United States Office of Research and EPA/600/BP-92/001c Environmental Protection Development August 1994 Agency Washington DC 20460 External Review Draft Sire Health Assessment Review tw, Document for 2,3,7,8- Draft pu. Tetrachlorodibenzo-p- (Do Not Dioxin (TCDD) and Cite or Related Compounds Quote) Volume III of III 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. WY ''''DRAFT EPA/600/BP-92/001c DO NOT QUOTE OR CITE August 1994 External Review Draft Health Assessment Document for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related Compounds Volume III of Ill 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of Research and Development U.S. Environmental Protection Agency Washington, D.C. && Printed on Recycled Paper ''DRAFT--DO NOT QUOTE OR CITE DISCLAIMER This document is an external draft for review purposes only and does not constitute Agency policy.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. HI-11 08/15/94 ''aA - YS DRAFT--DO NOT QUOTE OR CITE Health Assessment Document for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related Compounds TABLE OF CONTENTS - OVERVIEW Volume I DISPOSITION AND PHARMACOKINETICS MECHANISM(S) OF ACTION ACUTE, SUBCHRONIC, AND CHRONIC TOXICITY IMMUNOTOXICITY DEVELOPMENTAL AND REPRODUCTIVE TOXICITY CARCINOGENICITY OF TCDD IN ANIMALS Volume II EPIDEMIOLOGY/HUMAN DATA PART A. CANCER EFFECTS PART B. EFFECTS OTHER THAN CANCER DOSE-RESPONSE MODELING Volume III RISK CHARACTERIZATION OF 2,3,7,8-TETRACHLORODIBENZO-p-DIOXIN (TCDD) AND RELATED COMPOUNDS I-11 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE Health Assessment Document for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related Compounds CONTENTS - VOLUME III List Of ‘Tables. 0.0.0. cw ew Me Ot im Eda em ewe ee em III-vi List of Figures... 2... ee eee eee I-vi List of Abbreviations and Acronyms .......... 0.0.00 0c eee eee eee -vii PUGS pce A DEMO ERE we de III-xv Authors, Contributors, and Reviewers ...........0 000 eee eee eee III-xix 9. RISK CHARACTERIZATION OF 2,3,7,8-TETRACHLORODIBENZO-p-DIOXIN (TCDD) AND RELATED COMPOUNDS ccicesiieeiiateiabisiweanven 9-1 9.1. INTRODUCTION ........ 0.0.0.0 0006 cc eee 9-1 9.2. CHEMICAL STRUCTURE AND PROPERTIES .................... 9-3 9.3. ENVIRONMENTAL FATE ............ 00000 ee eee 9-7 9.4. SOURCES .. 10... 0.0.0. eee eee 9-8 9.4.1. Levels in the Environment and in Food ...............-.2005- 9-13 9.4.2. Background Exposure Levels ............. 0.0000 e ee eee 9-14 9.4.3. Highly Exposed Populations ........... 0.20.00... 000 eee eee 9-19 9.5. DISPOSITION AND PHARMACOKINETICS .............0.-..-0000- 9-23 9.6. MECHANISMS OF DIOXIN ACTION ............... 0.00000 00005 9-30 97, TOXIC EFFECTS OF DIOXIN ci csanss eet a wid ees iwi eRe ea H 9-36 9.7.1. General Comments ........ 0.0.00. eee 9-36 9.7.2. Chloracne .... 0.0.0... ee eee 9-38 9.7.3. Carcinogenicity 2... 0... . ee eee 9-39 9.7.4. Reproductive and Developmental Effects ..................04. 9-44 9.7.5. Immunotoxicity 2.0... 0.2 ee 9-48 9.7.6. Other Effects... 2... 0 eee 9-50 9.7.6.1. Circulating Reproductive Hormones ............... 9-51 9.7.6.2. Diabetes and Fasting Serum Glucose Levels ......... 9-51 D709. Brgyrie IngUCiO ows ce a RRR HE ER REES 9-51 II-iv 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE CONTENTS (continued) 9.7.6.4. Gamma Glutamyl Transferase (GGT) Activity .... 9.7.6.5. Endometriosis .......... 00.0.0 9.8. DOSE-RESPONSE CONSIDERATIONS «siaenroesaaewsanawiame 9.9. USE OF TOXICITY EQUIVALENCE .....................0.05. 9.10. KEY ASSUMPTIONS AND INFERENCES ..................... 9.11. OVERALL CONCLUSIONS REGARDING THE IMPACT OF DIOXIN AND RELATED COMPOUNDS ON HUMAN HEALTH ........... REFERENCES FOR CHAPTER 9 2... ccc cece eee renee een ees II-v 08/15/94 ''9-4. 9-5, 9-6. 9-2. DRAFT--DO NOT QUOTE OR CITE LIST OF TABLES Toxicity Equivalency Factors (TEF) for CDDs and CDFs_ .............. Effects of TCDD and Related Compounds in Different Animal Species .. 2... 2... ee ee eee ee eee 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 ............... Estimated Body Burdens of Experimental Animals and Humans Exposed to Dioxins: Responses in Humans Associated With Dioxin Exposure and Comparable Effects in Experimental Animals ...................00005 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 .................000000. Comparison of the Effects of TCDD Exposure on Human and Animal Tissue In Vitro .. 0... ee ne ee eee enna LIST OF FIGURES Dioxin and similar compounds--chemical structure .................004 Schematic representation of the complex sequence of molecular and biological events involved in dioxin-mediated toxicants ...............0.000005 HI-vi 08/15/94 ''ACTH Ah receptor AHH ALA ALT AOR APC AST ATPase BDD BDF BCF BGG bHLH bw cAMP CDC CDD CDF cDNA cl CMI CNS CSM CTL DCDD DRAFT--DO NOT QUOTE OR CITE LIST OF ABBREVIATIONS AND ACRONYMS Adrenocorticotrophic hormone Aryl hydrocarbon receptor Aryl hydrocarbon hydroxylase Aminolevulinic acid L-alanine aminotransferase Adjusted odds ratio Antigen-presenting cells L-aspartate aminotransferase Adenosine triphosphatase Brominated dibenzo-p-dioxin Brominated dibenzofuran Bioconcentration factor Bovine gamma globulin Basic helix-loop-helix Body weight Cyclic 3,5-adenosine monophosphate Centers for Disease Control and Prevention Chlorinated dibenzo-p-dioxin Chlorinated dibenzofuran