanism(s) of action.
Other reviews referenced in Chapter 2 provide additional background on the subject.
Discussion in this chapter will focus on aspects of our understanding of mechanism(s) of
dioxin action that are particularly important in understanding and characterizing dioxin risk,
including:
• Similarities at the biochemical level between humans and other animals with
regard to receptor structure and function;
" Relationship of receptor binding to toxic effects; and
• Role that the purported mechanism(s) of action might contribute to the
diversity of biological response seen in animals and, to some extent, in
humans.
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The remarkable potency of TCDD in eliciting its toxic effects in animals suggested
the possible existence of a receptor for dioxin. Biochemical and genetic evidence implicates
the TCDD-receptor in the biological responses to dioxin-like compounds. Electrophoretic
studies to evaluate the properties of specific proteins from inbred mouse strains reveal the
existence of several forms of the TCDD-binding protein. These observations imply the
existence of multiple alleles at the Ah locus in mice. The biochemical properties of the
different forms of the Ah receptor remain to be described. In particular, the extent to which
the different receptor forms affect the sensitivity to TCDD is not known.
Human cells contain an intracellular protein whose properties resemble those of the
Ah receptor in animals. Binding studies and hydrodynamic analyses have identified an Ah
receptor-like protein(s) in a variety of human tissues. Functional Ah receptors have been
found in many human tissues, including lymphocytes, liver, lung, and placenta. By analogy
with the existence of multiple receptor forms in mice, it is reasonable to anticipate that the
human population will also be polymorphic with respect to Ah receptor structure and
function. Therefore, it is also reasonable to expect that humans may differ from one another
in their susceptibilities to TCDD. The binding and hydrodynamic properties of the Ah
receptor differ relatively little across species and tissues yet responses vary widely; it is
impossible, therefore, to account for the diversity of TCDD's biological effects by
characteristics of the receptor alone.
TCDD acts via an intracellular protein (the Ah receptor), which is a ligand-dependent
transcription factor that functions in partnership with a second protein (known as Arnt);
therefore, from a mechanistic standpoint, TCDD's adverse effects appear likely to reflect
sustained alterations in gene expression. Mechanistic studies also indicate that several
proteins contribute to TCDD's gene regulatory effects and that the response to TCDD
probably involves a relatively complex interplay between multiple genetic and environmental
factors. Such mechanistic information imposes constraints on the possible models that can
plausibly account for TCDD's biological effects and, therefore, on the assumptions used
during the risk assessment process. Mechanistic knowledge of dioxin action may also be
useful in other ways. For example, knowledge of genetic polymorphisms that influence
TCDD responsiveness may allow the identification of individuals at particular risk from
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exposure to dioxin. In addition, mechanistic knowledge of the biochemical pathways that are
altered by TCDD may identify novel targets for the development of drugs that can antagonize
dioxin's adverse effects.
As described below, biochemical and genetic analyses of the mechanism by which
dioxin induces CYP1A1 gene transcription have revealed the outline of a novel regulatory
system whereby a chemical signal can alter the expression of specific mammalian genes.
The evidence to date implies that the Ah receptor participates in every biological response to
TCDD. For example, studies of structure-activity relationships among congeners of TCDD
reveal a correlation between a compound's specific binding affinity and its potency in
eliciting biochemical responses, such as enzyme induction. Furthermore, inbred mouse
strains in which TCDD binds with lower affinity to the receptor exhibit decreased sensitivity
to dioxin's biological effects, such as thymic involution, cleft palate formation, and hepatic
porphyria. While there are a few investigators who believe that dioxin may act directly on
specific cellular and biological processes without Ah-receptor mediation, the majority of
investigators believe that most, if not all, biological responses to dioxin and related
compounds are Ah-receptor mediated. A simplified diagram of this hypothesis is presented
in Figure 9-2. This hypothesis predicts that TCDD will be found to activate the transcription
of other genes via a receptor- and enhancer-dependent mechanism analogous to that described
for the cytochrome P4501A1 (CYP1A1) gene.
Compensatory changes, which occur in response to TCDD's primary effects, can
complicate the analysis of dioxin action in intact animals. For example, TCDD can produce
changes in the levels of steroid hormones, peptide growth factors, and/or their cognate
cellular receptors. In turn, such alterations have the potential to produce a series of
subsequent biological effects, which are not directly mediated by the Ah receptor.
Furthermore, the hormonal status of an animal appears to influence its susceptibility to the
hepatocarcinogenic effects of TCDD (Lucier et al., 1991). Likewise, exposure to other
chemicals can alter the developmental toxicity of TCDD (Couture et al., 1990). Therefore,
in some cases, TCDD may act in combination with other chemicals to produce its biological
effects. Such phenomena increase the difficulty of analyzing dioxin action in intact animals
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Dioxin Exposure
t
Free Dioxin in Tissues
Dioxin Binding to the
Ah Receptor in Tissue
I
Ah Receptor - Dioxin Complex
Binding with DNA
I
Gene Regulation
m-RNA Regulation
*
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*
Biochemical Alterations
f
Late (irreversible) Tissue
Response (cancer, terata)
Figure 9-2. Schematic representation of the complex sequence of molecular and biological
events involved in dioxin-mediated toxicants.
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and increase the complexity of risk assessment, given that humans are routinely exposed to a
wide variety of chemicals.
The fact that TCDD may induce a cascade of biochemical changes in the intact animal
raises the possibility that dioxin might produce a response such as cancer by mechanisms that
differ among tissues. These mechanisms are discussed in detail in Chapter 8, along with the
supporting biological data and dose-response models. One possible mechanism discussed in
Chapter 8 is that TCDD might activate a gene(s) that is directly involved in tissue
proliferation. A second mechanism involves TCDD-induced changes in hormone
metabolism, which may lead to tissue proliferation secondary to increased secretion of a
trophic hormone, and/or to changes in metabolism, which might lead to indirect mutagenic
effects. Thus, while this reassessment has identified a number of hypothetical mechanisms
for cancer induction by TCDD, there remains considerable uncertainty about which
mechanisms occur, with what levels of sensitivity, and in which species. Advances in
knowledge regarding the role of such activities in dioxin toxicity will facilitate the
development of more definitive biologically based models of dioxin action.
Under some circumstances, TCDD can protect against the carcinogenic effects of
polycyclic aromatic hydrocarbons in mouse skin; this may reflect the induction of detoxifying
enzymes by dioxin (Cohen et al., 1979; DiGiovanni et al., 1980). In other situations,
TCDD-induced changes in hormone metabolism may alter the growth of hormone-dependent
tumor cells, producing a potential anticarcinogenic effect (Spink et al., 1990). There is
considerable uncertainty about the magnitude and importance of these effects in relation to
both dose and response characteristics of dioxins in various species. Nonetheless, these (and
perhaps other) effects of TCDD complicate the risk assessment process for dioxin.
A substantial body of biochemical and genetic evidence indicates that the Ah receptor
mediates the biological effects of TCDD. This evidence implies that a response to dioxin
requires the formation of ligand-receptor complexes. TCDD-receptor binding appears to
obey the law of mass action and, therefore, depends on (1) the concentration of ligand in the
target cell; (2) the concentration of receptor in the target cell; and (3) the binding affinity of
the ligand for the receptor. In principle, some TCDD-receptor complexes will form even at
very low levels of dioxin exposure. However, in practice, at some finite concentration of
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TCDD, the formation of TCDD-receptor complexes may be insufficient to elicit detectable
effects. Furthermore, biological events subsequent to TCDD-receptor binding may or may
not exhibit a linear response to dioxin. However, recent studies in several laboratories have
indicated no evidence of a threshold for relatively simple responses to dioxin-like compounds
such as CYP1A1 induction and others. Further information will be required to determine if
other responses to dioxin-like compounds requiring gene transcription will also demonstrate
low-dose linear behavior.
While much of our understanding of TCDD impacts on genetic activity is derived
from studies on liver, studies of other tissues (e.g., skin, thymus) are likely to reveal
additional TCDD-responsive genes, which exhibit tissue-specific expression (Sutter et al.,
1991). Analyses of the mechanism of dioxin action in such systems appear likely to reveal
additional factors that influence the susceptibility of a particular tissue to TCDD. In
addition, studies of other TCDD-inducible genes, such as glutathione-S-transferase, quinone
reductase, and aldehyde dehydrogenase, may reveal whether differences in enhancer
structure, receptor-enhancer interactions, or promoter structure affect the responsiveness of
the target gene to TCDD (Whitlock, 1990).
Based on our understanding of dioxin mechanism(s) to date, it is accurate to say that
interaction with the Ah receptor is necessary, that humans are likely to be sensitive to the
effects of dioxin, and that there is likely to be a variation between and within species and
between tissue in individual species based on differential responses to receptor binding.
Although threshold mechanisms may exist for some of these responses, thresholds have yet
to be demonstrated. Further analyses of dioxin action may provide more insight into the
mechanisms by that TCDD and related compounds produce immunoJogical effects,
reproductive and/or developmental effects or cancer, effects which are of particular public
health concern. A major challenge for the future will be the establishment of experimental
systems in which such complex biological phenomena are amenable to study at the molecular
level.
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9.7. TOXIC EFFECTS OF DIOXIN
9.7.1. General Comments
It is clear from the evaluation of the toxicologic literature that dioxin and related
compounds have the ability to produce a wide spectrum of responses in animals and,
presumably, in humans, if the dose is high enough (Table 9-2). Relatively few chronic
effects related to exposure to dioxin-like compounds have been observed in humans. The
epidemiologic data are limited due to a number of possible factors: the absence of many,
specific individual measurements of dioxin exposure for the general population; a limited
number of cross-sectional and prospective studies of more highly exposed populations; the
limited ability of epidemiologic studies to detect significant differences between exposed and
relatively unexposed populations when the outcomes are relatively rare, the exposures are
low, and the population under study is small; and the difficulty in quantifying the impact of
all potentially confounding exposures. Evaluation of hazard and risk for dioxin and related
compounds must rely on a weight-of-the-evidence approach in which all available data
(animal and human) are examined together. This process often requires extrapolation of
effects across various animal species as well as to humans.
The reliability of using animal data to estimate human hazard and risk has often been
questioned for this class of compounds. Although human data are limited, evidence suggests
that animal models are appropriate for estimating human risk if all available data are
considered. As discussed in detail in Chapters 2 and 8, humans have a fully functional Ah
receptor and both in vivo and in vitro studies demonstrate comparability of biochemical
responses in humans and animals (see also Table 8-5). When comparing species and strains
for their responses to these compounds, a wide range of sensitivity to TCDD-induced
toxicities has been noted. Qualitatively speaking, however, almost every response can be
produced in every species if the appropriate dose is administered. Although outliers, i.e.,
species that are either very sensitive or refractory, can be identified for a particular response,
no species is consistently sensitive or refractory for all effects. In addition, sensitivity for a
given effect among the majority of species clusters within approximately one order of
magnitude (factor of 10). Therefore, despite a range of sensitivities across species, it is
reasonable to assume that humans will not be refractory to all effects nor that they will be as
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sensitive as the most sensitive responder for each effect. Humans are likely, because of
interindividual variability in response to a variety of toxic chemicals, which is generally
greater than that found in individual species of laboratory animals, to show a wide range of
sensitivities for various dioxin-induced toxicities. For purposes of the current assessment,
therefore, unless there are data to identify a particular species as being representative of
humans for a particular effect, average humans can be reasonably assumed to be of average
sensitivity for various'effects, recognizing that individuals in the population might vary
widely in their sensitivity to individual effects. The uncertainty introduced by this
assumption, i.e., that, on average, humans will respond as do average animal models for
individual effects of exposure to dioxin-like compounds and that an unknown range of
variability exists in the human population for individual effects, should be carefully
considered as results of this characterization are applied to individuals or specific
subpopulations.
9.7.2. Chloracne
Chloracne and associated dermatologic changes are widely recognized responses to
TCDD and other dioxin-like compounds in humans. Chloracne is a severe acne-like
condition that develops within months of first exposure to high levels of dioxin. For many
individuals, the condition disappears after discontinuation of exposure, despite serum levels
of dioxin in the thousands of parts per trillion; for others, it may remain for many years.
The duration of persistent chloracne is on the order of 25 years although cases of Chloracne
persisting over 40 years have been noted. There are very little human data from which to
determine definitively the doses at which chloracne is likely to occur. Data from
occupational studies suggest that persistent chloracne is more often associated with exposures
of high intensity, for long duration, and commencing at an early age. Acute exposures or
chronic lower level exposures, if resulting in chloracne, have generally resulted in a
condition that resolves itself in a matter of months to a few yearSi Details of chloracnegenic
response in occupationally exposed humans are described in detail in Chapter 7 of the Health
Assessment Document.
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Induction of chloracne in humans after exposure to dioxin and related compounds is
supported by studies in laboratory animals. Rabbits, monkeys, and hairless mice have all
proved useful in investigating this response. In addition, cellular systems provide a research
tool in elucidating the chloracne response at the cellular level. Keratinocytes, the principal
cell type in the epidermis, have been used as an in vitro model for studies of TCDD-induced
hyperkeratosis, a feature of chloracne, in human- and animal-derived cell cultures. The
response in these systems is analogous to the hyperkeratinization observed in vivo as a part
of chloracne.
