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   9                                   Exposure and Health Assessment for

  l0                             2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)

  l l                                            and Related Compounds

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  13                                                                   PART 3

  14                                   Integrated Summary and Risk Characterization for

  15                                         2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)

  16                                                       and Related Compounds

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  2I                                                       NOTICE

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  23                       THIS DOCUMENT IS A PRELIMINARY DRAFT.

 24            It has not been formally released by the U.S. Environmental Protection

  25           Agency and should not at this stage be construed to represent Agency

 26          policy.  It is being circulated for comment on its technical accuracy and

 27                                               policy implications

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  31                                National Center for Environmental Assessment

 32                                   Office of Research and Development

 33                                       U.S. Environmental Protection Agency

 34                                                Washington, D.C.

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  36

  37        TABLE of CONTENTS

 

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        1        List of Tables

       2   ..    List of Figures

       3         List of Equations

       4                  1.0 Introduction

       5                            1.1  Definition of Dioxin-like Compounds

       6                            1.2  The "Toxicity Equivalence" Concept

       7                            1.3 Understanding Exposure/Dose Relationships for Dioxin-like  Compounds

       8                  2.0 Effects Summary

       9                            2.1  Biochemical Responses

      10                           2.2 Adverse Effects in Humans and Animals

      11                             2.2.1  Cancer

      12                               2.2.1.1  Epidemiologic Findings

      13                               2.2.1.2 Animal Carcinogenesis

      14                               2.2.1.3 Other Data Relating to Carcinogenesis

      15                               2.2.1.4 Cancer Hazard Characterization

      16                            2.2.2 Developmental and Reproductive Effects

      17                              2.2.2.1  Epidemiologic Findings

      18                              2.2.2.2 Animal Findings

      19                              2.2.2.3 Other Data Related to Developmental and Reproductive Effects

      20                              2.2.2.4 Developmental and Reproductive Effects Hazard Characterization

      21                            2.2.3 Immunologic Effects

     22                              2.2.3.1  Epidemiologic Findings

     23                              2.2.3.2 Animal Findings

     24                              2.2.3.3 Other Data Related to Immunologic Effects

     25                              2.2.3.4 Immunologic Effects Hazard Characterization

     26                            2.2.4 Chloracne

     27                            2.2.5 Diabetes

     28                            2.2.6 Other Adverse Effects

 

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        l                  3.0 Mechanisms and Mode of Dioxin Action

        2                 4.0 Exposure Summary

        3                            4.1 Sources

        4                                      4.1.1 Inventory of Releases

        5                                      4.1.2 General Source Observations

        6                            4.2 Environmental Fate

        7                            4.3 Environmental Media and Food Concentrations

        8                            4.4 Background Exposures

        9                                      4.4.1 Tissue Levels

       10                                      4.4.2 Intake Estimates

       11                                      4.4.3 Variability in Intake Levels

       12                            4.5 Potentially Highly Exposed Populations or Developmental Stages

       13                            4.6 Environmental Trends

       14                  5.0 Dose-Response Summary

       15                            5.1  Dose Metrics

       16                            5.2 Empirical Modeling of Individual Data Sets

       17                                      5.2.1 Cancer

       18                                      5.2.2 Noncancer Endpoints

       19                            5.3 Mode-of-Action-based Dose-Response Modeling

      20                             5.4 Summary Dose-Response Characterization

      21                  6.0 Risk Characterization

22                                  7.0 References

 

 

 

 

 

 

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        1                                                               List of Tables

        2

        3        Table I-1. The TEF Scheme for I-TEQDv

        4        Table 1-2.  The TEF Scheme for TEQDrv-WHO94

          5        Table 1-3.  The TEF Scheme for TEQDn,-WHOgs

        6        Table 2-1.  Effects of TCDD and Related Compounds in Different Animal Species

        7        Table 4-1.  Confidence Rating Scheme

        8        Table 4-2.  Quantitative Inventory of Environmental Releases of TEQov-WHO98 in the U.S.

        9        Table 4-3. Preliminary Indication of the Potential Magnitude of TEQDv-WHO98 Releases

       10                         from "Unquantified" (i.e., Category D) Sources in Reference Year 1995

       11        Table 4-4.  Unquantified Sources

       12        Table 4-5.  Estimates of the range of typical background levels of dioxin-like compounds in

       13                           various environmental media

       14        Table 4-6.  Estimates of Typical Background Levels of Dioxin-like Compounds in Food

       15        Table 4-7.  Background Serum Levels in the US 1995- 1997

       16        Table 4-8.  Adult Contact Rates and Background Intakes of Dioxin-like Compounds

       17        Table 4-9.  The Variability in Average Daily TEQ Intake as a Function of Age

       18        Table 5-1.  Serum Dioxin Levels in the Background Population and Epidemiological

       19                          Cohorts (Back-calculated)

       20        Table 5-2.  Doses yielding 1% excess risk (95% lower confidence bound) based upon 2-year

       21                         animal carcinogenicity studies using simple multistage (Portier et. al, 1984)

       22                           models

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         I                                                               List of Figures

        2

        3        Figure 1-1.  Chemical Structure of 2,3,7,8-TCDD and Related Compounds

        4        Figure 2-1.  Generalized Model for Early Molecular Events in Response to Dioxin

        5        Figure 2-2.  Some of the Genes Whose Expression Is Altered By Exposure to TCDD

        6        Figure 4-1.  Estimated CDD/CDF I-TEQ Emissions to Air from Combustion Sources in the

        7                           United States; Period: 1995

        8        Figure 4-2. Comparison of Estimates of Annual I-TEQ Emissions to Air (grams I-TEQ/yr.)

        9                           for Reference Years 1987 and 1995

      10        Figure 5.1.  Dioxin Body Burden Levels in Background Populations and Epidemiological

      11                           Cohorts (Back-Calculated)

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       1                                                             List of Equations

       2

       3        Equation 1-1. Determination of TEQ

       4        Equation 3-1.  Ligand Binding Kinetics

       5        Equation 5-1. Calculating Slope Factors from Body Burdens at the ED01

       6        Equation 5-2. Calculating Upper Bound on Excess Risk at Human Background Body

       7                               Burden

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         1         1.0      INTRODUCTION

        2                  This document presents an integrated summary of available information related to

        3        exposure and possible health effects of dioxin and related compounds.  It also presents a short

        4        risk characterization which is a concise statement of dioxin science and the public health

        5        implications of both general population exposures from environmental "background" and

        6        incremental exposures associated with proximity to sources of dioxin and related compounds.

        7        While it summarizes key findings developed in the exposure and health assessment portions

        8        (Parts 1  and 2, respectively) of the Agency's Dioxin Reassessment effort, it is meant to be

        9        detailed enough to stand on its own for the average reader.  Readers are encouraged to refer to

       10        the more detailed documents for further information on the topics covered here and to see

       11       complete literature citations.  These documents are:

       12

       13       --       Estimating Exposure to Dioxin-like Compounds - This document, hereafter referred to as

      14         Part 1, the Exposure Document, is divided into four volumes:  1. Executive Summary, 2.

      15        Sources of Dioxin in the United States, 3.  Properties, Environmental Levels and

      16        Background Exposures, and 4. Site-Specific Assessment Procedures.

      17

      18         --      Health Assessment Document for 2, 3, 7,8-TCDD and Related Compounds - This

      19        document, hereafter referred to as Part 2, the Health Document, contains two volumes

      20        with nine chapters covering pharmacokinetics, mechanisms of action, epidemiology,

      21        animal cancer and various noncancer effects, toxicity equivalence factors (TEFs) and

      22        dose-response.

      23

     24                  Parts of this integrative summary and risk characterization go beyond individual chapter

     25        findings to reach general conclusions about the potential impacts of dioxin-like compounds on

     26        human health. It specifically identifies issues conceming the risks that may be occurring in the

     27        general population at or near population background exposure levels.  It articulates the  strengths

28               and weaknesses of the available evidence for possible sources, exposures and health effects, and

 

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       1        presents assumptions made and inferences used in reaching conclusions regarding these data.

       2        The filial risk characterization provides a synopsis of dioxin science and its implications for

       3        characterizing hazard and risk for use by risk assessors and managers inside and outside of EPA

       4        and by the general public.

       5                  This document (Part 3) is organized as follows:

       6                  1.0 Introduction - This section describes: the purpose/organization of, and the process

       7                  for developing, the report; defines dioxin-like compounds in the context of the EPA Re-

       8                  assessment; and explains the Toxicity Equivalency (TEQ) concept.

       9                  2.0 Effects Summary - This section summarizes the key findings of the Health

     10                  Document and provides links to relevant aspects of exposure, mechanisms and dose-

     11                  response.

     12                  3.0 Mechanisms and Mode of Dioxin Action - This section discusses the key findings

     l 3                  on effects in terms of mode-of action.  It uses the "Mode-of-Action Framework" recently

     14                  described by the VvS-IO/IPCS Harmonization of Approaches to Risk Assessment Project

      15                 and contained in the Agency's draft Guidelines for Carcinogen Risk Assessment as the

      16                  basis for the discussions.

      17                  4.0 Exposure Summary - This section summarizes the key findings of the Exposure

     18                  Document and links them to the effects, mechanisms and dose-response characterization.

     19                  5.0 Dose Response Summary - This section summarizes approaches to dose response

     20                  which are found in the Health Document and provides links to relevant aspects of

     21                  exposure and effects.

     22                  6.0 Risk Characterization - This section presents conclusions based on an integration of

     23                  the exposure, effects, mechanisms and dose response information. It also highlights key

     24                  assumptions and uncertainties.

     25                  The process for developing this risk characterization and companion documents has been

     26                 open and participatory.  Each of the documents have been developed in collaboration with

     27                 scientists from inside and outside the Federal Govemment.  Each document has undergone

     28                 extensive intemal and extemal review, including review by EPA's Science Advisory Board

     29                 (SAB).  In September 1994, drafts of each document, including an earlier version of this risk

 

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         l        characterization, were made available for public review and comment.  This included a 150-day

         2        comment period and 11 public meetings around the country, to receive oral and written comment.

        3        These comments along with those of the SAB have been considered in the drafting of this final

        4        document.  The Dose-Response Chapter of the Health Effects Document and an earlier version of

        5        this integrated Summary and Risk Characterization underwent peer review in 1997 and 1998,

        6        respectively, and comments have been incorporated.  In addition, as requested by the SAB, a new

        7        chapter on Toxicity' Equivalence has been developed and will undergo review in parallel with

        8        this document.  When complete, and following final SA.B review, the comprehensive set of

        9        background documents and this integrative summary and risk characterization will be published

       10       as final reports and replace the previous dioxin assessments as the scientific basis for EPA

       11       decision-making.

       12

       13        1.1  Definition of Dioxin-Like Compounds

       14                  As defined in Part 1, this assessment addresses specific compounds in the following

       15        chemical classes: polychlorinated dibenzodioxins (PCDDs or CDDs), polychlorinated

       16        dibenzofurans (PCDFs or CDFs),  polybrominated  dibenzodioxins (PBDDs or BDDs),

      17        polybrominated dibenzofurans (PBDFs or BDFs) and polychlorinated biphenyls (PCBs), and

      18        describes this subset of chemicals as "dioxin-like."   Dioxin-like refers to the fact that these

      19        compounds have similar chemical structure, similar physical-chemical properties, and invoke a

      20        common battery of toxic responses.  Due to their hydrophobic nature and resistance towards

      21         metabolism, these chemicals persist and bioaccumulate in fatty tissues of animals and humans.

      22        The CDDs include 75 individual compounds and CDFs include 135 different compounds.  These

      23        individual compounds are referred to technically as congeners.  Likewise, the BDDs include 75

      24        different congeners and the BDFs include an additional 135 congeners.  Only 7 of the 75

      25        congeners of CDDs, or of BDDs, are thought to have dioxin-like toxicity; these are ones with

     26        chlorine/bromine substitutions in, at a minimum, the 2, 3, 7, and 8 positions.  Only 10 of the 135

     27        possible congeners of CDFs or of BDFs are thought to have dioxin-like toxicity; these also are

     28        ones with substitutions in the 2, 3, 7, and 8 positions.  This suggests that 17 individual

 

 

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         I         CDDs/CDFs, and an additional  17 BDDs/BDFs exhibit dioxin-like toxicity.  The database on

         2        many of the brominated compounds regarding dioxin-like activity has been less extensively

         3        evaluated, and these compounds have not been explicitly considered in this assessment.

         4                  There are 209 PCB congeners.  Only 13 of the 209 congeners are thought to have dioxin-

         5        like toxicity; these are PCBs with 4 or more chlorines with just 1 or no substitution in the ortho

         O        position.  These compounds are sometimes relented to as coplanar, meaning that they can assume

         7        a flat configuration with rings in the same plane.  Similarly configured polybrominated biphenyls

         8        (PBBs) are likely to have similar properties.  However, the data base on these compounds with

         9        regard to dioxin-like activity has been less extensively evaluated, and these compounds have not

        10        been explicitly considered in this assessment.  Mixed chlorinated and brominated congeners of

        11        dioxins, furans and biphenyls also exist, increasing the number of compounds potentially

        12        considered dioxin-like within the definitions of this assessment.  The physical/chemical

       13        properties of each congener vary, according to the degree and position of chlorine and/or bromine

       14        substitution.  Very little is known about occurrence and toxicity of the mixed (chlorinated and

       15        brominated) dioxin, furan, and biphenyl congeners. Again, these compounds have not been

       16        explicitly considered in this assessment.  Generally speaking, this assessment focuses on the 17

       17        CDDs/CDFs and a few of the coplanar PCBs which are frequently encountered in source

       18        characterization or environmental samples.  While recognizing that other "dioxin-like"

       19        compounds exist in the chemical classes discussed above (e.g. brominated or

       20        chlorinated/brominated congeners) or in other chemical classes (e.g. halogenated naphthalenes or

       21        benzenes, azo- or azoxybenzenes), the evaluation of less than two dozen chlorinated congeners is

       22        generally considered sufficient to characterize environmental "dioxin."

       23                  The chlorinated dibenzodioxins and dibenzofurans are tricyclic aromatic compounds with

       24        similar physical and chemical properties. Certain of the PCBs (the so-called coplanar or mono-

       25        ortho coplanar congeners) are also structurally and conformationally similar.  The most widely

       26        studied of this general class of compounds is 2,3,7,8-tetrachlorodiben:zo-p-dioxin (TCDD).  This

       27        compound, often called simply "dioxin", represents the reference compound for this class of

       28        compounds.  The structure of TCDD and several related compounds is shown in Figure 1-1.

 

 

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        1        Although sometimes confusing, the term "dioxin" is often also used to refer to the complex

         2        mixtures of TCDD and related compounds emitted from sources, or found in the environment or

        3        in biological samples.  It can also be used to refer to the total TCDD "equivalents" found in a

        4        sample.  This concept of toxicity equivalence is discussed extensively in Part 2, Chapter 9 and is

        5        summarized below.

        6

        7        1.2 Toxicity Equivalence Factors

        8                  CDDs, CDFs and PCBs are commonly found as complex mixtures when detected in

        9         environmental media and biological tissues, or when measured as environmental releases from

       10        specific sources.  Humans are likely to be exposed to variable distributions of CDDs, CDF and

       11        dioxin-like PCB congeners that vary by source and pathway of exposures.  This complicates the

       12        human health risk assessment  that may be associated with exposures to variable mixtures of

 

 

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        1         dioxin-like compounds.  In order to address this problem, the concept of toxicity equivalence has

        2        been considered and discussed by the scientific community and toxic equivalency factors (TEFs)

        3        have been developed and introduced to facilitate risk assessment of exposure to these chemical

        4        mixtures.

        5                  On the most basic level, TEFs compare the potential toxicity of each dioxin-like

        6        compound comprising the mixture to the well-studied and understood toxicity of TCDD, the

        7        most toxic member of the group.  The background and historical perspective regarding this

        8        procedure is described in detail in Part2, Chapter 9 and in Agency documents (EPA 1987, 1989,

        9         1991a).   This procedure involves assigning individual toxicity equivalency factors (TEFs) to the

      10        2,3,7,8 substituted CDD/CDF congeners, and "dioxin-like" PCBs.  To accomplish this, scientists

      11         have reviewed the toxicological databases along with considerations of chemical structure,

      12        persistence and resistance to metabolism, and have agreed to ascribe specific, "order of

      13        magnitude" TEFs for each dioxin-like congener relative to TCDD which is assigned a TEF of

      14        1.0.  The other congeners have TEF values ranging from 1.0 to 0.00001.  Thus, these TEFs are

      15        the result of scientific judgment of a panel of experts using all of the available data and are

      16        selected to account for uncertainties in the available data and to avoid underestimating risk.  In

      17        this sense, they can be described as "public health conservative" values.  To apply this TEF

      18        concept, the TEF of each congener present in a mixture is multiplied by the respective mass

      19        concentration and the products are summed to represent the 2,3,7,8-TCDD Toxic Equivalence

      20        (TEQ) of the mixture as determined by Equation 1-1.

 

 

 

      22        Equation 1-1: Determination of TEQ

      23

     24                    The TEF values for PCDDs and PCDFs were originally adopted by intemational

     25        convention (U.S. EPA,  1989).  Subsequent to the development of the first intemational TEFs for

     26        CDD/Fs, these values were further reviewed and/or revised and TEFs were also developed for

     27        PCBs (Ahlborg et al.; 1994; van den Berg,  1998).  A problem arises in that past and present

     28        quantitative exposure and risk assessments may not have clearly identified which of three TEF

 

 

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         1         schemes were used to estimate the TEQ.  This reassessment introduces a new uniform TEQ

         2        nomenclature that clearly distinguishes between the different TEF schemes as well as identifies

         3        the congener groups included in specific TEQ calculations.  The nomenclature uses the following

         4        abbreviations to designate which TEF scheme was used in the TEQ calculation:

         5

         6        1.        I-TEQ refers to the Intemational TEF scheme adopted by EPA in 1989  (U.S. EPA,

         7                  1989). See Table 1- 1.

         8        2.        TEQ-WHO94 refers to the 1994 World Health Organization (WHO) extension of the 1-

         9                  TEF scheme to include 13 dioxin-like PCBs (Ahlborg et al.,  1994).  See Table 1-2.

       10        3.        TEQ-WHO98 refers to the 1998 WHO update to the previously established TEFs for

       11                  dioxins, furans, and dioxin-like PCBs (Van den Berg, et al.,  1998).  See Table 1-3.

       12

       13        The nomenclature also uses subscripts to indicate which family of compounds are included in

       14        any specific TEQ calculation.  Under this convention, the subscript D is used to designate

       15        dioxins, the subscript F to designate furans and the subscript P to designate PCBs.  As an

       16        example, "TEQDF-WHO98" would be used to describe a mixture for which only dioxin and furan

       17        congeners were determined and where the TEQ was calculated using the WHO98 scheme.  If

       18        PCBs had also been determined, the nomenclature would be "TEQDFP-WHO98."  Note that the

       19        designations TEQDF-WHO94 and I-TEQDF are interchangeable as the TEFs for dioxins and furans

      20        are the same in each scheme.  Note also that in the current draft of this document, I-TEQ

      21         sometimes appears without the D and F subscripts.  This indicates that the TEQ calculation

      22        includes both dioxins and furans.

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      24

      25

      26

      27

      28        Table 1-1. The TEF Scheme for I-TEQDF*

      29

 

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              Table 1-2.  The TEF Scheme for TEQDFp-WHO94.

         

                                                                                                        

      

 

 

 

 

 

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        1         equivalence to complex environmental mixtures for assessment and regulatory purposes.  Later

        2        sections of this document describe the mode(s) of action by which dioxin-like chemicals mediate

        3        biochemical and toxicological actions.  These data provide the scientific basis for the TEF/TEQ

        4        methodology.  In its twenty year history, the approach has evolved, and decision criteria

        5        supporting the scientific judgment and expert opinion used in assigning TEFs has become more

        6        transparent.  Numerous states, countries and several intemational organizations have evaluated

         7       and adopted this approach to evaluating complex mixtures of dioxin and related compounds (Part

         8        2, Chapter 9).  It has become the accepted methodology, although the need for research to explore

         9        altemative approaches is widely endorsed.   Clearly, basing risk on TCDD alone or assuming all

       10        chemicals are equally potent to TCDD is inappropriate based on available data.  While

       11        uncertainties in the use of the TEF methodology have been identified and are described later in

       12        this document and in detail in Part 2, Chapter 9, one must examine the use of this method in the

       13        broader context of the need to evaluate the potential public health impact of complex mixtures of

       14        persistent, bioaccumulative chemicals.  It can be generally concluded that the use of TEF

       15       methodology for evaluating complex mixtures of dioxin-like compounds decreases the overall

       16        uncertainties in the risk assessment process as compared to altemative approaches.  Use of the

       17        latest consensus values for TEFs assures that the most recent scientific information informs this

       18        "useful, interim approach" ( EPA, 1989; Kutz et al., 1990) to dealing with complex environmental

       19        mixtures of dioxin-like compounds.  As stated by the EPA Science Advisory Board (EPA, 1995),

       20        "The use of the TEFs as a basis for developing an overall index of public health risk is clearly

       21        justifiable, but its practical application depends on the reliability of the TEFs and the availability

       22        of representative and reliable exposure data."  EPA will continue to work with the intemational

       23        scientific community to update these TEF values and evaluate their use on a periodic basis.  One

       24        of the limitations of the use of the TEF methodology in risk assessment of complex environmental

       25        mixtures is that the risk from non-dioxin-like chemicals is not evaluated in concert with that of

       26        dioxin-like chemicals.  Future approaches to the assessment of environmental mixtures should

       27        focus on the development of methods that will allow risks to be predicted when multiple

       28        mechanisms are present due to a variety of contaminants.

       29        1.3 Understanding Exposure/Dose Relationships for Dioxin-like Compounds

       30                  Dose can be expressed as a variety of metrics (e.g., daily intake, serum concentrations,

 

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        1        steady-state body burdens, AUC).  Ideally, the best dose metric is that which is directly and

        2        clearly related to the toxicity of concem by a well-defined mechanism.  In the mechanism-based

        3        cancer modeling instantaneous values of a dose-metric, CYP1A2 or EGF receptor concentrations,

        4        are used as surrogates for mutational rates and growth rates within a two-stage cancer model.

        5        The utility of a particular metric will depend upon the intended application of the dose metric and

        6        the ability to accurately determine this dose metric.  For example, if concentration of activated Ah

        7        receptors in a target tissue was the most appropriate dose metric for a particular response, we

        8        presently have no means to determine these values in humans.

        9                  In this reassessment of the health effects of dioxins, dose is used to understand the animal

      10        to human extrapolations, comparing human exposure as well as comparing the sensitivity of

      11         different toxic responses.  Previous assessments of TCDD have used daily dose as the dose metric

      12        and applied either an allometric scaling factor or an uncertainty factor for species extrapolation.

      13        The present assessment uses steady-state body burdens as the dose metric of choice. One reason

      14        for the change in dose metrics is that recent data demonstrate that the use of either allometric

      15        scaling or uncertainty factors underestimates the species differences in the pharmacokinetic

      l 6        behavior of TCDD and related chemicals.  This is due to persistence and accumulation of dioxins

      17        in biological systems and to the large difference in half-lives (approximately 100 fold differences)

      18        between humans and rodents.  When extrapolating across species, steady-state body burden is the

      19        most appropriate dose metric. The choice of body burden as the dose metric is based on scientific

      20        and pragmatic approaches.  As stated earlier, the best dose metric is that which is directly and

      21        clearly related to the toxicity of concem.  For dioxins, there is evidence in experimental animals

      22        that tissue concentrations of dioxins is an appropriate dose metric for the developmental,

      23        immunological and biochemical effects of dioxins. Comparing target tissue concentrations of

      24        dioxins between animals and humans is impractical.  In humans, the tissues for which we have

      25        estimates of the concentration are limited to tissues which may not be the target tissue of concem

      26        such as serum, blood or adipose tissue. However, tissue concentrations are directly related to body

      27        burdens of dioxins.  Hence steady-state body burdens can be used as surrogates for tissue

      28        concentrations.

     29                  Body burdens have been estimated through two different methods.  Serum, blood or

 

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        1        adipose tissue concentrations of dioxins are reported as pg/g lipid.  Evidence supports the

        2        assumption that TCDD and related chemicals are approximately evenly distributed throughout the

        3        body lipid.  Using the tissue lipid concentrations and the assumption that TCDD is equally

        4        distributed based on lipid content, body burdens are calculated by multiplying the tissue

        5        concentration by the percent body fat composition.  One potential problem for estimating body

        6        burdens is the hepatic sequestration of dioxins.  In rodents, dioxins accumulate in hepatic tissue to

        7        a greater extent than predicted by lipid content.  This sequestration is due to CYP1A2 which binds

        8        dioxins.  There is also evidence in humans that dioxins are sequestered in hepatic tissue.

        9        Estimating body burdens on serum, blood or adipose tissue concentrations may under predict true

       10        body burdens of these chemicals.  This under prediction should be relatively small.   Since liver is

       11         approximately 5% of the body weight, even a 10-fold sequestration in hepatic tissue compared to

       12        adipose tissue would result in a 50% difference in the body burden estimated using serum, blood

       13        or adipose tissue concentrations.  In addition, the sequestration is dose-dependent and at human

       14        background exposures, hepatic sequestration should not be significant.

       15                  A second method for determining body burdens is based on estimates of the daily intake

       16        and half-life of dioxins. Limitations on estimating body burden through this method are dependent

       17        upon the accuracy of the estimates for intake and half-life.  Historically, intakes of dioxins have

      18         varied and there is some uncertainty about past exposures.  In addition, little is known about the

      19         half-life of dioxins at different life stages, although there is a relationship between fat composition

      20         and elimination of dioxins.  Finally, depending on the exposure scenario, using the half-life of

      21         TCDD for the TEQ concentrations may result in some inaccuracies.  While the chemicals that

      22         contribute most to the total TEQ, such as the pentachlorodioxins and dibenzofurans and PCB 126,

      23         have similar half-lives as TCDD, other contributors to the total TEQ have significantly different

      24         half-lives.  This document uses pharmacokinetic modeling in a number of places where it is

      25         assumed that the seven year half life for TCDD can be applied to the TEQDFP of a mixture of

      26         dioxins, furans and PCBs. The validity of this assumption was tested in the following way.  First,

      27         congener specific half-lives and intake rates were identified for each of the dioxin and furan

     28          congeners with nonzero TEFs.  These half lives and intakes were input into a one compartment,

     29          steady state pharmacokinetic model to get congener specific tissue concentrations.  The congener

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         1        specific tissue levels were summed to get an overall TEQDF tissue value.  Second, the

         2        pharmacokinetic model was run using the 7 yr half life and total TEQDF intake to get a TEQDF

         3        tissue concentration.  Both of these modeling approaches yielded very similar TEQDF tissue

         4        levels.  Although this exercise did not include PCBs (due to lack of half life estimates) and the

         5        congener specific half-lives for many of the dioxins and furans have limited empirical support, it

         6        provides some assurance that this is a reasonable approach (see full discussion in Part 1, Volume

         7        3, Chapter 4).

         8                  Body burdens also have an advantage as a dose metric when comparing the occupational

         9        or accidental exposures to background human exposures.   In the epidemiological studies, the

       10        extemal exposure and the rate of this exposure are uncertain.  The only accurate information we

       11        have is on serum, blood or adipose tissue concentrations.  Because of the long biological half-life

       12        of TCDD, these tissue concentrations of dioxins are better markers of past exposures than they are

       13        of present exposures.  Hence, body burdens allow for estimations of exposure in these

       14        occupational and accidentally exposed cohorts.  In addition, this dose metric allows us to compare

       15        these exposures with those of background human exposures.

       16                  The use of body burden, while not perfect, provides a better dose metric than daily dose.

       17        There is sufficient scientific evidence to support the use of body burden as a reasonable

       18        approximation of tissue concentrations.  Future efforts to seeking to better understand the dose-

       19        response relationships for the effects of dioxin-like chemicals should provide insight into

       20        determining better dose metrics for this class of chemicals.

       21

       22

       23

      24

      25        2.0      EFFECTS SUMMARY

      26

      27                  Since the identification of TCDD as a chloracnegen in 1957, over 5,000 publications have

      28        discussed its biological and toxicological properties. A large number of the effects of dioxin and

      29        related compounds have been discussed in detail throughout the chapters in Part 2 of this

 

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         1        assessment.  They illustrate the wide range of effects produced by this class of compounds.  The

         2        majority of effects have been identified in experimental animals; some have also been identified

         3        in exposed human populations.

         4                  Cohort and case-control studies have been used to investigate hypothesized increases in

         5        malignancies among the various 2,3,7,8-TCDD-exposed populations (Fingerhut et al., 1991a, b;

         6        Steenland et al., 1999; Manz et al., 1991; Eriksson et al., 1990).  Cross-sectional studies have been

         7        conducted to evaluate the prevalence or extent of disease ii1 living 2,3,7,8-TCDD-exposed groups

         8        (Suskind and Hertzberg,  1984; Moses et al.,  1984; Lathrop et al., 1984, 1987; Roegner et al.,

         9         1991; Grubbs et al.  1995; Sweeney et al.,  1989; Centers for Disease Control Vietnam Experience

       10        Study,  1988a; Webb et al.,  1989; Ott and Zober,  1994).  The limitations of the cross-sectional

       11         study design for evaluating hazard and risk is discussed in Part 2, Chapter 7b.  Many of the

       12        earliest studies were unable to define exposure-outcome relationships owing to a variety of

       13        shortcomings, including small sample size, poor participation, short latency periods, selection of

       14        inappropriate controls, and the inability to quantify exposure to 2,3,7,8-TCDD or to identify

       15        confounding exposures.  In more recent analyses of cohorts (NIOSH, Hamburg) and cross-

       16        sectional studies of U.S. chemical workers (Sweeney et al.,  1989), U.S. Air Force Ranch Hand

       17        personnel (Roegner et al.,  1991; Grubbs et al., 1995), and Missouri residents (Webb et al., 1989),

       18        serum or adipose tissue levels of 2,3,7,8-TCDD were measured to evaluate 2,3,7,8-TCDD-

       19        associated effects in exposed populations.  The ability to measure tissue or serum levels of

      20         2,3,7,8-TCDD for all or a large sample of the subjects confirmed exposure to 2,3,7,8-TCDD and

      21         permitted the investigators to test hypothesized dose-response relationships.