Complementary DNA Confidence level Cornell Medical Index Central nervous system Cerebrospinal malformation Cytotoxic T lymphocyte 2,7-Dichlorodibenzo-p-dioxin III-vii 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE LIST OF ABBREVIATIONS AND ACRONYMS (continued) DEN Diethylnitrosamine DHT 5a-Dihydrotestosterone DIS Diagnostic Interview Schedule DMBA Dimethylbenzanthracene DMSO Dimethyl] sulfoxide DNA Deoxyribonucleic acid DRE Dioxin-responsive enhancers DTH Delayed-type hypersensitivity ECs Concentration effective for 50% of organisms tested EC io0 Concentration effective for 100% of organisms tested ED, 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 UI -viii 08/15/94 ''GGT GnRH GST GVH HAH HCB HCDD HDL HLH HPAH HpCDD HpCDF HPLC HRB HRGC/HRMS HTL HxBB HxCB HxCDD HxCDF ICD-9 IDs I-TEF KVK LADD LD, DRAFT--DO NOT QUOTE OR CITE LIST OF ABBREVIATIONS AND ACRONYMS (continued) Gamma glutamyl transpeptidase Gonadotropin-releasing hormone Glutathione-S-transferase Graft versus host Halogenated aromatic hydrocarbons Hexachlorobenzene Hexachlorodibenzo-p-dioxin High density lipoprotein Helix-loop-helix Halogenated polycyclic aromatic hydrocarbon Heptachlorinated dibenzo-p-dioxin Heptachlorinated dibenzofuran High performance liquid chromatography Halstead-Reitan Battery High resolution gas chromatography/high resolution mass spectrometry Human tonsillar lymphocytes Hexabrom-bipheny] Hexachlorobiphenyl Hexachlorinated dibenzo-p-dioxin Hexachlorinated dibenzofuran International Classification of Diseases 9 Dose infective to 50% of recipients International TCDD-toxic-equivalency Kemisk Vaerk Koge Lifetime average daily dose Dose lethal to 50% of recipients (and all other subscripter dose levels) III-ix 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE LIST OF ABBREVIATIONS AND ACRONYMS (continued) LDH L-lactate dehydrogenase LH Luteinizing hormone LOL. 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-1,3,8-trichlorodibenzofuran MCF-7 (breast cancer cell) MCMI 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 N-methyl-N-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 II-x 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 PCBA Phenoxybutyric acid PCDD Polychlorinated dibenzodioxin PCDF Polychlorinated dibenzofuran PCP Pentachlorophenol PCPA Parachlorophenoxyacetic acid PCQ Quaterphenyl PCT Porphyria cutanea tarda PeCDD Pentachlorinated dibenzo-p-dioxin PeCDF Pentachlorinated dibenzo-p-dioxin III-xi 08/15/94 ''PEPCK PFC PGE, PGF,, PGST PGT PHA PIQ PKC PNS POMS ppb ppm ppt PRR PWM RNA RR SAR SB-IQ SCL-90-R SD SE SEA SGOT SGPT DRAFT--DO NOT QUOTE OR CITE LIST OF ABBREVIATIONS AND ACRONYMS (continued) Phosphoenol pyruvate carboxykinase Plaque-forming cell Prostaglandin E, Prostaglandin F,, Placental glutathione-S-transferase Placental glutathione transferase Phytohemagglutinin Performance IQ Protein kinase C Peripheral nervous system Profile of Mood States Parts per billion Parts per million Parts per trillion Prevalence risk ratio Pokeweed mitogen Ribonucleic acid Relative risk Structure-activity relationships Standford Binet IQ Self-Report Symptom Checklist-90-Revised Standard deviation Standard error Southeast Asia Serum glutamic oxaloacetic transaminase Serum glutamic pyruvic transaminase I-xii 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE LIST OF ABBREVIATIONS AND ACRONYMS (continued) SIR Standard incidence ratio SMR Standardized mortality ratio SRBC Sheep erythrocytes (red blood cells) STS Soft tissue sarcoma ty, Half-time TBB Tetrabromobipheny] 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 Tetrachlorodibenzofuran TCP Trichlorophenol TEF Toxic equivalency factors TEQ Toxic equivalents TGF Thyroid growth factor TI T helper cell independent TNF Tumor necrosis factor tPA Tissue plasminogen activator TPA Tetradecanoyl phorbol acetate TSH Thyroid-stimulating hormone TT Tetanus toxoid TIR Transthyretrin TxB, Thromboxane B, III-xiii 08/15/94 ''UDP UDPGT URO-D VIQ VLDL viv w/w WAIS WISC-R DRAFT--DO NOT QUOTE OR CITE LIST OF ABBREVIATIONS AND ACRONYMS (continued) Uridine diphosphate UDP-glucuronosyltransferase Uroporphyrinogen decarboxylase Verbal IQ Very low density lipoprotein Volume per volume Weight by weight Wechsler Adult Intelligence Scale Wechsler Intelligence Scale for Children, Revised Il-xiv 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE PREFACE In April 1991, the U.S. Environmental Protection Agency (EPA) announced that it would conduct a scientific reassessment of the health risks of exposure to 2,3,7,8- tetrachlorodibenzo-p-dioxin (TCDD) and chemically similar compounds collectively known as dioxin. The EPA has undertaken this task in response to emerging scientific knowledge of the biological, human health, and environmental effects of dioxin. Significant advances have occurred in the scientific understanding of mechanisms of dioxin toxicity, of the carcinogenic and other adverse health effects of dioxin in people, of the pathways to human exposure, and of the toxic effects of dioxin to the environment. In 1985 and 1988, the Agency prepared assessments of the human health risks from environmental exposures to dioxin. Also, in 1988, a draft exposure document was prepared that presented procedures for conducting site-specific exposure assessments to dioxin-like compounds. These assessments were reviewed by the Agency’s Science