There is little doubt that chloracne is a human condition often attributable to exposure
to dioxin and related compounds. The specific risk factors associated with this response are
still obscure. Recognition of chloracne has been associated with high-level exposure to these
compounds, and as such, may represent a biomarker of exposure. Because of the wide
variability of the chloracnegenic response in humans and its varied persistence, however, the
absence of chloracne is not a reliable indicator of low exposure to dioxin and related
compounds.
9.7.3. Carcinogenicity
Since the last EPA review of the human data base relating to the Carcinogenicity of
TCDD and related compounds in 1988, several new follow-up mortality studies have been
completed. Among the most important of these are a study of 5,172 workers by Fingerhut et
al. (1991), a study with 1,583 workers by Manz et al. (1991), a smaller study of 247
workers by Zober et al. (1990), and a study of over 18,000 workers by Saracci et al. (1991).
Although uncertainty remains in interpreting these studies because not all potential
confounders have been ruled out and coincident exposures to other carcinogens is likely, all
provide support for an association between exposure to dioxin and related compounds and
increased cancer mortality. With the exception of the study by Saracci et al. (1991), these
studies have some exposure information that permits an assessment of dose response. These
data have in fact served as the basis for fitting the additive and multiplicative risk models in
Chapter 8. In addition, more limited results have been presented recently on the Seveso
cohort (Bertazzi et al., 1993) and on women exposed to chlorophenoxy herbicides,
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chlorophenols, and dioxins (Kogevinas et al., 1993). While these two studies have
methodologic shortcomings that are described in Chapter 7, they provide findings,
particularly for exposure to women, that warrant additional follow-up.
While the data base from epidemiologic studies remains controversial, it is the view
of this reassessment that this body of evidence supports the laboratory data indicating that
TCDD probably increases cancer mortality of several types. Although not all confounders
were ruled out in any one study, positive associations between surrogates of dioxin exposure,
either length of occupational exposure or proximity to a known source combined with some
information on body burden, and cancer have been reported. These data alone suggest a role
for dioxin exposure to contribute to a carcinogenic response but do npt confirm a causal
relationship between exposure to dioxin and increased cancer incidence. Available human
studies alone cannot demonstrate whether a cause and effect relationship between dioxin
exposure and increased incidence of cancer exists. Therefore, evaluation of cancer hazard in
humans must include an evaluation of all of the available animal and in vitro data as well as
the data from exposed human populations.
The Peer Panel that rnet in September 1993 to review an earlier draft of the cancer
epidemiology chapter suggested that the epidemiology data alone were still not adequate to
implicate dioxin and related compounds as "known" human carcinogens but that the results
from the human studies were largely consistent with observations from laboratory studies of
dioxin-induced cancer and, therefore, should nqt be dismissed or ignored, Other scientists,
including those who attended the Peer Panel meeting, felt either rn.ore or less strongly about
the weight of the evidence from epidemiology studies, representing tfie range of opinion that
still exists on the interpretation of the cancer epidemiology studies,
Many of the earlier epidemiolqgical studies that suggested an association with soft
tissue sarcoma were criticized for a variety of reasons. Nonetheless, the incidence of soft
tissue sarcoma is elevated in several of the recent studies, supporting the findings from
previous studies. The fact that similar results were obtained in independent studies of
differing design and evaluating populations exposed to dioxin-like compounds under varying
conditions, along with the rarity of this tumor type, weighs in favor of a consistent and real
association. On the other hand, arguments regarding selection bias, differential exposure
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misclassification, confounding, and chance in each individual study have been presented in
the scientific literature which increase uncertainty around this association. In addition,
excess respiratory cancer was noted by Fingerhut, Zober, and Manz. These results are also
supported by significantly increased mortality from lung and liver cancers subsequent to the
Japanese rice oil poisoning accident where exposure to PCDFs and PCBs occurred. Again,
while smoking as a confounder cannot be totally eliminated as a potential explanation of
these results, analyses conducted to date suggest that smoking is not likely to explain the
entire increase in lung cancer. The question of confounding exposures, such as asbestos and
other chemicals, in addition to smoking, has not been entirely ruled but and must be
considered as potentially adding to the observed increases. Although increases of cancer at
other sites (e.g., non-Hodgkin's lymphoma, stomach cancer) have been reported, the data for
an association with exposure to dioxin-like chemicals are less compelling.
The comparison of the results of different investigations that examine the outcome of
similar exposures must always be evaluated in light of factors that may influence the outcome
of the study. A few of these factors include study design, potential confounding factors and
exposures (extraneous factors or exposures that relate to both outcome and exposure such as
age), biases that affect the selection and participation of the study population, differential
exposure misclassification, variation in age of the study population, (different conditions of
exposure (mode, intensity, duration, and route), and differences in methods used to assess
outcomes of interest. Such differences may result in some variation in the results of the
compared studies. Given that the studies are well conducted and the variations noted, what
is important is the within-study consistency of the results.
What emerges from an analysis of the epidemiology data is a view of dioxin-like
compounds as potentially multisite carcinogens in more highly exposed human populations
that have been studied, consisting primarily of adult males. There are currently very few
data for women and children exposed to dioxin-like compounds. Although uncertainty in this
view remains, the cancer findings are generally consistent with results from studies of
laboratory animals and appear to be plausible given what is known about mechanisms of
"i
dioxin action.
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While both past and more recent human studies have focused on males, there are
some limited data suggesting carcinogenic responses associated with dioxin exposure in
females. Because both laboratory animal data and mechanistic inferences suggest that males
and females may respond differently to dioxin-like activity, further data will be needed to
address this question of differential response.
An extensive data base on the carcinogenicity of dioxin and related compounds in
laboratory studies exists and is described in detail in Chapter 6r There is adequate evidence
that 2,3,7,8-TCDD is a carcinogen in laboratory animals based on long-term bioassays
conducted in both sexes of rats and mice. All studies have produced positive results, leading
to the conclusions that TCDD is a multistage carcinogen increasing the incidence of tumors
at sites distant from the site of treatment and at doses well below the maximum tolerated
dose. Since this issue was last reviewed by the Agency in 1988, TCDD has been shown to
be a carcinogen in hamsters, which are relatively resistant to the lethal effects of TCDD.
Recent data have also shown TCDD to be a liver carcinogen in the small fish, Medaka
(Johnson et al., 1992). Few attempts have been made to demonstrate the carcinogenicity of
other dioxin-like compounds, Other thaji a mixture of two isomgrs of
hexachlorodibenzodioxin (HCDDs), which produced liver tumors in both sexes of rats and
mice (NTP, 1980), the more highly chlorinated CDDs and CDFs hay§ not been studied in
long-term animal cancer bioassays, However, it is generally recognised that these
compounds bioaccumulate and exhibit toxiqities similar to TCDD and ar§, therefor©, also
likely to be carcinogens (U.S. EPA Science Advisory Bpard, 1989),
In addition to the demonstration of TCDD as a complete carcinogen in long-term
cancer bioassays, a number of dioxin-like PCDDs and PCDFs, aj vvell as several PCBs, have
also been demonstrated to be tumor promoters in two-stage (initiation-promotion) protocols in
rodent liver and skin. In addition, a recent study has demonstrated the ability of TCDD to
neoplastically transform immortalized human cells in culture at very low concentrations of
TCDD. While dioxin and related compounds are not generally considered to be "genotoxic"
in traditional terms, both empirical data and the results of modeling efforts suggest that they
may be functioning indirectly to produce irreversible genetic changes in exposed cells. All
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of these data add substantially to the weight of the evidence that dioxin and related
compounds are likely to be carcinogenic, at least under some circumstances, in humans.
Despite the relatively large number of bioassays on TCDD, the study of Kociba et al.
(1978) and those of the NTP (1982), because of their multiple dose groups and wide dose
range, continue to be the focus of additional review. Sauer (1990) re-evaluated the female
rat liver tumors in the Kociba study using the latest pathology criteria for such lesions. The
review confirmed only approximately one-third of the tumors of the: previous review (Squire,
1980). While this finding did not change the determination of carcinogenic hazard since
TCDD induced tumors in multiple sites in this study, it does have an effect on evaluation of
dose-response and on estimates of risk at low doses. These issues will be discussed in a later
section of this chapter.
One of the more interesting findings in the Kociba bioassay was reduced tumor
incidences of the pituitary, uterus, mammary gland, pancreas, and sidrenals. These findings,
coupled with the sex specificity of the TCDD-induced liver tumors in rats, emphasize that the
carcinogenic actions of TCDD involve a complex interaction of hormonal factors.
Moreover, it is hypothesized that cell-specific factors modulate TCDD/hormone actions
relevant to cancer. The findings of reduced tumor incidence in certain tissues suggest that
dioxin exposure may be exerting an anticarcinogenic effect under certain circumstances or in
certain tissues. The complex interplay between dioxin and hormones in terms of both
carcinogenic and anticarcinogenic responses will continue to be a matter of hypothesis until
specific data to address these issues are obtained.
In summary, publication of additional studies of human populations exposed to dioxin
and related compounds since the last EPA assessment (Fingerhut et al., 1991; Manz et al.,
1991; Zober et al., 1990; Saracci et al., 1991; Bertazzi et al., 1993; Kogevinas et al., 1993)
has strengthened the inference, based on all the evidence from mechanistic, animal, and
epidemiologic studies, that these compounds are appropriately characterized as probable
human carcinogens. While the data for 2,3,7,8-TCDD are particularly comprehensive, the
data on other congeners remain limited. This puts added emphasis on the assumptions and
inferences regarding toxicity equivalence in evaluating complex exposures to dioxin and
related compounds with regard to carcinogenicity. The evolving understanding of the
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complex interplay between dioxin-like compounds and hormones and other modulators of cell
growth and differentiation continues to complicate more precise determinations of cancer
hazard and risk.
9.7.4. Reproductive and Developmental Effects
The potential for dioxins and related compounds to cause reproductive and
developmental toxicity in animals has been recognized for many years, and the data base
regarding these effects is analyzed in Chapter 5. Recent laboratory studies have suggested
that altered development may be among the most sensitive TCDD end points in laboratory
animal systems although the likelihood and level of response in humans are much less clear.
Although the discussion of these effects in Chapter 5 is divided into developmental toxicity
and male and female reproductive toxicity, it is important to recognize the interrelatedness of
developmental and reproductive events at all levels of biological complexity. This point is
critical for understanding and fully characterizing the hazards and risks of dioxin and related
compounds. For example, effects of TCDD on circulating levels of sex hormones and/or on
responsiveness to sex hormones in laboratory animals or humans may be translated into
reproductive dysfunction if exposure occurs in adulthood as well as abnormal development
and/or reproductive dysfunction if exposure occurs prenatally. Therefore, a similar effect of
dioxin-like compounds may be manifest as a reproductive end point if exposure occurs to
adults or as a developmental and/or a reproductive end point if exposure occurs to the fetus.
Likewise, even though effects on organ structure and on growth are considered separate
developmental end points that are associated with pre- and postnatal exposure to TCDD in
laboratory animals, they are interrelated because effects on prenatal growth can significantly
disrupt the structural integrity of an organ system. It is important to note that adverse
developmental effects are a complex set of end points, many of which are caused by multiple
factors, requiring coincidence of a number of events.
In the current data base, developmental toxicity end points have been observed at
lower TCDD exposure levels than male and female reproductive toxicity end points in a
number of animal systems. The lowest effective TCDD egg burden for causing
developmental toxicity in fish and birds and the lowest effective maternal TCDD body
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burden for producing a wide range of developmental responses in mammals are summarized
in Chapter 5. Of particular interest to the risk assessment process is the fact that a wide
variety of developmental events, crossing three vertebrate classes arid several species within
each class, can be perturbed, suggesting that dioxin has the potential to disrupt a large
number of critical developmental events at specific developmental stages. Not only can these
changes lead to increases in embryo/fetal mortality, but they can disrupt organ system
structure and irreversibly impair organ function.
The laboratory studies demonstrating adverse health effects from prenatal exposures
often involved a single dose administered at a discrete time during pregnancy. The doses
that produced adverse effects, such as reproductive and developmental toxicity, can be
related to longer term body burdens produced by the single dose or to background body
burdens. Because the production of prenatal effects often requires exposures to occur during
certain critical times during fetal development, the uncertainties in the relationship with
steady-state body burdens must be carefully assessed. A single dose may cause a spike in
both maternal and fetal blood concentration related to the magnitude; of the dose, and the
concentrations will fall rapidly as the dioxin-like compounds are redistributed to adipose and
other tissues. Application of pharmacokinetic models described earlier in this chapter to
estimate blood concentrations at the critical time of development is expected to be a sound
method for relating chronic background exposures to the results obtained from single-dose
studies.
Because developmental toxicity following exposure to TCDD-like congeners occurs in
fish, birds, and mammals, it is likely to occur at some level in humans. It is not currently
possible to state exactly how or at what levels humans in the population will respond with
adverse impacts on development or reproductive function. Data analyzed in Chapter 5 and
Chapter 7 suggest, however, that adverse effects may be occurring a.t levels lower than
originally thought to represent a no observed adverse effect level (NOAEL) in animals.