      22                  A large number of effects of exposure to TCDD and related compounds have been

      23        documented in the scientific literature.  Although many effects have been demonstrated in

      24        multiple species (see Table 2-1), other effects may be specific to the species in which they are

      25        measured and may have limited relevance to the human situation. While this is an important

      26        consideration for character/zing potential hazard, all observed effects may be indicative of the

      27        fundamental level at which dioxin produces its biological impact and illustrate the multiple

      28        sequelae which are possible ,,,,'hen primary impacts are at the level of signal transduction and gene

      29        transcription.  While all observed effects may not be characterized as "adverse" effects (i.e. some

 

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          1        may be adaptive and of neutral consequence), they represent a continuum of response expected

          2        from the fundamental changes in biology caused by exposure to dioxin-like compounds.   As

          3        discussed in following sections, the dose associated with this plethora of effects is best compared

          4        across species using a common measurement unit of body burden of TCDD and other dioxin-like

          5        compounds, as opposed to the level or rate of exposure/intake.

          6                  The effects discussed in the following sections are focused on development of an

          7        understanding of dioxin hazard and risk.  This discussion is by its nature selective of findings that

          8        inform the risk assessment process.  Readers are referred to the more comprehensive chapters for

          9        further discussion of the Epidemiologic and toxicologic database.

        10

        11

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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         I         2.1. BIOCHEMICAl, RESPONSES (Cross reference Part 2, Chapters 2, 3, and 8)

        2                  Mechanistic studies can reveal the biochemical pathways and types of biological events

        3        that contribute to adverse effects from exposure to dioxin-like compounds.  For example, much

        4        evidence indicates that TCDD acts via an intracellular protein (the Ah receptor), which is a

        5        ligand-dependent transcription factor that functions in partnership with a second protein (known

        6        as Amt).  Therefore, from a mechanistic standpoint, TCDD’s adverse effects appear likely to

        7        reflect alterations in gene expression that occur at an inappropriate time and/or for an

        8        inappropriate length of time.  Mechanistic studies also indicate that several other proteins

        9        contribute to TCDD's gene regulatory effects and that the response to TCDD probably involves a

       10       relatively complex interplay between multiple genetic and environmental factors.  This model is

       11       illustrated in Figure 2-1. (From Part 2, Chapter 2)

       12                  Comparative data from animal and human cells and tissues suggest a strong qualitative

       13        similarity across species in response to dioxin-like chemicals.  This further supports the

       14        applicability to humans of the generalized model of early events in response to dioxin exposure.

       15        These biochemical and biological responses are sometimes considered adaptive and are often not

       16        considered adverse in and of themselves.  However, many of these biochemical changes are

       17        potentially on a continuum of the dose-response relationships which leads to adverse responses.

       18        At this time, caution must be used when describing these events as adaptive.

       19                  If, as we can infer from the evidence, TCDD and other dioxin-like compounds operate

      20        through these mechanisms, there are constraints on the possible models that can plausibly account

      21         for dioxin's biological effects and also on the assumptions used during the risk assessment

      22        process.  Mechanistic knowledge of dioxin action may also be useful in other ways.  For example,

      23        a further understanding of the ligand specificity and structure of the Ah receptor will likely assist

      24        in the identification of other chemicals to which humans are exposed that may either add to,

      25        synergize, or antagonize the toxicity of TCDD and other dioxin-like compounds.  Knowledge of

      26        genetic polymorphisms that influence TCDD responsiveness may also allow the identification of

      27        individuals at particular risk from exposure to dioxin.  In addition, knowledge of the biochemical

      28        pathways that are altered by dioxin-like compounds may help in the development of drugs that

      29        can prevent dioxin's adverse effects.

 

 

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       I         Figure 2-1.  Generalized Model for Early Molecular Events in Response to Dioxin

        2

        3                  As described in Part 2, Chapter 2, biochemical and genetic analyses of the mechanisms by

        4        which dioxin modulates particular genes have revealed the outline of a novel regulatory system

        5        whereby a chemical signal can alter cellular regulatory processes.  Future studies of dioxin action

        6        have the potential to provide additional new insights into mechanisms of mammalian gene

        7        regulation that are of relatively broad interest.  Additional perspectives on dioxin action can be

        8        found in several recent reviews (Bimbaum, 1994; Schecter, 1994; Hankinson, 1995; Schmidt and

        9        Bradfield,  1996; Rowlands and Gustafsson,  1997; Gasiewicz,  1997; Hahn, 1998; Denison et al.,

 

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        1         1998; Wilson and Safe, 1998).

        2                  The ability of TCDD and other dioxin-like compounds to modulate a number of

        3        biochemical parameters in a species-, tissue-, and temporal specific manner is well recognized.

        4        Despite the ever expanding list of these responses over the past 20 years and the elegant work on

        5        the molecular mechanisms mediating some of these, there still exists a considerable gap between

        6        our knowledge of these changes and the degree to which they are related to the more complex

        7        biological and toxic endpoints elicited by these chemicals. A framework for considering these

        8        responses in a mode-of action context is discussed later in this document.

        9                  TCDD-elicited activation of the Ah receptor has been clearly shown to mediate altered

       10        transcription of a number of genes, including several oncogenes and those encoding growth

      11         factors, receptors, hormones and drug metabolizing enzymes.   Figure 2-2 provides an illustrative

      12        list of gene products shown to be mediated by TCDD.  While this list is not meant to be

      13        exhaustive, if demonstrates the range of potential dioxin impacts.

      14

 

      15        Figure 2-2: Some of the Genes Whose Expression Is Altered By Exposure to TCDD

      16

      17                  As discussed in Volume 2, Chapter 2, it is possible that the TCDD-elicited alteration of

      18        activity of these genes may occur through a variety of mechanisms including signal transduction

      19        processes. These alterations in gene activity may be secondary to other biochemical events that

      20        may be directly regulated transcriptionally by the AhR.  Some of the changes may also occur by

      2l        post-transcriptional processes such as mMA stabilization and altered phosphorylation (Gaido et

      22        al.,  1992; Matsumura,  1994 ).  Thus, the molecular mechanisms by which many, if not most, of

      23        the biochemical processes discussed herein are altered by TCDD treatment remain to be

      24        determined.  Nevertheless, it is presumed, based on the cumulative evidence available, as

 

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         I         discussed earlier, that all of these processes are mediated by the binding of TCDD to the Ah

         2        receptor.  While the evidence for the involvement of the Ah receptor in all of these processes has

         3        not always been ascertained, structure-activity relationships, genetic data, and reports from the use

         4        of biological models like "knockout" mice which are lacking the Ah receptor (AhR-/-) are

         5        consistent with the involvement of the Ah receptor as the initial step leading to many of these

         6        biochemical alterations.  In fact, for every biochemical response that has been well studied, the

         7        data are consistent with the particular response being dependent on the Ah receptor.

         8                  The dioxin-elicited induction of certain drug metabolizing enzymes such as CYP1A1,

         9        CYP1A2, and CYP1B 1  is clearly one of the most sensitive responses observed in a variety of

       10        different animal species including humans, occurring at body burdens as low as 1-10 ng TCDD/kg

       11         in animals (See Part 2, Chapter 8).  These and other enzymes are responsible for the metabolism

       12        of a variety of exogenous and endogenous compounds.  Several lines of experimental evidence

       13        suggest that these enzymes may be responsible for either enhancing or protecting against

       14        (depending on the compounds and experimental system used) toxic effects of a variety of agents

       15        including known carcinogens as welt-as-endogenous substrates such as hormones. Several reports

       16        (Kadlubar et al., 1992; Esteller et al., 1997; Ambrosone et al., 1995; Kawajiri et al., 1993) provide

       17        evidence that human polymorphisms in CYPIA1  and CYPIA2 which result in higher levels of

       18        enzyme are associated with increased susceptibility to colorectal, endometrial breast, and lung

       19        tumors.  Also, exposure of AhR-deficient ("knockout") mice to Benzo[a]pyene (BaP) results in no

      20         tumor response, suggesting a key role for the Ah receptor, and perhaps, CYPIA1 and CYPIA2 in

      21         BaP carcinogenesis (Dertinger et al.,  1998; Shimizu et al., 2000). Modulation of these enzymes by

      22         dioxin may play a role in chemical carcinogenesis.  However, the exact relationship between the

      23         induction of these enzymes and any toxic endpoint observed following dioxin exposure has not

      24         been clearly established.

      25                  As with certain of the cytochrome P450 isozymes, there does not yet exist a precise

      26        understanding of the relationships existing between the alteration of specific biochemical

      27        processes and particular toxic responses observed in either experimental animals or humans

      28        exposed to the dioxins.  This is due predominantly to our incomplete understanding of the

      29        complex and coordinate molecular, biochemical, mid cellular interactions that regulate tissue

      30        processes during development and under not-real homeostatic conditions.  Nevertheless, a further

 

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        I         understanding of these processes and how TCDD may interfere with them remain important goals

        2        that would greatly assist in the risk characterization process.  In particular, knowledge of the

        3        causal association of these responses coupled with dose-response relationships may lead to a

        4        better understanding of sensitivity to various exposure levels of the dioxin-like compounds.

        5                  In contrast to what is known about the P450 isozymes, there exists some evidence from

        6        experimental animal data to indicate that the alteration of certain other biochemical events might

        7        have a more direct relationship to sensitive toxic responses observed following TCDD exposure.

        8        Some of these may be relevant to responses observed in humans, and further work in these areas is

        9        likely to lead to data that would assist in the risk characterization process.  For example, changes

       10       in epidermal growth factor (EGF) receptor have been observed in tissues from dioxin-exposed

       11       animals and humans (See Part 2, Chapters 3 and 6 ).  EGF and its receptor possess diverse

       12       functions relevant to cell transformation and tumorigenesis, and changes in these functions may

       13       be related to a number of dioxin-induced responses including neoplastic lesions, chloracne, and a

       14       variety of reproductive and developmental effects.  Likewise, the known abilit2,, of TCDD to

      15        directly or indirectly alter the levels and/or activity of other growth factors and hormones, such

      16        as estrogen, thyroid hormone, testosterone, gonadotropin-releasing hormone and their respective

      17        receptors, as well as enzymes involved in the control of the cell cycle (Safe, 1995 ), may affect

      18        growth patterns in cells/tissues leading to adverse consequences.  In fact, most of the effects that

      19        the dioxins produce at the cellular and tissues levels are due not to cell/tissue death but to altered

      20        growth patterns (Bimbaum, 1994 ).  Many of these may occur at critical times in development

      21         and/or maturation and thus may be irreversible.

      22                  From this brief discussion and that detailed in Part 2, Chapters 2 and 8, it seems clear that

      23        much work needs to be done to clarify the exact sequence and interrelations of those biochemical

      24        events altered by TCDD and how and at what point they might lead to irreversible biological

      25        consequences.  Nevertheless, it is important to recognize that many of the biochemical and

     26        biological changes observed are consistent with the notion that TCDD is a powerful growth

     27        dysregulator.  This notion may play a considerable role in the risk characterization process by

     28        providing a focus on those processes, such as development, reproduction and carcinogenesis,

     29        which are highly dependent on coordinate growth regulation.  Further understanding of these

     30        biochemical events in humans n-my provide useful biomarkers of exposure and responsiveness.

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        1         The use of these potential biomarkers may subsequently improve our understanding of the

        2        variation of responsiveness within an exposed population.

        3

        4        2.2 ADVERSE EFFECTS IN HUMANS AND ANIMALS

        5

        6        2.2.1 CANCER (Cross Reference, Volume 2: Chapters 6, 7 and 8)

        7

        8        2.2.1.1. Epidemiologic Studies

        9                  Since the last formal EPA review of the human data base relating to the carcinogenicity

       10        of TCDD and related compounds in 1988, a number of new follow-up mortality studies have

       11        been completed.  This body of information is described in Part 2, Chapter 7 of this assessment

       12        and has recently been published as part of an IARC Monograph (1997) and the ATSDR

       13        ToxProfile (ATSDR, 1999). Among the most important of these are the studies of 5,172 U.S.

       14        chemical manufacturing workers by Fingerhut et al. (1991a), Alyward (1996) and Steenland

       15        (1999); a study of 2,479 German workers involved in the production of phenoxy herbicides and

       16        chlorophenols by Becher et al. (1996, 1998) and by others in separate publications (Manz et

       17        al., 1991; Nagel et al., 1994; and Flesch-Janys et al, 1995,1998); a study of over 2,000 Dutch

       18        workers in two plants involved in the synthesis and formulation of phenoxy herbicides and

       19        chlorophenols (Bueno de Mesquita et al, 1993) and subsequent follow-up and expansion by

      20         Hooiveld et al, 1998); a smaller study of 247 workers involved in a chemical accident clean-up

      21         by Zober et al. (1990) and subsequent follow-up (Ott and Zober, 1996), and an intemational

      22         study of over 18,000 workers exposed to phenoxy herbicides and chlorophenols by Saracci et al.

      23        (1991) with subsequent follow-up and expansion by Kogevinas et al (1997).  Although

      24        uncertainty remains in interpreting these studies because not all potential confounders have been

      25        ruled out and coincident exposures to other carcinogens are likely, all provide support for an

      26        association between exposure to dioxin and related compounds and increased cancer mortality.

      27        One of the strengths of these studies is that each has some exposure information that permits an

      28        assessment of dose response.  Some of these data have, in fact, served as the basis for fitting the

      29        risk models in Chapter 8.  In addition, limited results have been presented on the non-

      30        occupational Seveso cohort (Bertazzi et al., 1993, 1997) and on women exposed to

 

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          1        chlorophenoxy herbicides, chlorophenols, and dioxins (Kogevinas et al., 1993).  While these two

          2        studies have methodologic shortcomings that are described in Chapter 7, they provide findings,

          3        particularly for exposure to women, that warrant additional follow-up.

          4                  Increased risk for all cancers combined was a consistent finding in the occupational

          5        cohort studies.  While the increase was generally tow (20-50%), it was highest in sub-cohorts

          6        with presumed heaviest exposure.  Positive dose-response trends in the German studies arid

          7        increased risk in the longer duration U.S. sub-cohort and the most heavily exposed Dutch

          8        workers support this view.

          9                  One of the earliest reported associations between exposure to dioxin-like compounds, in

         10        dioxin-contaminated phenoxy herbicides, and increased cancer risk involved an increase in soft

         11        tissue sarcomas (Hardell and Sandstrom, 1979; Eriksson et al., 1981; Hardell and Eriksson, 1988;

         12        Eriksson et al., 1990).  In this and other recent evaluations of the epidemiologic database, many

         13        of the earlier epidemiological studies that suggested an association with soft tissue sarcoma are

         14        criticized for a variety of reasons.  Arguments regarding selection bias, differential exposure

         15        misclassification, confounding, and chance in each individual study have been presented in the

         16        scientific literature which increase uncertainty around this association. Nonetheless, the

         17        incidence of soft tissue sarcoma is elevated in several of the most recent studies (refs), supporting

         18        the findings from previous studies.  The fact that similar results were obtained in independent

         19        studies of differing design and evaluating populations exposed to dioxin-like compounds under

        20        yawing conditions, along with the rarity of this tumor type, weighs in favor of a consistent and

         21         real association.

        22                  In addition to soft tissue sarcoma, other cancer sites have been associated with exposure

        23         to dioxin.  Excess respiratory cancer was noted by Fingerhut (1991), Zober (1994), and Manz

        24        (1991).  These results are also supported by significantly increased mortality from lung and liver

        25        cancers subsequent to the Japanese rice oil poisoning accident where exposure to high levels of

        26        PCDFs and PCBs occurred (Kuratsune et al., 1988).  Again, while smoking as a confounder

        27        cannot be totally eliminated as a potential explanation of the occupational studies results,

        28         analyses (Fingerhut, 1991; Ott and Zober, 1996) conducted to date suggest that smoking is not

        29        likely to explain the entire increase in lung cancer and may even suggest synergism between

        30        occupational exposure to dioxin and smoking.  These analyses have not been deemed entirely

 

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        1         satisfactory by some reviewers of the literature.  The question of confounding exposures, such as

        2        asbestos and other chemicals, in addition to smoking, has not been entirely ruled out and must be

        3        considered as potentially adding to the observed increases.  Although increases of cancer at other

        4        sites (e.g., non-Hodgkin's lymphoma, stomach cancer) have been reported (See Part 2, Chapter

        5        7a), the data for an association with exposure to dioxin-like chemicals are less compelling.

        6                  As mentioned above, both past and more recent human studies have focused on males.

        7        Although males comprise all the case-control studies arid the bulk of the cohort study analyses,

        8        animal and mechanism studies suggest that males and females might respond differently to

        9        TCDD. There are now, however, some limited data suggesting carcinogenic responses associated

       10        with dioxin exposure in females. The only reported female cohort with good TCDD exposure

       11         surrogate information was that of Manz et al. (1991), which had a borderline statistically

       12        significant increase in breast cancer.  While Saracci et al. (1991) did report reduced female breast

       13        and genital organ cancer mortality, this was based on few observed deaths and on chlorophenoxy

       14        herbicide, rather than TCDD, exposures.  In the later update and expansion of this cohort

       15        Kogevinas et al. (1997) provided evidence of a reversal of this deficit and produced a borderline

       16        significant excess risk of breast cancer in females.  Bertazzi et al. (1993, 1997, 1998) reported

       17        nonsignificant deficits of breast cancer and endometrial cancer in women living in geographical

       18        areas around Seveso contaminated by dioxin.  Although Kogevinas et al. (1993) saw an increase

       19        in cancer incidence among female workers most likely exposed to TCDD, no increase in breast

      20         cancer was observed in his small cohort. In sum, TCDD cancer experience for women may differ

      21        from that of men, but currently there are few data. Because both laboratory animal data and

      22        mechanistic inferences suggest that males and females may respond differently to the

      23        carcinogenic effects of dioxin-like chemicals, further data will be needed to address this question

      24        of differential response between sexes, especially to hormonally-mediated tumors. No

      25'        epidemiological data available to address the question of the potential impact of exposure to

      26        dioxin-like compounds on childhood cancers. However, recent studies of Brown et al. (1998)

      27        demonstrate that prenatal exposure of rats enhances their sensitivity as adults to chemical

      28        carcinogenesis.

      29                   Based on the analysis of the cancer epidemiology data as presented in Part 2, Chapters 7

      30        and 8, TCDD, and by inference, other dioxin-like compounds, are described as potentially

                                                                               30

 

 

 

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        1         multi-site carcinogens in more highly exposed human populations that have been studied,

        2        consisting primarily of adult males.  Although uncertainty remains, the cancer findings in the

        3        epidemiologic literature are generally consistent with results from studies of laboratory animals

        4        where dioxin-like compounds have clearly been identified as multi-site carcinogens.  In addition,

        5        the findings of increased risk at multiple sites appear to be plausible given what is known about

        6        mechanisms of dioxin action, and the fundamental level at which it appears to act in target

        7        tissues. While several studies exhibit a positive trend in dose-response and have been the subject

        8        of empirical risk modeling (Becher et al., 1998), the epidemiologic data alone provide little

        9        insight into the shape of the dose-response curve below the range of observation in these

       10       occupationally exposed populations.  This issue will be further discussed in Section 5.2.1.  The

       11       contribution of cancer epidemiology to overall cancer hazard and risk characterization is

       12       discussed in Section 6.

       13

       14       2.2.1.2. Animal Carcinogenicity (Cross reference, Part 2: Chapter 6 and 8)

       15                  An extensive data base on the carcinogenicity of dioxin and related compounds in

       16       laboratory studies exists and is described in detail in Chapter 6.  There is adequate evidence that

       17       2,3,7,8-TCDD is a carcinogen in laboratory animals based on long-term bioassays conducted in

       18       both sexes of rats and mice (U.S. EPA,  1985; Huff et al, 1991;Zeise et a1,1990; L&RC,1997).  All

       l 9       studies have produced positive results, leading to the conclusions that TCDD is a multistage

       20       carcinogen increasing the incidence of tumors at sites distant from the site of treatment and at

       21       doses well below tine maximum tolerated dose.  Since this issue was last reviewed by the Agency

      22        in 1988, TCDD has been shown to be a carcinogen in hamsters (Rao et al, 1988), which are

      23        relatively resistant to the lethal effects of TCDD.  Other data have also shown TCDD to be a

      24        liver carcinogen in the small fish, Medaka (Johnson et al., 1992).  Few attempts have been made

      25        to demonstrate the carcinogenicity of other dioxin-like compounds.  Other than a mixture of two

      26        isomers of hexachlorodibenzodioxin (HCDDs), which produced liver tumors in both sexes of rats

      27        and mice (NTP,  1980) when given by the gavage route, but not by the dermal route in Swiss

      28        mice (NTP, 1982) and a recent report (Rozman et al., 2000) attributing lung cancer in female rats

      29        to gavage exposures of 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin(HpCDD), neither the more

      30        highly chlorinated PCDDs/PCDFs nor the co-planar PCBs have been studied in long-term

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        1        animal cancer bioassays.   However, it is generally recognized that these compounds

        2        bioaccumulate and exhibit toxicities similar to TCDD and are, therefore, also likely to be

        3        carcinogens (U.S. EPA Science Advisory Board,' 1989).  The NTP currently has 4 (check status)

        4        congeners under test in cancer bioassays, alone and in combination.  These data should add

        5        significantly to our certainty regarding the carcinogenicity of these dioxin-like congeners when

        6        they are available.

         7                  In addition to the demonstration of TCDD as an animal carcinogen in long-term cancer

         8        bioassays, a number of dioxin-like PCDDs and PCDFs, as well as several PCBs, have also been

         9        demonstrated to be tumor promoters in two-stage (initiation-promotion) protocols in rodent liver,

       10        lung  and skin. These studies are described in some detail in Part 2, Chapter 6.  In that Chapter,

       11         TCDD is characterized as a non-genotoxic carcinogen since it is negative in most assays for

       12        DNA damaging potential, as a potent "promoter," and as a weak initiator or non-initiator in two

       13        stage initiation-promotion(I-P) models for liver and for skin.

       14                  The liver response is characterized by increases in altered hepatocellular foci (AHF)

       15        which are considered to be pre-neoplastic lesions since-increases in AHFs are associated with

       16        liver cancer in rodents. The results of the multiple I-P studies which are enumerated in Figure 6-

       17        X in Part 2, Chapter 6 have been interpreted as showing that AHF induced by TCDD are dose-

       18        dependent (Maronpot et al,  1993;Teegarden et al, 1999), are exposure-duration dependent

       19        (Dragan et al  1992; Teegarden et al,  1999; Walker et al, 2000), and reversible after cessation of

       20        treatment (Dragan et al, t992; Tritscher et al, 1995; and Walker et al, 2000).  Other studies

       21         indicate that other dioxin-like compounds have the ability to induce AHFs.  These studies show

       22        that the compounds demonstrate a rank-order of potency for AHF induction which is similar to

       23        that for CYP1A1  (Flodstrom and Ahlborg,  1992; Waem et al,  1991; and Schrenk et al, 1994).

       24        Non-ortho substituted, dioxin-like PCBs also induce the development of AHF according to their

       25        potency to induce CYP 1Al  (Hemming et al., 1993; van der Plas, 1999).  It is interesting to note

       26        that liver I-P studies carried out in ovariectomized rats demonstrate the influence that the intact

       27        hormonal system has on AHF development.  AHF are significantly reduced in ovariectomized

       28        female livers ( Graham et al., 1988; Lucier et al., 1991).

       29                  I-P studies on skin have demonstrated that TCDD is a potent tumor promoter in mouse

       30        skin as well as rat liver.  Early studies demonstrated that TCDD is at least two orders of

 

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        1         magnitude more potent than the "classic" promoter tetradecanoyl phorbol acetate (TPA) (Poland

        2        et al., 1982); that TCDD skin tumor promotion is Ah receptor dependent (Poland and Knutsen,

        3         1982); that TCDD had weak or no initiating activity in the skin system (DiGiovanni et al., 1977)

        4        and that TCDD's induction of drug metabolizing enzymes is associated with both metabolic

        5        activation as well as deactivation as described by Lucier et a1. (1979).  More recent studies show

        6        that the skin tumor promoting potencies of several dioxin-like compounds reflect relative Ah

        7        receptor binding and pharmacokinetic parameters (Hebert et al., 1990).

        S                  While few I-P studies have demonstrated lung tumors in rats or mice, the study of Clark

        9        et al. (1991) is particularly significant because of its use of ovariectomized animals.  In contrast

       10       to liver tumor promotion, lung tumors were seen only in initiated (diethylnitrosamine (DEN)),

       11       TCDD treated rats. No tumors were seen in DEN only, TCDD only, control, or DEN/TCDD

       12       intact rats.  Liver tumors are ovary dependent but ovaries appear to protect against TCDD-

       13       mediated tumor promotion in rat lung. Perhaps, use of transgenic animal models will allow

       14       further understanding of the complex interaction of factors associated with carcinogenesis in

       15       rodents as well as, presumably, in man.  Several such systems are being evaluated (Eastin et al.,

       16       1998; van Birgelen et al., 1999; Duston, 2000).

       17                  Several potential mechanisms for TCDD carcinogenicity are discussed in Part 2, Chapter

       18        6.  These include: indirect DNA damage; endocrine disruption/growth dysregulation/altered

       19        signal transduction; and cell replication/apoptosis leading to tumor promotion. All of these are

      20         biologically plausible as contributors to the carcinogenic process and none are mutually

       21        exclusive.  Several biologically-based models which encompass many of these activities are

       22        described in Part 2, Chapter 8.  Further work will be needed to elucidate a detailed mechanistic

       23        model for any particular carcinogenic response in animals or in humans.  Despite this lack of a

      24        defined mechanism at the molecular level, there is a general consensus that TCDD and related

      25        compounds are receptor-mediated carcinogens in that 1) interaction with the Ah receptor is a

      26        necessary early event; 2)TCDD modifies a number of receptor and hormone systems involved in

      27        cell growth and differentiation such as the epidermal growth factor receptor and estrogen

      28        receptor; and 3) sex hormones exert a profound influence on the carcinogenic action of TCDD.

      29

      30        2.2.1.3.  Other Data Related to Carcinogenesis

 

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        1                  Despite the relatively large number of bioassays on TCDD, the study of Kociba et al.

        2        (1978) and those of the NTP (1982), because of their multiple dose groups and wide dose range,

        3        continue to be the focus of dose-response modeling efforts and of additional review.  Goodman

        4        and Sauer (1992) reported a re-evaluation of the female rat liver tumors in the Kociba study

        5        using the latest pathology criteria for such lesions.  The review confirmed only approximately

        6        one-third of the tumors of the previous review (Squire, 1980).  While this finding did not change

        7        the determination of carcinogenic hazard since TCDD induced tumors in multiple sites in this

        8        study, it does have an effect on evaluation of dose-response and on estimates of risk at low doses.

        9        These issues will be discussed in a later section of this document.

       10                  One of the more intriguing findings in the Kociba bioassay was reduced tumor incidences

       11        of the pituitary, uterus, mammary gland, pancreas, and adrenals in exposed female rats as

       12        compared to controls (Kociba, 1978).  While these findings, coupled with evaluation of

       13        epidemiologic data, have led some authors to conclude that dioxin possesses "anticarcinogenic"

       14        activity (Kayajanian, 1997; Kayajanian, 1999), it should be noted that, in experimental studies,

       15        with the exception of the mammary gland tumors, the decreased-incidence of tumors is

       16        associated with significant weight loss in these rats.  Examination of the data from the National

       17        Toxicology Program also demonstrates a significant decrease in these tumor types when there is

       18        a concomitant weight loss in the rodents, regardless of the chemical administered (Haseman and

       19        Johnson, 1996).  Because the mechanism of the decreases in the tumors is unknown,

      20         extrapolation of these effects to humans is premature.   In considering overall risk, one must take

      21         into account the range of doses to target organs and hormonal state to obtain a complete picture.

      22         It is unlikely, however, that such data will be available to argue that dioxin exposure provides a

      23         net benefit to human health.

      24

      25        2.2.1.4 Cancer Hazard Characterization

      26                  TCDD, CDDs, CDFs and dioxin-like PCBs are a class of well studied compounds whose

      27        human cancer potential is supported by a large database including limited epidemiological

      28        support, unequivocal animal carcinogenesis, and biologic plausibility based on mode-of-action

      29        data. In  1985, EPA classified TCDD and related compounds as "probable" human carcinogens

      30        based on the available data. During the intervening years, the data base relating to the

 

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        1        carcinogenicity of dioxin and related compounds has grown and strengthened considerably.  In

        2        addition, EPA guidance for carcinogen risk assessment has evolved (EPA, 1996).  Under EPA's

                        3        current approach, TCDD is best characterized as a  "human carcinogen.   This means that, based

        4        on the weight of all of the evidence (human, animal, mode-of-action), TCDD meets the stringent

        5        criteria that allows EPA and the scientific community to accept a causal relationship between

                      6        TCDD exposure and cancer hazard.  The guidance suggests that "human carcinogen” is an

        7        appropriate descriptor of carcinogenic potential when there is an absence of conclusive

        8        epidemiologic evidence to clearly establish a cause and effect relationship between human

        9        exposure and cancer, but there is compelling carcinogenicity data in animals and mechanistic

       10       information in animals and humans demonstrating similar modes of carcinogenic action.  The

       1 l       "human carcinogen" descriptor is suggested for TCDD since all of the following conditions are

       12        met:

       13                  -  occupational epidemiologic studies show an association between TCDD exposure and

       14                      increases in cancer at all sites, in lung cancer, and, perhaps, at other sites, but the data

     _-5                      are insufficient on their own to demonstrate a causal association;

       16                  -  there is extensive carcinogenicity in both sexes of multiple species of animals at

       17                      multiple sites;

       18

       19

      20                  -  there is general agreement that the mode of TCDD's carcinogenicity is Ah receptor

      21                      dependent and proceeds through modification of the action of a number of receptor

      22                      and hormone systems involved in cell growth and differentiation such as the epidermal

      23                      growth factor receptor and estrogen receptor; and

      24                  -  equivalent body burdens in animals and in human populations expressing an

      25                      association between exposure to TCDD and cancer, and the determination of active Ah

      26                      receptor and dioxin responsive elements in the general human population.  There is no

      27                      reason to believe that these events would not occur in the occupational cohorts studied.