Advisory Board (SAB). At the time of the 1988 assessments, there was general agreement within the scientific community that there could be a substantial improvement over the existing approach to analyzing dose response, but there was no consensus as to a more biologically defensible methodology. The Agency was asked to explore the development of such a method. The current reassessment activities are in response to this request. The scientific reassessment of dioxin consists of five activities: Update and revision of the health assessment document for dioxin. Laboratory research in support of the dose-response model. Development of a biologically based dose-response model for dioxin. Update and revision of the dioxin exposure assessment document. “af YP he Research to characterize ecological risks in aquatic ecosystems. III-xv 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE PREFACE (continued) The first four activities have resulted in two draft documents (the health assessment document and exposure document) for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and related compounds. These companion documents, which form the basis for the Agency’s reassessment of dioxin, have been used in the development of the risk characterization chapter that follows the health assessment. The process for developing these documents consisted of three phases which are outlined in later paragraphs. The fifth activity, which is in progress at EPA’s Environmental Research Laboratory in Duluth, Minnesota, involves characterizing ecological risks in aquatic ecosystems from exposure to dioxins. Research efforts are focused on the study of organisms in aquatic food webs to identify the effects of dioxin exposure that are likely to result in significant population impacts. A report titled, Interim Report on Data and Methods for the Assessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) Risks to Aquatic Organisms and Associated Wildlife (EPA/600/R-93/055), was published in April 1993. This report will serve as a background document for assessing dioxin-related ecological risks. Ultimately, these data will support the development of aquatic life criteria which will aid in the implementation of the Clean Water Act. The EPA had endeavored to make each phase of the current reassessment of dioxin an open and participatory effort. On November 15, 1991, and April 28, 1992, public meetings were held to inform the public of the Agency’s plans and activities for the reassessment, to hear and receive public comments and reviews of the proposed plans, and to receive any current, scientifically relevant information. In the Fall of 1992, the Agency convened two peer-review workshops to review draft documents related to EPA’s scientific reassessment of the health effects of dioxin. The first workshop was held September 10 and 11, 1992, to review a draft exposure assessment titled, Estimating Exposures to Dioxin-Like Compounds. The second workshop was held September 22-25, 1992, to review eight chapters of a future draft Health II-xvi 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 noncancer, from exposure to dioxin, with an emphasis on the specific congener, 2,3,7,8-TCDD. As mentioned previously, completion of the health assessment and exposure documents involves three phases: Phase 1 involved drafting state-of-the-science chapters and a dose-response model for the health assessment document, expanding the exposure document to address dioxin related compounds, and conducting peer review workshops by panels of experts. This phase has been completed. Phase 2, preparation of the risk characterization, began during the September 1992 workshops with discussions by the peer-review panels and formulation of points to be carried forward into the risk characterization. Following the September 1993 workshop, this work was completed and was incorporated as Chapter 9 of the draft health assessment document. This phase has been completed. Phase 3 is currently underway. It includes making External Review Drafts of both the health assessment document and the exposure document available for public review and comment. Following the public comment period, the Agency’s Science Advisory Board (SAB) will review the draft documents in public session. Assuming that public and SAB comments are positive, the draft documents will be revised, and final documents will be issued. III-xvii 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. II-xviii 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE AUTHORS, CONTRIBUTORS, AND REVIEWERS This draft Health Assessment Document was prepared under the leadership and direction of the Office of Health and Environmental Assessment (OHEA) within EPA’s Office of Research and Development (ORD). The overall coordination and leadership of the activities associated with EPA’s reassessment of dioxin, which includes the development of this draft document, is Dr. William H. Farland, Director of OHEA. Authors and chapter managers for the Health Assessment Document are listed below. Early drafts of some chapters were prepared by Syracuse Research Corporation under EPA Contract No. 68-CO-0043. Other chapters were authored totally or in part by scientists within EPA and other agencies within the federal government. The ORD chapter managers were responsible for providing oversight, review, and technical editing of successive drafts, and incorporating comments from reviewers to develop a comprehensive and consistent document. In some cases, the chapter managers also authored sections or parts of the chapter. AUTHORS AND CHAPTER MANAGERS Chapter . Disposition and Pharmacokinetics EPA Chapter Manager/Author Jerry Blancato U.S. EPA Environmental Monitoring Systems Laboratory Las Vegas, NV Outside Author James Olson Department of Pharmacology and Therapeutics State University of New York Buffalo, NY . 