Traditional toxicology studies had led to the conclusion that the NOAEL was in the range of
intake values of 1 ng TEQ/kg/day. Current data suggest that the NOAEL in animals should
be lower. This issue will be discussed further in the dose-response section of this chapter.
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While human data on potential developmental effects of dioxin-like compounds are
limited, developmental effects in human infants exposed to a complex mixture of PCBs,
CDFs, and PCQs in the Yusho and Yu-Cheng poisoning episodes were probably caused by
the combined exposure to those PCB and CDF congeners that are Ah-receptor agonists.
However, it should be noted that not all effects that are seen are attributable only to dioxin-
like compounds. Similarity of the effects observed in human infants prenatally exposed to
this complex mixture with those reported in adult monkeys exposed only to TCDD increases
the probability that at least some of the effects in the Yusho and Yu-Cheng children are due
to the TCDD-like congeners in the contaminated rice oil ingested by the mothers of these
children. Most significant is a clustering of effects in organs derived from the ectodermal
germ layer, a syndrome referred to as ectodermal dysplasia. Included in this syndrome are
effects on the skin, nails, and meibomian glands that occur in both adult monkeys exposed to
TCDD and in Yusho and Yu-Cheng infants exposed transplacentally to PCB, CDF, and PCQ
contaminated rice oils. In addition, accelerated tooth eruption has been reported both in
human infants affected by the Yusho and Yu-Cheng exposures and in neonatal mice exposed
to TCDD. Yu-Cheng children exposed transplacentally to PCB, CDF, and PCQ
contaminated rice oil have also exhibited developmental and psychomotor delay during
developmental and cognitive tests (Chen et al., 1992). Some investigators believe that,
because these effects do not correlate with TEQ, the effects are exclusively due to nondioxin-
like PCBs or a combination of all congeners. However, monkeys pre- and postnatally
exposed to TCDD are also affected by a deficit in cognitive function. Recent studies
presented at Dioxin '93 (Hsu et al., 1993; Lai et al., 1993) have demonstrated that these,
effects persist throughout childhood, as does the growth retardation (Guo et al., 1994). The
concept that the ectodermal dysplasia syndrome in Yusho and Yu-Cheng infants may be
caused by the combination of PCB and CDF congeners in the rice oil that are Ah receptor
agonists but are less potent than TCDD is consistent with structure-activity results for various
developmental end points in different species of fish, birds, and mammals.
In mammals, postnatal functional alterations involving learning behavior and the
developing reproductive system appear to be the developmental events most sensitive to
prenatal dioxin exposure. The developing immune system may also be highly sensitive.
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Alterations in structures and diminished prenatal viability and growth begin to predominate at
maternal TCDD body burdens and/or daily TCDD doses during gestation that are above 100
ng/kg in virtually every species tested. These doses of TCDD are not maternally toxic.
Higher dose levels can be demonstrated to result in prenatal mortality. A general finding in
fish, bird, and mammalian species is that the embryo or fetus is more sensitive to TCDD-
induced mortality than the adult. Thus, the timing of TCDD exposure during the life history
of an animal can greatly influence its susceptibility to overt dioxin toxicity.
With respect to male and female reproductive end points, there are clear effects
following dioxin exposure of the adult animal. Such reproductive effects generally occur at
TCDD body burdens that are higher than those required to cause the more sensitive
developmental end points. For example, TCDD exposure of the adult male rodent causes
reduced testis and accessory sex organ weights, abnormal testis structure, decreased
spermatogenesis, reduced fertility, decreased testicular testosterone synthesis, reduced plasma
androgen concentrations, and altered regulation of pituitary LH secretion. However, in
laboratory animal studies, these effects are detectable only at TCDD exposure levels that are
overtly toxic to the animal. In the more limited studies focusing on female reproduction, the
primary effects include decreased fertility, inability to maintain pregnancy, and in the rat,
decreased litter size. Signs of ovarian dysfunction and alterations in hormone levels have
also been reported.
Exposure of female mice and rats to TCDD has an antiestrogenic effect on the uterus.
The dose of TCDD required to produce this response is generally higher than that needed to
cause the most sensitive signs of developmental toxicity in these speeies. More specifically,
hydronephrosis and cleft palate in mice and reductions in spermatogenesis in rats occur at
maternal doses of TCDD that are far less than those needed to exert a demonstrable
antiestrogenic effect when adult female mice and rats are exposed to dioxin. The precise
mechanism of TCDD's antiestrogenic effect is not fully understood. It may be caused by
both a decrease in available estrogen receptor number and/or by an increase in cytochrome
P-4501A-mediated estrogen metabolism within the target cell.
These studies indicate that while there is variability between species in the profile of
developmental responses elicited by TCDD, essentially all dioxin-like PCB, CDD, and CDF
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congeners that have Ah receptor affinity and intrinsic activity produce the same pattern of
developmental effects within a given vertebrate species if a sufficiently high dose of the
congener is given. Data to support these conclusions regarding reproductive and
developmental hazards of dioxin and related compounds continue to accumulate, but the
weight of the evidence is still a subject of much scientific debate.
9.7.5. Immunotoxicity
Concern over the potential toxic effects of chemicals on the immune system arises
from the critical role that the immune system plays in maintaining health. It is well
recognized that suppressed immunological function can result in increased incidence and
severity of infectious diseases as well as some types of cancer. Conversely, the
inappropriate enhancement of immune function or the generation of misdirected immune
responses may precipitate or exacerbate the development of allergic and autoimmune
diseases. Thus, suppression as well as enhancement of immune function are considered to
represent potential immunotoxic effects of chemicals.
Extensive evidence has accumulated over the past 20 years to demonstrate that the
immune system is a target for toxicity of TCDD and structurally related compounds,
including PCDDs, PCDFs, PCBs, and PBBs. This evidence is described in detail in Chapter
4. The evidence has derived from numerous studies in various animal species, primarily
rodents, but also guinea pigs, rabbits, monkeys, marmosets, and cattle. Epidemiological
studies also provide some evidence for the immunotoxicity of dioxin and related compounds
in humans. In animal studies, relatively high doses of HAH produce lymphoid tissue
depletion, except in the thymus where cellular depletion occurs at lower doses. Alterations
in specific immune effector functions and increased susceptibility to infectious disease have
been identified at doses of TCDD well below those that cause lymphoid tissue depletion.
Both cell-mediated and humoral immune responses are suppressed following TCDD
exposure, suggesting that there are multiple cellular targets within the immune system that
are altered by TCDD. Evidence also suggests that the immune system is indirectly targeted
by TCDD-induced changes in nonlymphoid tissues. In addition, in parallel with increased
understanding of the cellular and molecular mechanisms involved in immunity, studies on
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TCDD are beginning to establish biochemical and molecular mechanisms of TCDD
immunotoxicity.
The ability of an animal to resist and/or control viral, bacterial, parasitic, and
neoplastic diseases is determined by both nonspecific and specific immunological functions.
Decreased functional activity in any immunological compartment may result in increased
susceptibility to infectious and neoplastic diseases. In terms of risk assessment, host
resistance is often accorded the "bottom line" in terms of relevant immunotoxic end points.
Animal host resistance models that mimic human disease are available and have been used to
assess the effect of TCDD on altered host resistance. Results from host resistance studies
provide evidence that exposure to TCDD results in increased susceptibility to bacterial, viral,
parasitic, and neoplastic diseases. These effects are observed at relatively low doses and
likely result from TCDD-induced suppression of immunological function. The specific
immunological functions targeted by TCDD in each of the host resistance models remain to
be fully defined.
Despite considerable investigation, the cells that are altered by TCDD exposure,
leading to suppressed immune function, have not been unequivocally identified. Direct in
vitro effects of TCDD on purified B cell activity have been reported, while direct effects on
macrophages and T cells in vitro have not been described. The in vitro effects of TCDD on
lymphocytes, however, appear to be influenced by cell culture conditions, which may explain
the discrepancies in effects observed in different laboratories. Although the direct effects of
TCDD on T cells in vitro have not been demonstrated, it is clear that functional T cell
responses generated in vivo are compromised following in vivo exposure. TCDD may alter
immune function by indirect mechanisms. One potentially important indirect mechanism is
via effects on the endocrine system. Several endocrine hormones have been shown to
regulate immune responses, including glucocorticoids, sex steroids, thyroxine, growth
hormone, and prolactin. Importantly, TCDD and other related compounds have been shown
to alter the activity of these hormones.
It is important to consider that if an acute exposure to TCDD even temporarily raises
the TCDD body burden at the time when an immune response is initiated., there may be a
risk of adverse impacts even though the total body burden may indicate a relatively low
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average TCDD level. Furthermore, because TCDD alters the normal differentiation of
immune system cells, the human embryo may be very susceptible to long-term impairment of
immune function from in utero effects of TCDD on developing immune tissue. There are
currently no data to directly support this hypothesis. Concern arises as a consequence of
inferences derived from an understanding of dioxin action and observations in humans and
laboratory animals.
In summary, evidence has accumulated to demonstrate that the immune system is a
target for toxicity of TCDD and structurally related compounds. The evidence has derived
from numerous studies in various animal species. Animal studies suggest that some
immunotoxic responses may be evoked at very low levels of dioxin exposure.
Epidemiological studies also provide conflicting evidence for the immunotoxicity of these
compounds in humans. Few changes in the immune system in humans associated with dioxin
body burdens have been detected when exposed humans have been studied. Both direct and
indirect (e.g., hormonally mediated) impacts on the immune system have been hypothesized
to be the basis of dioxin immunotoxicity. While there is speculation that the developing
immune system may be particularly sensitive to the effects of exposure to dioxin and related
compounds, additional research will be needed to support this hypothesis.
9.7.6. Other Effects
A number of other effects of dioxin and related compounds have been discussed in
some detail throughout the chapters in this assessment. While they illustrate the wide range
of effects produced by this class of compounds, some may be specific to the species in which
they are measured and may have limited relevance to the human situation. On the other
hand, they may be indicative of the fundamental level at which dioxin produces its biological
impact and may represent a continuum of response expected from these fundamental changes.
While all may not be adverse effects (some may be adaptive and of neutral consequence),
several effects have been noted in human studies or in primates that deserve special mention.
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9.7.6.1. Circulating Reproductive Hormones
Two cross-sectional epidemiologic studies have detected an association between levels
of male reproductive hormones and exposure to TCDD. Decreased testosterone levels were
detected in two of the three studies where testosterone was evaluated and luteinizing hormone
(LH) was increased in one of the two studies evaluating that end point. The fact that the
results are based on a single sample rather than on the currently preferred series of three
samples adds to the uncertainty of these findings. Animal data are available to support the
plausibility of these findings. The mechanism(s) responsible for this effect are largely
unknown, but changes in receptor level or function and hormone metabolism and homeostasis
need to be investigated. If these data continue to hold up in future observations, their
clinical significance will need to be further evaluated. Follow-up studies are currently under
way.
9.7.6.2. Diabetes and Fasting Serum Glucose Levels
Epidemiologic evidence has been presented to suggest an increased risk of diabetes
and for an elevated prevalence of abnormal fasting serum glucose levels with dioxin
exposure. Three studies found that individuals with elevated serum levels of TCDD had a
slight but statistically significant or borderline significant increased risk for developing
diabetes or having elevated fasting serum glucose. There are virtually no animal data to
corroborate these findings although some data have indicated effects of TCDD on glucose
metabolism and insulin function. While the findings of a greater prevalence of elevated
fasting glucose may presage the development of diabetes, in the NIOSH study of chemical
workers, the traditional risk factors for diabetes (age, body mass index or weight, and family
history of diabetes) appear substantially more influential than TCDD exposure in the
development of the disease.
9.7.6.3. Enzyme Induction
One of the best characterized effects of exposure to dioxin-like compounds is the
induction of cytochrome P-450 1A1 (CYP1A1). CYP1A1 is one of a family of proteins
involved in the activation and detoxification of both endogenous and exogenous chemicals.
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Dioxin also increases the activity of a number of other enzymes involved in
biotransformation reactions. Increased activity of these enzymes has been implicated
mechanistically in the toxic responses seen in animals in response to dioxin-like compounds.
For example, it has been hypothesized that increases in UDP-glucuronyltransferases leads to
elimination of thyroxine and may lead indirectly to increased thyroid-stimulating hormone
synthesis by the pituitary and subsequent hyperplastic and hypertrophic responses by the
thyroid. There is speculation that such prolonged stimulation may lead to the thyroid tumors
seen in both rats and mice exposed to TCDD. Therefore, while changes in enzyme activity
in response to dioxin and related compounds may result in detoxification of certain
chemicals, examples exist in experimental animals of changed metabolism leading directly or
indirectly to adverse effects, some as severe as cancer. Data to confirm this effect of dioxin
and related compounds in humans are not available.