      28

      29                  Other dioxin-like compounds are characterized as "likely" human carcinogens primarily

      30        because of the lack of epidemiological evidence associated with their carcinogenicity, although

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        I         the inference based on toxicity equivalence is strong that they would behave in humans as TCDD

        2        does. Other factors, such as the lack of congener specific chronic bioassays also support this

        3        characterization.  For each congener, the degree of certainty is dependent on the available

        4        congener specific data and its consistency with the generalized mode-of-action which underpins

        5        toxicity equivalence for TCDD and related compounds.  Based on this logic, all complex

        6        environmental mixtures of TCDD and dioxin-like compounds would be characterized as "likely"

        7        carcinogens, but the degree of certainty of the cancer hazard would be dependent on the major

        8        constituents of the mixture.  For instance, the hazard potential, although still considered "likely,"

        9        would be characterized differently for a mixture whose TEQ was dominated by OCDD as

      10        compared to one which was dominated by pentaCDF.

      11

       1'_        2.2.2  REPRODUCTIVE AND DEVELOPMENTAL EFFECTS

       13                  Several sections of this reassessment (Part 2, Chapter 5 and Chapter 7b) have focused on

       14        the variety of effects that dioxin and dioxin-like agents can have on human reproductive health

      15        and development.  Emphasis in each of these chapters has been placed on the discussion of the

      16        more recent reports of the impact of dioxin-like compounds on reproduction and development.

      17        These have been put into context with previous reviews of the literature applicable in risk

      18        assessment (Hatch,  1984; Sweeney, 1994; Kimmel,  1988) to develop a profile of the potential for

      19        dioxin and dioxin-like agents to cause reproductive or developmental toxicity based on the

      20        available literature.  An earlier version of the literature review and discussion contained in Part 2,

      21         Chapter 5 has been previously published (Peterson et al., 1993).

      22                  The origin of concems regarding a potential link between exposure to chlorinated dioxins

      23        and adverse developmental events can be traced to early animal studies reporting increased

      24        incidence of developmental abnormalities in rats and mice exposed early in gestation to 2,4,5-

      25        Trichlorophenol (2,4,5-T) (Courtney and Moore,  1971).  2,4,5-T is a herbicide that contains

      26        dioxin and related compounds as impurities.  Its use was banned in the late 1970's but exposure

      27        to human populations continued as a result of past production, use, and disposal.

      28

      29        2.2.2.1  Human

      30                  The literature base with regard to potential human effects is detailed in Part 2, Chapter

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         I         7b.  In general, there is little epidemiological evidence that makes a direct association between

         2        exposure to TCDD or other dioxin-like compounds and effects on human reproduction or

         3        development. One effect that may illustrate this relationship is the altered sex ratio (increased

         4        females) seen ill the 6 years after the Seveso accident (Mocarelli et al., 1996).  Other sites have

         5        been examined for this effect of TCDD exposure with mixed results but with smaller numbers of

         6        offsp_5.ng (refs.)  Continued evaluation of the Seveso, Italy, population may provide other

         7        indications of impacts on reproduction and development but, for now, such data are very limited

         8        and further research is needed.   Positive human data on developmental effects of dioxin-like

         9        compounds are limited to a few studies of populations exposed to a complex mixture of

       10        potentially toxic compounds (e.g. developmental studies from the Netherlands and effects of

       11         ingestion of contaminated rice oil in Japan (Yusho) and Taiwan (Yu-Cheng)).   In the latter

       12        studies, however, all four manifestations of developmental toxicity (reduced viability, structural

       13        alterations, growth retardation and functional alterations) have been observed to some degree,

       14        following exposure to dioxin-like compounds as well as other agents. Data from the Dutch

       15        cohort of children exposed to PCBs and dioxin-like compounds (Huisman et al., 1995a,b;

       16        Koopman-Esseboom et al., 1994a-c; 1995a,b; I996; Pluim et al.,  1992, 1993, 1994; Weisglas-

       17        Kuperus et al., 1995; Patandin et al., 1998; Patandin et a1., 1999) suggest impacts of background

       18        levels of dioxin and related compounds on neurobehavioral outcomes, thyroid function, and liver

       19        enzymes (AST and ALT). While these effects can not be attributed solely to dioxin and related

      20         compounds, several associations suggest that these are, in fact, likely to be Ah-mediated effects.

      21         Likewise, it is highly likely that the developmental effects in human infants exposed to a

      22         complex mixture of PCBs, PCDFs, and polychlorinated quaterphenyls (PCQs) in the Yusho and

      23         Yu-Cheng poisoning episodes may have been caused by the combined exposure to those PCB

      24         and PCDF congeners that are Ah-receptor agonists (Lu and Wong, 1984; Kuratsune, 1989;

      25         Rogan, 1989).  However, it is not possible to determine the relative contributions of individual

      26         chemicals to the observed effects.   The incidents at Yusho and Yu-Cheng resulted in increased

      27         perinatal mortality and low birthweight in infants bom to women who had been exposed.  Rocker

      28         bottom heal was observed in Yusho infants, and functional abnormalities have been reported in

      29         Yu-Cheng children. Not all the effects that were seen are attributable only to dioxin-like

      30         compounds.  The similarity of effects observed in human infants prenatally exposed to this

 

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        1        complex mixture with those reported in adult monkeys exposed only to TCDD suggests that at

        2        least some of the effects in the Yusho and Yu-Cheng children are due to the TCDD-like

        3        congeners in the contaminated rice oil ingested by the mothers of these children.  The similar

        4        responses include a clustering of effects in organs derived from the ectodermal germ layer,

        5        referred to as ectodermal dysplasia, including effects on the skin, nails, and Meibomian glands;

        6        developmental and psychomotor delay during developmental and cognitive tests (Chen et al.,

        7        1992).  Some investigators believe that, because all of these effects ill the Yusho and Yu-Cheng

        8        cohorts do not correlate with TEQ, some of the effects are exclusively due to non-dioxin-like

        9        PCBs or a combination of all the congeners, it is still not clear to what extent there is an

       10       association between overt matemal toxicity and embryo/fetal toxicity in humans.

       11                  Of particular interest is the common developmental origin (ectodermal layer) of many of

       12        the organs and tissues that are affected in the human.  An ectodermal dysplasia syndrome has

       13        been clearly associated with the Yusho and Yu-Cheng episodes, involving hyperpigmentation,

       14        deformation of the fingemails and toenails, conjunctivitis, gingival hyperplasia and abnormalities

       15        of the teeth.  An investigation of dioxin exposure and tooth development was done in Finnish

       16        children as a result of studies of dental effects in dioxin-exposed rats, mice, and nonhuman

       17        primates (Chapter 5), and in PCB-exposed children (Rogan et al., 1988).   The Finnish

       18        investigators examined enamel hypomineralization of permanent first molars in 6-7 year old

       19        children (Alaluusua et al., 1996; Alaluusua et al., 1999).  The length of time which infants breast

       20        fed was not significantly associated with either mineralization changes, or with TEQ levels in the

       21        breast milk.  However, when the levels and length of breast feeding were combined in an overall

       22        score, a statistically significant association was observed ® = 0.3, p = 0.003, regression analysis).

       23        These data are discussed further in Part 2, Chapter 7b. The developmental effects that can be

       24        associated with the nervous system are also consistent with this pattern of impacts on tissues of

       25        ectodermal origin, since the nervous system is of ectodermal origin.  These data are limited but

       26        are discussed in Part 2, Chapter 7b.

       27                  Other investigations into non-cancer effects of human exposure to dioxin have provided

      28        human data on TCDD - induced changes in circulating reproductive hormones.  This was one of

      29        the effects judged as having a positive relationship with exposure to TCDD in Part 2, Chapter 7b.

      30        Levels of reproductive hormones have been measured with respect to exposure to 2,3,7,8-TCDD

 

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        l         in three cross-sectional medical studies.  Testosterone, LH, and FSH were measured in TCP and

        2        2,4,5-T production workers (Egeland et al., 1994), in Army Vietnam veterans (Centers for

        3        Disease Control Vietnam Experience Study, 1988d), and in Air Force personnel, known as

        4        "Ranch Hands," who handled and/or sprayed Agent Orange during the Vietnam War (Roegner et

        5        al., 1991; Grubbs et al., 1995).  The risk of abnormally low testosterone was two to four times

        6        higher in exposed workers with serum 2,3,7,8-TCDD levels above 20 pg/g than in unexposed

        7        referents (Egeland et al., 1994).  In both the 1987 and 1992 examinations, mean testosterone

        8        concentrations were slightly, but not significantly higher in Ranch Hands (Roegner et al., 1991;

        9        Grubbs et al., 1995).  FSH and LH concentrations were no different between the exposed and

       10       comparison groups.  No significant associations were found between Vietnam experience and

       11       altered reproductive hormone levels (Centers for Disease Control Vietnam Experience Study,

       12       1988d).  Only the NIOSH study found an association between serum 2,3,7,8-TCDD level and

       13       increases in serum LH.

       14                  The findings of the NIOSH and Ranch Hand studies are plausible given the

       15        pharmacological and toxicological properties of 2,3,7,8-TCDD in animal models which are

       16        discussed in Part 2, Chapters 5 and 7.  One plausible mechanism responsible for the effects of

       17        dioxins may involve their ability to influence hormone receptors.  The Ah receptor, to which

       18        2,3,7,8-TCDD binds, and the hormone receptors are signaling pathways which regulate

       19        homoeostatic processes.  These signaling pathways are integrated at the cellular level and there is

       20        considerable "cross-talk" between these pathways.  For example, studies suggest that 2,3,7,8-

       21        TCDD modulates the concentrations of numerous hormones and/or their receptors, including

       22        estrogen (Retakes and Safe, 1988; Retakes et al., 1987), progesterone (Retakes et al., 1987),

       23        glucocorticoid (Ryan et al, 1989) and thyroid hormones (Gorski and Rozman, 1987).

       24                  In summary, the results from both the NIOSH and Ranch Hand studies are limited by the

       25        cross-sectional nature of the data and the type of clinical assessments conducted.  However, the

       26        available data provide evidence that alterations in human male reproductive hormone levels are

       27        associated with serum 2,3,7,8-TCDD.

       28

       29        2.2.2.2 Experimental Animal

      30                  The extensive experimental animal data base with respect to reproductive and

 

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        1         developmental toxicity of dioxin and the dioxin-related agents has been discussed in Part 2,

        2        Chapter 5  Dioxin exposure has been observed to result in both male and female reproductive

        3        effects, as well as effects on development.  These latter effects are among the most responsive

        4        health endpoints to dioxin exposure (See Part 2, Chapter 8).  In. general, the prenatal and

        5        developing postnatal animal is more sensitive to the effects of dioxin than the adult.  In several

        6        instances ( e.g fetotoxicity in hamsters, rats, mice, and guinea pigs), the large species differences

        7        seen in acute toxicity are greatly reduced when developing animals are evaluated.  Most of the

        8        data reviewed is from studies of six genera of laboratory animals.  While much of the data comes

        9        from animals exposed only to TCDD, more recent studies of animals exposed to mixtures of

       10        PCDD;PCDF isomers provide results which are consistent with the studies of TCDD alone

       1l         (refs).

       12

       13

       14

       15         Developmental Toxicity

 

       16                  Dioxin exposure results in a wide variety of developmental effects and these are observed

       17        in three different vertebrate classes and in several species within each class.  All four of the

       18        manifestations of developmental toxicity have been observed following exposure to dioxin,

       19        including reduced viability, structural alterations, growth retardation and functional alterations.

      20        As summarized previously (Peterson et al., 1993), increased prenatal mortality (rat and monkey),

      21         functional alterations in leaming and sexual behavior (rat and monkey), and changes in the

      22        development of the reproductive system (rat) occur at the lowest exposure levels (See also Part 2,

      23        Chapter 8).

      24                  Dioxin exposure results in reduced prenatal or postnatal viability in virtually every

      25        species in which it has been tested.  Previously, increased prenatal mortality appeared to be

      26        observed only at exposures that also resulted in matemal toxicity.  However, the studies of Olson

      27        and McGarrigle (1991  Gary-Cheek this date in your chapt, and ref.-19907) in the hamster and

      28        Schantz et al. (1989) in the monkey were suggestive that this was not the case in all species.

      29        Although the data from these two studies were limited, prenatal death was observed in cases

      30        where no matemal toxicity was evident.  In the rat, Peterson's laboratory, (Bjerke et al.,1994a,

 

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        1         1994b, Roman et al.,  1995) reported increased prenatal death following a single exposure to

        2        TCDD during gestation which did not cause matemal toxicity, and Gray et al. (1995a) observed a

        3        decrease in postnatal survival under a similar exposure regimen. While identifying the presence

        4        or absence of matemal toxicity may be instructive as to the specific origin of the reduced prenatal

        5        viability, it does not alter the fact that pre- and postnatal death were observed.  In either case, the

        6        Agency considers these effects as being indicators of developmental toxicity in response to the

        7        exposure (U.S. EPA, 1991)

        8                  Some of the most striking findings regarding dioxin exposure relate to the effects on the

        9        developing reproductive system.  Only a single, low-level exposure to TCDD during gestation is

       10        required to initiate these developmental alterations.  Mably et al. (1992 a, b, c) originally

       11         reported that a single exposure of the Holtzman matemal rat to as low as 0.064 ug/kg could alter

       12        normal sexual development in the male offspring.  A dose of 0.064 ug/kg in these studies results

       13        in a body burden in the matemal animal of 64 ng/kg during critical windows in development.

       14        More recently, these findings of altered normal sexual development have been further defined

       15        (Bjerke et al,  1994; Gray et al., 1995a;; Roman et al.,  1995), as well as extended to females and

       16        another strain and species (hamster) (Gray et al, 1995b).  In general, the findings of these later

       17        studies have produced qualitatively similar results that define a significant effect of dioxin on the

       18        developing reproductive system.

       19                  In the developing male rat, TCDD exposure during the prenatal and lactational periods

      20        results in the delay of the onset of puberty as measured by age at preputial separation.  There is a

      21         reduction in testis weight, sperm parameters, and sex accessory gland weights.  In the mature

      22        male exposed during the prenatal and lactational periods, there is an alteration of normal sexual

      23        behavior and reproductive function.  Males exposed to TCDD during gestation are

      24        demasculinized.  Feminization of male sexual behavior and a reduction in the number of

      25        implants in females mated with exposed males have also been reported, although these effects

      26        have not been consistently found.  These effects do not appear to be related to reductions in

      27        circulating androgens, which were shown in the most recent studies to be normal.  Most of these

      28        effects occur in a dose-related fashion, some occurring at 0.05 ug/kg and 0.064 ug/'kg, the lowest

29                TCDD doses tested (Mably et al.  1992c; Gray et al.  1997a).

      30                  In the developing female rat, Gray and Ostby (1995) have demonstrated altered sexual

 

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       1        differentiation in both the Long Evans and Holtzman strains.  The effects observed depended on

       2        the timing of exposure.  Exposure during early organogenesis altered the cyclicity, reduced

       3        ovarian weight and shortened the reproductive life span.  Exposure later in organogenesis

       4        resulted in slightly lowered ovarian weight, structural alterations of the genitalia and a slight

       5        delay in puberty.  However, cyclicity and fertility were not affected with the later exposure.  The

       6        most sensitive dose-dependent effects of TCDD in the female rat were structural alterations of

       7        the genitalia that occurred at 0.20 .ug TCDD/kg administered to the dam (Gray et al.  1997b).

       8                  As described above, studies demonstrating adverse health effects from prenatal exposures

       9        often involved a single dose administered at a discrete time during pregnancy.  The production of

      10       prenatal effects at a given dose appears to require exposure during critical times in fetal

      11       development.  This concept is well supported by a recent report (Hurst et al., 1998 Need full

      12       paper citation) which demonstrated the same incidence of adverse effects in rat pups bom to

      13       dams with a single exposure of 0.2 ug TCDD/kgBW on gestation day 15 (GD  15) versus  1.0 _g

      14       TCDD/kgBW on gestation day 8 (GD 8).  Both of these experimental paradigms result in the

      15       same fetal tissue concentrations and body burdens during the critical window of sensitivity.-For

      16       example, exposure to 0.2 ug TCDD/kgBW on GD 15 results in 13.2 pg TCDD/g fetal tissue on

      17       GD 16; exposure to  1.0 ug TCDD/kgBW on gestation GD 8 resulted in 15.3 pg TCDD/g fetus on

      18       GD 16.  This study demonstrates the appropriateness of the use of body burden to describe the

      19       effects of TCDD when comparing different exposure regimens.   The uncertainties introduced

      20       when trying to compare studies with steady-state body burdens with single dose studies may

      21        make it difficult to determine a lowest effective dose.  Application of pharmacokinetics models,

      22       described earlier in Parts  1 and 2, to estimate body burdens at the critical time of development is

      23       expected to be a sound method for relating chronic background exposures to the results obtained

      24       from single-dose studies.

      25                  Structural malformations, particularly cleft palate and hydronephrosis, occur in mice

      26        administered doses of TCDD.  The findings, while not representative of the most sensitive

      27        developmental endpoints, indicate that exposure during the critical period of organogenesis can

      28        affect the processes involved in normal tissue formation.  The TCDD-sensitive events appear to

      29        require the Ah receptor.  Mouse strains that produce Ah receptors with relatively high-affinity for

      30        TCDD respond to lower doses than strains with relatively low-affinity receptors.  Moreover,

 

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        1        congeners with a greater affinity for the .Ah receptor are more developmentally toxic than those

       2        with a lower affinity.  This is consistent with the rank-ordering of toxic potency based on

        3        affinity for the receptor as discussed in Part 2, Chapter 9.

        4

        5        Adult Female Reproductive Toxicity

        6                  The primary effects of TCDD on female reproduction appear to be decreased fertility,

        7        inability to maintain pregnancy for the full gestational period, and in the rat, decreased litter size.

        8        In some studies of rats and of primates, signs of ovarian dysfunction such as anovulation and

        9        suppression of the estrous cycle have been reported (Kociba et al., 1976; Barsotti et al., 1979;

       10        Allen et al, 1979; Li et al.,  1995a,  1995b).

       11

       12         Adult Male Reproductive Toxicity

       13                  TCDD and related compounds decrease testis and accessory sex organ weights, cause

       14        abnormal testicular morphology, decrease spermatogenesis, and reduce fertility when given to

       15        adult animals in doses sufficient to reduce feed intake and/or body weight.  In the testis of these

       16        different species, TCDD effects on spermatogenesis are characterized by loss of germ ceils, the

       17        appearance of degenerating spermatocytes and mature spermatozoa within the lumens of

       18        seminiferous tubules, and a reduction in the number of tubules containing mature spermatozoa

       19        (Allen and Lalich, 1962; Allen and Carstens, 1967; McConnell et al., 1978; Chahoud et al.,

      20        1989).  This suppression of spermatogenesis is not a highly sensitive effect when TCDD is

      21         administered to postweanling animals, since all exposure of 1 _g/kg/day over a period of weeks

      22        appears to be required to result in these effects.

      23

      24        2.2.2.3 Other Data Related to Developmental and Reproductive Effects

      25        Endometriosis

      26                  The association of dioxin with endometriosis was first reported in a study of Rhesus

      27        monkeys which had been exposed for four years to dioxin in their feed and then held for an

      28        additional ten years.  There was a dose related increase in both the incidence and severity of

      29        endometriosis in the exposed monkeys as compared to controls.  Follow-up on this group of

      30        monkeys revealed a clear association with the total TEQ.  A study in which Rhesus monkeys

 

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         I         were exposed to PCBs for 6 years(?) and then held for one year(?) longer failed to show any

        2        enhanced ircidence of endometriosis.  However, many of these monkeys were no longer cycling,

        3        and the time may not have been adequate to develop the response.  In the TCDD monkey study,

        4        it took 7 years before the first endometriosis was noted.  A recent study in Cynomolgus monkeys

        5        has shown promotion of surgically induced endometriosis by TCDD within one year after

        6        surgery.  Studies using rodents models for surgically' induced endometriosis have also shown that

        7        ability of TCDD to promote the lesions in a dose/related manner. This response takes at least two

        8        months to be detected. Another study in mice which failed to detect dioxin-promotion of

        9        surgically-induced endometriosis only held the mice for one month, not long enough to detect a

       10        response.   Prenatal exposure to mice also enhanced the sensitivity of the offspring to the

       11         promotion of surgically induced endometriosis by TCDD.  This response appears to be Ah

       12        receptor mediated as demonstrated in a study using the mouse model for endometriosis, in which

       13        Ah receptor ligands were able to promote the lesions, while non-Ah ligands, including a non-

       14        dioxin-like PCB, had no effect on surgically induced endometriosis.  Dioxin has also been shown

       15        to result in endometriosis in human endometrial tissue implanted in nude mice.

       16                  Data on the relationship of dioxins to endometriosis in people is intriguing, but

       17        preliminary'.  Studies in the early 1990s suggested that women with higher levels of persistent

       18        organochlorines were at increased risk for endometriosis.  This was followed by the observation

       19        that Belgian women, who have the highest levels of dioxins in their background population, had

      20        higher incidences of endometriosis than reported from other populations.  A study from Israel

      21         then demonstrated that there was a correlation between detectable TCDD in women with

      22        surgically confirmed endometriosis, in comparison to those with no endometriosis.  Recent

      23        studies from Belgium have indicated that women with higher body burdens, based on serum

      24        TEQ determinations, are at greater risk for endometriosis.  No association was seen with total

      25        PCBs in this study.  A small study in the United States, which did not involved surgically

      26        confirmed endometriosis, saw no association between TCDD and endometriosis.  Likewise, a

      27        study in Canada saw no association between total PCBs mid endometriosis.  The negative

      28        association with total PCBs is not surprising since the rodent studies have indicated that this

      29        response is .Ah receptor mediated.  Preliminary results from Seveso suggest a higher incidence of

      30        endometriosis in the women from the two highly exposed zones (A and B) as compared to the

 

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        1        background incidence in Italy.

                _..         The animal results lend biological plausibility to the epidemiology findings.

        3        Endometriosis is not only an endocrine disorder, but is also associated with immune system

        4        alterations.  Dioxins are known to be potent modulators of the immune system, as well as

        5        affecting estrogen homeostasis.  Further studies are clearly needed to provide additional support

        6        to this association of endometriosis and dioxins, as well as demonstrate causality.

        7        Androgenic Deficiency

        8                  The effects of TCDD on the male reproductive system when exposure occurs in

        9        adulthood m-e believed to be due in part to an androgenic deficiency.  This deficiency is

       10        characterized in adult rats by decreased plasma testosterone and DHT concentrations, unaltered

       11         plasma LH concentrations, and unchanged plasma clearance of androgens and LH (Moore et al.,

       12         1985, 1989; Mebus et al., 1987; Moore and Peterson, 1988; Bookstaff et al., 1990a).  The cause

       13        of the androgenic deficiency was believed to be due to decreased testicular responsiveness to LH

       14        and increased pituitary responsiveness to feedback inhibition by androgens and estrogens (Moore

       15        et al., 1989, 199!; Bookstall et al., 1990a,b; Kleeman et al., 1990).  The single dose used in some

       16        of those earlier studies (15 ugTCDD/kgBW) is now known to effect Leydig cells (Johnson et al.,

       17         1994).

       18

       19        2.2.2.4  Developmental and Reproductive Effects Hazard Characterization

       20                    There is limited direct evidence addressing the issues of how or at what levels humans

      21         will begin to respond to dioxin-like compounds with adverse impacts on development or

      22        reproductive function. The series of published Dutch studies suggest that pre- and early post-

      23        natal exposures to PCBs and other dioxin-like compounds may impact developmental milestones

      24        at levels at or near current average human background exposures.  While it is unclear whether

      25        these measured responses indicate a clearly adverse impact, if humans respond to TCDD

      26        similarly to animals in laboratm3, studies, there are indications that exposures at relatively low

      27        levels might cause developmental effects and at higher exposure levels might cause reproductive

      28        effects.  There is especially good evidence for effects on the fetus from prenatal exposure.  The

      29        Yusho and Yu-Cheng poisoning incidents are clear demonstrations that dioxin-like compounds

      30        can produce a variety of mild to severe developmental effects in humans that resemble the effects

 

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        1        of exposure to dioxins and dioxin-like compounds in animals,   Humans do not appear to be

        2        particularly sensitive or insensitive to effects of dioxin exposure in comparison to other animals.

        3        Therefore it is reasonable to assume that human responsiveness would lie across the middle

        4        ranges of observed responses.  This still does not address the issues surrounding the potentially

        5        different responses humans (or animals) might have to the more complex and variable

        6        environmental mixtures of dioxin-like compounds.

        7                  TCDD and related compounds have reproductive and developmental toxicity potential in

        8        a broad range of wildlife, domestic and laboratory animals. Many of the effects have been shown

        9        to be TCDD dose-related.  The effects on perinatal viability and male reproductive development

       10        are among the most sensitive effects reported, occurring at a single prenatal exposure range of as

 

       11         little as 0.(/5-0.075 =g/kg, resulting in calculated fetal tissue concentrations of 3-4 ng/kg.  In

       12        these studies, effects were often observed at the lowest exposure level tested, thus a no-observed

       13        adverse effect level (NOAEL) has not been established for several of these endpoints. In general,

       14        the structure-activity results are consistent with an Ah receptor-mediated mechanism for the

       15        developmental effects that are observed in the low dose range.  The structure-activity relationship

       16        in laboratory mammals appears to be similar to that for Ah receptor binding.  This is especially

       17        the case with cleft palate in the mouse.

       18                  It is assumed that the responses observed in animal studies are indicative of the potential

       19        for reproductive and developmental toxicity in humans.  This is an established assumption in the

 

       20        risk assessment process for developmental toxicity (U.S. EPA,  1991b).  It is supported by the

       21         number of animal species and strains in which effects have been observed.  The limited human

       22        data are consistent with an effect following exposure to TCDD or TCDD-like agents.  In

       23        addition, the phylogenetic conservation of the structure and function of the Ah receptor also

       24        increases our confidence that these effects may occur in humans.

       25                  While there is evidence in experimental animals that exposure to dioxin-like chemicals

       26        during development produces neurobehavioral effects, the situation in humans is more complex.

       27        Studies in humans demonstrate associations between dioxin exposure and alterations in

       28        neurological development.  These same studies often show similar associations between

       29        exposure to non-dioxin-like PCBs and these same effects.  Based on the human studies, it is

       30        possible that the alterations in neurological development are due to an interaction between the

 

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        1         dioxins and the non-dioxin-like PCBs.  At present there is limited data which defines the roles of

        2        the dioxins vs the non-dioxin-like PCBs in these effects on neurological development.

        3                  In general, the structure-activity results on dioxin-like compounds are consistent with an

        4        Ah receptor-mediated mechanism for many of the developmental effects that are observed.  The

        5        structure-activity relationship in laboratory mammals appears to be similar to that for Ah

        6        receptor binding.  This is especially the case with cleft palate in the mouse.  However, a direct

        7        relationship with Ah binding is less clear for other effects, including those involving the nervous

8               system.

9                

        9        2.2.3 IMMUNOTOXICITY

       10

       11         2.2.3.1  Epidemiologic Finding

       12                  The available epidemiologic studies on immunologic function in humans relative to

       13        exposure to 2,3,7,8-TCDD do not describe a consistent pattern of effects among the examined

       14        populations.  Two studies of German workers, one exposed to 2,3,7,8-TCDD and the other to

       15        2,3,7,8-tetrabrominated dioxin and furan, observed dose-related increases of complements C3 or

       16        C4 (Zober et al., 1992; Ott et al.,  1994), while the Ranch Hands continue to exhibit elevations in

       17        immunoglobulin A (IgA) (Roegner et al., 1991; Grubbs et al., 1995).  Other studies of groups

       18        with documented exposure to 2,3,7,8-TCDD have not examined complement components to any

       19        great extent or observed significant changes ill IgA.  Suggestions of immunosuppression have

      20        been observed in a small group of exposed workers as a result of a single test (Tonn et al., 1996),

      21         providing support for a testable hypothesis to be evaluated in other exposed populations.

      22                  Comprehensive evaluation of immunologic status and function of the NIOSH, Ranch

      23        Hand, and Hamburg chemical worker cohorts found no consistent differences between exposed

      24        and unexposed groups for lymphocyte subpopulations, response to mitogen stimulation, or rates

      25        of infection (Halperin et al., 1998; Michalek et al., 1999; Jung et al., 1998; Emst et al., 1998).

      26        However, in a single study, T cell response to Inferon-y in TCDD-exposed workers was

      27        unaffected when tested in isolated peripheral blood mononuclear lymphocytes; but was impaired

      28        in the highly exposed population when examined in diluted whole blood (Emst et al., 1998).

      29                  More comprehensive evaluations of immunologic function with respect to exposure to

      30        2,3,7,8-TCDD and related compounds are necessary to assess more definitively the relationships

      31         observed in nonhuman species.  Longitudinal studies of the maturing human immune system

 

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       I         may provide the greatest insight, particularly because animal studies have found significant

       2        results in immature animals, and human breast milk is a source of 2,3,7,8-TCDD and other

       3        related compounds.  Additional studies of highly exposed adults may also shed light on the

       4        effects of long-term chronic exposures.  Therefore, there appears to be too little information to

       5        suggest definitively that 2,3,7,8-TCDD, at the levels observed, causes long-term adverse effects

       6        on the immune system in adult humans.

       7

        8        2.2.3.2 Animal Findings

       9                  Cumulative evidence from a number of studies indicates that the immune system of

      10        various animal species is a target for toxicity of TCDD and structurally related compounds,

      11        including other PCDDs, and PCDFs and PCBs.  Both cell-mediated and humoral immune

      12        responses are suppressed following TCDD exposure, suggesting that there are multiple cellular

      13        targets within the immune system that are altered by TCDD.  Evidence also suggests that the

      14        immune system is indirectly targeted by TCDD-induced changes in non-lymphoid tissues.