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 . Acute, Subchronic, and Chronic Toxicity Debdas Mukerjee Environmental Criteria and Assessment Office/OHEA Cincinnati, OH Ulf G. Ahlborg Karolinska Institute Stockholm, SWEDEN III-xix 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) Chapter EPA Chapter Manager/Author 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 Outside Author 4. Immunotoxicity Ralph Smialowicz 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 IlI-xx 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) Chapter 8. Dose-Response Modeling EPA Chapter Manager/Author Steven P. Bayard U.S. EPA Human Health Assessment Group OHEA Washington, DC Outside Author Dioxin Dose-Response Modeling Workgroup Michael Gallo (Co-chair), Keith Cooper, Panos Georgopolous, and Lynne McGrath UMDNS-Robert Wood Johnson Medical School Environmental and Occupational Health Sciences Institute (EOHSI) Piscataway, NJ George Lucier (Co-chair) and Christopher Portier National Institute of Environmental Health Sciences Research Triangle Park, NC Melvin Andersen and Michael DeVito U.S. EPA Health Effects Research Laboratory Research Triangle Park, NC Steven Bayard and Paul White U.S. EPA Office of Health and Environmental Assessment Washington, DC Lorrene Kedderis University of North Carolina Chapel Hill, NC Jeremy Mills Chemical Industry Institute of Toxicology Research Triangle Park, NC Ellen Silbergeld University of Maryland Baltimore, MD 9. Risk Characterization of 2,3,7,8-TCDD and Related Compounds William H. Farland (See Chapter 2) III-xxi 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) *Involved with an early draft, but no longer working on the reassessment project. CONTRIBUTORS Linda Birnbaum Marilyn Fingerhut Dorothy Patton Peter W. Preuss Dwain Winters REVIEWERS Director, Environmental Toxicology Division, Health Effects Research Laboratory, U.S. Environmental Protection Agency, Research Triangle Park, NC Chief, Industry-wide Studies Branch, National Institute for Occupational Safety and Health, Cincinnati, OH 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 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 II-xxii 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) Phillip Enterline Emeritus Professor of Biostatistics, University of Pittsburgh, Pittsburgh, PA Mark Feeley Toxicity Evaluation Division, Bureau of Chemical Safety, Health, and Welfare, Ottawa, Ontario, Canada Thomas A. Gasiewicz Department of Biophysics, University of Rochester, Rochester, NY James Gillette Laboratory of Chemical Pharmacology, National Heart, Lung, and Blood Institute, National Institutes of Health, Bethesda, MD Claude Hughes Duke University Medical Center, Durham, NC Curtis D. Klaassen Department of Pharmacology, Toxicology and Therapeutics, The University of Kansas Medical Center, Kansas City, KS Daniel Krewski Biostatistics and Computer Applications, Environmental Health Centre, Ottawa, Ontario, Canada Suresh Moolgavkar Professor of Epidemiology and Biostatistics, The Fred Hutchinson Cancer Research Center, Seattle, WA Jay Silkworth Wadsworth Center for Laboratories and Research, New York State Department of Health, Albany, NY Thomas Webster 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 IlI-xxili 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) Germaine Buck Clinical Assistant Professor, Department of Social and Preventive Medicine, State University of New York, Buffalo, NY Harvey Checkoway Professor, Department of Environmental Health, University of Washington, Seattle, WA Phillip Enterline Emeritus Professor of Biostatistics, University of Pittsburgh, Pittsburgh, PA M. Gerald Ott Director of Epidemiology, BASF Corporation, Parsippany, NJ Allan H. Smith Professor of Epidemiology, University of California, Berkeley, CA Anne Sweeney Assistant Professor of Epidemiology, School of Public Health, University of Texas, Houston, TX Karen Webb Medical Director, HealthLine Corporation Health, St. Louis, MO In addition, during the development of this draft Health Assessment Document, selected sections, chapters, or volumes were peer reviewed by scientists and experts within EPA and other federal agencies, as well as by experts in academia and the private sector. A draft of Chapter 9, the risk characterization, was reviewed by an interagency workgroup comprising scientists from the following agencies of the federal government: Department of Agriculture Department of Defense Department of Health and Human Services” “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 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE AUTHORS, CONTRIBUTORS, AND REVIEWERS (continued) Department of Labor (Occupational Safety and Health Administration) Department of Veterans Affairs Executive Office of the President Office of Science and Technology Policy Council of Economic Advisors Domestic Policy Council II-xxv 08/15/94 ''''DRAFT--DO NOT QUOTE OR CITE 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 9-1 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-2 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE questions concerning the science behind concerns for dioxins and should be useful in developing strategies for risk communication. 