9.7.6.4. Gamma Glutamyl Trans/erase (GGT) Activity
GOT is one of the many hepatic enzymes that are measured in human serum to
evaluate liver toxicity. Of these, GGT is the only hepatic enzyme found in a number of
human studies to be chronically elevated in adults exposed to high levels of 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. There are
few animal data to support these findings.
9.7.6.5. Endometriosis
Endometriosis is a serious disorder of the female reproductive system that is of
unknown etiology and a major cause of infertility in women. A recent study has determined
that chronic exposure to TCDD increases the risk of endometriosis in rhesus monkeys (Reier
et al., 1993). The incidence and severity of the disease were dose dependent. Additional
studies are under way to further evaluate these observations in rhesus monkeys, and studies
are planned to evaluate women exposed to TCDD after the accident at Seveso for any
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correlation between dioxin body burden and incidence or severity of endometriosis. Further
evaluation of this health end point awaits reports from these studies.
9.8. DOSE-RESPONSE CONSIDERATIONS
The current efforts to evaluate the risks of dioxin and related compounds have focused
on the understanding of the biological basis of response as well as evaluation of the weight of
the empirical observations on inferences regarding hazard and risk. Previous sections have
discussed the relationship of binding of this class of compounds to a specific receptor and
subsequent events. It is generally accepted that all well-studied responses to dioxin appear to
be mediated by receptor binding. This situation is not unlike the signal transduction
pathways that have been described for hormone action, particularly exemplified by the well-
studied family of steroid hormones, although the dioxin receptor does not belong to the
steroid receptor family.
The fact that much of the biological activity of this class of compounds follows the
rank order of binding affinity of the congeners to the Ah-receptor supports the concept that
these earliest steps play a determining role in the probability that later responses will occur.
This does not imply that a simple proportional relationship between receptor binding and
biological response can explain the diversity of biological responses described for dioxin and
related compounds. It is likely that differences in response will be due to tissue and cell-
specific factors that modulate the qualitative relationship between receptor binding, or more
precisely, occupancy and response. It is expected that there may be markedly different dose-
response relationships for different effects of dioxin, depending on the respective roles of
modulating activities. Coordinated biological responses, such as TCDD-mediated increases
in cell proliferation, likely involve numerous cellular factors and hormone systems. This
means that the dose-response for relatively simple sequelae of the early binding events such
as cytochrome (CYP1A1) induction may not accurately predict dose-response relationships
for more complex responses such as cancer. Much additional knowledge will be required
before we can accurately predict these complex dose-response relationships.
Development of biologically based dose-response models for dioxin and related
compounds as a part of this reassessment has led to considerable and valuable insights
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regarding both mechanisms of dioxin action and dose-response relationships for dioxin
effects. These are described in some detail in Chapter 8. These efforts have provided
additional perspectives on traditional methods such as the linearized multistage (LMS)
procedure for estimating cancer potency or the uncertainty factor approach for estimating
levels below which noncancer effects are not likely to occur. These methods have also
provided a biologically based rationale for what had been primarily statistical approaches.
The development of models like those in Chapter 8 allows for an iterative process of data
development, hypothesis testing, and model development. These efforts have resulted in
incorporation of more of the available biological data into models to predict human risk at
low increments of exposure.
Tables 9-3 through 9-6 summarize estimated body burdens and effect levels for a
variety of species, including the lowest observed effect levels (LOELs) for some of the more
sensitive indicators of biological response induced by dioxin and related compounds.
Important assumptions used in deriving these values are included as part of this table. It is
particularly important to note that the estimated body burdens associated with several of
these experimental doses are quite low relative to background body burdens in the general
human population. The implications of this observation will be discussed later in this
chapter.
Dose-response modeling efforts in Chapter 8 for liver cancer in female rats and for
lung cancer and all cancers combined in humans have produced results that can be used to
estimate risk-specific doses and risk estimates. Estimates from these efforts differ with
models based on the human data, providing somewhat higher risk estimates than the animal-
based estimates. The risk estimates resulting from these models have uncertainties that cause
their ranges to overlap, and all models produce fits that are consistent with the linearized
multistage model commonly used for cancer risk estimates. By the definitions of mechanistic
modeling given in Chapter 8, both modeling efforts fall short of completely explaining
conditions of biology or exposure. However, because the animal modeling establishes a
better mechanistic basis for extrapolation to low doses and the animal data have greater
certainty in terms of causal association for cancer (especially considering that smoking may
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Table 9-3. .Estimated Body Burdens of Experimental Animals and Humans Exposed to
Dioxins: Responses in Humans Causally Associated With Exposure to Dioxins and
Comparable Effects in Experimental Animals
Effect
Chloracne
Chloracne
Chloracne
Chloracne
Decreased Birth
Weight
Decreased Growth
Decreased Growth
Delayed
Developmental
Milestones
Object Learning
Down Regulation
of EGFR in
Placenta (Maximal
Effect)
Species
Humans
Monkey
Rabbits
Mice
Humans
Humans
Rats
Humans
Monkey
Humans
Experimental
Dose
1,000 ng/kg
4 ng/kg
5d/wk/4wk
4,000 ng/kg
3d/wk/2wk
Mother's body
burden
Mother's body
burden
400 ng/kg
maternal dose
gd 15
1.26 ng/kg/d
Body Burden
45-3,000 ng/kg
1,000 ng/kg
220 ng/kg
" 14,000 ng/kg
1,460 ng/kg
1,460 ng/kg
400 ng/kg
1,460 ng/kg
19 ng/kg
1,460 ng/kg
Ref./Note
Ryanetal.,
1990; Beck et
al., 1989/a,b
McNulty,
1985/c
Schwetz et al.,
1973/d
Puvel and
Sakamoto,
1988/e
Lucier, 1991/f
Guo et al. ,
1994/f
Mably et al. ,
1992a/g
Rogan et al. ,
1988/f
Schantz and
Bowman,
1989/h
Lucier, 1991/f
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Table 9-3. (continued)
Effect
Down Regulation
of EGFR in Liver
(Maximal Effect)
Increase in
Placental CYP1A1
(Maximal Effect)
Increase in Liver
CYP1A1 (Maximal
Effect)
Enzyme Induction
CYP1A1 (LOEL)
Enzyme Induction
CYP1A1/1A2
(LOEL)
Hepatic
Sequestration
Hepatic
Sequestration
Background
Background
Species
Rats
Humans
Rats
Rats
Mice
Human
Rats
Human
Mouse
Experimental
Dose
125 ng/kg/d
30 weeks
125 ng/kg/d
30 weeks
1 1 ng/kg
single dose
sac 24 hr
1.5 ng/kg/d
5 d/wk 13 wk
60 TEQ ppt in
serum
Body Burden
1,600 ng/kg
1,460 ng/kg
1,600 ng/kg
1 ng/kg
23 ng/kg
150 ng/kg
300 ng/kg
9 ng/kg
4 ng/kg
Ref./Note
Sewallet al.,
1993/i
Lucier, 199 1/f
Trltscher et
al,, 1992/i
Van den
Heuvel et al. ,
1993/j
DeVito etal.,
1994/k
Carrier etal.,
submitted/1
Carrier etal,,
submitted/1
m
n
Notes:
a. All human data assume a background level of 60 ppt TEQs in serum (lipid adjusted) in
addition to the dioxin levels presented in the referenced papers, Dioxins are assumed to
be distributed in the body lipid. Thus the concentration of dioxins in serum expressed as
lipid adjusted are assumed to be equivalent to the concentration of dioxins in total body
lipids. In addition, the average person is assumed to weigh 70 kg with 15% of the
weight from body fat. Hence a person with background levels of 60 ppt TEQs in serum
(lipid-adjusted levels) has a body burden of 9 ppt or 9 ng/kg. Although unpublished
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Table 9-3. (continued)
studies in our laboratory indicate that untreated 150-day-old mice also have background
levels of dioxins and dibenzofurans of approximately 4 ng TEQ/kg, these values were
not included in body-burden estimates for the effects seen in experimental animals.
b. The lower value, 45 ng/kg, is from a patient with chloracne who had the lowest reported
serum dioxin level for any patient with chloracne (Ryan et al., 1990). In this patient
adipose tissue levels at the time of exposure, and the development of chloracne, are
estimated by the authors (Ryan et al., 1990) based on the patient's adipose tissue level of
dioxins of 237 ppt and assuming a half-life of dioxin of 7.1 yeairs. The higher of the
two values is from Ryan et al., 1989 and represents the average; body burden of dioxins
in persons from Yu-Cheng who developed chloracne (Beck et al., 1989). Estimates of
body burdens from the Yu-Cheng patients were determined by Ryan et al. in Beck et al.
(1989).
c. Animal administered 1 /tg/kg TCDD and it is assumed that essentially no TCDD was
eliminated when the animal developed a chloracnegenic response. This is a LOEL dose;
no lower doses were tested.
d. Assumes the same rate of elimination as the rat and that the animals weighs 2.5 kg
throughout the experiment. This is a LOEL dose and no lower doses were tested.
e. Assumes a half-life of 11 days and an average weight of the animal at 25 grams. This is
a LOEL dose, and no lower doses were administered.
f. According to the author (Lucier, 1991), in highly exposed patients from Yu-Cheng, there
is a decrease in birth weights of children born from these patients compared to
unexposed control populations. In addition, there is an association between placenta!
levels of dioxins and alterations in placenta! epidermal growth factor receptor (EGFR)
and CYP1A1. In addition, the author suggested that the changes in placental EGFR and
CYP1A1 in these patients were maximal. Body burdens determined based on levels of
2,3,4,7,8-pentachlorodibenzofuran (TEF=0.1) and 1,2,3,4,7,8-hexachlorodibenzofuran
in placenta tissue. Assumes placenta is 1 % lipid (Beck et al., 1994) and that women
have a fat content of 21% of body weight (Ganong, 1982). Also used these body
burdens to estimate body burden of mothers of the children with decreased growth (Guo
et al., 1994) and delayed developmental milestones (Rogan et al., 1988). All patients
are from the Yu-Cheng rice oil poisoning.
g. Assumes pups exposed to an equal dose of TCDD as are the dams on a weight basis and
that the pups do not eliminate any of the TCDD. For decreased body weight in pups
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Table 9-3. (continued)
400 ng/kg is the LOEL, a dose of 64 ng/kg to the dam was the NOEL for this response.
For decreased sperm count, the LOEL is 64 ng/kg, and no lower doses were tested.
h. Assumes a single first-order elimination rate constant and a half-life for the whole body
elimination of 400 days (McNulty, 1985) and a gastrointestinal absorption of 86% (Rose
et al., 1976). This is the LOEL from this study; no lower doses tested.
i. From Tritscher et al. (1992) and Maronpot et al. (1993). Liver levels measured in study
at approximately 30 ppb. Liver and body weights were reported in 40. Assumes animal
is 20% body fat by weight and that at this dose, the liver has four times the
concentration of TCDD than adipose tissue. The body-burden calculation assumes that
liver and fat account for 90% of the body burden in these animals. For tumor
promotion, this is the LOEL in these animals. A NOEL for tumor promotion was
observed at a dose of 35 ng/kg/d. For induction of CYP1A1 and downregulation of
EGF-R, this body burden produces a maximal response.
j. Animals received a single dose and were sacrificed 24 hours later. Assumes no TCDD
eliminated at this time. CYP1A1 induction determined by RT-PCR. This is the LOEL
for this response; a NOEL from this study is 0.1 ng/kg.
k. Animals received 1.5 ng/kg/d, 5 d/wk for 13 wk. Animals sacrificed 3 days after last
dose. Hepatic, dermal, and pulmonary EROD activity induced at this dose. Tissue
levels measured in liver, skin, and fat. Assumes 100% of the body burden is in liver,
skin, and fat. This is the LOEL from this study; no lower doses were tested.
1. Body burdens are estimated by Zinkl et al. (1973) for the increased accumulation of
PCDD/PCDF in liver compared to adipose tissue.
m. Assumes a background TEQ of 60 ppt for dioxins, dibenzofurans, and PCBs. Also
assumes a body weight of 70 kg with 15% body fat.
n. Data from DeVito and Birnbaum. TEQ for TCDD 1,2,3,7,8-PCDD; 2,3,7,8-TCDF;
1,2,3,7,8-PCDF; 2,3,4,7,8-PCDF; and OCDF in 150 day old female B6C3F1 mice.
Chemicals were determined in liver, fat and skin of these animals. Assumes that 100%
of the body burden is in liver, fat, and skin.