      15        TCDD exposure of experimental animals results irt decreased host resistance following challenge

      16        with certain infectious agents, which likely result from TCDD-induced suppression of

      17        immunological functions.

      18                  The primary antibody response to the T cell-dependent antigen, sheep red blood cells

      19        (SRBCs), is the most sensitive immunological response that is consistently suppressed in mice

      20        exposed to TCDD and related compounds.  The degree of immunosuppression is related to the

      21        potency of the dioxin-like congeners.  There is remarkable agreement among several different

      22        laboratories for the potency of a single acute dose of TCDD (i.e., suppression at a dose as low as

      23        0.1  pg TCDD/kg with an average 50% immunosupressive dose (ID_0) value of approximately 0.7

      24        g TCDD/kg) to suppress this response in Ah responsive mice.  Results of studies that have

      25        compared the effects of acute exposure to individual PCDD, PCDF, and PCB congeners, that

      26        differ in their binding affinity for the AhR, on this response have provided critical evidence that

      27        certain dioxin-like congeners are also immunosuppressive.  The degree of immunosuppression

      28        has been found to be related to potency of the dioxin-like congeners.  Antibody responses to T

      29        cell-independent antigens, such as trinitrophenyl-lipopolysaccharide (TNP-LPS), and the

      30        cytotoxic T lymphocyte (CTL) response are also suppressed by a single acute exposure to

 

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       1         TCDD, albeit at higher doses than those which suppress the SRBC response.  A limited number

       2        of studies reveal that dioxin-like congeners also suppress these responses, with the degree of

       3        suppression by the congeners related to their A1LR binding affinity.  Although a thorough and

       4        systematic evaluation of the immunotoxicity of TCDD-like congeners in different species and for

       5        different immunological endpoints has not been performed, it can be inferred from the available

       6        data that dioxin-Iike congeners are immunosuppressive.

       7                  Perinatal exposure of experimental animals to TCDD results in suppression of primarily

        8        T cell immune functions, with evidence of suppression persisting into adulthood.  In mice, the

        9        effects on T cell functions appear to be related to the fact that perinatal TCDD exposure alters

      10        thymic precursor stem ceils ill the fetal liver and bone marrow, and thymocyte differentiation in

      11         the thymus.  These studies suggest that perinatal development is a critical and sensitive period

      12        for TCDD-induced immunotoxicity.  Efforts should be made to determine the consequences of

       13        perinatal exposure to TCDD and related compounds and mixtures on immune system integrity.

       14

       15        2.2.3.3 Other Data Related to Immunologic Effects

       16                  In addition to the TCDD-Iike congener results, studies using strains of mice which differ

       17        in the expression of the AhR have provided critical evidence to support a role for Ah-mediated

       18        immune suppression following exposure to dioxin-like compounds.  Recent in vitro work also

       19        supports a role for Ah-mediated immune suppression.  Other in vivo and in vitro data, however,

      20        suggest that non-A.h-mediated mechanisms may also play some role in immunotoxicity induced

       21         by dioxin-like compounds.  However, more definitive evidence remains to be developed to

       22        support this latter view.

       23                  While the immunosuppressive potency of individual dioxin-like compounds in mice is

       24        related to their structural similarity to TCDD, this pattern of suppression is observed only

       25         following exposure to an individual congener.  The immunotoxicity of TCDD and related

       26        congeners can be modified by co-exposure to other congeners in simple binary or more complex

       27        mixtures resulting in additive or antagonistic interactions.  There is a need for the generation of

       28        dose response data of acute, subchronic and chronic exposure to the individual congeners in a

       29        mixture and for the mixture itself in order to fully evaluate potential synergistic, additive or

       30        antagonistic effects of environmentally relevant mixtures.

 

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     1                  Animal host resistance models that mimic human disease have been used to assess the

     2        effects of TCDD on altered host susceptibility.  TCDD exposure increases susceptibility to

     3        challenge with bacteria, viruses, parasites and tumors.  Mortality is increased in TCDD-exposed

     4        mice challenged with certain bacteria.  Increased parasitemia occurs in TCDD-exposed mice and

     5        rats challenged with parasitic infections.  Low doses of TCDD also alter resistance to virus

     6        infections in rodents.  Increased susceptibility to infectious agents is an important benchmark of

     7        immunosuppression; however, the role that TCDD plays in altering immune-mediated

     8        mechanisms important in murine resistance to infectious agents remains to be elucidated.  Also,

     9        since little is known about the effects that dioxin-like congeners have on host resistance, more

    10        research is recommended in tiffs area.

   11                  Studies in nonhuman primates exposed acutely, subchronically or chronically to

    12        halogenated aromatic hydrocarbons (HAH) have revealed ,,,affable alterations in lymphocyte

   13        subpopulations, primarily T lymphocytes subsets.  In three separate studies in which monkeys

    14        were exposed subchronical!y or chronically to PCBs, the antibody response to SRBC was

    15        consistently found to be suppressed.  These results in nonhuman primates are important because

    16        they corroborate the extensive database of HAH-induced suppression of the antibody response to

   17        SRBC in mice and thereby provide credible evidence for immunosuppression by HAHs across

   18        species.  In addition, these data indicate that the primary antibody response to this T ceil-

   19        dependent antigen is the most consistent and sensitive indicator of HAH-induced

   20,       imnnunosuppression.

   2t                  The available database derived from well-controlled animal studies on TCDD

   22        immunotoxicity can be used for the establishment of no-adverse-effect levels.  Since the antibody

   23        response to SRBCs has been shown to be dose-dependently suppressed by TCDD and related

   24        dioxin-like compounds, this database is best suited for the development of dose-response

   25        modeling.

   26

   27        2.2.3.4 Immunologic Effects Hazard Characterization

   28                  Accidental or occupational exposure of humans to TCDD and/or related compounds

   29        variably affects a number of immunological parameters.  Unfortunately, the evaluation of

   30        immune system integrity' in humans exposed to dioxin-like compounds has provided data which

 

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        1         is inconsistent across studies.  However, the broad range of "normal" responses in humans due to

       2        the large amount of variability inherent in such a heterogenous population, the limited number

        3        and sensitivity of tests performed, and poor exposure characterization of the cohorts in these

       4        studies compromise any conclusions about the ability of a given study to detect immune

        5        alterations.  Consequently, there are insufficient clinical data from these studies to fully assess

        6        human sensitivity to TCDD exposure.  Nevertheless, based on the results of the extensive animal

        7        work, the database is sufficient to indicate that immune effects could occur in the human

        8        population from exposure to TCDD and related compounds at some dose level.  At present, it is

        9        EPA's scientific judgment that TCDD and related compounds should be regarded as non-specific

      10        immunosuppressants and immunotoxicants until better data to inform this judgment are

      11         available.

      12                  It is interesting that a common thread in several human studies is the observed reduction

      13        in CD4+ T helper cells, albeit generally within the "normal" range, in cohorts exposed to dioxin-

      14        like compounds  While these reductions may not translate into clinical effects, it is important to

       15        note that these cells play an important role in regulating immune responses and that their

       16        reduction in clinical diseases is associated with immunosuppression.  Another important

       17        consideration is that a primary antibody response following immunization was not evaluated in

       18        any of the human studies.  Since this immune parameter has been revealed to be the most

       19        sensitive in animal studies, it is recommended that TCDD and related compounds be judged

      20        immunosupressive and that this parameter be included in future studies of human populations

      21         exposed to TCDD and related compounds.  It is also recommended that research focused on

      22        delineating the mechanism(s) underlying dioxin-induced immunotoxicity mad

      23        immunosuppression continue.

      24

      25        2.2.4 CHLORACNE

      26                  Chloracne and associated dermatologic changes are widely recognized responses to

      27  .      TCDD and other dioxin-like compounds in humans.  Along with the reproductive hormones

       28        discussed above and gamma glutamyl transferase (GGT) levels, which are discussed below,

      29        chloracne is one of the noncancer effects which has a strong positive association with exposure to

      30        TCDD in humans (See Part 2, Chapter 7b). Chloracne is a severe acne-like condition that

 

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     I         develops within months of first exposure to high levels of dioxin and related compounds.  For

     2        many individuals, the condition disappears after discontinuation of exposure, despite initial

     3        serum levels of dioxin in the thousands of parts per trillion; for others, it may remain for many

     4        years.  The duration of persistent chloracne is on the order of 25 years although cases of

     5        chloracne persisting over 40 years have been noted. (See Chapter 7, Epidemiology).

     6                  In general, chloracne has been observed in most incidents where substantial dioxin

     7        exposure has occurred, particularly among trichlorophenol (TCP) production workers (Goldman,

     8         1972; May, 1973; Bleiberg et al., 1964; Bond et al., 1987; Suskind and Hertzberg, 1984; Moses

     9        et al., 1984; Zober et al.,  1990) and Seveso residents (Reggiani, 1978; Caramaschi et al., 1981;

    10        ideo et aJ.,  1985; Mocarelli et al.,  1986; Asse_mato et al.,  1989). The amount of exposure

    11        necessary' for development of chloracne has not been resolved, but studies suggest that high

    12        exposure (both high acute and long-term exposure) to 2,3,7,8-TCDD increases the likelihood of

    13        chloracne, as evidenced by chloracne in TCP production workers and Seveso residents who have

    14        documented high serum 2,3,7,8-TCDD levels (Beck et al.,  1989; Fingerhut et al.,  1991a;

    15        Mocarelli et al., 1991; Neuberger et al., 1991) or in individuals who have a work history with

    16        long duration of exposure to 2,3,7,8-TCDD-contaminated chemicals (Bond et al., 1989).  In

    17        earlier studies, chloracne was considered to be a "hallmark of dioxin intoxication" (Suskind,

    18         1985).  However, only ii,. two studies were risk estimates calculated for chloracne.  Both were

    19        studies of different cohorts of TCP production workers (Suskind and Hertzberg, 1984; Bond et

    20        al., 1989); one group was employed in a West Virginia plant, the other in a plant in Michigan.

    21         Of the 203 West Virginia workers, 52.7% (p<0.001) were found to have clinical evidence of

    22        chloracne, and 86.3% reported a history of chloracne (/2<0.001) (Suskind and Hertzberg,  1984).

    23        None of the unexposed workers had clinical evidence or reported a history of chloracne.  Among

    24        the Michigan workers, the relative risk for cases of chloracne was highest for individuals with the

    25         longest duration of exposure (a 60 months; RR = 3.5, 95% CI = 2.3-5.1), those with the highest

    26        cumulative dose of TCDD (based on duration of assignment across and within 2,3,7,8-TCDD-

    27        contaminated areas in the plant) (RR = 8.0, 95% CI = 4.2-15.3), and those with the highest

    28        intensity of 2,3,7,8-TCDD exposure (RR = 71.5, 95% CI=32.1-159.2) (Bond et al., 1989).

    29                  Studies in multiple animal species have been effective in describing the relationship

   30        between 2,3,7,8-TCDD and chloracne, particularly in rhesus monkeys (McNulty, 1977; Allen et

 

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          I         al.,  1977; McCormell et al., 1978).  Subsequent to exposure to 2,3,7,8-TCDD, monkeys

          2        developed chloracne and swelling of the meibomian glands, modified sebaceous glands in the

          3        eyelid.  The histologic changes in the meibomian glands are physiologically similar to those

          4        observed in human chloracne (Dunagin,  1984).

          5                  In summary, the evidence provided by the various studies convincingly supports what is

          6        already presumed, that chloracne is a common sequela of high levels of exposure to 2,3,7,8-

          7        TCDD and related compounds.  More information is needed to determine the level and frequency

          8        of exposure to dioxin-like compounds needed to cause chloracne and whether personal

          9        susceptibility plays a role in the etiology'.  Finally, it is important to recall that the absence of

        10        chloracne does not imply lack of exposure (Mocarelli et al., 1991).

        11

        12        2.2.5 DIABETES

        13                  Diabetes mellitus is a heterogeneous disorder that is a consequence of alterations in the

        14        number or function of pancreatic beta cells responsible for insulin secretion and carbohydrate

        15        metabolism. Diabetes and fasting serum glucose levels were evaluated in cross-sectional medical

        16        studies because of the apparently high prevalence of diabetes and abnormal glucose tolerance

        17        tests in one case report of 55 TCP workers (Pazderova-Vejlupkova et al., 1981). Recent

        18        epidemiology studies, as well as early case reports, have indicated an association between serum

        19        (blood) levels (body burden) of dioxin and diabetes.  This association was first noted in the early

        20        90s when a decrease in glucose tolerance was seen in the NIOSH cohort.  This was followed by a

        21         report of an increase in diabetes in the Ranch Hand cohort.  Several reports from other

        22        occupational cohorts, as well as the Seveso population and the Asian rice oil poisonings, then

        23        followed.  There was not a significant increase in diabetes in the NIOSH mortality study,

        24        although 6 of the 10 most highly exposed workers did have diabetes. The recent paper by

        25        Longnecker and Michalek (2000) demonstrated an association between diabetes and dioxin

        26         levels within Air Force Veterans who never had contact with dioxin-contaminated herbicides and

        27        whose blood levels are within the range of the background population.  The most recent update

        28        of the Ranch Hand study also shows a 47% excess of diabetes in the most heavily exposed group

       29        of veterans.

       30                  Much of the data suggests that the diabetes is Type II, or adult-onset, diabetes, rather than

 

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      1         insulin dependent, or Type 1.  Aging and obesity are the key risk factors for this form of diabetes.

     2        However, dioxins may shift the distribution of sensitivity, putting people at risk at younger ages

     3        or with less weight.  Dioxin alters lipid metabolism in multiple species, including people.  Dioxin

     4        also alters glucose uptake into both human and animal cells in culture. Mechanistic studies have

      5        demonstrated that dioxin affects glucose transport, a property under the control of the hypoxia

      6        response pathway.  A key regulatory protein in this pathway is the partner of the Ah receptor,

      7        AMT (also known as HIF 1-beta).  Activation of the .Kit receptor by dioxin may compete with

      8        other pathways, such as the HIF pathway, for AMT.  Dioxin has also been shown to down

      9        regulate the insulin growth factor receptor.  These three issues - altered lipid metabolism, altered

    I0        glucose transport, and alterations in the insulin signaling pathway - all provide biological

    1.1        plausibility to the association of dioxins with diabetes.

    12                  While there appears to be a relatively consistent association between diabetes and dioxin

     13        body' burdens, causality has _lot been established.  It is possible that the higher level of dioxin in

    14        people with diabetes is an effect, not a cause.   Does diabetes alter the pharmacokinetics of

    15        dioxin?  Diabetes is known to alter the metabolism of several drugs in people.  However, these

     16        drugs are not metabolized by the enzymes known to be induced by dioxins.  Since adult-onset

     17        diabetes is also associated with overweight, and body composition has been shown to modify the

     18        apparent half-life of dioxin, could the rate of elimination of dioxins be lowered in people with

     19        diabetes, causing them to have higher body burdens?  This may be relevant to the background

    20        population, but is hardly likely to be an explanation in the highly exposed populations.  Key

    2I         research needs are two-fold.  The first is to develop an animal model in which to study the

    22        association between dioxins and diabetes.  Several rodent models for Type 2 diabetes exist and

    23        may be able to be utilized.  The second is to conduct incidence studies.  Type II diabetes is often

    24        not the cause of death and therefore the association would not be noted in a mortality study.

    25

    26        2.2.6 OTHER ADVERSE EFFECTS

    27                  Elevated GGT - As mentioned above, there appears to be a consistent pattern of

    28        increased GGT levels among individuals exposed to 2,3,7,8-TCDD-contaminated chemicals.

    29        Elevated levels of serum GGT have been observed within a year after exposure in Seveso

    30        children (Caramaschi et al., 1981; Mocarelli et al., 1986) and 10 or more years after cessation of

 

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     1         exposure among TCP and 2,4,5-T production workers (May,  1982; Martin, 1984; Moses et al.,

    2         1984; Calvert et al., 1992) and among Ranch Hands (Roegner et al., 1991; Grubbs et al., 1995).

    3        All of these groups had a high likelihood of substantial exposure to 2,3,7,8-TCDD.  In addition,

    4        for those studies that evaluated dose-response relationships with 2,3,7,8-TCDD levels, the effect

     5        was observed only at the highest levels or categories of 2,3,7,8-TCDD. In contrast, although

     6        background levels of serum 2,3,7,8-TCDD suggested minimal exposure to Army Vietnam

     7        veterans, GGT was increased, at borderline significance, among Vietnam veterans compared to

     8        non-Vietnam veterans (Centers for Disease Control Vietnam Experience Study, 1988a).  In

     9        addition, despite the increases observed in some occupational cohorts, other studies of TCP

   10        production workers from West Virginia or Missouri residents measured but did not report

   11         elevations in GGT levels (Suskind and Hertzberg, 1984; Webb et al., 1989).

   12                  In clinical practice, GGT is often measured because it is elevated in almost all

   13        hepatobiliary diseases and is used as a marker for alcoholic intake (Guzelian, 1985).  In

   14        individuals with hepatobiliary disease, elevations in GGT are usually accompanied by increases

   15        in other hepatic enzymes, e.g., AST and-ALT, and metabolites, e.g., ufo- and coproporphyrins.

   16        Significant increases in hepatic enzymes other than GGT and metabolic products were not

   17        observed in individuals whose GGT levels were elevated 10 or more years after exposure ended,

   18        suggesting that the effect may be GGT-specific.  These data suggest that in the absence of

   19        increases in other hepatic enzymes, elevations in GGT are associated with exposure to 2,3,7,8-

   20        TCDD, particularly among individuals who were exposed to high 2,3,7,8-TCDD levels.

   21                  The animal data with respect to 2,3,7,8-TCDD-related effects on GGT are sparse.

   22        Statistically significant changes in hepatic enzyme levels, particularly AST, ALT, and ALK,

   23        have been observed after exposure to 2,3,7,8-TCDD in rats and hamsters (Gasiewicz et al., 1980;

   24        Kociba et al.,  1978; Olson et al.,  1980).  Only one study evaluated GGT levels (Kociba et al.,

   25         1978).  Moderate but statistically nonsignificant increases were noted in rats fed 0.10 ug,/kg

   26        2,3,7,8-TCDD daily for 2 years, and no increases were observed in control animals.

   27                  In summary, GGT is the only hepatic enzyme examined that was found in a number of

   28        studies to be chronically elevated in adults exposed to high levels of 2,3,7,8-TCDD.  The

   29        consistency of the findings in a number of studies suggest that the elevation may reflect a true

   30        effect of exposure but its clinical significance is unclear.  Long-term pathologic consequences of

 

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    1         elevated GGT have not been illustrated by excess mortality from liver disorders or cancer, or in

    2        excess morbidity in the available cross-sectional studies.

    3                  It must be recognized that the absence of an effect in a cross-sectional study, for example,

    4         liver enzymes, does not obviate tine possibility that the enzyme levels may have increased

    5        concurrent to the exposure but declined after cessation.  The apparently transient elevations in

    6        ALT levels among the Seveso children suggest that hepatic enzyme levels other than GGT may

    7        react in this mariner to 2,3,7,8-TCDD exposure.

    8                  Thyroid Function - Many effects of 2,3,7,8-TCDD exposure in animals resemble signs

    9        of thyroid dysfunction or significant alterations of thyroid-related hormones.  In the few human

   10        studies that examined the relationship between 2,3,7,8-TCDD exposure and hormone

   11         concentrations in adults, the results are mostly equivocal (Centers for Disease Control Vietnam

   12        Experience Study,  1988a; Roegner et al.,  1991; Grubbs et al., 1995; Suskind and Hertzberg,

   13         1984).  However, concentrations of thyroid binding globulin (TBG) appear to be positively

   14        correlated with current levels of 2,3,7,8-TCDD in the BASF accident cohort (Ott et al., 1994).

   15        Little additional information on thyroid hormone levels has been reported for production workers

   16        and none for Seveso residents, two groups with documented high serum 2,3,7,8-TCDD levels.

   17                  Thyroid hormones play important roles in the developing nervous system in of all

   18         vertebrates species, including humans.  In fact, thyroid hormones are so important in

   19        development that in the U.S. all infants are tested for hypothyroidism shortly after birth.  Several

   20        studies of nursing infants suggest that ingestion of breast milk with a higher dioxin TEQ may

   21         alter thyroid function (Plium et al  1993; Koopman-Esseboom et al 1994c; Nagayam et al., 1997).

   22        These findings suggest a possible shift in the distribution of thyroid hormones, particularly T4,

   23        and point out the need for collection of longitudinal data to assess the potential for long-term

   24        effects associated with developmental exposures. The exact processes accounting for these

   25        observations in humans are unknown, but when put in perspective of animal responses, the

   26        following might apply: dioxin increases the metabolism and excretion of thyroid hormone,

   27        mainly T4, in the liver.  Reduced T4 levels stimulate the pituitary to secrete more TSH, which

   28        enhances thyroid hormone production.  Early in the disruption process, the body can

   29        overcompensate for the loss ofT4, which may result in a small excess of circulating T4 to the

   30        increased TSH.  In animals, given higher doses of dioxin, the body is unable to maintain

 

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         3         homeostasis, and TSH levels remain elevated and T4 levels decrease.

         2                  Cardiovascular Disease - Elevated cardiovascular disease has been noted in several of

         3        the occupational cohorts and in Seveso, as well as in the rice oil poisonings.  This appears to be

         4        associated with ischemic heart disease and in some cases with hypertension.  In fact, recent data

         5        from the Ranch Hand study indicates that dioxin may be a risk factor for the development of

         6        essential hypertension.  Elevated blood lipids have also been seen in several cohorts.  The

         7        association of dioxins with heart disease in people has biological plausibility given the data in

         8        animals.  First is the key role of hypoxia in heat disease, and the potential for involvement of the

         9        activated Ah receptor in blocking an hypoxic response.  Dioxin has been shown to perturb lipid

        10        metabolism in multiple laboratory species.   The heart, in fact, the entire vascular system, is a

        11         clear target for the adverse effects of dioxin in fish and birds. Dioxin has recently been shown to

        12        disrupt blood flow in mammals, dioxin has been shown to disturb heart rhythms at high doses in

        13        guinea pigs.

        14                  Oxidative Stress - Several investigators have hypothesized that the some of the adverse

        15        effects of dioxin and related compounds may be associated with oxidative stress.  Induction of

        16        CYPIA isoforms has been shown to be associated with oxidative DNA damage (Park et al.,

        17         1996).  Altered metabolism of endogenous molecules such as estradiol can lead to the formation

        18        ofquinones and redox cycling.  This has been hypothesized to play a role in the enhanced

        19        sensitivity of female rats to dioxin-induced liver tumors ( Tritscher et al., 1996). Lipid

        20        peroxidation, enhanced DNA single strand breaks, and decreased membrane fluidity have been

        21         shown in liver as well as in extrahepatic tissues following exposure to high doses of TCDD

        22        (Stohs,  1990). A dose- and time-dependent increase in superoxide anion is caused in peritoneal

        23        macrophages by exposure to TCDD (Alsharif et al.,  1994).  A recent report that low dose (0.45

        24        ng TCDD/kg/day) chronic exposure can lead to oxidative changes in several tissues in mice

        25        (Slezak et al., 2000) suggests that this mechanism or mode of toxicity deserves further attention.

        26

        27

        28        3.0      MECHANISMS AND MODE OF DIOXIN ACTION

        29

       30                      Mechanistic studies can reveal the biochemical pathways and types of biological and

 

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         1         molecular events that contribute to dioxin's adverse effects.  For example, much evidence

         2        indicates that TCDD acts via an intracellular protein (the aryl hydrocarbon receptor; Ah

         3        receptor), which functions as a ligand-dependent transcription factor in partnership with a second

         4        protein (known as the .Ah receptor nuclear translocator; Amt).  Therefore, from a mechanistic

         5        standpoint, TCDD's adverse effects appear likely to reflect alterations in gene expression that

         6        occur at an inappropriate time and/or fol' an inappropriately long time.  Mechanistic studies also

         7        indicate that several other proteins contribute to TCDD's gene regulatory effects and that the

         8        response to TCDD probably involves a relatively complex interplay between multiple genetic

         9        and environmental factors.  If TCDD operates through such a mechanism, as all evidence

        10        indicates, then there are certain constraints on the possible models that can plausibly account for

        1 l         TCDD's biological effects and, therefore, on the assumptions used during the 14sk assessment

        12        process (e.g. Poland,  1996; Limbird and Taylor, 1998).  Mechanistic knowledge of dioxin action

        13        may also be useful in other ways.  For example, a further understanding of the ligand specificity

        14        and structure of the Ah receptor will likely assist in the identification of other chemicals to which

        15        humans are exposed that may either add to, synergize, or block the toxicity of TCDD.

        16        Knowledge of genetic polymorphisms that influence TCDD responsiveness may also allow the

         17        identification of individuals at greater risk from exposure to dioxin.  In addition, -knowledge of

         18        the biochemical pathways that are altered by TCDD may help identify novel targets for the

         19        development of drugs that can antagonize dioxin's adverse effects.

        20                  As described below, biochemical and genetic analyses of the mechanisms by which

        21         dioxin may modulate particular genes have revealed the outline of a novel regulatory system

        22        whereby a chemical signal can alter cellular regulatory processes.  Future studies of dioxin action

        23        have the potential to provide additional insights into mechanisms of mammalian gene regulation

        24        that are of a broader interest.  Additional perspectives on dioxin action can be found in several

        25        recent reviews (Bimbaum,  1994a,b; Schecter, 1994; Hankinson, 1995; Schmidt and Bradfield,

        26         1996; Gasiewicz,  1997; Rowlands and Gustafsson, 1997; Denison et al., 1998; Hahn, 1998;

        27        Wilson and Safe,  1998).

        28                  Knowledge of the mode(s) of action by which the broad class of chemicals known as

        29        dioxins act may facilitate the risk assessment process by imposing bounds on the models used to

        30        describe possible responses of humans resulting from exposure to mixtures of these chemicals.

 

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          1        The relatively extensive data base on TCDD, as well as the more limited data base on related

         2        compounds, has been reviewed with emphasis on the role of the specific cellular receptor for

         3        TCDD and related compounds, the Ah receptor, ill the mode(s) of action. The present discussion

         4        will focus on summarizing the elements of the mode(s) of dioxin action that are relevant for

          5         understanding and characterizing dioxin risk for humans.  These elements include:

         6                    -  similarities between humans and other animals with regard to receptor structure and

          7                      function;

          8                   -  the relationship between receptor binding and toxic effects; and

          9                   -  the extent to which the purported mechanism(s) or mode(s) of action might contribute

        10                      to the diversity of biological responses seen in animals and, to some extent, in humans.

        11

        12        In addition, this Section will identify important and relevant knowledge gaps and uncertainties in

        13        the understanding of the mechanism(s) of dioxin action, and wilt indicate how these may affect

        14        the approach to risk characterization.

        15

        16        3.1 Mode Versus Mechanism of Action

        17                  In the context of revising its Cancer Risk Assessment Guidelines, the EPA has proposed

        18        giving greater emphasis to use of all of tine data in hazard characterization, dose-response

        19        characterization, exposure characterization and risk characterization (EPA,  1996). One aid to the

        20        use of more information in risk assessment has been the definition of mode versus mechanism of

        21         action.  Mechanism of action is defined as the detailed molecular description of a key event in

        22        the induction of cancer or other health endpoints.  Mode-of-action refers to the description of key

        23        events and process, starting with interaction of an agent with the cell, through functional and

        24        anatomical changes, resulting in cancer or other health endpoints.  Despite a desire to construct

        25        detailed biologically-based toxicokinetic and toxicodynamic models to reduce uncertainty in

        26        characterizing risk, few examples have emerged.  Use of mode-of-action approach recognizes

        27        that, although all of the details may not have been worked out, prevailing scientific thought

        28        supports moving forward using a hypothesized mode-of action supported by data.  This approach

       29        is consistent with advice offered by the National Research Council in its report entitled,  Science

       30        and Judgment in Risk Assessment (NRC,  1994). Mode-of-action discussions help to provide

       31         answers to the questions: How does the chemical produce its effect?; Are there mechanistic data

       32        to support this hypothesis?; Have other modes of action been considered and rejected?  In order

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            1        to demonstrate that a particular mode-of-action is operative it is generally necessary to outline

           2        the hypothesized sequence of events leading to effects, identify key events that can be measured

            3        and outline the information that is available to support the hypothesis and also discuss those data

           4        which are inconsistent with the hypothesis or which support an alternative hypothesis, and weigh

            5        the information to determine if there is a causal relationship between key, precursor events

            6        associated with the mode-of- action and cancer or other toxicological endpoint.

            7

            8        3.2 Generalized Model for Dioxin Action

            9                  Dioxin and related compounds are generally recognized to be receptor-mediated

          10        toxicants.  The generalized model has evolved over the years to appear as in Figure 2-1.  Events

          11         embodied in this model of dioxin's mode-of-action include:

          1_2

          28

          29        These events are discussed in detail in Part2, Chapter2.

 

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       1

       2        THE RECEPTOR CONCEPT

       3                  One of the fundamental concepts that influences our approach to risk assessment of

       4        dioxin and related compounds is the receptor concept.  The idea that a drug, hormone,

       5        neurotransmitter, or other chemical produces a physiological response by interacting with a

       6        specific cellular target molecule, i.e., a "receptor," evolved from several observations.  First,

       7        many chemicals elicit responses that are restricted to specific tissues.  This observation implies

       8        that the responsive tissue (e.g., the adrenal cortex) contained a "receptive" component whose

       9        presence is required for the physiologic effect (eo.,., cortisol secretion).  Second, many chemicals

      l0        are quite, potent.  For example, picomolar to nanomolar concentrations of numerous hormones

      11        and growth factors elicit biological effects.  This observation suggests that the target cell contains

      12        a site(s) to which the particular chemical binds with high affinity.  Third, stereoisomers of some

      13        chemicals (e.g., catecholamines, opioids) differ by orders of magnitude in their ability to produce

      14        the same biological response.  This observation indicates that the molecular shape of the

      15        chemical strongly influences its biological activity.  This, in turn, implies that the binding site on

      16        or in the target cell also has a specific, three-dimensional configuration.  Together, these types of

      17        observations predict that the biological responses to some chemicals involve stereospecific, high-

      18        affinity binding of the chemicals to specific receptor sites located on or in the target cell. Many

      19        of these characteristics were noted for TCDD and related compounds.