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-p-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 9-3 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE aINJONS [eowsyd--spunodwioo sJepiwis pue ulxo1q "T-6 nsIy Auaydiqosojyourjuad-s‘.p'p'.£'C fo) 19 19 19 ID 19 jAuaydiqosojysexoq-.s‘s‘y'p'.e'c UEINJOZUaQ|POIO|YIEWUId-8'L C'S ores Ujxo|p-d-oZUaq|POLOIYSeIUad-B" Le'z's KO UeINJOZUSq!POJO}YIeI}91-8'L'C'Z I | Tr ma eas ujxo}p-d-ozueq! polo] 4yoeje 1 -9'2'C'Z ° CL Lx 1S 1D ° 08/15/94 9-4 ''DRAFT--DO NOT QUOTE OR CITE Table 9-1. Toxicity Equivalency Factors (TEF) for CDDs and CDFs Compound TEF Mono-, Di-, and Tri-CDDs 0 2,3,7,8-TCDD 1 Other TCDDs 0 2,3,7,8-PeCDD 0.5 Other PeCDDs 0 2,3,7,8-HxCDD 0.1 Other HxCDDs 0 2,3,7,8-HpCDD 0.01 Other HpCDDs 0 OCDD 0.001 Mono-, Di-, and Tri-CDFs 0 2,3,7,8-TCDF 0.1 Other TCDFs 0 1,2,3,7,8-PeCDF 0.05 2,3,4,7,8-PeCDF 0.5 Other PeCDFs 0 2,3,7,8-HxCDF 0.1 Other HxCDFs 0 2,3,7,8-HpCDF 0.01 Other HpCDFs 0 OCDF 0.001 Source: EPA, 1989. 9-5 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 ecotoxicological 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 9-6 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 compounds are sometimes referred to as coplanar, meaning that they can assume a flat configuration with rings in the same plane. Similarly configured polybrominated 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. 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 9-7 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-8 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 (Hites, 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 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: a 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. 9-9 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 2 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. 2 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 9-10 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-11 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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/m?-yr is typical of remote areas and that 2-6 ng TEQ/m?-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. 9-12 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 resuspension 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. 9-13 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 9-14 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 et al. (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 (Gilman 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 9-15 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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). a Age Effects: Tissue concentrations of CDD/Fs were found to increase with age. a 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. 9-16 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE = 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 PCB 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 9-17 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-18 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 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 an 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 9-19 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. Ona 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 9-20 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 blood 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. Dioxin-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 9-2] 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-22 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 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 9-23 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 while 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 umol/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 hour'). 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 9-24 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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). 9-25 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 control was 5 pg/g (ppt). In addition, 79% of the Ranch Hand personnel and 2% of the controls hac 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 9-26 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 and 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 biotransforming 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, and 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. 0.97 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 EGF 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). 9-28 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 9-29 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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: a Similarities at the biochemical level between humans and other animals with regard to receptor structure and function; 2 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. 