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Table 9-4. Estimated Body Burdens of Experimental Animals and Humans Exposed
to Dioxins: Responses in Humans Associated With Dioxin Exposure and Comparable
Effects in Experimental Animals
Effect
Cancer
Cancer
Cancer
Cancer
Liver Tumor
Promotion
Skin Tumor
Promotion
Decreased
Testosterone
Decreased
Testosterone
Decreased Testis
Size
Altered Glucose
Tolerance
Altered Glucose
Tolerance
Species
Humans
Hamsters
Rats
Mice
Rats
Mice
Humans
Rats
Humans
Humans
Humans
Experimental
Dose
100 /ig/kg
6 doses
(600 ^g/kg
total dose)
100 ng/kg/d
for 2 years
400 ng/kg/d
for 2 years
125 ng/kg/d
30 weeks
7.5 ng/kg/wk
for 20 wks
dermal
exposure
12,500 ng/kg
sac day 7
Body Burden
109-7,000
ng/kg ;
500 ng/kg
1,400 ng/kg
1,000 ng/kg
1,600 ng/kg
1,100 ng/kg
83 ng/kg
10,200 ng/kg
14 ng/kg
110 ng/kg
14 ng/kg
Ref./Note
Fingerhut et al.,
1991; Bertazzi et al.,
1993/a
Raoetal., 1988/b
Kocibaetal., 1978/c
NTP, 1982/d
Maronpot et al.,
1993/e
Poland etal., 1982/f
Egeland etal.,
1994/g
Moore etal., 1985/h
Air Force Study,
1991/i
Sweeney etal.,
1992/j
Wolfe etal., 1992/i
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Table 9-4. (continued)
Effect
Decreased
Glucose Uptake
Adipocytes
Decreased
Serum Glucose
Background
Background
Species
Guinea
Pigs
Rats
Human
Mouse
Experimental
Dose
30 ng/kg
sac day 1
100 ng/kg/d
30 days
60 TEQ ppt
in serum
Body Burden
30 ng/kg
1,900 ng/kg
9 ng/kg
4 ng/kg
Ref./Note
Enanetal., 1992/k
Zinkletal., 1973/1
m
n
Notes:
a. Estimated highest body burden at time of last exposure. Calculations based on measured
TCDD levels in serum (lipid adjusted) and assuming a first-order elimination kinetics
and a half-life for elimination of 7.1 years. Also assumes a body weight of 70 kg and
22% body fat. Calculations for estimated serum concentrations at last time of exposure
performed by authors (Fingerhut et aL, 1991; Bertazzi et al,, 1993),
b. Animals administered 100 jug/kg six times over a 4-week period. Assumes a half-life of
23.4 days and that animals are sacrificed at 10 months after the first dose. This is the
LOEL; however, no other doses tested in this study.
c. Assumes a single first-order elimination rate constant and a half-life for the whole body
elimination of 23.7 days (Rose et a!,, 1976) and a gastrointestinal tract absorption of
86% (Rose et al., 1976). This is the LOEL of the study; a dose of 10 ng/kg/d was also
tested with no significant increase in tumors,
d. Body burden estimated from animals treated with 450 ng/kg/d for 90 days (DeVito and
Birnbaum, unpublished results).
e. From Tritscher et al. (1992) and Maronpot et al. (1993). Liver levels measured in study
at approximately 30 ppb. Liver and body weights were reported in White and Gasiewicz
(1993). Assumes animal is 20% body fat by weight and that at this dose, the liver has
four times the concentration of TCDD than adipose tissue. The body-burden calculation
assumes that liver and fat account for 90% of the body burden in these animals. For
tumor promotion, this is the LOEL in these animals. A NOEL for tumor promotion was
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Table 9-4. (continued)
observed at a dose of 35 ng/kg/d. For induction of CYP1A1 and downregulation of
EGF-R, this body burden produces a maximal response.
f. Assumes an elimination rate of 11 days and a body weight of 20 grams.
g. From Egeland et al. (1994) in which workers with half-life extrapolated levels of TCDD
of 496-1,860 ppt have a greater percentage of workers with low testosterone levels.
Extrapolation performed by Egeland et al. (1994) assuming a half-life of 7.1 years.
Also assumed that the background TEQ was 60 ppt so that the total serum TEQ was 496
ppt + 60 ppt = 556 ppt (lipid adjusted). Average worker was male weighing 70 kg
with 15% body fat.
h. Animals received single exposure of 12.5 /*g/kg (LOAEL) and sacrificed 7 days after
dosing. Assumes a half-life of 23.4 days and body burden corrected for elimination. A
dose of 6.25 pig/kg was tested and is the NOEL for this study.
i. From Ranch Hand study (Sweeney et al., 1992), assumes that high exposed group (> 33
ppt) had a background of 60 TEQ ppt. Thus, this group had at least 93 TEQ ppt.
Assumes average Ranch Hand patient was male weighing 70 kg with 15% body fat.
j. Same assumptions in note g except average serum level in affected workers is 640 ppt.
k. Guinea pigs received 0.03 fig TCDD/kg i.p. and sacrificed 24 hours after dose.
Assumes that no TCDD was eliminated at this time. This is a LOEL; no other doses
tested.
1. Animals were treated with 0.1 /xg/kg/day for 30 days and assumes half-life of TCDD in
the rat is 23.4 days.
m. Assumes a background TEQ of 60 ppt for dioxins, dibenzofurans, and PCBs. Also
assumes a body weight of 70 kg with 15% body fat.
n. Data from DeVito and Birnbaum (1994). TEQ for TCDD, 1,2,3,7,8-PCDD- 2378-
TCDF; 1,2,3,7,8-PCDF; 2,3,4,7,8-PCDF; and OCDF in 150-day-old female B6C3F1
mice. Chemicals were determined in liver, fat, and skin of these animals. Assumes that
100% of the body burden is hi liver, fat, and skin.
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Table 9-5. Estimated Body Burdens of Experimental Animals and Humans Exposed to
Dioxins: Low-Dose Effects in Animals Exposed to Dioxins and Their Relationship to
Background Human Exposure
Effect
Decreased
Offspring
Viability
Altered
Lymphocyte
Subsets
Altered
Lymphocyte
Subsets
Enhanced Viral
Susceptibility
Endometriosis
Decreased Sperm
Count
Background
Background
Species
Rhesus
Monkeys
Rhesus
Monkeys
Marmosets
Mice
Monkeys
Rats
Human
Mouse
Experimental
Dose
25 ppt in diet
for 4 years
25 ppt in diet
for 4 years
0.3 ng/kg/wk
for 24 weeks
1.5 ng/kg/wk
for 12 weeks
10 ng/kg
sac day 7
5 ppt in diet
4 years
64 ng/kg
maternal dose
gd!5
60 TEQ ppt
in serum
Body Burden
270 ng/kg
270 ng/kg
6-8 ng/kg
7 ng/kg
54 ng/kg
64 ng/kg
9 ng/kg
4 ng/kg
Ref./Note
Hong et al. ,
1989/a
Hong et al.,
1989/a
Neubert et al. ,
1992/b
Burelson et al. ,
1994/c
Reier et al.,
1993/a
Mably et al.,
1992b/d
e
f
Notes:
a. Assumes a single first-order elimination rate constant and a half-life for the whole body
elimination of 400 days (McNulty, 1985) and a gastrointestinal absorption of 86% (Rose
et al., 1976). This is the LOEL from this study; no lower doses tested.
b. Assuming a single first-order elimination rate constant and a half-life of 6-8 wks. Body
burdens calculated by authors (Neubert et al., 1992).
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Table 9-5. (continued)
c. Body burden determined in these animals (Diliberto et al., submitted). Approximately
70% of the body burden remains at 7 days after dosing. This is the LOEL from this
study. A dose of 5 ng/kg was also tested in this study with no effect (NOEL).
d. Assumes pups exposed to an equal dose of TCDD as are the dams on a weight basis and
that the pups do not eliminate any of the TCDD. For decreased! body weight in pups
400 ng/kg is the LOEL, a dose of 64 ng/kg to the dam was the NOEL for this response.
For decreased sperm count, the LOEL is 64 ng/kg, and no lower doses were tested.
e. Assumes a background TEQ of 60 ppt for dioxins, dibenzofurans, and PCBs. Also
assumes a body weight of 70 kg with 15% body fat.
f. Data from DeVito and Birnbaum (1994). TEQ for TCDD, 1,2,3,7,8-PCDD; 2378-
TCDF; 1,2,3,7,8-PCDF; 2,3,4,7,8-PCDF; and OCDF in 150-day-old female B6C3F1
mice. Chemicals were determined in liver, fat, and skin of these animals. Assumes that
100% of the body burden is in liver, fat, and skin.
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Table 9-6. Comparison of the Effects of TCDD Exposure on Human and Animal Tissue
In Vitro
Effect
Binding to CYP1A1
DRE
Binding to CYP1A1
DRE
Binding to ER DRE
Binding to ER DRE
Induction CYP1A1
Induction CYP1A1
Cytotoxicity
Cytotoxicity
Cytotoxicity
Altered Lymphocyte
Subsets
Altered Lymphocyte
Subsets
Inhibition of
Proliferation
Inhibition of
Proliferation
Inhibition of
Proliferation
Inhibition of
Proliferation
Cell Line/Species
Hepa-lclc7/mouse
LS180/human
Hepa IclcT/mouse
MCF-7/human
Lymphocytes mouse
Lymphocytes human
Embryonic palate
mouse
Embryonic palate rat
Embryonic palate
human
Peripheral
lymphocytes
marmoset
Peripheral
lymphocytes human
Thymocytes mouse
Thymocytes human
Tonsilar lymphocytes
human
Splenic lymphocytes
murine
Concentration
2nM
10 nM
15.5 nM
5.6 nM
1.3 nM
1.8nM
0.1 nM
100 nM
100 nM
0.0001 nM
0.0001 nM
0.1 nM
0.1 nM
0.3 nM
3.0 nM
Ref./Note
Probst etal., 1993/a
Probst etal., 1993/a
White and Gasiewicz,
1993/b
White and Gasiewicz,
1993/b
Clark etal., 1992/c
Clark etal., 1992/c
Abbott and
Birnbaum, 199 1/d
Abbott and
Birnbaurn, 1991/d
Abbott and
Birnbaum, 1991/d
Neubert etal.,
199 1/e
Neubert etal.,
1991/e
Greenlee et al. ,
1985/f
Cook etal., 1987/f
Wood etal., 1993/g
Wood etal., 1993/g
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Table 9-6. (continued)
Effect
Inhibition of IgM
Secretion
Inhibition of IgM
Secrection
Cell Line/Species *
Splenic lymphocytes
marine
Tonsilar lymphocytes
human
Concentration
3.0 nM
0.3 nM
Ref./Note
Woodetal., 1993/g
Woodetal., 1993/g
Notes:
a.
b.
c.
d.
e.
f.
Using gel retardation assay, Probst et al. (1993) compared the ability of the Ah receptor
isolated from either murine or human cell lines to bind to a dioxin response element
(DRE). The authors used only one concentration of TCDD for either cell type, 2 nM
for murine cells and 10 nM for human cells.
White and Gasiewicz (1993) compared the ability of Ah receptors isolated from either
murine or human cell lines to bind to a DRE found in the human estrogen receptor (ER)
structural gene. Concentration values are binding affinities to this DRE.
Splenic lymphocytes from C57B1/6 mice and peripheral blood lymphocytes were
isolated, cultured, and exposed to TCDD. EROD activity, a marker for CYP1A1, was
determined following TCDD exposure. Data presented are EC50.
Abbott and Birnbaum (1991) compared the cytotoxic effects of TCDD on organ culture
of human, mouse, and rat embryonic palatal shelves. Embryonic palates from human
mouse and rat were grown in the same organ culture system and exposed to TCDD.
Cytotoxicity was detected using transmission electron microscopy. Concentrations are
the lowest observable effect level.
Neubert et al. (1991) isolated lymphocytes from human and primates and determined
lymphocyte subsets following antigen stimulation in cells treated with TCDD. The
concentration is the level at which the authors describe a consistent effect on lymphocyte
subsets in this system. ;
Thymocytes were isolated from either human or murine sources ajid cocultured with a
human thymic epithelium culture (human thymocytes) or with murine thymic epithelium
(murine thymocytes). The incorporation of tritiated thymidine was determined in cells
treated with TCDD following antigen stimulation. Data presented are LOEL for both
cell lines.
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Table 9-6. (continued)
g. Human tonsilar lymphocytes and murine splenic lymphocytes were used to isolate B
cells. Human and murine B cells were grown under identical conditions and exposed to
TCDD. Proliferation and IgM secretion were determined in response to different
concentrations of TCDD ranging from 0.3 to 30 nM. Values presented are LOELs from
Wood et al. (1993).
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have a potentiating effect with dioxin for lung cancer in the human studies), the animal
model will be the focus of estimates of cancer risk.
i
Many scientists agree that the cancer modeling efforts should continue to focus on the
animal studies in the absence of better quantitative human data. Others suggest that there is
no compelling reason to conclude that estimates derived from the human data are any more
uncertain than the estimates based on the rodent bioassay. In both cases, modeling efforts
have indicated the sensitivity of certain model parameters to choice of data sets and/or
assumptions. Analyses in Chapter 8 illustrate that the slope of the dose-response curve for
surrogate markers of low-dose response such as enzyme induction or indirect mutagenic
activity on estimates of cancer risk using animal data are highly dependent on the
assumptions that go into the modeling. Dependent on assumptions, use of the obvious dose
surrogates could predict very different low-dose risks, differing by orders of magnitude from
the estimates described above. For gene expression of biological markers, the major factor
controlling this broad range of low-dose risk estimates is the mechanism by which dioxin
modifies constitutive expression. However, as expressed in Appendix D of Chapter 8,
reasonable assumptions concerning constitutive expression of the biochemical markers will
result in low-dose linearity and risk estimates consistent with that obtained using the
linearized multistage approach.