      20                  The availability of compounds of high specific radioactivity has permitted quantitative

      21         analyses of their binding to cellular components in vitro.  To qualify as a potential "receptor," a

      22        binding site for a given chemical must satisfy several criteria: (1) the binding site must be

      23        saturable, i.e., the number of binding sites per cell should be limited; (2) the binding should be

      24        reversible; (3) the binding affinity measured in vitro should be consistent with the potency of the

      25        chemical observed in vivo; (4) if the biological response exhibits stereospecificity, so should the

      26        in vitro binding; (5) for a series of structurally related chemicals, the rank order for binding

      27        affinity should correlate with the rank order for biological potency; and (6) tissues that respond

      28        to the chemical should contain binding sites with the appropriate properties.

     29                  The binding of a chemical ("ligand") to its specific receptor is assumed to obey the law of

     30        mass action; that is, it is a bin2olecular, reversible interaction.  The concentration of the liganded,

 

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          1         or occupied, receptor [RL] is a function of both the ligand concentration [L] and the receptor

2               concentration IR] as shown in Equation3-1:

3                

          3                                                                   k_

 

                                                                                h,

                  4                                                IL] + [R]      "         [RI..]

 

5                                                                                                                           k2

6                                                                                                                            

          6        Equation 3-1.  Ligand Binding Kinetics

          7

          8        Inherent in this relationship is that the fractional occupancy (i.e. [RL] / [R]) is a function of

          9        ligand concentration [L] and the apparent equilibrium dissociation constant Ko, which is a

         10        measure of the binding affinity of the ligand for the receptor, that is, [RL] / [Ps] =  [L] / (Ko +

         11         [L]), where Ko = [L] [R] / [LR] = k: / k,.  Therefore, the relationship between receptor ccupancy

         12        and ligand concentration is hyperbolic.  At low ligand concentrations (where [L]<<KD), a small

         13        increase it,. [L] produces an approximately linear increase in fractional receptor occupancy.  At

         14        high ligand concentration (where [L]>>KD), the fractional occupancy of the receptor is already

         15        vets, close to 1, that is, almost all receptor sites are occupied.  Therefore, a small increase in [L]

         16        is likely to produce only a slight increase in receptor occupancy.  These issues are discussed in

         17        regard to TCDD binding to the Ah receptor and dose response in Part 2, Chapter 8.

         18                  Ligand binding constitutes only one aspect of the receptor concept.  By definition, a

         19        receptor mediates a response, arid the functional consequences of the ligand-receptor binding

        20        represent an essential aspect of the receptor concept.  Receptor theory attempts to quantitatively

        21         relate ligand binding to biological responses.  The classical "occupancy" model of Clark (1933)

        22        postulated that (1) the magnitude of the biological response is directly proportional to the

        23        fraction of receptors occupied and (2) the response is maximal when all receptors are occupied.

        24        However, analyses of numerous receptor-mediated effects indicate that the relationship between

        25        receptor occupancy and biological effect is not as straightforward as Clark envisioned.  In certain

        26        cases, no response occurs even when there is some receptor occupancy.   This suggests that there

        27        may be a threshold phenomenon that reflects the biological "inertia" of the response (Ariens et

        28        al., 1960).  In other cases, a maximal response occurs well before all receptors are occupied, a

 

 

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        I        phenomenon that reflects receptor "reserve" (Stephenson, 1956).  Therefore, one cannot simply

        2        assume that the relationship between fractional receptor occupancy and biological response is

        3        linear.  F'urthermore, for a ligand (such as TCDD) that elicits multiple receptor-mediated effects,

        4        one can not assume that the binding-response relationship for a simple effect (such as enzyme

         5        induction) will necessarily be identical to that for a different and more complex effect (such as

         6        cancer).  The cascades of events leading to different complex responses (e.g., altered immune

         7        response to pathogens or development of cancer) are likely to be different, and other rate-limiting

         8        events likely influence the final biological outcome resulting in different dose-response curves.

         9        Thus, even though ligand binding to the same receptor is the initial event leading to a spectrum

       10        of biological responses, ligand-binding data may not always mimic the dose-effect relationship

       11        observed for particular responses.

       12                  Another level of complexity is added when one considers different chemical ligands that

       13        bind to the same receptor.  Relative potencies are determined by two properties of the ligand:

        14       affinity for the receptor, and capacity to confer a particular response in the receptor (e.g., a

        15        particular conformational change), also called efficacy (Stephenson, 1956).  Ligands with

        16        different affinities arid the same degree of efficacy would be expected to produce parallel dose-

        17        response curves with the same maximal response within a particular model system..  However,

        18        ligands of the same affinity with different efficacies may result in dose-response curves that are

        19        not parallel or that differ in maximal response.  Many of these issues may apply to dioxin-

       20        receptor interactions.  To the extent that they do occur, they may present complications to use of

       21         the toxicity equivalence approach, particularly for extrapolation purposes.  As described

       22        previously, this argues strongly for the use of ail available information in setting TEFs and

       23        highlights the important role that scientific judgment plays in the face of incomplete mechanistic

       24        understanding to address uncertainty.

        25

       26        A FRAMEWORK TO EVALUATE MODE-OF-ACTION

        27                  The U.S. EPA in its revised, proposed cancer guidelines (EPA, 1999) recommends the

        28         use of a structured approach to evaluating mode-of--action.  This approach is similar to and

       29        builds upon an approach developed within the World Health Organization's (WHO) International

       30        Programme on Chemical Safety's Harmonization Project (WHO, 2000).  Fundamentally, the

 

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          I         approach uses a modification of the "Hill Criteria" (Hill, 1965) which have been used in the field

          2        of epidemiology for many years to examine causality between associations of exposures and

          3        effects.  The framework calls for a summary description of the postulated mode-of-action,

          4        followed by the identification of key events which are thought to be part of the mode-of-action.

          5        These key events are then evaluated as to strength, consistency and specificity of association

          6        with the endpoint under discussion.  Dose-response relationships between the precursor, key

          7        events are evaluated and temporal relationships are examined to be sure that "precursor" events

          8        actually precede the induction of the endpoint.  Finally, biological plausibility and coherence of

          9        the data with the biology is examined and discussed.  All of these "criteria" are evaluated and

         10        conclusions are drawn with regard to postulated mode-of-action.

         11                  In the case of dioxin and related compounds, elements of such an approach are found for

         12        a number of effects including cancer in Part 2.  Application of the framework to dioxin auld

         13        related compounds would now stop short of evaluating the association between the chemical or

         14        complex mixture and clearly adverse effects.  Instead, the approach would apply to early events

         15        e.g. receptor binding and intermediate events such as enzyme induction or endocrine impacts.

         16        Additional data will be required to extend the framework to most effects but several have data

         17        which would support a framework analysis.  Several of these are discussed below.

         18

         19        MECHANISTIC INFORMATION, MODE-OF-ACTION AND RISK ASSESSMENT

         20                  A substantial body of evidence from investigations using experimental animals indicates

         21         that the Ah receptor mediates the biological effects of TCDD.  Although studies using human

        22        tissues are much less extensive, it appears reasonable to assume that dioxin's mode-of-action to

        23        produce effects in humans includes receptor-mediated key events.  Studies using human organs

        24        and cells in culture are consistent with this hypothesis.  A receptor-based mode-of action would

        25        predict that, except in cases where the concentration of TCDD is already high (i.e., [TCDD]~KD),

        26        incremental exposure to TCDD will lead to some increase in the fraction of Ah receptors

        27        occupied,  However, it cannot be assumed that an increase in receptor occupancy will necessarily

        28        elicit a proportional increase in all biological response(s), because numerous molecular events

        29        (e.g., cofactors, other transcription factors, genes) contributing to the biological endpoint are

        30        integrated into the overall response.  That is, the final biological response should be considered

 

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        1         as an integration of a series of dose-response curves with each curve dependent on the molecular

        2        dosimetry for each particular step.  Dose-response relationships that will be specific for each

        3        endpoint must be considered when using mathematical models to estimate the risk associated

        4        with exposure to TCDD.  It remains a challenge to develop models that incorporate all the

        5        complexities associated with each biological response.  Furthermore, the parameters for each

        6        mathematical model may only apply to a single biological response within a given tissue m_d

        7        species.

        8                  Given TCDD's widespread distribution, its persistence, and its accumulation within the

        9        food chain, it is likely that most humans are exposed to some level of dioxin; thus, the population

       10        at potential risk is large and genetically heterogensous.  By analogy with the findings in inbred

       11         mice, polymorphisms in the Ah receptor probably exist in humans.  Therefore, a concentration of

       12        TCDD that elicits a particular response in one individual may not do so in another.  For example,

       13        studies of humans exposed to dioxin following an industrial accident at Seveso, Italy, fail to

       14        reveal a simple and direct relationship between blood TCDD levels and development of

       15        chloracne (Mocarelli et al.,  1991).  These differences in responsiveness to TCDD may reflect

       16        genetic variation either in the ,<_ receptor or in some other component of the dioxin-responsive

       17        pathway.  Therefore, analyses of human polymorphisms in the Ah receptor and Arnt genes have

       18        the potential to identify genotypes associated with higher (or lower) sensitivities to dioxin-related

       19        effects.  Such molecular genetic information may be useful in the future for accurately predicting

       20        the health risks dioxin poses to humans.

       21                  Complex responses (such as cancer) probably involve multiple events and multiple genes.

       22        For example, a homozygous recessive mutation at the hr (hairless) locus is required for TCDD's

       23        action as a tumor promoter in mouse skin (Poland et al., 1982).  Thus, the hr locus influences the

       24        susceptibility of a particular tissue (in this case skin) to a specific effect of dioxin (tumor

       25        promotion).  An analogous relationship may exist for the effects of TCDD in other tissues,  For

       26        example, TCDD may produce porphyria cutanea tarda only in individuals with inherited

       27        uroporphyrinogen decarboxylase deficiency (Doss et al., 1984).  Such findings suggest that, for

       28        some adverse effects of TCDD, the population at risk may be limited to individuals with a

       29        particular genetic predisposition.

       30                  Other factors can influence an organism's susceptibility to TCDD.  For example, female

 

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        1         rats are more prone to TCDD-induced liver neoplasms than are males; this phenomenon is

        2        related to the hormonal status of the animals (Lucier et al., 1991).  In addition, hydrocortisone

        3        and TCDD synergize in producing cleft palate in mice.  Retinoic acid and TCDD produce a

        4        similar synergistic teratogenic effect (Couture et al., 1990).  These findings indicate that, in some

        5        cases, TCDD acts in combination with hormones or other chemicals to produce adverse effects.

        6        Such phenomena might also occur in humans.  If so, the difficulty in assessing risk is increased,

        7        given the diversity among humans in hormonal status, lifestyle (e.g., smoking, diet), and

        8        chemical exposure.

        9                  Dioxin's action as a tumor promoter and developmental toxicant presumably reflects its

       10        ability to alter cell proliferation and differentiation processes.  There are several plausible

       11         mechanisms by which this could occur.  First, TCDD might activate a gene (or genes) that is

       12        directly involved in tissue proliferation.  Second, TCDD-induced changes in hormone

       13        metabolism may lead to tissue proliferation (or lack thereof) and altered differentiation secondary

       14        to altered secretion of a trophic hormone.  Third, TCDD-induced changes in the expression of

       15        growth factor or hormone receptors may alter the sensitivity of a tissue to proliferative stimuli.

       16        Fourth, TCDD-induced toxicity may lead to cell death, followed by regenerative proliferation.

       17        These mechanisms likely differ among tissues and periods of development, and might be

       18        modulated by different genetic and environmental factors.  As such, this complexity increases the

       19        difficult3, associated with assessing the human health risks form dioxin exposure.

       20                  Under certain circumstances, exposure to TCDD may elicit beneficial effects.  For

       21         example, TCDD protects against the carcinogenic effects of PAH’s in mouse skin, possibly

       22        reflecting induction of detoxifying enzymes (Cohen et al., 1979; DiGiovanni et al., 1980).  In

       23        other situations, TCDD-induced changes in estrogen metabolism may alter the growth of

       24        hormone-dependent tumor cells, producing a potential anticarcinogenic effect (Spink et al.,  1990;

       25        Gierthy et al.,  1993).  However, several recent studies in mice indicate that the Ah receptor has

       26        an important role in the genetic damage and carcinogenesis caused by components in tobacco

       27        smoke such as benzo[a]pyrene through its ability to regulate CYP1A] gene induction (Derringer

       28        et al., 1998; Shilnizu et al., 2000).  TCDD's biological effects likely reflect a complicated

       29        interplay between genetic and environmental factors. These issues complicate the risk assessment

       30        process for dioxin.

 

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        1          4.0      EXPOSURE CHARACTERIZATION

        2                  This section summarizes key findings developed in the exposure portion of the Agency's

        3        Dioxin Reassessment Effort.  The findings are developed in the companion document entitled

        4         "Part 1.  Estimating  Exposure to Dioxin-like Compounds."  This document is divided into four

         5        volumes:  I Executive Summary, 2. Sources of Dioxin in the United States; 3. Properties,

         6        Environmental Levels and Background Exposures and 4. Site-Specific Assessment Procedures.

         7        Readers are encouraged to examine the more detailed companion document for further

         8        information on the topics covered here and to see complete literature citations.  The

         9        characterization discussion provides cross references to help readers find the relevant portions of

        10        the companion document.

       11                  This discussion is organized as follows:  1. Sources, 2. Fate, 3. Environmental Media and

        12        Food Concentrations, 4. Background Exposures, 5. Potentially Highly Exposed Populations and

        13        6. Trends.  The key findings are presented in italics.

       14

        15        4.1. Sources  (cross reference: Part 1, Volume II: Sources of Dioxin-Like Compounds in the

                   u.s.)

      17                  The CDD/CDFs have never been intentionally produced other than on a laboratory scale

        18        basis For use in scientific analysis.  Rather, they are generated as unintended byproducts in trace

        19        quantities in various combustion, industrial and biological processes.  PCBs on the other hand,

        20        were commercially produced in large quantities, but are no longer commercially produced in the

        21         U.S.  The EPA has classified sources of dioxin-like compounds into five broad categories:

        22

        23        ·         Combustion Sources: CDD/CDFs are formed in most combustion systems. These can

        24                  include waste incineration (such as municipal solid waste, sewage sludge, medical waste,

        25                  and hazardous wastes), burning of various fuels (such as coal, wood, and petroleum

        26                  products), other high temperature sources (such as cement kilns), and poorly or

        27                  uncontrolled combustion sources (such as forest fires, building fires, and open burning of

        28                  wastes).

        29

        30        ·         Metals Smeitin2, Refining and Processing Sources:  CDD/CDFs can be fon'ned during

 

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        1                  various types of primary and secondary metals operations including iron ore sintering,

        2                  steel production, and scrap metal recovery.

        3

        4        ·         Chemical Manufacturing,: CDD/CDFs can be formed as by-products from the

        5                  manufacture of chlorine bleached wood pulp, chlorinated phenols (e.g.,

        6                  pentachlorophenol - PCP), PCBs, phenoxy herbicides (e.g., 2,4,5-T), and chlorinated

        7                  aliphatic compounds (e.g., ethylene dichloride).

        8

        9        .         Biological and Photochemical Processes:  Recent studies suggest that CDD/CDFs can be

       10                  fon-ned under certain environmental conditions (e.g., composting) from the action of

       11                  microorganisms on chlorinated phenolic compounds.  Similarly, CDD/CDFs have been

       12                  reported to be formed during photolysis of highly chlorinated phenols.

       13

       14        ·         Reservoir Sources: Reservoirs are materials or places that contain previously formed

       15                  CDD/CDFs or dioxin-like PCBs and have the potential for redistribution and circulation

       16                  of these compounds into the environment.  Potential reservoirs include soils, sediments,

       17                  biota, water and some anthropogenic materials.  Reservoirs become sources when they

       18                  have releases to the circulating environment.

       19                   Development of release estimates is difficult because only a few facilities in most

       20        industrial sectors have been tested for CDD/CDF emissions.  Thus an extrapolation is needed to

       21        estimate national emissions.  The extrapolation method involves deriving an estimate of

       22        emissions per unit of activity at the tested facilities and multiplying this by the total activity level

       23        in the untested facilities.  In order to convey the level of uncertainty in both the measure of

       24        activity and the emission factor, EPA developed a qualitative confidence rating scheme.  The

       25        confidence rating scheme, presented in Table 4-1, uses qualitative criteria to assign a high,

       26        medium, or low confidence rating to the emission factor and activity, level for those source

       27        categories for which emission estimates can be reliably quantified. The overall "confidence

       28        rating" assigned to a quantified emission estimate was determined by the confidence ratings

       29        assigned to the corresponding "activity level" and "emission factor."  If the lowest rating

       30        assigned to either the activity level or emission factor terms is "high," then the category rating

 

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        1         assigned to the emission estimate is high (also referred to as "A").  If the lowest rating assigned

        2        to either the activity level or emission factor terms is "medium," then the category rating

        3        assigned to the emission estimate is medium (also referred to as "B").  If the lowest rating

        4        assigned to either the activity level or emission factor terms is "low," then the category rating

        5        assigned to the emission estimate is low (also referred to as "C").  For many source categories,

         6        either the emission factor information or activity level information were inadequate to support

         7        development of reliable quantitative release estimates for one or more media.  For some of these

         8        source categories, sufficient information was available to make preliminary estimates of

         9        emissions of CDD/CDFs or dioxin-like PCBs, however, the confidence in the activity level

       10        estimates or emission factor estimates was so low that the estimates cannot be included in the

       11         sum of quantified emissions from sources with confidence ratings of A, B and C.  These

       12        estimates were given an overall

       13

        14

        15

        16

        17

        18

        19

       20        Table 4-1.  Confidence Rating Scheme

        21

                                                                                                       .        ..        ,.          .

 

 

                    

 

 

 

 

 

 

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         12        4.1.1 Inventory of Releases

         13                  The Dioxin Reassessment has produced an Inventor)' of source releases for the U.S.

        14        (Table 4-2).  The Inventor, was developed by considering all sources identified in the published

 

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    I         literature and numerous individual emissions test reports.  U.S. data were always given first

    '_        priority for developing emission estimates.  Data from other countries were used for making

    3        estimates in only a few source categories where foreign technologies were judged similar to

    4        those found in the U.S. and the U.S. data were inadequate.  The Inventory is limited to sources

    5        whose releases can be reliably quantified (i.e. those with confidence ratings of A, B or C as

    6        defined above).  Also, it is limited to sources with releases that are created essentially

    7        simultaneously with formation.  This means that the reservoir sources are not included.  As

     8        discussed below, this document does provide preliminary estimates of releases from these

     9        excluded sources (i.e. reservoirs and Class D sources) but they are presented separately from the

   10        inventory'.

   11                  The Inventory presents the environmental releases in terms of two reference years:  1987

   I2        and 1995.  1987 was selected primarily because little empirical data existed for making source

   13        specific emission estimates.  1995 represents the latest time that could practically be addressed

   14        consistent with the time table for producing the rest of this document.

   15

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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   35

   36                  Figure 4-1 displays the emission estimates to air for sources included in the Inventory and

   37        shows how the emission factors and activity levels were combined to generate emission

   38        estimates.  Figure 4-2 compares the animal mean TEQDF)_-WHO98 emission estimates to air for the

   39        two reference years (i,e., 1987 and 1995).

   4O

 

 

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      I         The following conclusions are made for sources of dioxin-like compounds included in the

      2        Inventory:

      3

      4        ·         EPA 's best estimates of releases of CDD/CDFs to air, water and land from reasonably

      5                  quantifiable sources were approximately 2.800 gram (g) TEQDF-WHO98 in 1995 and

      6                  13,500 g TEQDF- WHO98  in 1987.

      7

      8        ·         The decrease in estimated releases of CDD/CDFs between 1987 and 1995

      9                  (approximately 80%) was due primarily to reductions in air emissions from municipal

     10                  and medical waste incinerators.  For both categories, these emission reductions have

     11                  occurred from a combination of improved combustion and emission controls and from

     12                  the closing of a number of facilities.   Regulations recently promulgated or under

     13                  development should result in some additional reduction in emissions from major

     14                  combustion sources.

     15

     16        ·         The environmental releases of CDD/Fs in the U.S. occur from a wide variety of sources,

     17                  but are dominated by, releases to the air from combustion sources.  The current (1995)

     18                  inventory indicates emissions from combustion sources are over an order of magnitude

     19                  greater than emissions from the sum of emissions from all other categories.

    20

    21         ·        Insufficient data are available to comprehensively estimate point source releases of

    22                  dioxin-like compounds to water.  Sound estimates of releases to water are only available

    23                  for chlorine bleached pulp and paper mills and manufacture of ethylene dichloride/vinyl

    24                  chloride monomer.   Other releases to water bodies that cannot be quantified on the basis

    25                  of existing data include effluents from POTWs and most industrial/commercial sources.

    26

    27        ·         Based on the available information, the inventory  includes only a limited set of activities

    28                  that result in direct environmental releases to land.  The only releases to land quantified

    29                  in the inventory, are land application of sewage sludge and pulp and paper mill

    30                  wastewater sludges.  Not included in the Inventor's definition of an environmental

 

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    1                  release is the disposal of sludges and ash into approved landfills.

    2

    3        ·         The inventory is likely to underestimate total releases.  A number of investigators have

    4                  suggested that national inventories may underestimate emissions due to the possibility of

    5                  unknown sources.  These possibilities have been supported with mass balance analyses

    6                  suggesting that deposition exceeds emissions.  The uncertainty, however, in both the

    7                  emissions and deposition estimates for the U.S. prevent the use of this approach for

    8                  reliably evaluating the issue.  As explained below, this document has instead, evaluated

    9                  this issue by making preliminary estimates of poorly characterized sources and listing

   10                  other sources which have been reported to emit dioxin-like compounds but cannot be

   11                  characterized on even a preliminary basis.

   12

   13        4.1.2. General Source Observations

   14                  The preliminary release estimates for contemporary formation sources and reservoir

   15        sources are presented in Table 4-3.  Table 4-4  lists all the sources which have been reported to

   16        release dioxin-like compounds but cannot be characterized on even a preliminary basis.

   17        For any given time period, releases from both contemporary formation sources and reservoir

   18        sources determine the overall amount of the dioxin-like compounds that are being released to the

   19        open and circulating environment.  Because existing information is incomplete with regard to

   20        quantifying contributions from contemporary and reservoir sources, it is not currently possible to

   21         estimate the total magnitude of release for dioxin-like compounds into the U.S. environment

   22        from all sources.  For example, in terms of 1995 releases from reasonably quantifiable sources,

   23        tiffs document estimates releases of 2800 g WHO98 TEQDF  for contemporary formation sources

   24        and 2900 g WHO98 TEQDF for reservoir sources.  In addition, there remains a number of

   25        unquantifiable and poorly quantified sources.  No quantitative release estimates can be made for

   26        agricultural burning or for most dioxin/furan reservoirs or for any dioxin-like PCB reservoirs.

   27        The preliminary estimate of 1995 poorly characterized contemporary formation sources is 1900 g

   28        WHO98 TEQDF

  

 

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29        Table 4-3.   Preliminary Indication of the Potential Magnitude of TEQDF-WHOo8 Releases

  30        from "Unquantified" (i.e., Category D) Sources in Reference Year 1995

     25

    26

    27

    28

    29

    30

    31

    32

   

 

 

 

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33        Table 4-4.  Unquantified Sources

 

 

             7

         8

         9

        10                  Additional observations and conclusions about all sources of dioxin-like compounds are

        11         summarized below:

        12

        13        ·         The contribution of dioxin-like compounds to waterways from nonpoint source reservoirs

        14                  is likely to be greater than the contributions from point sources.  Current data are only

        15                  sufficient to support preliminary estimates of nonpoint source contributions of dioxin-like

        16                  compounds to water (i.e., urban storm water run off and rural soil erosion).  These

        17                  estimates suggest that, on a nationwide basis, total nonpoint releases are significantly

        t 8                  larger than point source releases.

        19        ·         Current emissions of CDD/Fs to the U.S. environment result principally from

       20                  anthropogenic activities.  Evidence which supports this finding include: matches in time

 

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      1                  of rise of environmental levels with time when general industrial activity began rising

      2                  rapidly (see trend discussion in Section 4.6), lack of any identified large natural sources

      3                  and observations of higher CDD/F body burdens ill industrialized vs. less industrialized

      4                  countries (see discussion on human tissue levels in Section 4.4).

      5

      6        ·         Although chlorine is an essential component for the formation of CDD/Fs in combustion

      7                  systems, the empirical evidence indicates that for commercial scale incinerators, chlorine

      8                  levels m feed are not the dominant controlling factor for rates of CDD/F stack emissions.

      9                  important factors which can affect the rate of dioxin formation include the overall

     10                 combustion efficiency, post combustion flue gas temperatures and residence times, and

    11                  the availability of surface catalytic sites to support dioxin synthesis.  Data from bench,

     12                 pilot and commercial scale combustors indicate that dioxin format/on can occur by a

     13                 number of mechanisms.  Some of these data, primarily from laboratory and pilot scale

     14                 combustors, have shown direct correlation between chlorine content in fuels and rates of

     15                 dioxin formation.  Other data, primarily from commercial scale combustors, show little

     16                 relation with availability of chlorine and rates of dioxin formation. The conclusion that

     17                 chlorine in feed is not a strong determinant of dioxin emissions applies to the overall

     18                 population of commercial scale combustors.  For any individual commercial scale

     19                 combustor, circumstances may exist in which changes in chlorine content of feed could

    20                  affect dioxin emissions.  For uncontrolled combustion, such as open burning of house-

    21                  hold waste, chlorine content of wastes may play a more significant role in affecting levels

    22                  of dioxin emissions than observed in commercial scale combustors.

    23

    24        ·         No significant release of newly formed dioxin-like PCBs is occurring in the U.S.  Unlike

    25                  CDD/CDFs, PCBs were intentionally manufactured in the U.S. in large quantities from

    26                  1929 until production was banned in 1977.  Although it has been demonstrated that small

    27                  quantities of coplanar PCBs can be produced during waste combustion,  no strong

    28                  evidence exists that the dioxin-like PCBs make a significant contribution to TEQ releases

    29                  during combustion. The occurrences of dioxin-like PCBs in the U.S. environment most

    30                  likely reflects past releases associated with PCB production, use and disposal.  Further

 

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        I                  support of this finding is based on observations of reductions since 1980s in PCBs in

       2                  Great Lakes sediment and other areas.

       3

       4        ·         It is unlikely that the emission rates of CDD/CDFs from known sources correlate

       5                  proportionally with general population exposures.  Although the emissions inventory

       6                  shows the relative contribution of various sources to total emissions, it cannot be assumed

       7                  that these sources make the same relative contributions to human exposure.  It is quite

        8                  possible that the major sources of dioxin in food (see discussion in Section 2.6 indicating

        9                  that the diet is the dominant exposure pathway for humans) may not be those sources that

      10                  represent the largest fractions of total emissions in the U.S.  The geographic locations of

      11                  sources relative to the areas from which much of the beef, pork, milk, and fish come, is

      12                  important to consider.  That is, much of the agricultural areas which produce dietary

      13                  animal fats are not located near or directly down wind of the major sources of dioxin and

      14                  related compounds.

      15

      16        ·         The contribution of reservozr sources to human exposure may be significant.  Several

      17                  factors support this finding.  First, human exposure to the dioxin-like PCBs is thought to

      18                  be derived almost completely from reservoir sources.  Since one third of general

      19                  population TEQ exposure is due to PCBs, at least one third of the overall risk from

      20                  dioxin-like compounds comes from reservoir sources.  Second, CDD/CDF releases from

      21                  soil via soil erosion and runoff to waterways appear to be greater than releases to water

      22                  from the primary sources included in the inventory.  CDD/CDFs in waterways can

      23                  bioaccumulate in fish leading to human exposure via consumption of fish which makes

      24                  up about one third of the total general population CDD/CDF TEQ exposure.  This

      25                  suggests that a significant portion of the CDD/CDF TEQ exposure could be due to

      26                  releases from the soil reservoir.  Finally, soil reservoirs could have vapor and particulate

      27                  releases which deposit on plants and enter the terrestrial food chain.  The magnitude of

      28                  this contribution, however, is unknown.

      29        4.2.  Environmental Fate  (cross reference: Part l, Volume III, Chapter 2)

      30                  Dioxin-like compounds are widely distributed in the environment as a result of a number

      31         of physical and biological processes.  The dioxin-like compounds are essentially insoluble in

      32        water, generally classified as semi-volatile and tend to bioaccumulate in animals.  Some evidence

 

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      I         has shown that these compounds can degrade in the environment, but in general they are

      2        considered very, persistent and relatively immobile in soils and sediments. These compounds are

      3       'transported through the atmosphere as vapors or attached to air-borne particulates and can be

      4        deposited on soils, plants, or other surfaces (by wet or dry deposition).  The dioxin-like

      5        compounds enter water bodies primarily via direct deposition from the atmosphere, or by surface

      6        run off and erosion.  From soils, these compounds can reenter the atmosphere either as

      7        resuspended soil particles or as vapors.  In water, they can be resuspended into the water column

      8        from sediments, volatilized out of the surface waters into the atmosphere or become buried in

      9        deeper sediments.  Immobile sediments appear to serve as permanent sinks for the dioxin-like

     10        compounds. Though not always considered an environmental compartment, these compounds are

     11         also found in anthropogenic materials (such as pentachlorophenol) and have the potential to be

     12        released from these materials into the broader environment.