9-30 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-31 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 (CYPIA1) 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 9-32 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE Dioxin Exposure V Free Dioxin in Tissues v Dioxin Binding to the Ah Receptor in Tissue v Ah Receptor - Dioxin Complex Binding with DNA V Gene Regulation V m-RNA Regulation V Protein Synthesis v Biochemical Alterations v Early Cellular Responses (cell growth stimulation) v Late (irreversible) Tissue Response (cancer, terata) < -— < _—— Interactions <---- of Multiple Target Genes Se --- <- see Figure 9-2. Schematic representation of the complex sequence of molecular and biological events involved in dioxin-mediated toxicants. 233 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-34 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 CYPIA1 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 immunological 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. 9-35 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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, 1.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 9-36 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE “BIep OU = S][e0 YURI *poaresgo jou = 0) “SYNSOI -/+ JO ‘JUSIXS POITUNT] 0} poATesgo = -/+ “poalesqo = + visejdoddy MOLIBU oUOg Aydone iepnonsey euopy Anorxoioiedey eukydiog $102]J° oluesouseIoyyD AylorueZourosea Aqorxojounurwy soo}jo ouLDOpuq Aytpeyour ‘Ay101xo} [PlaJ/stsoussojyese |, owoipuds sunsen Ayypempey cinoy uononpul ewAzuq (as0ueyus) Fuad om 0} xejdwop yyy :AdOL Jo Burpurg WV JO souescig wey nqqey mod Jo}sure yy asnoy wey voumsy Aayuoy ueumyy IPHA satedgs [BUUY JUatazjIq Ul spunodui0; payejay pue GAOL JO $12JJA *7-6 AGEL 08/15/94 9-37 ''DRAFT--DO NOT QUOTE OR CITE 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 years. Details of chloracnegenic response in occupationally exposed humans are described in detail in Chapter 7 of the Health Assessment Document. 9-38 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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, 9-39 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 not 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 met 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 not be dismissed or ignored. Other scientists, including those who attended the Peer Panel meeting, felt either more or less strongly about the weight of the evidence from epidemiology studies, representing the range of opinion that still exists on the interpretation of the cancer epidemiology studies. Many of the earlier epidemiological 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 9-40 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 out 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 chiidren 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 dioxin action. 9-41 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 6. 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 than a mixture of two isomers of hexachlorodibenzodioxin (HCDDs), which produced liver tumors in both sexes of rats and mice (NTP, 1980), the more highly chlorinated CDDs and CDFs have not been studied in long-term animal cancer bioassays, However, it is generally recognized that these compounds bioaccumulate and exhibit toxicities similar to TCDD and are, therefore, also likely to be carcinogens (U.S. EPA Science Advisory Board, 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, as well 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 9-42 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 adrenals. 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 9-43 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-44 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 and 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 at 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. 9-45 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 9-46 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 species. 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 9-47 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-48 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-49 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 9-50 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 1Al (CYPIA1). CYPIA1 is one of a family of proteins involved in the activation and detoxification of both endogenous and exogenous chemicals. 9-51 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 Transferase (GGT) Activity GGT 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 9-52 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-53 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-54 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 Experimental Effect Species Dose Body Burden Ref./Note Chloracne Humans 45-3,000 ng/kg | Ryan et al., 1990; Beck et al., 1989/a,b Chloracne Monkey 1,000 ng/kg 1,000 ng/kg McNulty, 1985/c Chloracne Rabbits 4 ng/kg 220 ng/kg Schwetz et al., 5d/wk/4wk 1973/d Chloracne Mice 4,000 ng/kg 14,000 ng/kg | Puvel and 3d/wk/2wk Sakamoto, 1988/e Decreased Birth Humans | Mother’s body 1,460 ng/kg Lucier, 1991/f Weight burden Decreased Growth Humans | Mother’s body 1,460 ng/kg Guo et al., burden 1994/f Decreased Growth Rats 400 ng/kg 400 ng/kg Mably et al., maternal dose 1992a/g gd 15 Delayed Humans 1,460 