The two-stage modeling of the Kociba et al. (1978) female rat liver tumor data in
Chapter 8 incorporates data from earlier events in the carcinogenic process into the
estimation of model parameters. In fact, the results using the two-stage model incorporating
dioxin-altered hepatic foci data to estimate mutation and growth parameters provide nearly
the same low-dose estimates as the linearized multistage model using only the tumor data.
When using the default species extrapolation from animals to humans (body weight ratio to
the 3/4 power), both models yield oral intake risk-specific doses of slightly less than 0.01 pg
TCDD/kg/day, corresponding to unit risk estimates of 1X W4 per pg TCDD/kg/day.
Chapter 8 discusses other potential models that might fit these data as well as the best-fitting
model (Appendix C, Chapter 8). These analyses indicate that, unless a protective effect of
TCDD on mutation rates occurs at low doses, low-dose risk will remain proportionate to
exposure and consistent with the linearized multistage model. If protective effects are
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allowed in the model, the low-dose risks may be substantially reduced. The focal lesion data
and the biochemical markers generally agree and do not suggest the protective effect
discussed above. These models assume that the POST foci are precursors to cancer. Other
hepatic focal lesion markers could be used in this context and may lead to different dose-
response curves for tumor response (see discussion in Chapter 8).
Uncertainty in estimates of human half-life for dioxin arid related compounds
represents an important factor in comparing human-based risk estimates versus animal-based
risk estimates. For instance, if the dose-dependent pharmacokinetic model of Carrier (1991)
is correct, exposures in the occupational studies must have been greater than the fixed half-
life model would suggest, so that the estimated risk per unit of exposure may well have been
lower. However, this reduction will be relatively small and is unlikely to move the risks
outside the range of risk estimated by the linearized multistage model.
An additional consideration regarding the evaluation of dose response for dioxin and
related compounds involves the ubiquity of background exposure to these compounds. Body
burdens of these compounds have been discussed previously in several parts of this
assessment. In all studies, both in laboratory animals and in humans, incremental exposures
are being added onto an existing body burden that is present at birth and appears to increase
with age. This background is often insignificant from the standpoint of added dose in
experimental studies or for highly exposed human cohorts. On the other hand, it has real
implications relative to the detectability of response at low incremental exposures and may
have implications for the use of models that assume additivity to ongoing processes that may
have been stimulated by background levels. Modeling estimates suggest that, if dioxin and
related compounds are adding to human cancer burden, current background exposures may
result in upper bound population cancer risk estimates attributable to exposure to dioxin and
related compounds in the range of 1 in 10,000 (10'4) for the average population exposures to
1 in 1,000 (10'3) for more highly exposed members of the population (e.g., individuals
consuming high levels of dioxin-containing foods). Actual risk for individuals exposed to
background levels in the population is likely to be less than these upper bound estimates and,
for some, may even be zero. More highly exposed populations with exposures to specific
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sources of dioxin and related compounds such as those exposed under occupational or
accident conditions may, on average, experience proportionately higher risk.
Background levels also complicate the evaluation of no observed or low observed
adverse effect levels (NOAELs or LOAELs). Incremental exposures must be considered in
light of existing body burdens in determining whether increased probability of effects having
biological thresholds is likely. The concept that an incremental exposure is below an
experimental threshold is moot unless the combined background and incremental exposure
are below the threshold level. This has important consequences for the assessment of
compounds like dioxin where certain biochemical alterations can be detected at or near
equivalent human background body burden levels.
9.9. USE OF TOXICITY EQUIVALENCE
The concept of toxicity equivalence in evaluating mixtures of dioxin-like compounds
is fundamental to many of the conclusions reached in this characteri2ation. This is based on
the fact that most data described in this and preceding chapters were obtained using 2,3,7,8-
TCDD as the experimental compound. More limited data exist as individual congeners are
evaluated. Nonetheless, estimates of body burden as derived in this reassessment suggest
that greater than 90% of the total dioxin equivalence is due to dioxin-like compounds other
than 2,3,7,8-TCDD. While there are empirical bases for the toxicity equivalence factors
assigned to dioxin-like compounds relative to 2,3,7,8-TCDD, they generally represent order
of magnitude estimates of relative toxicity and are not meant to be used precisely. The
potency for most, if not all, of the toxic end points is determined by the number and position
of the halogen (chlorine or bromine) atoms on the dioxin-like molecule. This appears, based
on a substantial body of evidence, to be a function of relative ability to bind to a specific
cellular receptor that mediates most, if not all, of the toxic end points of this class of
compounds. This inference is based on experimental evidence, primarily in rodents but
involving some other species, that for some toxic effects, the potency of the effect itself is
proportional to receptor binding as measured by either binding studies or a sensitive measure
of receptor binding, AHH induction. When ED50s for effects versus binding are plotted
logarithmically, good linear correlations are obtained (Safe, 1990). This approach constitutes
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a "relative ranking" scheme based on 2,3,7,8-TCDD. Because the data base for effects for
individual congeners is incomplete and because the concept is based on responsiveness of
humans to these compounds in a manner similar to that of animals and high to low dose
extrapolation, the TEQ approach is considered a useful but uncertain procedure.
In addition to the idea of "relative ranking," there is a second aspect to the TEQ
approach. This is the concept of additivity. The hypothesis is that one can estimate the
toxicity of a mixture of dioxin-Uke compounds by adding together the products of the
concentrations of the individual congeners and their TEFs. This hypothesis has not been
extensively tested although data addressing this issue are generally supportive of additivity.
Some data collected using high levels of different congeners have suggested the potential for
interactions (mostly, antagonism) between congeners. There is currently general acceptance
of the concept of additivity with the recognition that issues such as congener interactions,
presence of "spare" receptors, and the unavoidable presence of other dietary constituents that
react with the dioxin receptor must be considered to add uncertainty to the concept.
The points discussed above describe the basis of the TEQ concept and indicate some
of the assumptions on which they are based. A more detailed description of these issues is
contained in U.S. EPA (1989). In addition to scientific grounds, the use of TEQs can be
justified on a practical basis, not the least of which is the sheer enormity of the task of
attempting to conduct appropriate studies on all toxic end points for all of the congeners.
They continue to be described by the EPA and others as an "interim" approach, and the
extent of their current use should not detract from the expressed need for more data to
further validate their use.
9.10. KEY ASSUMPTIONS AND INFERENCES
One of the primary functions of the risk characterization is to present key assumptions
and inferences that are used to reach conclusions in the absence of definitive information.
Not all scientists may agree with the use of these specific assumptions and inferences. The
degree to which there is disagreement will have profound effects on the acceptance of this
analysis. While many of these assumptions and inferences are discussed in previous sections,
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it is important that they be recognized in order to put our overall conclusions in a proper
perspective. Key assumptions and inferences are listed below.
The limited information on sources, fate, and transport in the environment provides a
reasonable basis for predicting human exposure. While data are limited and, therefore,
uncertain, information from a variety of studies in industrialized countries coupled with our
detailed knowledge of physicochemical properties for this class of compounds allows
reasonable assumptions to be made regarding relative ranking of sources with regard to their
contribution to environmental loading, the persistence of this class of compounds under
specific environmental conditions, and the likelihood that the chemical will be transferred
from the environment to biological systems. Nonetheless, these are assumptions that are
arguable and that will be refined as more data become available. Additional data will be
required to validate the numerous hypotheses that go into assembling models for
environmental release, fate, and transport for this complex mixture of individual chemical
congeners.
The air to food hypothesis is plausible and is supported by enough data to warrant its
use in the absence of more complete information. The air-to-food fyrpothesis is founded on
data evaluating deposition, environmental transport, bioaccumulation, and consumption
patterns. It is supported by studies from Europe and Canada. While; individual measurement
data are still quite limited, the consistency of the evidence supporting the validity of the
hypothesis is compelling. The hypothesis has been accepted by a larjge segment of the
knowledgeable scientific community. Because airborne dioxin may come from direct
releases to air or from recycling of dioxin-like compounds released into various
environmental media from a number of sources, this hypothesis provides a perspective on
how dioxin-like compounds move through the environment to humans but does not allow
attribution of exposure to particular sources.
Toxicity equivalence is a valid, interim method for assessing exposure to a complex
mixture of dioxin and related compounds and predicting likely health outcomes. The EPA
and the international scientific community have agreed that the use of toxicity factors to
predict relative toxicities of mixtures of this class of compounds has aji empirical basis, is
theoretically sound, and, in the absence of more complete data sets on the toxicity of
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individual members of this class, is a useful procedure. This is not to say that the use of
TEFs is a certain procedure. Since 1986 when the first Agency-wide consensus on the use of
TEFs was published, additional refinements to the data bases and to the use of TEFs have
occurred. Published revisions in accord with international agreement appeared in 1989. In
the course of this reassessment, critical data were collected, and agreement was reached
regarding the contribution of dioxin-like PCBs to overall TEQs. Additional validation of the
TEQ concept in predicting effects of this class of compounds on wildlife species lends further
support to the use of this approach. It must be recognized that this relatively simple,
additive approach does not take into account interactions between dioxin-like compounds and
other chemical exposures. These interactions may result in either an. overestimate or an
underestimate of likely effects of the complex mixture. While generally accepted as useful
for evaluating intakes of various dioxin-like compounds, the application of this approach to
the evaluation of measured body burdens remains even more uncertain,
Use of one-half the nondetect level for estimating low levels of exposure is a
reasonable but conservative approach to evaluating limited blood and. tissue level data. For
some data sets, use of zero values for nondetects could result in significantly lower estimates
and, therefore, use of the current procedure may be overestimating blood or tissue levels.
However, it is widely held that use of zero values for nondetects would most likely
underestimate true levels of exposure, particularly where nondetects do not dominate
measured values. Similar estimates of TEQs derived from different data sets, developed by
different investigators in several countries, strengthen the probability that this inference
represents the exposure of the general population in industrialized countries to dioxin and
related compounds.
The limited data available from studies of levels of dioxin and related compounds in
humans provide an adequate basis to infer general population body burdens. Although there
are still limited measurements of general population body burdens, the data provide a
consistent picture of background body burdens for industrialized countries. While additional
data will help refine the range of general population body burdens as a function of location,
human activity, age, and the like, there are adequate data to estimate current body burdens in
the general population for the purposes of this assessment. It is highly unlikely that these
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estimates would represent a sensitive parameter in estimating margins of exposure within an
order of magnitude.
Laboratory animal studies provide useful information in evaluating potential human
responses to dioxin and related compounds. Based on our knowledge of the biochemical and
biological similarities between laboratory animals and humans, our understanding of some of
the fundamental impacts of this class of compounds on biological systems, and comparable
responses from animal and human studies both in vitro and in vivo, our decision to use
laboratory animal data to contribute to weight-of-the-evidence conclusions on human hazard
and risk is reasonable. Humans do not appear to be an unusual responder for dioxin effects,
that is, we do not, on average, appear to be either refractory to or exquisitely sensitive to the
effects of dioxin-like compounds. While positive human data are preferable for ascribing
hazard or risk, the lack of adequate human data to demonstrate causiility for many suspected
dioxin effects is assumed not to negate the findings from laboratory ianiinal and in vitro
studies. Although some scientists may disagree, in our estimation, the data base on dioxin
and related compounds is one of the most comprehensive among all environmental chemicals.
The fundamental understanding of mechanisms of dioxin action provides a unifying theory
for the mechanisms for observed effects in laboratory animals and humans and for using a
weight-of-the-evidence approach considering all relevant data to infer the human health
impacts of dioxin and related compounds.
Observations of effects from exposure to dioxin and related compounds in humans and
other animals suggest that fundamental changes in cellular biochemistry and biology may be
related to frankly adverse effects, which can be more readily observed at higher levels of
exposure. Observations described in this assessment suggest a continuum of response to
exposure to dioxin-like chemicals. By a continuum of response we suggest that as dose
increases, the probability of occurrence of individual effects increases and the severity of
collective effects increases. This continuum provides a basis for inferring a relationship
between some early events that are not necessarily considered to be adverse effects with later
events that are adverse effects. Considerable uncertainty remains in inferring how these
events are related, although we know more about how dioxin-like compounds may elicit
effects than we know about the mechanisms of action for most chemicals. This inference
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may be the most contentious of all, and it is likely that a wide range of opinion will be
provided by the scientific community regarding the relationship of these mechanistic
observations and prediction of potential for adverse effects in exposed humans. This range
of opinion must be carefully weighed to assure that the proper perspective concerning the
relative likelihood of adverse effects in humans exposed to environmental levels is
maintained.
9.11. OVERALL CONCLUSIONS REGARDING THE IMPACT OF DIOXIN AND
RELATED COMPOUNDS ON HUMAN HEALTH
An extensive data base provides information pertinent to the evaluation of
exposure of humans to dioxin and related compounds. An even larger data base of
equal quality suggests that exposure to dioxin results in a broad spectrum of biochemical
and biological effects in animals and, based on limited data, some of these effects occur
in humans. Relatively speaking, these exposures and effects are observable at very low
levels in the laboratory and in the environment when compared with other
environmental toxicants. Despite the large amount of information available on exposure
and effects of dioxin and related compounds, this risk characterization serves to
highlight significant data gaps and identifies information needed to reduce uncertainty in
its conclusions.