     1 3                   Atmospheric transport and deposition of the dioxin-like compounds are a primary

     14        means o[dispersal of these compounds throughout the environment.  The dioxin-like compounds

     15        can be measured in wet and dry deposition in most locations including remote areas.  Numerous

     16        studies have shown that they are commonly found in soils throughout the world.  Industrialized

     17        countries tend to show similar elevated concentrations in soil and detectable levels have been

     l 8        found in nonindustrialized countries.  The only satisfactory explanation available for this

     19        distribution is air transport and deposition.  Finally, by analogy these compounds would be

     20        expected to behave similarly to other compounds with similar properties and this mechanism of

     21        global distribution is becoming widely accepted for a variety of persistent organic compounds.

     22                  The two primary pathways for the dioxin-like compounds to enter the ecological food

     23        chains and human diet are: air-to-plant-to-animal and water/sediment-to-fish.  Vegetation

     24        receives these compounds via atmospheric deposition in the vapor and particle phases.  The

    25        compounds are retained on plant surfaces mad bioaccumulated in the fatty tissues of animals that

    26        feed on these plants. Vapor phase transfers onto vegetation have been experimentally shown to

    27        dominate the air-to-plant pathway for the dioxin-like compounds, particularly for the lower

    28        chlorinated congeners  In the aquatic food chain, dioxins enter water systems via direct discharge

    29        or deposition and runoff' from watersheds. Fish accumulate these compounds through their direct

    30        contact with water, suspended particles, bottom sediments and through the consumption of

 

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    1         aquatic organisms.  Although these two pathways are thought to normally dominate contribution

    2        to the commercial food supply, others can also be important.  Elevated dioxin levels in cattle

    3        resulting from animal contact with pentacholorophenol treated wood have been documented by

    4        the USDA.  Animal feed contamination episodes have led to elevations of dioxins in poultry in

    5        the United States, milk in Germany, and meat/dairy products in Belgium.

    6

    7        4.3.   Environmental Media and Food Concentrations (cross reference: Part 1, Volume III,

    8        Chapter 3-

    9                  Estimates of the range of typical background levels of dioxin-like compounds in various

   10        environmental media are presented in Table 4-5 below:

   11

   12        Table 4-5.  Estimates of the range of typical background levels of dioxin-like compounds in

   13        various environmental media

   14

   21

   22        Estimates for background levels of dioxin-like compounds in environmental media are based on

   23        a variety of studies conducted at different locations in North America.  Of the studies available

   24        for this compilation, only those conducted in locations representing "background"  were selected.

   25        The amount and representativeness of the data varies, but in general these data lack the statistical

   26        basis to establish true national means.   The environmental media concentrations were consistent

   27        among the various studies, mad were consistent with similar studies in Western Europe.  These

   28        data are the best available for comparing site specific values to national background levels.

  29        Because of the limited number of locations examined, however,  it is not known if these ranges

 

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       l         adequately capture the full national variability, if significant regional variability exists making

       2        national means of limited utility, or if elevated levels above this range could still be the result of

       3        background contamination processes.  As new data are collected these ranges are likely to be

       4        expanded and refined.  The limited data on dioxin-like PCBs in environmental media are

       5        summarized in the document (Part 1, Volume I1I, Chapter 4), but were not judged adequate for

       6        estimating background levels.

       7                  Estimates of typical background levels of dioxin-like compounds in food are presented in

       8         Table 4-6 below:

       9

    

 

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     1                  Estimates for levels in food are based on data from a variety of studies conducted in

     2        North America.  Beef, pork and poultry were derived from statistically based national surveys.

     3        Milk estimates were derived from a survey of a nationwide milk sampling network.  Dairy

     4        estimates were derived from milk fat concentrations, coupled with appropriate assumptions for

     5        the amount of milk fat in dairy products.  Egg samples were grab samples from retail stores.

     6        Fish data were collected from a combination of field and retail outlets and were normalized so

     7        that all concentrations ':,,ere expressed on the basis of fresh weight in edible tissue.  As with

      8       environmental media, food levels found in the United States are similar to levels found in

      9       Europe.

    I0

    11        4.4.  Background Exposures  (cross reference: Part l, Volume III, Chapter 4)

    12

    13        4.4.1 Tissue Levels

    14                  The average CDD/CDF tissue level for the general adult U.S. population appears to be

    15        declining and the best estimate of current (late ] 990s) levels is 25 ppt (TEQDFp-WH098,  lipid

    16        basis).   The tissue samples collected in North America in the late 1980s and early 1990s showed

    17        an average TEQDFP, WHO98,)s level of about 55 pg/g lipid. This finding is supported by a number of

    18        studies which measured dioxin levels in adipose, blood and human milk, all conducted in North

    19        America.  The number of people in most of these studies, however, is relatively small and the

    20        participants were not statistically selected in ways that assure their representativeness of the

    21         general U.S. adult population. One study, the 1987 National Human Adipose Tissue Survey

    22        (NHATS), involved over 800 individuals and provided broad geographic coverage, but did not

    23        address coplanar PCB s. Similar tissue levels of these compounds have been measured in Europe

    24        and Japan during similar time periods.

    25                  Because dioxin levels in the environment have been declining since the 1970s (see trends

    26        discussion), it is reasonable to expect that levels in food, human intake and ultimately human

    27        tissue have also declined over this period.  The changes in tissue levels are likely to lag the

    28        decline seen in environmental levels and the changes in tissue levels cannot be assumed to occur

    29        proportionally with declines in environmental levels.  ATSDR (1999) summarized levels of

    30        CDDs, CDFs and PCBs in human blood collected during the time period 1995 to 1997.  The

 

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      1         individuals sampled were alt US residents with no -known exposures to dioxin other than normal

      2        background.  The blood was collected from 400 individuals in seven different locations with an

      3        age range of 20 to 70 years.  All TEQ calculations were made assuming nondetects were equal to

      4        half the detection limit.  While these samples were not collected in a manner that can be

      5        considered statistically representative of the national population and lack wide geographic

      6        coverage, they are judged to provide a better indication of current tissue levels in the US than the

      7        earlier data (see Table 4-7 ). PCBs 105,  118, and 156 are missing from the blood data for the

      8        comparison populations reported in the Calcasieu study (ATSDR, 1999).  These congeners

      9        account for 62% of the total PCB TEQ estimated in the early 1990's.  Assuming that the missing

     10        congeners from the Calcasieu study data contribute the same proportion to the total PCB TEQ as

     11         in earlier data, they would increase our estimate of current body burdens by another 3.7 pgTEQ/g

     12        lipid for a total PCB TEQ of 5.9 pg/g lipid and a total DFP TEQ of 25 pg/g lipid.

     13                  This finding regarding current tissue levels is further supported by the observation that

     14        this mean tissue level is consistent with our best estimate of current intake, i.e. 1 pg/kg-d in

     15        TEQDFP WHO98.  Using this intake in a one compartment, steady-state pharmacokinetic model,

     16        yields a tissue level estimate of about 16 pg TEQDFP WHO98/g lipid (assumes TEQ DFP has an

     17        effective half life of 7 yr, 80% of ingested dioxin is absorbed into the body and lipid volume is

     18        19 L).  Since intake rates appear to have declined in recent years and steady state is not likely to

     19        have been achieved, it is reasonable to observe higher measured tissue levels than predicted by

     20        the model.

     21                  Characterizing national background levels of dioxins in tissues is uncertain because the

     22        current data cannot be considered statistically representative of the general population.  It is also

     23        complicated by the fact that tissue levels are a function of both age and birth year.  Because

     24        intake levels have varied over time, the accumulation of dioxins in a person who turned 50 years

     25        old in 1990 is different than in a person who turned 50 in 2000.  Future studies should help

     26        address these uncertainties.  The National Health and Nutrition Examination Survey 0N'-I-IANES)

     27        began a new national survey in 1999 which will measure dioxin blood levels in about 1700

     28        people per year (see http:www.cdc.gov/nchs/nhanes.htm).  The survey is conducted  at  15

     29        different locations per year and is designed to select individuals statistically representative of the

     30        civilian US population in terms of age, race and ethnicity.  These new data should provide a

 

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        I         much better basis for estimating national background tissue levels and evaluating trends than the

        2        currently available data.

        3

        4        

 

12           4.4.2. Intake Estimates

13                      Adult daily intakes of CDD/CDFs and dioxin-like PCBs are estimated to average 45 and

14           25 pg TEQDFP-WHO98/day, respectively, for a total intake of 70 pg./day TEQDFP-WHO98.  Daily

15           intake is estimated by combining exposure media concentrations (food, soil, air) with contact

16           rates (ingestion, inhalation).  Table 4-8 below summarizes the intake rates derived by this

17           method.

18            

       19             

 

 

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     10

     11

     12

     13                  The intake estimate is supported by an extensive database on food consumption rates and

     14        food data (as discussed above).  Pharmacokinetic (PK) modeling provides further support for the

     15        intake estimates.  Applying a simple steady-state PK model to an adult average CDD/CDF

     16        adipose tissue level of 18.8 ppt TEQDFWHO98 (on a lipid basis) yields a daily intake of 110 pg

     17        TEQDFWHO98/day.  Insufficient half-life data are available for making a similar intake estimate

     18        for the dioxin-like PCBs.  This PK modeled CDD/CDF intake estimate is about 2.5 times higher

     19        than the direct intake estimate of 45 pg TEQDFWHO98/day.  This difference is to be expected

     20        with this application of a simple steady-state PK model to current average adipose tissue

     21        concentrations.  Current adult tissue levels reflect intakes from past exposure levels which are

     22        thought to be higher than current levels (see Trends Section 2.6).  Since the direction and

     23        magnitude of the difference in intake estimates between the two approaches are understood, the

     24        PK derived value is judged supportive of the pathway derived estimate.  It should be recognized,

     25        however, the pathway derived value will underestimate exposure if it has failed to capture all

     26        significant exposure pathways.

 

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      1         4.4.3. Variability in Intake Levels

      2                  CDD/CDF and dioxin-like PCB intakes for the general population may extend to levels at

      3        least three times higher than the mean.  Variability in general population exposure is primarily

      4        the result in the differences in dietary choices that individuals make.  These are differences in

      5        both quantity and types of food consumed.  A diet which is disproportionately high in animal fats

      6        will result in art increased background exposure over the mean. Data on variability of fat

      7        consumption indicate that the 95th percentile is about twice the mean and the 99th percentile is

      8        approximately 3 times the mean.  Additionally, a diet which substitutes meat sources that are low

      9        in dioxin (i.e. beef, pork or poultry) with sources that are high in dioxin (i.e. fresh water fish)

     10       could result in exposures elevated over three times the mean.  This scenario may not represent a

     11       significant change in total animal fat consumption, even though it results in an increased dioxin

     12       exposure.

     13                  Intakes of CDD/Fs and dioxin-like PCBs are over three times higher for a young child as

     14        compared to that of an adult, on a body weight basis.   Using age-specific food consumption rate

     15        and average food concentrations, as was done above for adult intake estimates, the following

     16        Table 4-9 describes the variability in average intake values as a function of age.

     17

     18       

    26

    27                  Only four of the 17 toxic CDD/CDF congeners and one of the 11 toxic PCBs account for

    28        most of the toxicity in human tissue concentrations:  2378-TCDD,  12378-PCDD,  123678-

 

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      1        HxCDD. 23478-PCDF and PCB 126.  This finding is derived directly from the data described

     2        earlier on human tissue levels, and is supported by intake estimations which indicate that these

      3        congeners are also the primary contributors to dietary dose.  These five compounds make up over

      4        half of the total TEQ tissue level.

      5

      6        4.5.  Potentially Highly Exposed Populations or Developmental Stages (cross reference:

      7        Part I, Volume III, Chapter 6)

      8                  As discussed earlier, background exposures to dioxin-like compounds may extend to

      9        levels at least three times higher than the mean.  This upper range is assumed to result from the

    10        normal variability of diet and human behaviors.  Exposures from local elevated sources or

    11        exposures resulting from unique diets would be in addition to this background variability.  Such

    12        elevated exposures may occur in small segments of the population such as individuals living near

     13       discrete local sources, or subsistence or recreational fishers.  Nursing infants represent a special

     14       case where, for a limited portion of their lives, these individuals may have elevated exposures on

     15       a body weight basis when compared to non-nursing infants and adults.

     16                  Dioxin contamination incidents involving the commercial food supply have occurred in

     17       the U.S. and other countries.  For example, in the U.S., contaminated ball clay was used as an

     18       anti-caking agent in soybean meal and resulted in elevated dioxin levels in some poultry and

     19       catfish. This incident involved less than 5% of the national poultry production and has since been

    20        eliminated. Elevated dioxin levels have also been observed in a few beef and dairy animals

    21        where the contamination was associated with contact with pentachlorophenol treated wood.

    22        Evidence of this kind of elevated exposure was not detected in the national beef survey.

    23        Consequently its occurrence is likely to be low, but it has not been determined. These incidents

    24        may have led to small increases in dioxin exposure to the general population.  However, it is

    25        unlikely that such incidents have led to disproportionate exposures to populations living near

    26        where these incidents have occurred, since, in the U.S., meat and dairy products are highly

    27        distributed on a national scale.  If contamination events were to occur in foods that are

    28        predominantly distributed on a local or regional scale, then such events could lead to highly

    29        exposed local populations.

    30                  Elevated exposures associated with the workplace or industrial accidents have also been

 

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       1        documented.  U.S. workers in certain segments of the chemical industry, had elevated levels of

       2        TCDD exposure, with some tissue measurements in the thousands of ppt TCDD.  There is no

       3        clear evidence that elevated exposures are currently occurring among U.S. workers. Documented

       4        examples of past exposures for other groups include certain Air Force personnel exposed to

       5        Agent Orange during the Vietnam War and people exposed as a result of industrial accidents in

       6        Europe and Asia.

       7                  Consumption of breast milk by nursing infants may lead to higher levels of exposure

       8        compared to the intake of non-nursing infants and dietary intakes later in life.  A number of

       9        studies have measured levels of the dioxin-like compounds in human breast milk, yielding an

      10        average of 35 ppt TEQDFP-WHO98.  Based on a six month nursing scenario, the average daily

      11         intake for an infant is about 100 times higher than the adult daily intake on a body weight basis:

      12        the adult intake is 1 pg TEQDFP-WHO98/kg-d, while the infant intake while breast feeding would

      13        be 100 pg TEQDFP-WHO98/kg-d.  The differences in body burden between nursing infants and

      14        adults are expected to be much less than the differences in daily intake. On a mass basis, the

      15        cumulative dose to the infant under this scenario is about 9% of the lifetime intake.

      16                  Consumption of unusually high amounts of fish, meat, or dairy products containing

      17        elevated levels of dioxins and dioxin-like PCBs can lead to elevated exposures in comparison to

      18       the general population.   Most people eat some fish from multiple sources, both fresh and salt

      19       water  The typical dioxin concentrations in these fish and the typical rates of consumption are

     20        included in the mean background calculation of exposure. People who consume large quantities

     21        of fish at typical contamination levels may have elevated exposures since the concentration of

     22        dioxin-like compounds in fish are generally higher than in other animal food products. These

     23        kinds of exposures are addressed within the estimates of variability of background and are not

     24        considered to result in highly exposed populations.   If high-end consumers obtain their fish from

     25        areas where the concentration of dioxin-like chemicals in the fish is elevated, they may constitute

     26        a highly exposed subpopulation.  Although this scenario seems reasonable, no supporting data

     27        could be found for such a highly exposed subpopulation in the U.S. One study measuring dioxin-

     28        like compounds in blood of sports fishers in the Great Lakes area showed elevations over mean

     29        background, but within the range of normal variability.  Elevated CDD/CDF levels in human

     30        blood have been measured in Baltic fishermen.  Similarly elevated levels of coplanar PCBs have

 

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         1         been measured in the blood of fishers on the north shore of the Gulf of the St. Lawrence River

         2        who consumed large amounts of seafood.

         3                  Similarly, high exposures to dioxin-like chemicals as a result of consuming meat and

         4        dairy products would only occur in situations where individuals consume large quantities of

         5        these foods and the level of these compounds is elevated.  Most people eat meat and dairy

         6        products from multiple sources and, even if large quantities are consumed, they are not likely to

         7        have unusually high exposures.  Individuals who raise their own livestock for basic subsistence

         8        have the potential for higher exposures if local levels of dioxin-like compounds are high.  One

         9        study in the U.S. showed elevated levels in chicken eggs near a contaminated soil site.  European

        10       studies at several sites have shown elevated CDD/CDF levels in milk and other animal products

        I 1       near combustion sources.

        12

        13        4.6.  Environmental Trends  (cross reference: Part I, Volume III, Chapter 6)

        14                  Concentrations of CDD/CDFs and PCBs m the U.S. environment were consistently Iow

        15       prior to the 1930s.  Then, concentrations rose steadily until about 1970. At this time, the trend

        16       reversed and the concentrations have declined to the present.

        17                  The most compelling supportive evidence of this trend for the CDD/CDFs and PCBs

        18       comes from dated sediment core studies. Sediment concentrations in these studies are generally

        19       assumed to be an indicator of the rate of atmospheric deposition.  CDD/CDF and PCB

        20       concentrations in sediments began to increase around the 1930s, and continued to increase until

        21       about 1970.  Decreases began in 1970 and have continued to the time of the most recent sediment

        22       samples (about 1990).  Sediment data from 20 U.S. lakes and rivers from seven separate research

        23       efforts consistently support this trend. Additionally, sediment studies in lakes located in several

        24       European countries have shown similar trends.

        25                  It is reasonable to assume that sediment core trends should be driven by a similar trend in

        26       emissions to the environment.  The period of increase generally matches the time when a variety

        27       of industrial activities began rising and the period of decline appears to correspond with growth

        28        in pollution abatement.  Many of these abatement efforts should have resulted in decreases in

        29        dioxin emissions, i.e. elimination of most open burning, particulate controls on combustors,

        30        phase out of leaded gas, and bans on PCBs, 2,4,5-T, hexachlorophene, and restrictions on use of

 

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    1       pentachlorophenol.  Also, the national source inventory of this assessment documented a

   2        significant decline in emissions from the late 1980s to the mid-1990s. Further evidence of a

   3        decline in CDD/CDF levels in recent years is emerging from data, primarily from Europe,

   4        showing declines in foods and human tissues.

    5                  In addition to the congener-specific PCB data discussed earlier, a wealth of data on total

    6        PCBs and Aroclor mixtures exist which also supports these trends.  It is reasonable to assume

    7        that the trends for dioxin-like PCBs are similar to those for PCBs as a class because the

    8        predominant source of dioxin-like PCBs is the general production of PCBs in Aroclor mixtures.

    9        PCBs were intentionally manufactured in large quantities from 1929 until production was banned

  10        in the U.S. in 1977.  U.S. production peaked in 1970, with a volume of 39,000 metric tons.

  11        Further support is derived from data showing declining levels of total PCBs in Great Lakes

  12        sediments and biota during the 1970s and 1980s.  These studies indicate, however, that during

  13        the 1990s the decline is slowing and may be leveling off.

  14                  Past human exposures to dioxins were most likely higher than current estimates.  This is

  15        supported by a study which applied a non-steady state pharmacokinetic model to data on

  16        background U.S. tissue levels of 2,3,7,8-TCDD from the 1970s and 80s. Various possible intake

  17        histories (pg/kg-day over time) were tested to see which best fit the data.  An assumption of a

  18        constant dose over time resulted in a poor fit to the data.  The "best fit" (statistically derived) to

  19        the data was found when the dose, like the sediment core trends, rose through the 60s into the

  20        70s, and declined to current levels.  Some additional support for this finding comes from a

  21         limited study of preserved meat samples from several decades in the twentieth century.  One

  22        sample, from before 1910, showed very low concentrations of dioxins and coplanar PCBs.

  23        Thirteen other samples, from the 1940s until the early 1980s, consistently showed elevated levels

  24        of all dioxin-like compounds as compared to food surveys conducted during the 1990s.

  25

  26        5.0      DOSE-RESPONSE CHARACTERIZATION

  27

  28        Previous sections of this integrated summary have focused on characterizing the hazards of and

  29        exposure to dioxin-like compounds. In order to bring these issues together and provide an

  30        adequate characterization of risk, the relationships of exposure to dose and, ultimately, to

 

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         1       response must be evaluated.  Key questions to be asked include:  1) What can be said about the

        2        shape of the dose-response function in the observable range and what does this imply about

        3        dose-response in the range of environmental exposures? 2) What is a reasonable limit (critical

        4        dose or point of departure) at the lower end of the observable range and what risk is associated

        5        with this exposure?  In addition, one can address the issue of extrapolation beyond the range of

         6       the data in light of the answers to the above questions.  While extrapolation of risks beyond the

         7       range of observation in animals and/or humans is an inherently uncertain enterprise, it is

        8        recognized as an essential component of the risk assessment process (NAS, 1983).  The level of

         9       uncertainty is dependent on the nature (amount and scope) of the available data and on the

       I0        validity of the models which have been used to characterize dose-response.  These form the bases

       11       for scientific inference regarding individual or population risk beyond the range of current

       12       observation (NAS 1983, 1994)

       13                  In Part 2, Chapter 8, the body of literature concerning dose-response relationships of

       14        TCDD has been presented. This Chapter addresses the important concept of selecting an

       15        appropriate metric for cross-species scaling of dose and presents the results of empirical

       16        modeling for many of the available data sets on TCDD exposures in humans and in animals.

       17        Although not all human observations or animal experiments are amenable to dose-response

       18        modeling, over 200 data sets were evaluated for shape, and an effective dose (ED) value

       19        expressed as a percent response for the endpoint being evaluated is presented e.g. ED01  equals an

       20        effective dose for a 1% response.  The analysis of dose-response relationships for TCDD,

       21        considered within the context of toxicity equivalence, mechanism of action and background

       22        human exposures, helps to elucidate the common ground and the boundaries of the science and

       23        science policy components inherent in this risk characterization for the broader family of dioxin-

       24        like compounds,  For instance, the dose-response relationships provide a basis to infer a point of

       25        departure for extrapolation for cancer and noncancer risk for a complex mixture of dioxin-like

       26        congeners given the assumption of toxicity equivalence as discussed in Part 2, Chapter 9.

      27         Similarly, these relationships provide insight into the shape of the dose-response at the point of

       28        departure which can help inform choices for extrapolation models for both TCDD and total TEQ.

       29                  In evaluating the dose-response relationships for TCDD as a basis for assessing this

      30        family of compounds, both empirical dose response modeling approaches as well as mode-of-

 

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      I         action based approaches have been developed and applied (See Part2, Chapter 8; Portier et al,

      2        1996).  Empirical models have advantages and disadvantages relative to more ambitious

      3        mechanism-based models.  Empirical models provide a simple mathematical model that

      4        adequately describes the pattern of response for a particular data set and can also provide the

      5        means for hypothesis testing and interpolation between data points. In addition, they can provide

      6        qualitative insights into underlying mechanisms. However, the major disadvantage is their

      7        inability to quantitatively link data sets in a mechanistically meaningful manner.  On the other

      8        hand, mechanism-based modeling can be a powerful tool for understanding and combining

      9        information on complex biological systems.  Use of a truly mechanism-based approach can, in

     10        theory, enable more reliable and scientifically sound extrapolations to lower doses and between

     11        species. However, any scientific uncertainty about the mechanisms that the models describe is

     12        inevitably reflected in uncertainty about the predictions of the models.

     13             Physiologically-based pharmacokinetic (PBPK) models have been validated in the

     14        observable response range for numerous compounds in both animals and humans.  The

     15        development of PBPK models for disposition of TCDD in animals has proceeded through

     16        multiple levels of refinement, with newer models showing increasing levels of complexity by

     17        incorporating data for disposition of TCDD, its molecular actions with the Ah receptor and other

     18        proteins, as well as numerous physiological parameters (Part 2, Chapter 1).  These have provided

     19        insights into key determinants of TCDD disposition in treated animals. The most complete PBPK

     20        models give similar predictions about TCDD tissue dose metrics.  The PBPK models have been

     2l         extended to generate predictions for early biochemical consequences of tissue dosimetry of

     22        TCDD such as induction of CYP1A1.  Nevertheless, extension of these models to more complex

     23        responses are more uncertain at this time.  Differences in interpretation of the mechanism of

     24        action lead to varying estimates of dose-dependent behavior for similar responses.  The shape of

     25        the dose-response curves governing extrapolation to low doses are determined by these

     26        hypotheses and assumptions.   At this time, the knowledge of the mechanism of action of dioxin

     27        and receptor theory, and the available data of dose-response, do not firmly establish a scientific

     28        basis for replacing a linear procedure for estimating cancer potency.  Consideration of this same

     29        information indicates that the use of different procedures to estimate the risk of exposure for

    30        cancer and noncancer endpoints may not be appropriate.  Both the cancer and noncancer effects

 

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    I         of dioxin appear to result from qualitatively similar modes-of-action of action.  Initial steps in the

    2        process of toxicity are the same and many early events appear to be shared.  Thus, the inherent

    3        potential for low dose significance of either 'type' of effect (cancer or noncancer) should be

    4        considered equal and evaluated accordingly. In the observable range around 1% excess response,

     5       the quantitative differences are relatively small. Below this response, the different mechanisms

     6       can diverge rapidly. The use of predicted biochemical responses as dose metrics for toxic

     7       responses is considered as a potentially useful application of these models.  However, greater

     8       understanding of the linkages between these biochemical effects and toxic responses is needed to

     9       reduce the potentially large uncertainty associated with these predictions.

   10        5.1  Dose Metric(s)

   11                  One of the most difficult issues in risk assessment is the determination of the dose metric

   12        to use for animal-to-human extrapolations.  To provide significant insight into differences in

   13        sensitivity among species, the appropriate animal-to-human extrapolation of tissue dose is

   14        required.  The most appropriate dose metric should reflect both the magnitude and frequency of

   15        exposure, and should be clearly related to the toxic endpoint of concern by a well-defined

   16        mechanism. This is, however, often difficult because human exposures with observable

   17        responses may be very. different from highly controlled exposures in animal experiments.  In

   18        addition, comparable exposures may be followed by very different pharmacokinetics (absorption,

   19        distribution, metabolism and/or elimination) in animals and humans.  Finally, the sequelae of

   20        exposure in the form of a variety of responses related to age, organ-, and species-sensitivity

   21        complicate the choice of a common dose metric.  Despite these complexities, relatively simple

   22        default approaches including body surface or body weight scaling of daily exposures have

   23        often been recommended (EPA, 1992; EPA, 1996).

   24                  Given the data available on dioxin and related compounds, dose can be expressed in a

   25        multitude of metrics (Devito et al, 1995) such as daily intake (ng/kg/d), current body burden

   26        (ng/kg), average body burden over a given period of time, plasma concentration, etc.  Examples

   27        of other dose metrics of relevance for TCDD and related compounds can be found in the

   28        literature including concentration of occupied Ah receptor (Jusko, 1995), induced CYP1A2

   29        (Andersen et al, 1997; Kohn, 1993) and reduced epidermal growth factor receptor (EGFR)

   30        (Portier and Kohn, 1996). Considering the variety of endpoints seen with TCDD, and expected

   31        with other dioxin-like chemicals, in different species, it is unlikely that a single dose metric will

 

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      t         be adequate for interspecies extrapolation for all of these endpoints.  The issue of an appropriate

      2        dose metric for developmental effects considering the potential for a narrow time window of

      3        sensitivity, for instance, has been discussed in a number of places in this document.

      4        Furthermore, the use of different dose metrics with respect to the same endpoint may lead to

      5        widely diverse conclusions.  This latter point is discussed in more detail in Part 2, Chapter 8.

      6        Nevertheless, it is possible to express dose in a form that allows for comparison of responses for

      7        selected endpoints and species.  This can be done by either choosing a given exposure and

      8        comparing responses (e.g. unit risk level (URL)) or choosing a particular response level and

      9        comparing the associated exposures (e.g. effective dose (ED)).

    10                  As discussed above: dose can be expressed in a number of ways.  For TCDD and other

    11        dioxin-like compounds, attention has focused on the consideration of dose expressed as daily

    12        intake (ng/kg/day), body burden (ng/kg), or area under the plasma concentration versus time

    13        curve (AUC)  (Devito et al, 1995; Aylward et al, 1996).  While the AUC may be a more precise

     14       dose metric, the concept of physiological time (lifetime of an animal) complicates the

    15        extrapolation, as the appropriate scaling factor is uncertain for toxic endpoints.  Because body

    16        burden incorporates differences between species in TCDD half-life (these differences are large

     17       between rodent species and humans (See Table 8.2)), this dose metric appears to be the most

     18       practical for this class of compounds (Devito et al, 1995).  Average lifetime body burden is best

    19        suited for steady-state conditions, with difficulties arising when this dose metric is applied to

    20        evaluation of acute exposures, such as those occurring in the 1976 accidental exposure of some

    21        people living in Seveso, Italy (Bertazzi and di Domenico, 1994).  In cases such as this, increased

    22        body burden associated with the acute exposure event is expected to decline (half-life for TCDD

    23        is approximately 7 years) until it begins to approach a steady state level associated with the much

    24        smaller daily background intake. However, this issue of acute exposure is not a major factor in

    25        the current analyses.  In general, daily excursions in human exposure are relatively small and

    26        have minor impact on average body burden. Instead, physiologically-based pharmacokinetic

    27        (PBPK) models suggest that human body burdens increase over time and begin to approach

    28        steady state after approximately 25 years with typical background doses.  Occupational

    29        exposures represent the middle ground where daily excursions during the working years can

    30        significantly exceed daily background intakes for a number of years, resulting in elevated body

 

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          1        burdens.  This is illustrated in Table 5-1.  Estimation of the range and mean or median of

          2        "attained" body burden in accidentally- or occupationally exposed cohorts is presented and

          3        compared to body burdens based on background exposures.  These data are presented graphically

          4        in Figure 5-1. As discussed earlier, using background of total body burden (TEQDFP-WHO98) as a

          5        point of

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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       I         comparison, these often- termed "highly exposed" populations have maximum body burdens

       2        which are relatively close to general population backgrounds at the time.  When compared to

       3        background body burdens of the late 1980s, many of the median values and some of the mean

       4        values fall within a range of one order of magnitude (factor of 10) and all fall within a range of

       5        two orders of magnitude (factor of 100).  General population backgrounds at the time are likely

       6        to have been higher.  Since these are attained body burdens, measured at the time of the Seveso

       7        accident or back-calculated to the time of last known elevated exposure, being compared to

       8        background, average lifetime body burdens in these cohorts will be even closer to lifetime

       9        average background levels. This will be important if, as demonstrated for some chronic effects in

      10       animals and as assumed when relying on average body burden as a dose metric, cancer and other

      11        noncancer effects are a consequence of average tissue levels over a lifetime.  Body burdens begin

      12        to slowly decline soon after elevated exposure ceases. Some data in humans and animals suggest

      13        that elimination half-lives for dioxin and related compounds may be dose dependent, with high

      14        doses being eliminated more rapidly than lower doses.  Nonetheless, the use of an approximately

      15        7 year half-life of elimination presents a reasonable approach for evaluating both back-calculated

      16        and average lifetime levels since for most cohorts the exposure is primarily to TCDD.