ng/kg | Rogan et al., Developmental 1988/f Milestones Object Learning Monkey 1.26 ng/kg/d 19 ng/kg Schantz and Bowman, 1989/h Down Regulation Humans 1,460 ng/kg Lucier, 1991/f of EGFR in Placenta (Maximal Effect) 9-55 08/15/94 ''Table 9-3. (continued) DRAFT--DO NOT QUOTE OR CITE Experimental Effect Species Dose Body Burden Ref./Note Down Regulation Rats 125 ng/kg/d 1,600 ng/kg | Sewall et al., of EGFR in Liver 30 weeks 1993/i (Maximal Effect) Increase in Humans 1,460 ng/kg | Lucier, 1991/f Placental CYP1A1 (Maximal Effect) Increase in Liver Rats 125 ng/kg/d 1,600 ng/kg | Tritscher et CYP1A1 (Maximal 30 weeks al., 1992/i Effect) Enzyme Induction Rats 1 ng/kg 1 ng/kg Van den CYP1A1 (LOEL) single dose Heuvel et al., sac 24 hr 1993/j Enzyme Induction Mice 1.5 ng/kg/d 23 ng/kg DeVito et al., CYP1A1/1A2 5 d/wk 13 wk 1994/k (LOEL) Hepatic Human 150 ng/kg Carrier et al., Sequestration submitted/1 Hepatic Rats 300 ng/kg Carrier et al., Sequestration submitted/1 Background Human | 60 TEQ ppt in 9 ng/kg m serum Background Mouse 4 ng/kg 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 9-56 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 years. 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 wg/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 placental levels of dioxins and alterations in placental epidermal growth factor receptor (EGFR) and CYPIA1. In addition, the author suggested that the changes in placental EGFR and CYPI1A1 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 9-57 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 9-58 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 Experimental Effect Species Dose Body Burden Ref./Note Cancer Humans 109-7,000 Fingerhut et al., ng/kg 1991; Bertazzi et al., 1993/a Cancer Hamsters 100 pg/kg 500 ng/kg Rao et al., 1988/b 6 doses (600 ug/kg total dose) Cancer Rats 100 ng/kg/d 1,400 ng/kg | Kociba et al., 1978/c for 2 years Cancer Mice 400 ng/kg/d 1,000 ng/kg | NTP, 1982/d for 2 years Liver Tumor Rats 125 ng/kg/d 1,600 ng/kg | Maronpot et al., Promotion 30 weeks 1993/e Skin Tumor Mice 7.5 ng/kg/wk 1,100 ng/kg | Poland et al., 1982/f Promotion for 20 wks dermal exposure Decreased Humans 83 ng/kg Egeland et al., Testosterone 1994/g Decreased Rats 12,500 ng/kg 10,200 ng/kg | Moore et al., 1985/h Testosterone sac day 7 Decreased Testis Humans 14 ng/kg Air Force Study, Size 1991/i Altered Glucose Humans 110 ng/kg Sweeney et al., Tolerance 1992/j Altered Glucose Humans 14 ng/kg Wolfe et al., 1992/i Tolerance 9-59 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE Table 9-4. (continued) Experimental Effect Species Dose Body Burden Ref./Note [© Yl Decreased Guinea 30 ng/kg 30 ng/kg Enan et al., 1992/k Glucose Uptake Pigs sac day | Adipocytes Decreased Rats 100 ng/kg/d 1,900 ng/kg | Zinkl et al., 1973/1 Serum Glucose 30 days Background Human 60 TEQ ppt 9 ng/kg m in serum Background Mouse 4 ng/kg n Notes: 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). Animals administered 100 g/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. Assumes a single first-order elimination rate constant and a half-life for the whole body elimination of 23.7 days (Rose et al., 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. Body burden estimated from animals treated with 450 ng/kg/d for 90 days (DeVito and Birnbaum, unpublished results). 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 9-60 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 wg/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 pg/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 ng 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 wg/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; 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. 9-61 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 Experimental Effect Species Dose Body Burden Ref./Note Decreased Rhesus 25 ppt in diet 270 ng/kg Hong et al., Offspring Monkeys for 4 years 1989/a Viability Altered Rhesus 25 ppt in diet 270 ng/kg Hong et al., Lymphocyte Monkeys for 4 years 1989/a Subsets Altered Marmosets | 0.3 ng/kg/wk 6-8 ng/kg Neubert et al., Lymphocyte for 24 weeks 1992/b Subsets 1.5 ng/kg/wk for 12 weeks Enhanced Viral Mice 10 ng/kg 7 ng/kg Burelson et al., Susceptibility sac day 7 1994/c Endometriosis Monkeys 5 ppt in diet 54 ng/kg Reier et al., 4 years 1993/a Decreased Sperm Rats 64 ng/kg 64 ng/kg Mably et al., Count maternal dose 1992b/d gd 15 Background Human 60 TEQ ppt 9 ng/kg e in serum Background Mouse 4 ng/kg f Notes: 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. 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). 9-62 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE Table 9-5. (continued) Cc. 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). 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. Assumes a background TEQ of 60 ppt for dioxins, dibenzofurans, and PCBs. Also assumes a body weight of 70 kg with 15% body fat. Data from DeVito and Birnbaum (1994). 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. 