An extensive data base detailing dioxin emissions and dioxin levels in environmental
media and in human serum and tissue indicates widespread, low-level human exposure.
Much of the public concern for this potential exposure revolves around the characterization
of these compounds as among the most toxic "man-made" chemicals ever studied. These
compounds, which are generally unwanted by-products of chemical reactions, are extremely
potent in producing a variety of effects in experimental animals based on traditional
toxicology studies at levels hundreds or thousands of times lower than most synthetic
chemicals of environmental interest. In addition, human studies demonstrate that exposure to
dioxin and related compounds is associated with subtle biochemical and biological changes
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whose clinical significance is as yet unknown and, at higher levels, with chloracne, a serious
skin condition. Laboratory studies suggest that exposure to dioxin-like compounds may be
associated with other serious health effects, including cancer. Human data, while limited in
their ability to answer questions of hazard and risk, are consistent with some observations in
animals. The ability to determine the expression in humans of adverse effects noted in
laboratory studies or to detect these effects in human population studies is dependent on the
dose absorbed and the intrinsic sensitivity of humans to these compounds. The large data
base on exposure coupled with toxicity data from animal experiments, as well as more
limited human information, forms the basis for the risk characterization of dioxin and related
compounds.
A large variety of sources of dioxin have been identified and others may exist.
Because dioxin-like chemicals are persistent and accumulate in biological tissues,
particularly in animals, the major route of human exposure is through ingestion of foods
containing minute quantities of dioxin-like compounds. This results in widespread, low-
level exposure of the general population to dioxin-like compounds. Certain segments of
the population may be exposed to additional increments of exposure by being in
proximity to point sources or because of dietary practices.
Dioxin-like compounds are released to the environment in a variety of ways and in
varying quantities, depending on the source. Despite a growing body of literature from
laboratory, field, and monitoring studies examining the environmental fate and environmental
distribution of CDDs, CDFs, and PCBs, the fate of these environmentally ubiquitous
compounds is not yet fully understood. The available information suggests that the presence
of dioxin-like compounds in the environment has occurred primarily as a result of industrial
practices and is likely to reflect changes in release over time. Further work to confirm
declining concentrations in environmental samples and to relate these data to human
exposures will be required.
The principal identified sources of environmental release of CDDs and CDFs may be
grouped into four major types: combustion and incineration sources; chemical
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manufacturing/processing sources; industrial/municipal processes; and reservoir sources.
PCBs were produced in relatively large quantities for use in such commercial products as
dielectrics, hydraulic fluids, plastics, and paints. They are no longer produced in the United
States but continue to be released to the environment through the use and disposal of these
products. A similar situation exists for the commercially produced PBBs that are produced
for a number of uses such as flame retardants. Additional measurement data will be needed
to gain an adequate appreciation for the nature and magnitude of major U.S. sources and
releases of CDDs, CDFs, and polyhalogenated biphenyls.
CDDs, CDFs, and PCBs have been found throughout the world in practically all
media, including air, soil, water, sediment, fish and shellfish, and agricultural food products
such as meat and dairy products. The highest levels of these compounds are found in soils,
sediments, and biota; very low levels are found in water and air. The widespread occurrence
observed, particularly in industrialized countries, is not unexpected, considering the
numerous sources that emit these compounds into the environment and the overall resistance
of these compounds to biotic and abiotic transformation. The levels of dioxin and related
compounds in environmental media and in food in North America are based on few samples
and must be considered quite uncertain. However, they seem reasonably consistent with
levels measured in a number of studies in Western Europe and Canada. The consistency of
these levels across industrialized countries provides reassurance that the U.S. estimates are
reasonable. Collection of additional data to reduce uncertainty in U.S. estimates of dioxin-
like compounds in the environment and in food represents an important data need.
This assessment adopts the hypothesis that the primary mechanism by which dioxin-
like compounds enter the terrestrial food chain is via atmospheric deposition. Dioxin and
related compounds enter the atmosphere directly through air emissions or indirectly, for
example, through volatilization from land or water or from resuspension of particles.
Deposition can occur directly onto soil or onto plant surfaces. At present, it is unclear
whether atmospheric deposition represents primarily current contributions of dioxin and
related compounds from all media reaching the atmosphere or whether it is past emissions of
dioxin and related compounds that persist and recycle in the environment. Understanding the
relationship between these two scenarios will be particularly important in understanding the
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relative contributions of individual point sources of these compounds to the food chain and in
assessing the effectiveness of control strategies focused on either current or past emissions of
dioxins in attempting to reduce the levels in food.
Throughout this document, concentrations of dioxin and related compounds have been
presented as TCDD equivalents (TEQs). Total TEQs are the sum of the products of
concentrations of individual dioxin-like compounds in a complex environmental mixture times
the corresponding TCDD toxicity equivalence factor (TEF) for that compound [Total
TEQs = £ Ccongener x TEFcongener]. The strengths and weaknesses as well as the
uncertainties associated with the TEF/TEQ approach have been discussed in this chapter. As
noted, the use of the TEQ approach is fundamental to the evaluation of this group of
compounds and, as such, represents a key assumption on which many of the conclusions in
this characterization hinge.
The term "background" exposure has been used throughout this reassessment to
describe exposure of the general population that is not exposed to readily identifiable point
sources of dioxin-like compounds. Data on human tissue levels suggest that body burden
levels among industrialized nations are reasonably similar (Schecter, 1991). These data can
also be used to estimate background exposure through the use of pharmacokmetic models.
Using this approach, exposure levels to 2,3,7,8-TCDD in industrialized nations are estimated
to be about 0.3-0.6 pg TCDD/kg body weight/day1. This is generally consistent with the
estimates derived using diet-based approaches to estimate total TCDD intake.
Pharmacokinetic approaches have not been applied to estimate exposures to CDDs or CDFs
other than TCDD, which contribute substantially to the body burden of dioxin-like
compounds. Estimates of exposure to dioxin-like CDDs and CDFs based on dietary intake
are in the range of 1-3 pg TEQ/kg body weight/day. Estimates basisd on the contribution of
dioxin-like PCBs to toxicity equivalents raise the total to 3-6 pg TEQ/kg body weight/day.
This range is used throughout this characterization as an estimate of average background
exposure to dioxin-like CDDs, CDFs, and PCBs. This average background exposure leads
to body burdens in the human population that average 40-60 pg TEQ/g lipid (40-60 ppt)
!Since 2,3,7,8-TCDD is the reference compound for the TEF/TEQ approach, 1.0 pg
TCDD = 1.0 pg TEQ.
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when all dioxins, fiirans, and PCBs are included. High-end estimates of body burden of
individuals in the general population (approximately the top 10% of the general population)
may be greater than three times higher.
In addition to general population exposure, some individuals or groups of individuals
may also be exposed to dioxin-like compounds from discrete sources or pathways locally
within their environment. Examples of these "special" exposures include occupational
exposures, direct or indirect exposure to local populations from discrete sources, exposure to
nursing infants from mother's milk, or exposures to subsistence or recreational fishers.
These exposures have been discussed previously in terms of increased exposure due to
dietary habits (see Exposure Document) or due to occupational conditions or industrial
accidents (see Chapter 7). Although exposures to these populations may be significantly
higher than to the general population, they usually represent a relatively small percentage of
the total population. Inclusion of their levels of exposure in the general population estimates
would have little impact on average population estimates. Simply evaluating these exposures
as average daily intakes prorated over a lifetime might obscure the potential significance of
elevated exposures for these subpopulations, particularly if exposures occur for a short period
of time during critical windows of biological sensitivity.
The scientific community has identified and described a series of common
biological steps that are necessary for most if not all of the observed effects of dioxin
and related compounds in vertebrates, including humans. Binding of dioxin-like
compounds to a cellular protein called the "Ah receptor" represents the first step in a
series of events attributable to exposure to dioxin-like compounds, including
biochemical, cellular, and tissue-level changes in normal biological processes. Binding to
the Ah receptor appears to be necessary for all well-studied effects of dioxin but is not
sufficient, in and of itself, to elicit these responses. This reassessment concludes that the
effects elicited by exposure to 2,3,7,8-TCDD are shared by other chemicals that have a
similar structure and Ah receptor-binding characteristics. Consequently, the biological
system responds to the cumulative exposure of Ah receptor-mediated chemicals rather
than to the exposure to any single dioxin-like compound.
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Based on our understanding of dioxin mechanism(s) to date, it is accurate to say that
interaction with the Ah receptor is necessary, that humans are likely to be sensitive to many
of the effects of dioxin demonstrable in laboratory animals, and that there is likely to be a
variation between and within species and between tissues in individual species based on
differential responses "downstream" from receptor binding. Further analyses of dioxin action
may provide more insight into the mechanisms by which TCDD and related compounds
produce effects that are of particular public health concern. A major challenge for the future
will be the establishment of experimental systems in which complex biological phenomena
associated with these effects are amenable to study at the molecular level.
The concept of toxicity equivalence based on a unifying mechanism of action within
this class of compounds and the use of toxicity equivalence factors as described in this
document and elsewhere have been extensively reviewed and are widely used. While some
uncertainty remains,with regard to the.additivity of complex mixtures of these compounds
and with the impacts of co-exposure to nondioxin-like compounds, the use of this approach is
consistent with the Agency's guidance on the evaluation of complex mixtures in the absence
of data on the impact of the actual mixture. This approach to the evaluation of dioxin and
related compounds, while considered an interim procedure to be usexi in the absence of more
specific data, is an integral part of this reassessment. Additional validation studies to reduce
uncertainty in the use of TEFs/TEQs will be very important.
There is adequate evidence based on all available information, including studies
in human populations as well as in laboratory animals and from ancillary experimental
data, to support the inference that humans are likely to respondl with a broad spectrum
of effects from exposure to dioxin and related compounds, if exposures are high enough.
These effects will likely range from adaptive changes at or near background levels of
exposure to adverse effects with increasing severity as exposure increases above
background levels.
Enzyme induction, changes in hormone levels, and indicators of altered cellular
function represent examples of effects of unknown clinical significance and which may or
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may not be early indicators of toxic response. Induction of activating/metabolizing enzymes
at or near background levels, for instance, may be adaptive or may be considered adverse
since induction may lead to more rapid metabolism and elimination of potentially toxic
compounds, or may lead to increases in reactive intermediates and may potentiate toxic
effects. Demonstration of examples of both of these situations is available in the published
literature.
Clearly adverse effects, including perhaps cancer, may not be detectable until
exposures exceed background by one or two orders of magnitude. The mechanistic
relationships of biochemical and cellular changes seen at very low levels of exposure to
production of adverse effects detectable at higher levels remain uncertain and controversial.
Individual species vary in their sensitivity to any particular dioxin effect. However,
the evidence available to date indicates that humans most likely fall in the middle of the
range of sensitivity for individual effects among animals rather than at either extreme. In
other words, evaluation of the available data suggests that humans, in general, are neither
extremely sensitive nor insensitive to the individual effects of dioxin-like compounds.
Human data provide direct or indirect support for evaluation of likely effect levels for several
of the end points discussed in previous sections, although the influence of variability among
humans remains difficult to assess. Discussions in previous chapters have highlighted certain
prominent, biologically significant effects of TCDD and related compounds. These
biochemical, cellular, and organ-level end points have been shown to be affected by TCDD,
but specific data on these end points do not generally exist for other congeners. Despite this
lack of congener-specific data, there is reason to infer that these effects may occur for all
dioxin-like compounds, based on the concept of toxicity equivalence.
Some of the effects of dioxin and related compounds, such as enzyme induction,
changes in hormone levels, and indicators of altered cellular function, have been observed in
laboratory animals and humans at or near levels to which people in the general population
are exposed. Other effects are detectable only in highly exposed populations, and there may
or may not be a likelihood of response in individuals experiencing lower levels of exposure.
Evaluation of effects in this health assessment document is based on the concept that .lipid-
adjusted serum levels approximate the body burden of dioxin and related compounds and that
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there will be a dose-response relationship between effects and body burden. Adverse effects
associated with temporary increases in dioxin blood levels based on short-term high-level
exposures, such as those that might occur in an industrial accident or in infrequent contact
with highly contaminated environmental media, may be dependent on exposure coinciding
with a window of sensitivity of biological processes. It is reasonable to assume that
developing organisms may be particularly sensitive to adverse impacts from temporary
increases above average background exposure levels. Such exposures may also lead to
higher tissue levels over the long term because of the long half-life for elimination of dioxin
and related compounds.