      17                  The ability to detect effects in epidemiologic study is dependent on a sufficient difference

      18        between control and exposed populations. The relatively small difference (<10-100 fold)

      19        between exposed and controls in these studies makes exposure characterization in the studies a

      20        particularly serious issue. This point also strengthens the importance of measured blood or tissue

      21        levels in the epidemiologic analyses, despite the uncertainties associated with calculations

      22        extending the distribution of measured values to the entire cohort and assumptions involved in

      23        back-calculations.

      24                   Characterization of the risk of exposure of humans today remains focused on the levels

      25        of exposure that occur in the general population, with particular attention given to special

      26        populations (See Part 1).  For evaluation of multiple endpoints and considering the large

      27        differences in half-lives for TCDD across multiple species, it is generally best to use body burden

      28        rather than daily intake as the dose metric for comparison unless data to the contrary is presented.

      29        Further discussion of this point which provides the rationale for this science-based policy choice

      30        is presented in Part 2, Chapters  1  and 8.

      31

 

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        1         Calculations of Effective Dose (ED)

        2                  Comparisons across multiple endpoints, multiple species and multiple experimental

        3        protocols are too complicated to be made on the basis of the full dose-response curve.  As

        4        discussed above, comparisons of this sort can be made by either choosing a given exposure and

        5        comparing tine responses, or choosing a particular response level and comparing the associated

         6       exposures, in the analyses contained in Chapter 8 and elsewhere in the reassessment,

         7       comparison of responses are made using estimated exposures associated with a given level of

         8       excess response or risk.  To avoid large extrapolations, this common level of excess risk was

         9       chosen such that for most studies, the estimated exposure is in or near the range of the exposures

       I0        seen in the studies being compared with extra weight given to the human data.  A common

       11       metric for comparison is the effective dose or ED, which is the exposure dose resulting in an

        I2       excess response over background iii the studied population.  The USEPA has suggested this

        13      approach in calculating Benchmark Doses (BMD) (Allen et al, 1994) and in its proposed

        14      approaches to quantifying cancer risk (EPA, 1996).  While effective dose evaluation at the 10%

        15        response level (ED_0 or lower bound on ED_0 (LED_0)) is somewhat the norm, given the power of

        16        most chronic toxicology studies to detect an effect, this level is actually higher than those

        17        typically observed in the exposed groups in studies of TCDD impacts on humans.  To illustrate,

        18        h.tng cancer mortality has a background lifetime risk of approximately 4% (smokers and

        19        nonsmokers combined), so that even a relative risk of 2.0 ( 2 times the background lifetime risk)

       20        represents approximately a 4% increased lifetime risk.  Based upon this observation and

       21         recognizing that many of the .TCDD-induced endpoints studied in the laboratory include 1%

       22        effect levels in the experimental range, Chapter 8 presents effect/ye doses of 1% or EDm.  The

       23         use of  ED values below  10% is consistent with the Agency's guidance on the use of mode-of-

       24        action in assessing risk, as described in the evaluation framework discussed in Section 3.3., in

       25         that the observed range for many "key events" extends down to or near the 1% response level.

       26        Deten'nining the dose at which key events for dioxin toxicity begin to be seen in a heterogeneous

       27        human population provides important information for decisions regarding risk and safety.

       28

       29        5.2 Empirical Modeling of Individual Data Sets

       30                  As described in Chapter 8, empirical models have advantages and disadvantages relative

 

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        I         to more ambitious mechanism-based models.  Empirical models provide a simple mathematical

        2        model that adequately describes the pattern of response for a particular data set and can also

        3        provide the means roi' hypothesis testing and interpolation between data points. In addition, they

        4        can provide qualitative insights into underlying mechanisms. However, the major disadvantage is

        5        their inability to quantitatively link data sets in a mechanistically meaningful manner.  Data

        6        available for several biochemical and toxicological effects of TCDD, and on the mechanism of

        7        action of this chemical, indicate that there is good qualitative concordance between responses in

        8        laborato_5' animals and humans (see Table 1).  For example, human data on exposure zmd cancer

        9        response appear to be qualitatively consistent with animal-based risk estimates derived from

       10        carcinogenicity bioassays (See Part 2, Chapter 8). These and other data presented tl'u'oughout this

       11         reassessment would suggest that animal models are generally an appropriate basis for estimating

       12        human responses.  Nevertheless, there are clearly differences in exposures and responses between

       13        animals and humans, and recognition of these is essential when using animal data to estimate

       14        human risk.  The level of confidence ir! any prediction of human risk depends on the degree to

       15        which the prediction is based on an accurate description of these interspecies extrapolation

       16        factors.  See Chapter 8 for a further discussion of this point.

        17                  Almost all data are consistent with the hypothesis that the binding of the TCDD to the Ah

       18        receptor is the first step in a series of biochemical, cellular, and tissue changes that ultimately

       19        lead to toxic responses observed in both experimental animals and humans (See Part 2, Chapter

       20        2).  As such, an analysis of dose-response data and models should use, whenever possible,

       21         information on the quantitative relationships between ligand (i.e. TCDD) concentration, receptor

       22        occupancy, and biological response.  However, it is clear that multiple dose-response

       23        relationships are possible when considering ligand-receptor mediated events.  For example,

       24        dose-response relationships for relatively simple responses, such as enzyme induction, may not

       25        accurately predict dose-response relationships for complex responses such as developmental

       26        effects and cancer.  Cell- or tissue-specific factors may determine the quantitative relationship

       27        between receptor occupancy and the ultimate response.  Indeed, for TCDD there is much

       28        experimental data from studies using animal and human tissues to indicate that this is the case.

       29        This serves as a note of caution as empirical data on TCDD are interpreted in the broader context

       30        of complex exposures to mixtures of dioxin-like compounds as well as to non-dioxin-like

 

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    I         toxicants.

    2                  As for other chemical mechanisms where high biological potency is directed ttu'ough the

    3        specific and high affinity interaction between chemical and critical cellular target, the

    4        supposition of a response threshold roi' receptor-mediated effects is a subject for scientific

    5        debate.  The basis of this controversy has been recently sun'u'narized (Sewall and Lucier, 1996).

    6                  Based on classic receptor theory, the occupancy assumption states that the magnitude of

    7        biological response is proportional to the occupancy of receptors by drug molecules.  The

    8        'typical' dose-response curve for such a receptor-mediated response is sigmoidal when plotted on

    9        a semilog graph or hyperbolic if plotted on a arithmetic plot.  Implicit in this relationship is

   10        low-dose linearity (0-10°_, fi'actional response) ti'u'ough the origin. Although the law of mass

   11         action predicts a single molecule of ligand can interact with a receptor, thereby inducing a

   12        response, it is also stated that there must be some dose that is so low that receptor occupancy is

   13        trivial and therefore no perceptible response is obtainable.

   14                  Therefore, the same receptor occupancy assulnption of the classic receptor theow is

   15        interpreted by different parties as support for and against the existence of a threshold.  It has been

   16        stated that the occupancy assulnption cannot be accepted or rejected on experinaental or

   17        theoretical grounds (Goldstein et al,  1974).  To detennine the relevance of receptor interaction

   18        for TCDD-mediated responses, one must consider (1) altematives as well as limitations of the

   19        occupancy theory; (2) molecular factors contributing to  measured endpoints; (3) limitations of

   20        experimental methods; (4) contribution of measured effect to a relevant biological/toxic

   21         endpoint; and (5) background exposure.

   22                  Throughout this reassessment, each of these considerations has been explored within the

   23        cul-rent context of the understanding of the mechanism of a action of TCDD, of the

   24        methods for analysis of dose-response for cancer and noncancer endpoints, and of the available

   25        data sets of TCDD dose and effect for several rodent species, and humans that were

   26        occupationally exposed to TCDD at levels exceeding the exposure of the general population.

   27        5.2.1  Cancer

   28                  As described in Section 2.2.2.4 above, TCDD has been classified as a known human

   29        carcinogen, mad is a carcinogen in all species and strains of laboratory animals tested.  The

   30        epidemioiogical database for TCDD, described in detail in Part 2, Chapter 7a, suggests that

 

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    1         exposure may be associated with increases in all cancers combined, in respiratory tumors, and,

    2        perhaps, in soft-tissue sarcoma.  Although there are sufficient data in animal cancer studies to

    3        model dose-response for a number of tumor sites, as with mm2y chemicals, it is generally

    4        difficult to find human data with sufficient information to model dose-response relationships.

    5        For TCDD, there exist tI'n'ee studies of human occupational exposure which provide enough

    6        information to perfon-n a quantitative dose-response analysis. These are the NIOSH Study

    7        (Fingerhut et al, 1991), the Hamburg Cohort Study (Manz et al, 1991), and the BASF Cohort

    8        Study (Zober et al, 1990). In Part 2, Chapter 8, simple empirical models were applied to these

    9        studies for which exposure-response data for TCDD are available in human populations.

   10                  Modeling cancer in humans uses slightly different approaches than those used in

   11         modeling animal studies. The modeling approach used in the analysis of the human

   12        epidemiology data for all ca]cers comb/ned and lung cancer involves applying estimated human

   13        body burden to cancel' rcsponse, and estimating parameters in a linear risk model for each data

   14        set.  A linear risk model was used since the number of exposure groups available for analyses

   15        was too small to support more complicated models.  Because of this, evaluating the shape of the

   16        dose-response data for the human studies was not done.  Access to--th_ raw data may make it

   17        possible to use more complicated mathematical forms which allow for the evaluation of shape.

   18        In the one case in which this has been done, the dose-response shape suggested  response which

   19        was less than linear (dose raised to a power <1) (Becher et al, 1998).  For these studies, there are

   20        several assumptions and uncertainties involved in the modeling of the data including

   21         extrapolation of dosage, both in back-calculation and in elimination ldnetics, and the type of

   22        extrapolation model employed,

   23                   As described in Part 2, Chapter 8, the data used in the analyses are from Aylward et al.

   24        (1996) for the NIOSH study, Flesch-lanys et al. (1998) for the Hamburg cohort, and Ott and

   25        Zober (1996a;I996b) for the BASF cohort. The limited information available from these studies

   26        is in the form of standard mortality ratios (SMRs) and/or risk ratios by exposure subgroups with

   27        some estimate of cumulative subgroup exposures. Exposure subgroups were defined either by

   28        number of years of exposure to dioxin-yielding processes or by extrapolated TCDD levels. No

   29        study sampled TCDD blood serum levels for more than a fraction of their cohort and these

   30        samples were generally taken decades after last known exposure. In each study, serum fat or

 

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        i         body fat levels of TCDD were back calculated using a first-order model.  The assumed half-life

       2        of TCDD used in the model varied from study to study.  Aylward et al. used the average TCDD

       3         levels of those sampled in an exposure subgroup to represent the entire subgroup. Flesch-Janys et

       4        al. and Ott and Zober performed additional calculations, using regression procedures with data

        5        on time spent at various occupational tasks to estimate TCDD levels for all members of their

       6        respective cohorts. They then divided the cohorts into exposure groups based on the estimated

        7        TCDD levels The information presented in the literature cited above was used to calculate

        8        estimated average TCDD dose levels in Chapter 8.

        9                  To provide EDc;_ estimates for comparison in Chapter 8, Poisson regression (Breslow and

      10        Day,  1987) was used to fit a linear model to the data described above. Analysis of animal cancer

      11         data suggests a mixture of linear and non-linear responses with linear shape parameters

      12        predominating; complex responses to TCDD, both cancer and noncancer, are more often than

      13        not nonlinear.  Besides the issue of use of a linear model, several other important uncertainties

       14        discussed in Chapter 8 are the representativeness and precision of the dose estimates that were

       15        used, the choice of half-life and whether it is dose dependent, and potential interactions between

      16        TCDD and smoking or other toxicants.  Nevertheless, with these qualifications, it is possible to

      17        apply simple empirical models to studies in which exposure data for TCDD are available in

      18        human populations.

      19                  The analysis of these three epidemiological studies of occupationally exposed individuals

      20        suggest an effect of TCDD on all cancers, and hmg cm-tcers in the adult human male.  The EDs0_

      21         based upon average excess body burden of TCDD ranged from 6 ngTCDD/kg to 161 ngTCDD/kg

      22        in humans. The lower bounds on these doses (based on a modeled 95% C.I.) range from 3.5

      23        ngTCDD/kg to 77 n.gTCDD,q<g. For the effect of TCDD on lung cancers, the only tumor site

      24        increased in both rodents and humans, the human EDso_ ranged fi'om 24 ng/kg to 161 rig/kg.

      25        The lower bounds on these doses (based on a modeled 95% C.I.) range from 10.5 ngTCDD/kg

      26        to 77 ngTCDD/kg. These estimates of EDs0_ are compared to animal estimates later in this

      27        discussion.

      28                  Both empirical and mechanistic models were used to examine cancer dose-response in

      29        animals.  Port/er et a1.(1984) used a simple multistage model of carcinogenesis with up to two

      30        mutation stages affected by exposure to model the five tumor types observed to be increased in

 

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       1        the 2 year feed study of Kociba et al. (Sprague-Dawley rats) (1978) and the eight tumor types

       2        observed to be increased in the 2 year gavage cancer study conducted by the National Toxicology

       3        Program (Osbome-Mendel rats and B6C3F_ mice) (1982).  The findings from this analysis which

       4        examined cancer dose-response within the range of observation are presented in Table 8.3.2.

       5        which is reproduced with slight modifications as Table 5-2 below.  All but one of the estimated

       6        EDs0_ are above the lowest dose used in the experiment  (approximately 1 ngTCDD/kg/day in

       7        both studies) and are thus interpolations rather than extrapolations.  The exception, liver cancer

       8        in [emale rats from the Kociba study, is very near the lowest dose used in this study and is only a

       9        small extrapolation (from 1 ngTCDD/kg/day to 0.77 ngTCDD/kg/day).  Steady-state body

      10        burden calculations were also used to derive doses for comparison across species.  Absorption

      11         was assumed to be 50% for the Kociba et al. study (feed experiment) and 100% for the NTP

      12        study  (gavage experiment).  Also presented in Table 5-2 are the shapes of the dose-response

      13        curves as detemlined by Portier et al (1984).

      14

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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        1         this does not imply that a non-linear model such as the quadratic or cubic would not fit these

        2        data.  In fact, it is unlikely that in m'_y one case, a linear model or a quadratic model could be

        3         rejected statistically for these cases. These studies had only three experimental dose groups hence

        4        these shape calculations are not based upon sufficient doses to guarantee a consistent estimate;

        5        they should be viewed with caution. The ED0_ steady state body burdens range from a Iow value

        6        of 14 ng/kg based upon the linear model associated with liver tumors in female rats to as high as

        7         1190 ng/kg based upon a cubic model associated with thyroid follicular cell adenomas in female

        8        rats. Lower bounds on the steady state body burdens in the animals range from 10 ngTCDD/kg to

        9        224 ng/kg. The corresponding estimates of daily intake level at the ED0_ obtained from an

       10        empirical linear model range fi'om 0.8 to 43 ngTCDD/kg body weight/day depending on the

       11         tumor site, species and sex of the animals investigated. Lower confidence bounds on the

       12        estimates of daily intake level at the ED0_ in the animals range from 0.6 to 14 ngTCDD/kg body

       13        weight/day.  In addition, using a mechanistic approach to modeling, Portier and Kohn (1996)

       14        combined the biochemical response model of Kolm et al. (1993) with a single initiated

       15        phenotype two stage model of carcinogenesis to estimate liver tumor incidence in female

       16        Sprague-Daw!ey rats fi'om the two-year cancer bioassay of Kociba et al. (197'8-57. By way of

       17        comparison, the ED0t estimate obtained from this linear mechanistic model of liver tm'nor

       18        induction in female rats was 0.15 ngTCDD/kg body weight/day based on intake, which is

       19        equivalent to 2.7 ngTCDD/kg steady state body burden. No lower bound on this modeled

       20        estimate of steady state body burden ,,vas provided

       21                  As discussed in Part 2, Chapter 8, different dose metrics can lead to widely diverse

       22        conclusions.  For example, as described in Chapter 8, the ED0: intake for the animal tumor sites

       23        presented above ranges [rom less than one to tens of ng/kg/day m'_d the lowest dose with an

       24        increased tumorigenic response (thyroid tun'mrs) in a rat is 1.4 ng/kg/day (NTP, 1982). The daily

       25        intake of TCDD in hun-tans is estimated to be 0.14 to .4 pgTCDD/kg/day. This implies that

       26        humans are exposed to doses 3,500 to 10,000 times lower thml the lowest tumorigenic dose in rat

       27        thyroid.  However, 1.4 ng&g/d in the rat leads to a steady state body burden of approximately

 

 

 

 

 

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        1         25 ng/kg, assuming a half-life of TCDD of 23 days and absorption from feed of 50%'.  If the

       2        body burden of TCDD in humans is approximately 5 ngTCDD/kg lipid or 1.25 ng/kg body

       3        Weight (assuming about 25% of body weight is lipid),  humans are exposed to about 20 times

       4        less TCDD than the minimal carcinogenic dose for the rat.  If total TEQ is considered the

        5        difference is even less, approaching only a factor of 2 difference.  The difference between these

        6        two estimates is entirely due to the approximately 100 fold difference in the half-life between

        7        humans and rats.  At least for this comparison, if cancer is a function of average levels in the

        8        body, the most appropriate metric for comparison is the average or steady state body burden

        9        since the large differences in animal to human half-life are accounted for.

       10                  Comparisons of human and animal EDs0_ from Part 2, Chapter 8 for cancer response on a

       11         body burden basis show approximately equal potential for the carcinogenic effects of TCDD.  In

       12        humans, restricting the analysis to log-linear models ill Part 2, Chapter 8, resulted in cancer

       13        EDso_ ranging fi'om 6 ng/kg to  161  ng/kt,.  This was similar to the empirical modeling estimates

       14        from the animal studies, which ranged from 14 ng/kg to 1190 ng/kg  (most estimates were in the

       15        range from 14 to 500 ng&g).  The lower bounds on the human body burdens at the EDso, (based

       16        on a modeled 95% C.I.) range from 3.5 ngTCDD/kg to 77 ngTCDD/kg. Lower bounds On the

       17        steady state body burdens m the animals range from 10 ngTCDD/kg to 224 ng/kg..  The estimate

       18        for the single mechanism-based model presented earlier (2.7 ng/kg) was approximately 2 times

       19        lower than the lower end of the range of human ED01 estimates and less than one times less than

       20        the lower bound on the ED0, (LED0_).  The same value was approximately 5 times lower than the

       21         lower end of the range of animal ED0, estimates and less than 4 times less than the LEDm.

       22                  Using human and animal cancer EDs0_, their lower bound estimates, and the value of 2.7

       23        ngTCDD/kg from the single mechanism based model, slope factors and comparable risk

       24        estimates for a human background body burden of approximately 5 ngTEQ/kg (20 ngTEQ/kg

       25         lipid) can be calculated using the following equations:

       26

 

 

                  I

 

                                   steady state body burden (ng/kg) =( daily dose (nw'kg/day) * (half-life)/'Ln(2)) (f)    where fis the fraction

                  absorbed from the exposure route (umtless) and half-life is the half-life in days.

 

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        1         Equation 5-1 Calculating Slope Factors from Body Burdens at the ED0_

        2

        3        Slope Factor (per pgTEQ/kgBW/day) = Risk at ED0_ / Intake (pgTEQ/kgBW/day) Associated

        4        with Human Equivalent Steady State Body Burden at ED0_

        5        where:

        0        Risk at ED_: = 0.01; and

        7        Intake (pgTEQ/kgBW/day) = [Body Burden at ED01 (ngTEQ/kg)*halflife (days')] * f

        8                                                                             Ln(2)

        9        half life = 2593 days in humans and 25 days in rats (See Table 8.1 in Part 2, Chapter 8)

      10        f= fraction of dose absorbed; it is assumed to be 50% for absorption from food (Kociba et al.,

      11         1976) and 100% for other routes.

      12

       13        Equation 5-2 Calculating Upper Bound on Excess Risk at Human Background Body

       14        Burden

      15

       16        Upper Bound on Excess Risk at Humm_ Background Body Burden = ( Human-13aekground Body

       17        Burden ( ng/kg))(risk at EDu_)/lower bound on Human Equivalent Steady State Body Burden

      18        (ng/kg) at EDoE

      19        where:

      20        Risk at ED0_ =0.01

      21

      22                  Use of these approaches reflects methodologies being developed within the context of the

      23        revised Cancer risk assessment guidelines.  Slopes are estimated by a simple proportional

      24        method at the" point of departure" (LED0_) at the Iow end of the range of experimental

      25        observation.  As discussed below, these methods can be compared to previous approaches using

      26        the hnearized multistage (LMS) procedure to determine if the chosen approach has significantly

      27        chm_ged the estimation of slope.  The estimates of ED0_/LED0_ represent the human-equivalent

      28        body burden for 1% excess cancer risk based on exposure to TCDD and are assumed for

      29        purposes of this analysis to be equal for TCDD equivalents (total TEQ).  This assumption is

      30        based on the toxicity equivalence concept discussed tlu-oughout this report and in detail on Part

 

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     1         2, Chapter 9. All cancer slope factors can be compared to the Agency's previous slope factor of

     2         1.6 X 10.4

                             per pgTCDD/kgBW/day (EPA,  1985).

     3

     4        Estimates of slope factors and risk at current background body burdens based on human

     5        data

     6                  Estimates of upper bound slope factors( pet' pgTCDD/kgBW/day) calculated from the

     7        human EDs0_ presented on Table 8.3.1 range fi'om 5.3 X 10'-_, if the LED0,  for all cancer deaths in

     8        the Hamburg cohort is used, to 2.4 X I0'4

                                                                                                 if the ED0_ for lung cancer deaths in the smaller BASF

     9        cohort is used.  ,4.11 of the other slope factors for all cancer deaths or lung cancer deaths in the

    10        thJ'e_: cohorts would fall within that range. LEDs0_ for all cancer deaths span approximately an

    11         order of magnitude and would generate slope factors in the range of 5 X I04 to5 X10'4.  Slightly

    12        smaller slope factors are generated when LEDs0, roi' lung cancer are used.  The largest slope

    13        factors based on LEDso_ come fi'om the Hamburg cohort (5.3 X 10.3 and 1.8 X 10.3 respectively

    I4        for all cancer deaths and lung cancer deaths.) These estimates compare well with the estimates of

    15        risk associated with TCDD exposure in the Hamburg cohort published by Becher (Becher etal.,

    16         1998).  The risk estimates of Becher et al. derived from data on TCDD exposure to male workers

    17        with a ten year latency and taking greater caution over other factors affecting risk including

    18        choice of model, latency, job category, dose metric and concmTent exposures.  These estimates

    19        range from 1.3 X10'_ to 5.6 X 10'3 per pg TCDD,q;gBW/day. In this analysis all excess cancers

    20        are attributed to TCDD exposure despite significant levels of other dioxin-like compounds in

    21        blood measurements of this cohort (see Table 5~1).  Although risk estimates using TCDD alone

    22         in this cohort  might suggest an overestimate of risk, no evidence for this emerged from the

    23        analysis and assuming that TCDD will still dominate total TEQ, differences in slope factor

    24        estimates are likely to be less than a factor of two and may not be discemable. Taking into

    25        account different sources of variation, Becher et al.(1998) suggest a range of 10'3 to 10'_  for

    26        additional lifetime cancer risk for a daily intake of 1 pgTCDD/kg BW/day.  By inference, that

    27        range couid also apply to total TEQ intake. As described in Section 4.4.2, current estimates of

    28        intake in the U.S. are estimated to be approximately 1 pgTEQ/kg BW/day.  Using Equation 5-2,

    29        the upper bound range of risks estimated from current human body burdens of 5 ngTEQ/kgBW

   30        (which equates to a serum level of 20 pg/g lipid (see Table 4.7)) based on all cancer deaths in the

 

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       1         tl'u'ee cohorts ranged from 1.4 X 10-2 tO  1.3 X 10'3; based on lung cancer deaths, the upper bound

       2        on the estimates of excess risk extended to 6 X10_.  The range of these estimates provides further

       3        support for the perspective on risk provided by Becher et al.  Uncertainties associated with these

       4        estimates fi'om human studies are discussed in Part 2, Chapter 8 and in Becher et al. (1998).

       5

       6         Estimates of slope factors and risk at current background body burdens based on animal

       7        data

        8                  Upper bound slope factors ( per pgTCDD/kgBW/day) for human cancer risk calculated

        9        from lower bounds on EDs0t (LEDs0.) for the animal cancers presented in Table 5-2 range from

      10         1.9  X 10'" to 8.4 X10'5.  This spans a range front being 12 times gTeater than the previous upper

      11         bound estimate on cancer slope ( 1.6 X10-4 (EPA,  1985)) to 2 times less.  The largest slope

      12         factor is derived from the same study as the 1985 estimate; that is, the slope factor derived from

      13        the female liver cancer in the Kociba et a1.(1978) study continues to give the largest slope factor.

      14        In attempting these comparisons, two issues became apparent. First, the body burden and the

       15         intake at the ED0_ from Pottier et al. (1994) does not result in the same slope factor as EPA

       16        (1985). Despite the use of the same study results, a slope factor of 1.8 X10'5 per

       17        pgTCDD,'kgBW/day using the LMS approach.  This is a factor of approximately 10 lower than

       18        the EPA (1985) estimate of the slope.  The differences are attributable to the aims of the

       19        respective calculations at the time. Pottier et al. (1984) calculated "virtually safe doses"

      20        assuming that rodent and human doses scaled on a mg/kg basis and he used the original tumor

      21         counts fi-om the study.  EPA (1985), on the other hand used (BW)TM

                                                                                                                                                                       to arrive at a human

      22        equivalent dose and used the pathology results from a re-read of the original Kociba study (

      23        Albert, 1980).  In addition, rrm'lot counts were adjusted for early mortality in the study.  The

      24        factor to adjusi for (BW)3/4-scaling in the rat is 5.8.  The correction for early mortality can be

       25        accounted for with a factor of 1.6 ( this is the ratio of the intake values at the ED0_ with and

      26        without the early mortality correction).  If the Portier et al. slope factor (1.8 X10'Sper

      27        pgTCDD/kgBW/day) is multiplied by these two factors, a slope of 1.7 X10'4 per

       28        pgTCDD/kgBW/day is calculated.  This is equivalent to the EPA (1985) estimate of 1.6 X10'qper

      29        pgTCDD/kgBW/day. Reconciling these issues is important to assure appropriate comparisons of

      30        slope factor estimates.

 

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      1         More important is tine calculation of slope factor estimates using current methods of analysis

      2        which recognize the importance of the dose metric and the differences in halfqife of dioxins in

      3        the bodies of laboratmLy animals and humans (See part 2, Chapter 8 for detailed discussion). The

      4        major difference between the approaches used to calculate risks in the mid-1980's (Portier et al,

      5         1994; EPA,  1985) and the current approach is the use of body burden as the dose metric for

      6        animal to human dose equivalence.  All things being equal, the use of body burden accounts for

      7        the approximately  100-fold difference between half-lives of TCDD in humans and rats (2593

      8        days versus 25 days (see Part 2, Table 8.1  )). Use of Equation 5-1 results in an estimated body

      9        burden at the LEDo_  of  6.1  ng TEQ/Kg to be derived from the EPA (1985) Kociba tumor counts.

     10        This compares favorably with the Portier estimate of 10 ng TEQ/Kg found in Table 5-2.  The

     11         difference is entirely accounted for by the early deaths adjustment by EPA (1985).  Use of these

     12        body burdens at the LED0_ result in slope factor estimates of 1.9 X10'3 per pgTCDD/kgBW/day

     13         and 3.1  X 10" per pgTCDD/kgBW/day roi' the Chapter 8 re'id the newly derived body burden,

     14        respectively.  Again, the difference is due solely to the adjustment for early mortality and EPA

     15        believes this provides a better estimate of upper bound lifetime risk than does the unadjusted.

     16        EPA's new slope· factor (3.1  X 10'_ per pgTCDD/kgBW/day) is 19 times lower than the slope

     17        factor from  1985.

     18                  A second issue with the modeling of the Kociba data relates to the appropriate tumor

     19        counts to use.  As mentioned in Section 2, Goodman and Sauer (1992) reported a second

    20        re-evaluation of the female rat liver tumors in the Kociba study using the latest pathology c_iteria

    21         for such lesions. Results of this review are discussed in more detail in Part 2, Chapter 6.  The

    22        review confin'ned only approximately one-third of the tumors of the previous review (Albert,

    23         1980).  While this finding did not change the determination of carcinogenic hazard since TCDD

    24        induced tumors in multiple sites in this study, it does have an effect on evaluation of

    25        dose-response and on estimates of risk.. Since neither the original EPA (1985) slope factor

    26        estimate nor that of Portier et al. (1984) reflect this re-read, it is important to factor these results

    27        into the estimate of the ED0_ and slope factor.  Using the LMS procedure which was used by the

    28        EPA in 1985 and the tumor counts as reported in Part 2, Chapter 6, Table 6.2, the revised slope

    29        factor is reduced by approximately 3.6-fold to yield a slope factor of 4.4 X 10'5  per

    30        pgTCDD/kgBW/day.  However, since the original estimates used a (BW)3/4-scaling, this n'mst be

 

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        1         adjusted to use body burden and obtain an appropriate result.  When dose is adjusted and

        2        Equation 5-1  is used, all L ED0_ of 22.2 ngTEQ/kg is derived and a slope factor of 8.3 X10'4 per

        3        pgTCDD/kgBW/day.  This represents EPA's most current upper bound estimate of human

        4        cancer risk based on animal data.  It is 5.2 times larger than the slope factor calculated in

 

        5        EPA(1985).  This number reflects the increase in slope factor based on use of the body burden

 

        6        dose metric (19 times greater) and the use of the Goodman and Sauer (I 992) pathology (3.6

        7        times less).