9-63 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE Table 9-6. Comparison of the Effects of TCDD Exposure on Human and Animal Tissue In Vitro Effect Cell Line/Species Concentration Ref./Note | Binding to CYP1A1 Hepa-1c1c7/mouse 2 nM Probst et al., 1993/a DRE Binding to CYP1A1 LS180/human 10 nM Probst et al., 1993/a DRE Binding to ER DRE Hepa 1clc7/mouse 15.5 nM White and Gasiewicz, 1993/b Binding to ER DRE MCF-7/human 5.6 nM White and Gasiewicz, 1993/b Induction CYP1A1 Lymphocytes mouse 1.3 nM Clark et al., 1992/c Induction CYP1A1 Lymphocytes human 1.8 nM Clark et al., 1992/c Cytotoxicity Embryonic palate 0.1 nM Abbott and mouse Birnbaum, 1991/d Cytotoxicity Embryonic palate rat 100 nM Abbott and Birnbaum, 1991/d Cytotoxicity Embryonic palate 100 nM Abbott and human Birnbaum, 1991/d Altered Lymphocyte Peripheral 0.0001 nM Neubert et al., Subsets lymphocytes 1991/e marmoset Altered Lymphocyte Peripheral 0.0001 nM Neubert et al., Subsets lymphocytes human 1991/e Inhibition of Thymocytes mouse 0.1 nM Greenlee et al., Proliferation 1985/f Inhibition of Thymocytes human 0.1 nM Cook et al., 1987/f Proliferation Inhibition of Tonsilar lymphocytes | 0.3 nM Wood et al., 1993/g Proliferation human Inhibition of Splenic lymphocytes 3.0 nM Wood et al., 1993/g Proliferation murine 9-64 08/15/94 ''Table 9-6. (continued) DRAFT--DO NOT QUOTE OR CITE Effect Cell Line/Species Concentration Ref./Note Inhibition of IgM Splenic lymphocytes 3.0 nM Wood et al., 1993/g Secretion murine Inhibition of IgM Tonsilar lymphocytes | 0.3 nM Wood et al., 1993/g Secrection human Notes: 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 C57BI/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 ECso. 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 and 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. 9-65 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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). 9-66 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 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 1 x10“ 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 9-67 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 PGST 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 and 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 | in 10,000 (10%) for the average population exposures to 1 in 1,000 (10°) 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 9-68 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 characterization. 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 ED. , for effects versus binding are plotted logarithmically, good linear correlations are obtained (Safe, 1990). This approach constitutes 9-69 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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-like 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, 9-70 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 hypothesis 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 large 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 an empirical basis, is theoretically sound, and, in the absence of more complete data sets on the toxicity of 9-71 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-72 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 causality for many suspected dioxin effects is assumed not to negate the findings from laboratory animal 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 9-73 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-74 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-75 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-76 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 = 2 Coongener * TEFongener]- 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 pharmacokinetic 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/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 body weight/day. Estimates based 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. 9-77 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE when all dioxins, furans, 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. 9-78 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 used 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 respond 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 9-79 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-80 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 and/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- 9-81 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-82 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 data 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 intake 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 9-83 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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 9-84 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE specificity to evaluate the impact of immunotoxic responses to dioxin and related compounds in humans. Despite the possibility that these compounds may be immunotoxic 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° 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 9-85 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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°) 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% to 10% (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 9-86 08/15/94 ''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 toa 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 in 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 9-87 08/15/94 ''DRAFT--DO NOT QUOTE OR CITE 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. 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