In TCDD-exposed men, subtle changes in biochemistry and physiology, such as
enzyme induction, altered levels of circulating reproductive hormones, or reduced
glucose tolerance, have been detected in a limited number of available studies. These
findings, coupled with knowledge derived from animal experiments, suggest the potential
for adverse impacts on human metabolism and developmental amd/or reproductive
biology and, perhaps, other effects in the range of current human exposures. Given the
assumption that TEQ intake values represent a valid comparison with TCDD exposure,
some of these adverse impacts may be occurring at or within one order of magnitude of
average background TEQ intake or body-burden levels (equal to 3-6 to 60 pg TEQ/kg
body weight/day or 40-60 to 600 ppt in lipid). As body burdens increase within and
above this range, the probability and severity as well as the spectrum of human
noncancer effects most likely increase. It is not currently possible to state exactly how
or at what levels humans in the population will respond, but the margin of exposure
(MOE) between background levels and levels where effects are detectable in humans in
terms of TEQs is considerably smaller than previously estimated,
Average human daily intakes of TCDD are in the range of 0.3-0.6 pg TCDD/kg body
weight/day. Using the TEQ approach, average human daily intakes of dioxin and related
compounds, including the dioxin-like PCBs, are in the range of 3-6 pg TEQ/kg body
weight/day. This intake results in average body burdens estimated to be in the range of 30-
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60 pg TEQ/g lipid (30-60 ppt) or 5-10 ng TEQ/kg body weight. Subtle changes in
biochemistry and physiology described above and discussed in detail in previous chapters are
seen with TCDD exposures at or just several fold above these average TEQ levels. Since
exposures within the general population are thought to be log-normally distributed,
individuals at the high end of the general population range (with body burdens estimated to
be three, and perhaps as high as seven, times higher than the average) may be experiencing
some of these effects. These facts and assumptions lead to the inference that some more
highly exposed members of the general population or more highly exposed, special
populations may be at risk for a number of adverse effects, including developmental toxicity
based on the inherent sensitivity of the developing organism to changes in cellular
biochemistry and/or physiology, reduced reproductive capacity in males based on decreased
sperm counts, higher probability of experiencing endometriosis in women, reduced ability to
withstand an immunological challenge, and others. This inference that more highly exposed
members of the population may be at risk for various noncancer effects is supported by
observations in animals, by some human information from highly exposed cohorts, and by
scientific inference.
The deduction that humans are likely to respond with noncancer effects 'from exposure
to dioxin-like compounds is based on the fundamental level at which these compounds affect
cellular regulation and the broad range of species that have proven to respond with adverse
effects. Since, for example, developmental toxicity following exposure to TCDD-like
congeners occurs in fish, birds, and mammals, it is likely to occur at some level in humans.
It is not currently possible to state exactly how or at what levels people will respond with
adverse impacts on development or reproductive function, fortunately, there have been few
human cohorts identified with TCDD exposures in the high end of the exposure range, and
when these cohorts have been examined, few clinically significant effects were detected. The
lack of adequate human information and the focus of most currently available epidemiologic
studies on occupationally TCDD-exposed adult males make difficult the evaluation of the
inference that noncancer effects associated with exposure to dioxin-like compounds may be
occurring. It is important to note, however, that when exposures to very high levels of
dioxin-like compounds have been studied, such as in the Yusho and Yu-Cheng cohorts, a
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spectrum of adverse effects has been detected in men, women, and children. Some have
argued that to deduce that a spectrum of noncancer effects will occur in humans in the
absence of better human data overstates the science; most scientists involved in the
reassessment as authors and reviewers have indicated that such inference is reasonable given
the weight of the evidence from available data. As presented, this logical conclusion
represents a testable hypothesis that may be evaluated by further dalta collection.
The likelihood that noncancer effects may be occurring in the human population at
environmental exposure levels is often evaluated using a margin of exposure approach. A
MOE is calculated by dividing the human-equivalent animal lowest observed adverse effect
level or no observed adverse effect level with the human exposure level. MOEs in the range
of 100 to 1,000 are generally considered adequate to rule out the likelihood of significant
effects occurring in humans based on sensitive animal responses. The average levels of
intake of dioxin-like compounds in terms of TEQs in humans described above would be well
within a factor of 100 of levels representing lowest observed adverse effect levels in
laboratory animals exposed to TCDD or TCDD equivalents. For several of the effects noted
in animals, a MOE of less than a factor of 10, based on intake levels or body burdens, is
likely to exist.
The previous basis for MOE calculations was the observation that exposure in the
range of 1-10 ng TEQ/kg/day represented a no observed adverse effect level for a sensitive
noncancer end point in laboratory animals and, therefore, that an inlake of up to 10 pg
TEQ/kg/day might represent an adequate MOE for all other noncancer effects in humans.
Recent data suggest that "high-end" average exposures in the general population are likely to
approach this intake level and that several effects, both subtle and frank, can be demonstrated
to occur in animals at intake values significantly lower than 1-10 ng TEQ/kg/day. This
information, coupled with limited human data suggesting measurable; effects, which may or
may not be considered adverse, at or near average background intake levels, makes it highly
unlikely that a margin of exposure of 100 or more currently exists for these effects at
background intake levels, at least for some members of the human population. Whether the
current MOE is adequate to protect public health is beyond the purview of this document and
represents a risk management decision. The reassessment points to the need to continue to
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monitor trends in human intake and body burden for dioxin and related compounds. If levels
are declining, the relationship of background body burdens to observed effect levels in
animal and human studies will need to be reevaluated.
Another approach that has been used to evaluate the likelihood of noncancer effects of
environmental chemicals is the reference dose (RfD). The EPA has frequently defined a
reference dose for toxic chemicals to represent a scientific estimate of the dose below which
no appreciable risk of noncancer effects is likely to occur following chronic exposures. In
the case of dioxin and related compounds, calculation of an RfD based on human and animal
data and including standard uncertainty factors to account for species differences and
sensitive subpopulations would likely result in reference intake levels on the order of 10 to
100 times below the current estimates of daily intake in the general population. For most
compounds where RfDs are applied, the compounds are not persistent and background
exposures that are generally low are not taken into account. Dioxin and related compounds
present an excellent example of a case where background levels in the general population are
likely to have significance for evaluation of the relative impact of incremental exposures
associated with a specific source. Since RfDs refer to the total chronic dose level, the use of
the RfD in evaluating incremental exposures in the face of a background intake exceeding the
RfD would be inappropriate and make the calculation of an Rfd for dioxin-like compounds of
doubtful significance.
In addition to the concern for various noncancer health end points discussed above,
the potential immunotoxicity of dioxin and related compounds represents a special situation.
Impairment of the immune system can be considered an adverse outcome in its own right,
being responsible for induced pathologies. At the same time, immunotoxicity can function as
a modulator of the disease process. It has been clearly established that TCDD is
immunotoxic and that it can impair normal immune function in laboratory animals at very
low levels (see Table 9-5). Epidemiological studies provide conflicting evidence for the
immunotoxicity of these compounds in humans. Few changes in the immune system in
humans associated with dioxin body burdens have been detected when exposed adult males
have been studied. It is possible that humans may be less sensitive than certain animal
models to dioxin immunotoxicity, or that available studies have lacked the power or the
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specificity to evaluate the impact of immunotoxic responses to dioxin and related compounds
in humans. Despite the possibility that these compounds may be imtmunotoxic at some level
in humans, the impact of dioxin and related compounds on the immune system and
implications for characterizing risk are largely unknown at this time.
With regard to carcinogenicity, a weight-of-the-evidence evaluation suggests that
dioxin and related compounds (CDDs, CDFs, and dioxin-like PCBs) are likely to present
a cancer hazard to humans. While major uncertainties remain, efforts of this
reassessment to bring more data into the evaluation of cancer potency have resulted in a
risk-specific dose estimate (1 x 10'6 risk or one additional cancer in one million exposed)
of approximately 0.01 pg TEQ/kg body weight/day. This risk-specific dose estimate
represents a plausible upper bound on risk based on the evaluation of animal and
human data. "True" risks are not likely to exceed this value, may be less, and may
even be zero for some members of the population.
Based on bioavailability and uptake studies, a cancer hazard is likely by oral,
inhalation, and dermal routes of exposure. As daily doses through these routes and
subsequent body burdens approach those seen in occupational studies, the uncertainty of the
hazard characterization is reduced. The epidemiological data alone are not yet deemed
sufficient to characterize the cancer hazard of this class of compounds as being "known."
However, combining suggestive evidence of recent epidemiology studies with the unequivocal
evidence in animal studies and inferences drawn from mechanistic data supports the
characterization of dioxin and related compounds as likely cancer hazards, that is, likely to
produce cancer in some humans under some conditions. It is important to distinguish this
statement of cancer hazard from the evaluation of cancer risk. The extent of cancer risk will
depend on such parameters as route and level of exposure, overall body burden, dose to
target tissues, individual sensitivity, and hormonal status.
The current evidence suggests that both receptor binding and most early biochemical
events such as induction of CYP1A1 and CYP1A2, as described in Chapter 8, are likely to
demonstrate low-dose linearity. The mechanistic relationship of these early events to the
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complex process of carcinogenesis remains to be established. If these findings imply low-
dose linearity in biologically based cancer models under development, then the probability of
cancer risk will be linearly related to exposure to TCDD at low doses. Until the mechanistic
relationship between early cellular responses and the parameters in biologically based cancer
models is better understood, the shape of the dose-response curve for cancer in the low-dose
region can only be inferred with uncertainty. Associations between exposure to dioxin and
certain types of cancer have been noted in occupational cohorts with average body burdens of
TCDD approximately two orders of magnitude (100 times) higher than average TCDD body
burdens in the general population. The average body burden in these occupational cohorts is
within one to two orders of magnitude (10 to 100 times) of average background body
burdens in the general population in terms of TEQ. Thus, there is no need for large-scale
low-dose extrapolations. Nonetheless, the relationship of apparent increases in cancer
mortality in these populations to calculations of general population risk remains uncertain.
With regard to average intake, humans are currently exposed to background levels of
dioxin-like compounds on the order of 3-6 pg TEQ/kg body weight/day, including dioxin-like
PCBs. This is more than 500-fold higher than the EPA's 1985 risk-specific dose associated
with a plausible upper-bound, one in a million (1 x 10~6) risk of 0.006 pg TEQ/kg body
weight/day and several hundredfold higher than revised risk-specific dose estimates presented
in Chapter 8 of this reassessment. Plausible upper-bound risk estimates for general
population exposures to dioxin and related compounds, therefore, may be as high as 10"4 to
10"3 (one in ten thousand to one in a thousand).
The fact that dioxin-like compounds are ubiquitous in the environment may have
further implications for low-dose risk assessment. Special populations may receive
identifiable, incremental exposures, based on proximity to specific sources or specific human
activity patterns such as consumption of higher amounts of foods containing average or
higher levels of dioxin-like compounds. The additive background model of Crump et al.
(1976) implies that the addition of an incremental dose to an existing background exposure
would support the use of a dose response model containing the assumption of linearity. This
assumption is particularly appropriate, in the absence of more definitive data on dose
response, if the exposure range (i.e., background exposure plus the added incremental
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DRAFT-DO NOT QUOTE OR CITE
exposure) is within one to two orders of magnitude (10 to 100 times) of the range of
observation of purported dioxin-induced tumors in highly exposed humans. In other words,
the proximity of background exposures to the range of observation of tumors in animals and
humans provides added support for the assumptions of additivity to background and linearity
of response.
TCDD has been clearly shown to increase malignant tumor incidence in laboratory
animals. In addition, a number of studies analyzed in Chapter 8 elucidate other biological
effects of dioxin related to the process of carcinogenesis. These studies have been used to
develop biologically based models of the pharmacokinetics of dioxin, of binding to the Ah
receptor, and of induction of various proteins that may be involved in the carcinogenic
process. In addition, bioassay data on TCDD reported by Kociba have been analyzed using
»
the two-stage clonal expansion model of carcinogenesis. There is evidence to suggest that
hormones and growth factors may be involved in TCDD carcinogenesis. The role of such
factors warrants additional study. Ideally, a biologically based model for cancer induction by
TCDD should explicitly consider hormonal influences. Initial attempts to construct a
biologically based model for certain dioxin effects as a part of this reassessment will need to
be continued and expanded to accommodate more of the available biology and to apply to a
broader range of potential health effects associated with exposure to dioxin-like compounds.
Based on all of the data reviewed in this reassessment and scientific inference, a
picture emerges of TCDD and related compounds as potent toxicants in animals with the
potential to produce a spectrum of effects. Some of these effects; may be occurring hi
humans at very low levels and some may be resulting in adverse impacts on human
health.
The potency and fundamental level at which these compounds act on biological
systems are analogous to several well-studied hormones. Dioxin and related compounds have
the ability to alter the pattern of growth and differentiation of a number of cellular targets by
initiating a series of biochemical and biological events resulting in the potential for a
spectrum of responses in animals and humans. Despite this potential, there is currently no
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clear indication of increased disease in the general population attributable to dioxin-like
compounds. The lack of a clear indication of disease in the general population should not be
considered strong evidence for no effect of exposure to dioxin-like compounds. Rather, lack
of a clear indication of disease may be a result of the inability of our current data and
scientific tools to directly detect effects at these levels of human exposure. Several factors
suggest a need to further evaluate the impact of these chemicals on humans at or near current
background levels. These are the weight of the evidence on exposure and effects, an
apparently low margin of exposure for noncancer effects, and potential for additivity to
background processes related to carcinogenicity.
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