        8

        9        Estimates of slope factors and risk at current background body burdens based on a

       10        mechanistic model

       1 l                  As discussed above, Portier and Kohn (1996) combined the biochemical response model

       12        of Kohn et al. (1993) with a single initiated phenotype two stage model of carcinogenesis to

       13        estimate Iix,et' tumor incidence in female Sprague-Dawley rats from the Kociba et al. (1978)

       14        bioassay.  The model is described in more detail in Part 2, Chapter 8.  This model adequately fit

 

       15        the tumor data, although it overestimated the observed tumor response at the lowest dose in the

       16        Kociba study.  The shape of the dose-response curve was approximately linear and the eg it-_d

       17        ED0_ value for this model was 1.3 ng/kg/day.  The corresponding body burden giving a  1%

       18        increased effect was 2.7 ng/kg.  The model authors believe that the use of cYP1A2 as a dose

        19        metric for the first mutation rate is consistent with its role as the major TCDD-inducible estradiol

       20        hydrolase in liver and with it hypothesized role in the production of estrogen metabolites leading

       21         to increased oxidative DNA damage and increased mutation (Yager and Liehr, 1996; Hayes et al.,

       22         1996; Dan.nan et al., 1986;Roy et al., 1992).  Although no lower bound estimate of the ED0t is

       23        calculated, a maximum likelihood estimate of the slope factor can be calculated.  It is 7.1 X 10.3

 

       24        per pgTCDD&gBW/day. This estimate represents an example of the type of modeling, based on

       25        key events in a mode-of-action for carcinogenesis, which is consistent with future directions in

       26        dose-response modeling described in EPA's revised draft Cancer Risk Assessment Guidelines

       27        (EPA, 1999).  While a number of uncertainties remain regarding structure and parameters of the

       28        model, the slope estimate is consistent with those derived from humans and animals.  More

       29        detain on this model can be found in Part2, Chapter 8.

       30

 

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       1         5.2.2 Non Cancer Endpoints

       2                  At this point, sufficient data are not available to model noncancer endpoints in humans.

 

       3        Many studies are available to estimate ED0_ values for noncancer endpoints in animals. However,

       4        there are a number of difficulties and uncertainties that should be considered when comparing the

 

        5        same or different endpoints across species.  Some of these include differences in sensitivity of

 

        6        endpoints, times of exposure, exposure routes, species and strains, use of multiple or single

        7        doses, and variability between studies even for the same response. The estimated EDs0_ may be

 

        8        influenced by experimental design, suggesting that caution should be used in comparing values

 

        9        fi'om different designs. In addition, caution should be used when comparing studies that

 

       10        extrapolate EDs0_ outside the experimental range.  Furthermore, it may be difficult to compare

       11         values across endpoints.  For example, the hm-nan health risk for a 1% change ofbody weight

       12        may not be equivalent to a 1% change in enzyme activity.  Finally, background exposures are not

 

       13        often considered in these calculations simply because they were not known.  The latter

       14        consideration is particularly important since the inclusion of these may alter the shape of the

       15        dose-response curve, possibly increasing the shape parameter so that the responses would

       16        demonstrate more threshold-like effects.[Chris- This needs to be explained more fully.]

       17                  Nevertheless, given these considerations several general trends were observed arid

       18        discussed m Part 2, Chapter 8.  The lowest EDs0_ tended to be for biochemical effects, followed

       19        by hepatic responses, immune responses arid responses in tissue weight.  An analysis of shape

       20        parm'neters implies that many dose-response curves are consistent with linearity over the range of

 

       21         doses tested.  This analysis does not imply that the cuD'es would be linear outside this range of

 

       22        doses but it does inform the choices for extrapolation.  This is particularly true when body

       23        burdens or exposures at the lower end of the observed range are close to body burdens or

       24        exposures of interest for humans, wkich is the case with dioxin-like chemicals.

       25        Overall, shape parameter data suggest that biochemical responses to TCDD are more likely to be

       26        linear within the experimental dose range, while the more complex responses are more likely to

       27        assume a nonlinear shape. However, a large number (greater than 40%) of the more complex

       28        responses have shape paralneters that are more consistent with linearity than non-linearity.

       29        The tissue weight changes seen  for animals (using only data sets with good or moderate

       30        empirical fits to the model) yielded a median ED0_ at average body burdens of 510 ng/kg in the

 

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     1         multidose studies (range;  11 to 28000 ng/kg) and a median EDm of 160 ng/kg ( range 0.0001 to

    2        9700 ng/kg)in the single dose studies. Toxicity endpoints from the single dose studies resulted in

     3        a median value at average body burdens of 4300 ng/kg  (range 1.3 to 1,000,000 ng/kg).  For

     4        tissue weight changes, 43% of the dose response curves exhibited linear response. In contrast, the

     5         toxicity endpoints fi'om the single dose studies exhibited predominantly non-linear responses

     6        (80%). All multi-dose studies demonstrated a greater degree of linear response (41%) than did

     7        single dose studies (37%), especially for tissue weight changes m_d toxicity endpoints. (50%

     8        linear for multidose versus 34% for single dose). In general, it is not possible to dissociate the

     9        differences between ca'_cer and non-cancer dose response as being due to differences in endpoint

   10        response or simply due to differences ill the length of dosing and exposure. Also, a greater

    11         percentage of the non-cancer EDs0_ were extrapolations below the lower range of the data (42%)

    12        than was the case for the cancer endpoints (8% in animals and no extrapolations in humans).

    13

    14        5.3 Mode-of-Action Based Dose-Response Modeling

    15                  As described in Chapter 8, mechanism-based modeling can be a powerful tool for

    16        understanding and combining infom_ation on complex biological systems.  Use of a truly,

    17        mechanism-based approach can, in theory, enable reliable and scientifically sound extrapolations

    18        to lower doses and between species. However, any scientific uncertainty about the mechanisms

    19        that the models describe is inevitably reflected in uncertainty about the predictions of the models.

   20        The assumptions and uncertainties involved in the mechanistic modeling described in Chapter 8

   21         are discussed at length in that Chapter and in cited publications.

   22                  The development and continued refinement of physiologically based pharmacokinetic

   23        (PBPK) models of the tissue dosimetry of dioxin have provided important infom_ation

   24        conceming the relationships between administered does and dose to tissue compartments (section

   25        8.2).  Aspects of these models have been validated in the observable response range for multiple

   26        tissue compartments, species, and class of chemical.  These models will continue to provide

   27        important new information for future revisions of this health assessment document.  Such

   28        information will likely include improved estimates of tissue dose for liver and other organs

   29        where toxicity has been observed, improved estimates of tissue dose(s) in humans, and improved

   30        estimates of tissue dose for dioxin related compounds.

 

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        1                   As a part of this reassessment, the development of biologically based dose-response

       2        (pharmacodynamic) models for dioxin and related compounds has lead to considerable and

        3        valuable insights regardingboth mechanisms of dioxin action and dose-response relationships for

        4        dioxin effects.  These efforts, described in some detail in Chapter 8, have provided additional

        5        perspectives on traditional methods such as the linearized multistage procedure for estimating

        6        cancer potency or tine uncertainty factor approach for estimating levels below which noncancer

        7        effects are not likely to occur.  These methods have also provided a biologically based rationale

        8        for what had been primarily statistical approaches.  The development of models like those in

        9        Chapter 8 allows for an iterative process of data development, hypotheses testing and model

       10        development.

       11

       12        5.4 Summary Dose-Response Characterization

       13                  All humans tested contain detectable body burdens of TCDD and other dioxin-like

       14        compounds that are likely to act tl'u'ough the same mode-of-action.  It is possible that any

       15        additional exposure above cun'ent background body burdens will be additive to ongoing

       16        responses.  The magnitude ofthe additional response will be a function of the toxicitY

       17        equivalence of the incremental exposure.  This observation, the relatively small margin of

       18        exposure fol' "key events", and the high percentage of observed linear responses suggests that a

       19        proportional model should be used when extrapolating beyond the range of the experimental

       20        data.  Short of extrapolating to estimate risk in the face of uncertainties described above, a simple

       21        margin of exposure approach may be useful to decision-makers when discussing risk

       22        management goals.  However, this decision would have to be based upon a policy choice since

       23        this analysis does not strongly support either choice.

       24                  Because human data for cancer dose-response analysis were available and because of a

       25        strong desire to stay within the range of responses estimated by these data, the risk chosen for

       26        deten'nining a point of departure was the 1% excess risk.  Doses and exposures associated with

       27        this risk (the EDs0t ) were estimated from the available data using both mechanistic and empirical

       28        models.  Comparisons were made on the basis ofbody burdens to account for differences in

       29        half-life across the numerous species studied.

       30                  In humans, restricting the analysis to log-linear models resulted in cancer EDs0_ ranging

 

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      1         from 6 ng/kg to 161 ng/kg.  This ,,,,,as similar to the estimates, from empirical modeling, fi'om the

      2         animal studies which ranged from 14 ng,q(g to 1190 ng/kg  (most estimates were in the range

      3         from 14 to 500 ng/kg), and 2.7 ng/kg for the single mechanism-based model.  Lower bounds on

      4        these ED0_ estimates were used to calculate upper bound slope factors and risk estimates for

      5         average background body burdens.  These estimates are presented above. Upper bound slope

      6        factors allow the calculation of the probability of cancer risk for the highly vulnerable in the

      7        population (estimated to be the top 5% or greater.  V_qfile there may be individuals in the

      S        population who might experience a higher cancer risk on the basis of genetic factors or other

      9        determinants of cancer risk not accounted for in epidemiologic data or animal studies, the vast

     10        majority of the population is expected to have less risk per unit of exposure and some my have

     11         zero risk.  Based on these slope factor estimates (per pgTEQ/kgBW/day), average current

     12        background body burdens (5 ng/kgBW) which result from average intakes of approximately 3

     13        pgTEQ/kgBW/day are in the range of 10'3 to 10-2.  A very small percentage of the population

     14        (< 1%) may' experience risk which are 2-3 times higher than this if they are both the most

     15        vulnerable and the most highly exposed (among the top 5%) based on dietary intake of dioxin

     16        and related compounds. This range of upper bound risk for the general population has increased         ............

     17        an order of magnitude from the risk described at background exposure levels based on EPA's

     18        draft of this reassessme[lt (10'4-  10'3) (EPA,  1994).

     ] 9                  Estimates for non-cancer endpoints showed much greater variability,  ranging over 10

     20        orders of magnitude.  In general, the rloncancer endpoints displayed lower EDs0_ for short-term

     21         exposures versus longer tem-_ exposures, and for simple biochemical endpoints versus more

     22        complex endpoints such as  tissue weight changes or toxicity.  In addition, the noncancer

     23        endpoints generally displayed higher estimated EDs0_ than the cancer endpoints, with most

     24        estimates ranging from  100 ng/kg to  100,000 ng/'kg.  The mechanism-based models for

     25        noncancer endpoints gave a lower range of EDs0_ (0.17 to 105 ng/kg). While most of these

     26        estimates were based upon a single model the estimate from the hepatic zonal induction model

     27        gave an ED0_  for CYP1A2 induction of 51 ng/kg and hence was within the same range.

    28        These estimates, although highly variable, suggest that  any choice of body burden, as a

     29        point-of-departure, above 100 ng/kg would likely yield greater than 1% excess risk for some

    30        endpoint in humans.  Also, choosing of a point-of-departure below 1 ng/kg would likely be an

 

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       1         extrapolation below the range of these data and would likely represent a risk of less than 1%.

 

      2        Any choice in the middle range of l ngfkg to  100 ng/kg, would be supported by the analyses,

       3        although the data provide the greatest support in the range of 10rig/kg to 50 n_d'kg.

       4

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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       1         6.0      RISK CHAILatCTERIZATION

       2

       3                  Characterizing risks from dioxin and related compounds requires the integration of

       4  .      complex data sets and the use of science-based inferences regarding hazard, mode-of-action,

       5        dose-response and exposure.  It also requires consideration of incremental exposures in the

       6        context of'an existing background exposure which is, for the most part, independent of local

       7        sources and dominated by exposure tl-u-ough the food supply.  Finally, this characterization must

        8        consider risks to special populations and developmental stages (subsistence fishers, children,

        9        etc.) as well as the general population.  It is important that this characterization convey the

      10        current understanding of the scientific community regm'ding these issues, highlight uncertainties

      11         in this understanding, and speci_ where assumptions or inferences have been used in the absence

      12        of data.  Although characterization of risk is inherently a scientific exercise, by its nature, it must

      13        go beyond empirical observations and draw conclusions in areas which are untested.  In some

      14        cases, these conclusions are, in fact, untestable given the current capabilities in mlalytical

      15        chemistry, toxicology and epidemiology.  This situation should not detract from our confidence

       16        in a well structured and documented characterization of risk but should se_'e to confirm the

      17        importance of considering risk assessment as an iterative process which benefits from evolving

       18        methods and data collection.

       19

      20        Dioxin and related compounds can produce a wide variety of effects in animals and might

      21         produce many of the same effects in humans.

      22                  There is adequate evidence based on all available information discussed in Parts  1 and 2

      23        of this reassessment, as well as that discussed in this Integ-rated Sun'unary, to support the

      24        inference that humans are likely to respond with a broad spectrum of effects from exposure to

      25        dioxin and related compounds.  These effects will likely range from biochemical changes at or

      26        near background levels of exposure to adverse effects with increasing severity as body burdens

      27        increase above background levels.  Enzyme induction, chm'lges in hormone levels and indicators

      28        of altered cellular function seen in humans and laboratory animals represent examples of effects

      29        of unlcnown clinical significance but which may be early indicators of toxic response.  Induction

      30        of activating/metabolizing enzymes at or near background levels, for instance, may be adaptive,

      31         and in some cases, beneficial, or may be considered adverse.  Induction may lead to more rapid

 

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     1         metabolism and elimination of potentially toxic compounds, or may lead to increases in reactive

     2         intermediates and may potentiate toxic effects.  Demonstration of examples of both of these

     3        situations is available in the published literature and events of this type formed the basis for a

     4        biologically based model which is discussed in Section 5. Subtle effects, such as the impacts on

     5         neurobehav_ora- outcomes, thyroid function, and liver enzymes (AST and ALT) seen in the

     6        Dutch chl!dren exposed to background levels of dioxin and related compounds, or changes in

     7        circulating reproductive hormones in men exposed to TCDD, illustrate the types of responses

     8        that support the finding of arguably adverse effects at or near background body burdens. Clearly

     9        adverse effects including, perhaps, cancer may not  be detectable until exposures contribute to

    10        body burdens which exceed back_ound by one or two orders of magnitude (10 or 100 times).

    11         The mechanistic relationships of biochemical and cellular changes seen at or near background

    12        body burden levels to production of adverse effects detectible at higher levels remains uncertain

    13        but data are accumulating to suggest mode-of-action hypotheses for further testing.

    14                  It is well known that individual species vary in their sensitivity to any particular dioxin

    15        effect.  However, the evidence available to date indicates that humans most likely fall in the

    16        middle of the range of sensitivity for individual effects among animals rather than at either

    17        extreme.  In other words, evaluation of the available data suggests that humm'm, in general, are

    la        _cithc_ extrcnx:ly sensitive nor insensitive to the individual effects of dioxin-like compounds.

    19        Human data provide direct or indirect support for evaluation of likely effect levels for several of

    20        the endpoints discussed in the reassessment although the influence of variability among hm'nans

    21         remains difficult to assess.  Discussions have highlighted certain prominent, biologically

    22        significant effects of TCDD and related compounds.  In TCDD-exposed men, subtle changes in

    23        biochemistry and physiology such as enzyme induction, altered levels of circulating reproductive

    24        hormones, or reduced glucose tolerar_ce and, perhaps, diabetes, have been detected in a limited

    25        number of epidemiologic studies.  These findings, coupled with knowledge derived from animal

    26        experiments, suggest the potential for adverse impacts on human metabolism, and developmental

    27        and/or reproductive biology, and, perhaps, other effects in the range of current human exposures.

    28        These biochemical, cellular, and  organ-level endpoints have been shown to be affected by

    29        TCDD, but specific data on these endpoints do not generally exist for other congeners.  Despite

    30        this lack of congener specific data, there is reason to infer that these effects may occur  for all

    31         dioxin-like compounds, based on the concept of toxicity equivalence.

 

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       1                   In this volume, dioxin and related compounds are characterized as carcinogenic,

       2        developmental, reproductive, immunological and endocrinological hazards.  The deduction that

 

       3        humans are likely to respond with noncancer effects from exposure to dioxin-like compounds is

 

       4        based on the fundamental level at which these compounds impact cellular regulation and the

 

       5        broad range of species which have proven to respond with adverse effects.  Since, for example,

       6        developmental toxicity following exposure to TCDD-iike congeners occurs in fish, birds, and

 

       7        mammals, it is likely to occur at some level in humans.  It is not currently possible to state

 

        8        exactly how or  at what levels individuals will respond with specific adverse impacts on

 

        9        development or reproductive function, but analysis of the Dutch cohort data and laboratory

      10        animal studies suggest that some effects may occur at or near background levels.  Fortunately,

      11         there have been few human cohorts identified with TCDD exposures high enough to raise body

       12        burdens significantly over background levels (See Table 5- 1 and Figure 5-1 in Section 5), and

 

       13        when these cohorts have been examined, relatively few clinically significant effects were

 

       14        detected.  'Fine lack of exposure gradients, adequate human information and the focus of most

 

       15        currently available epidemiologic studies on occupationally, TCDD-exposed adult males makes

 

       16        evaluation of the inference, that noncancer effects associated with exposure to dioxin-like

 

       17        compounds may be occurring, difficult. It is important to note, however, that when exposures to

 

       18        very high levels of dioxin-like compounds have been studied, such as in the Yusho and Yu- ·

 

       19        Cheng cohorts, a spectrLtm of adverse effects have been detected in men, women and children.

 

       2[)        Some ha_ c argued that to deduce that a spectrum of noncancer effects will occur in humans in

 

       21         the absence of better human data overstates the science; most scientists involved in the

 

       22        reassessment as authors and reviewers have indicated that such inference is reasonable given the

 

       23        weight-of-the-evidence fi'om available data.  As presented, this logical conclusion  represents a

 

       24        testable hypothesis which may be evaluated by further data collection. The EPA, its Federal

       25        colleagues and others in tine general scientific community are continuing to fill critical data gaps

       26        which will reduce our uncertainty regarding both hazard and risk characterization for dioxin and

 

       27        related compounds.

       28

       29        Dioxin and related compounds are structurally related and elicit their effects through a

       30        common mode-of-action.

 

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      1                  The scientific community has identified and described a series of common biological

      2        steps that are necessary for most if not all of the observed effects of dioxin and related

      3        compounds in vertebrates including humans.  Binding of dioxin-like compounds to a cellular

      4        protein called the "Ah receptor" represents the first step in  a series of events attributable to

      5        exposure to dioxin-)ike compounds including biochemical, cellular and tissue-level changes in

      6        normal biological processes.  Binding to the .Kh receptor appears to be necessary, for all well-

      7        studied effects of dioxin but is not sufficient, in and ofitself, to elicit these responses.  There

      8        remains some uncertainty as to whether every dioxin response is Ah receptor-mediated. Sensitive

      9        biological tools like the aryl hydrocarbon receptor deficient (Ahr'/-) mice indicate a small residual

     10        of effects to exposure to TCDD which does not allow us to rule out altemative pathways which

     11         are receptor-independent. The well documented effects elicited by exposure of animals, and in,

     12        some cases, humans, to 2,3,7,8-TCDD are shared by other chemicals which have a similar

     13        structure and Ah receptor binding characteristics.  In the last five years, significant data has

     14        accumulated which supports the concept of toxicity equivalence, a concept which is at the heart

     15        of risk assessment for the complex mixtures of dioxin and related compounds which are

     16        encountered in the envirom'nent. These data have been analyzed and summarized in Part 2,

     17        Chapter 9. This Chapter has been added to the EPA's dioxin re-assessment effort to address

     18        questions raised by the Agency's Science Advisory Board (SAB) in 1995.  The SAB suggested

     19        that, since the TEQ approach was a critical component of risk assessment for dioxin and related

     20        compounds, the Agency should be explicit in its description of the history and application of the

     21         process and go beyond reliance on the Agency's published reference documents on the subject

     22        (EPA 1987,  1989).

     23

     24        EPA and The Intemational Scientific Community have Adopted Toxicity Equivalence of

     25        Dio×in and Related Compounds as Prudent Science Policy.

     26                  Dioxin and related compounds always exist in nature as complex mixtures.  As discussed

     27        in the Exposure Document, these complex mixtures can be characterized through analytic

     28        methods to determine concentrations of individual congeners.  Dioxin and related compounds

     29        can be quantified and biological activity of the mixture can be estimated using relative potency

     30        values and an assumption of dose additivity.  Such an approach has evolved over time to form

 

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     1        the basis for the use of"toxicitv equivalence" (TEQ) in risk assessment for this group of

    2        compounds. While such an approach is dependent on critical assumptions and scientific

     3        judgement, it has been characterized as a "useful, interim" way to deal with the complex mixture

     4        problem and has been accepted by numerous countries and several intemational organizations.

     5        Altemative approaches, including the assumption that all congeners carry the toxicity

     6        equivalence of 2,3,7,8-TCDD, or that all congeners other than 2,3,7,8-TCDD can be ignored,

     7        have been generally rejected as inadequate for risk assessment purposes.

     8                  Significant additional literature is now available on the subject of toxicity equivalence of

     9        dioxin and related compounds and Chapter 9 provides the reader with a summary which is up to

    10        date through  1999. A recent intemational evaluation of all of the available data ( van den Berg et

    11         al., t 998) has re-affirmed the TEQ approach and has provided the scientific community with the

    12        latest values for toxicity equivalence factors (TEFs) for PCDDs, PCDFs and dioxin-like PCBs.

    13        Consequently, we can infer with greater confidence that humans will respond to the cumulative

    14        exposure of Ah receptor-mediated chemicals. The position taken in this Reassessment is that

    15        these 1998 TEFs should be adopted for use by the Agency.  Future research will be needed to

    16        address remaining uncertainties inherent in the current approach.  The WHO has suggested that

    17        the TEQ scheme be re-evaluated on a periodic basis and that TEFs and their application to risk

    18        assessment be re-analysed to account for emerging scientific information.

    19

   20        Complex Mixtures of Dioxin and Related Compounds are Highly Potent, "Likely"

   21         Carcinogens.

   22                  With regard to carcinogenicity,  a weight-of-the-evidence evaluation suggests that

   23        mixtures of dioxin and related compounds (CDDs, CDFs, and dioxin-like PCBs) are strong

   24        cancer promoters, weak direct or indirect initiators and likely to present a cancer hazard to

   25        humans.  Since dioxin and related compounds always occur in the environment and in humans as

   26        complex mixtures of individual congeners, it is appropriate that the characterization apply to the

   27        mixture.  According to the Agency's revised draft Cancer Guidelines, the descriptor, likely, is

   28        appropriate when the available tumor effects and other key data are adequate to demonstrate

   29        carcinogenic potential to humans.  Adequate data are recognized to span a wide range. The data

   30        for complex mixtures of dioxin and related compounds represents a case which, according to the

   31         draft Guidelines, would approach the strong evidence end of the adequate data spectrmn.

 

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      i         Ep_demiologic obse_'ations of an association between exposure and cancer responses (TCDD),

     2        unequivocal positive responses in both sexes, multiple species and different routes in lifetime

     3        bioassays or initiation-promotion protocols or other shorter-term in vivo systems e.g. transgenic

     4        models CI'CDD plus numerous PCDDs, PCDFs, dioxin-like PCBs), and mechmfistic or mode-of

     5        action data that are assumed to be relevant to human carcinogenicity (PCDDs, PCDFs, dioxin-

     6        like PCBs) all support the description ofcomplex mixtures of dioxin and related compounds as

     7        likely human carcinogens.

      8                  g,,_ile the data base from cancer epidemiologic studies remains controversial, it is the

      9        view of this reassessment that this body of evidence is supported by the laboratory data

    10        indicating that TCDD probably increases cancer mortality of several types.  Although not all

    11         confounders were ruled out ill any one study, positive associations between surrogates of dioxin

    12        exposure, cithcr length of occupational exposure or proximity to a known source combined with

    13        some information based on measured blood levels, and cancer have been reported.  These data

     14        suggest a role for dioxin exposure to contribute to a carcinogenic response but do not confirm a

    15        causal relationship between exposure to dioxin and increased cancer incidence,  Available human

     16        studies alone cannot demonstrate whether a cause and effect relationship between dioxin

    17        exposure and increased incidence of cancer exists.  Therefore, evaluation of cancer hazard in

    18        humans must include an evaluation of all of the available animal and in vitro data as well as the

    19        data fi'om exposed human populations.

    20                  As discussed earlier in Section 2.2.1.4, under EPA's current approach, individual

    21         congeners can also be characterized as to their carcinogenic hazard.  TCDD is best characterized

    22        as "carcinogenic to humans."  This means that, based on the weight of all of the evidence

    23        (human, animal, mode-of-action), TCDD meets the criteria that allows EPA and the scientific

    24        community to accept a causal relationship between TCDD exposure and cancer hazard.  The

    25        guidance suggests that "carcinogenic to humans" is an appropriate descriptor of human

    26        carcinogenic potential when there is an absence of conclusive epidemiologic evidence to clearly

    27        establish a cause and effect relationship between human exposure and cancer, but there is

    28        compelIing carcinogenicity in animals and mechanistic information in animals and humans

    29        den-tonstrating similar modes of carcmogerfic action·  The "carcinogenic to humans" descriptor is

    30        suggested for TCDD since all of the following conditions are met:

 

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     1

     2                  there is evidence from occupational epidemiologic studies for an association between

     3                  TCDD exposure and increases in cancer at all sites combined, in lung cancer, and,

     4                  perhaps, at other sites, but the data are insufficient on their own to demonstrate a causal

     5                  association;

     6

      7                  there is extensive carcinogenicity ill both sexes of multiple species at multiple sites;

      8

      9        -         there is general agreement that the mode ofTCDD's carcinogenicity is Ah receptor

    10                  dependent and proceeds tkrough modification of the action of a number of receptor and

    11                   honnone systems involved in cell growth and differentiation such as the epidermal

    12                  growth factor receptor and estrogen receptor; and

     13

     14                  key events such as equivalent body burdens in animals mad in human populations

     15                  expressing an association between exposure to TCDD and cancer, mid the determination

     16                  of active ,ah receptor and dioxin responsive elements in the general human population.

     17                  There is no reason to believe that these events would not occur in the occupational

    18                  cohorts studied.

     19

    20                  Other individual dioxin-like compounds are characterized as "likely" human carcinogens

    21         primarily because of the lack of epidemiological evidence associated with their carcinogenicity

    22        although the inference based on toxicity equivalence is strong that they would behave in humans

    23         as TCDD does.  Other factors, such as the lack of congener specific ckronic bioassays also

    24        support this characterization.  For each congener, the degree of certainty is dependent on the

    25        available congener specific data and its consistency With the generalized mode-of-action which

    26        underpins toxicity equivalence for TCDD and related compounds.  Based on this logic, complex

    27        enviroru'nental mixtures of TCDD and dioxin-like compounds should be characterized as "likely"

    28        carcinogens, with the degree of certainty of the characterization being dependent on the

    29        constituents of the n-fixture, when known.  For instance, the hazard potential, although "likely,"

    30        would be characterized differently for a mixture whose TEQ was dominated by OCDD as

    31         compared to one which was dominated by pentaCDF.

 

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                      While uncertainties remain regarding quantitative estimates of upper bound cancel' risk

             From dioxin and related compounds, efforts of this reassessment to bring more data into the

             evaluation of cancer potency have resulted irl evaluation of the slope of the dose response curve

             at the low end of the observed range (using the LED00 using a simple proportional (linear) model

             and a calculation of both upper bound risk and margm of exposure (MOE) based on human

             equivalent background exposures and associated body burdens.  Evaluation of shape parameters

             (which is used to estimate degree of linearity or non-linearity of dose-response within the range

             of observation) for biochemical effects indicates that many of these biochemical effects can be

   9        hypothesized to be to key events in a generalized, dioxin mode-of action model.  These analyses

   0        do not argtte for significant departures from linearity below a calculated ED0_ for endpoints

  11         potentially related to cancer response, for at least one to two orders ofmagnitude lower exposure.

  12        Risk estimates for intakes associated with background body burdens or incremental exposures

  13        based on this slope factor represent a plausible upper bound on risk based on the evaluation of

  14        animal and human data.  The slope factors based on the most sensitive cancer responses, both

  15        animal and human, calculated in Section 5, fall in a range of 5 X 10'3 to 5 X 10'4 per

  16        pgTEQ/kgBW/day.  The ranges of estimates of upper bound cancer potency calculated from the

  17        human and animal data analyzed in Part 2, Chapter8 overlap and the range above is bounded on

  18        the upper end by the estimate of slope fi'om the Hamburg cohort epidemiology study and on the

  19        lower end by' the estimate fi'om the re-analyzed Kociba study.  Consequently, the Agency,

  20        although fully recognizing this range and the public health conservative nature of the slope

  21         factors that make up the range, suggests the use of 5 X I0-3 per pgTEQ/kgBW/day as an

  22        estimator of upper bound cancer risk for both background intakes and incremental intakes above

  23        background.  Slope factors allow the calculation of the probability ofcancer risk for the highly'

  24        vulnerable in the population (estimated to be the top 5% or zeater).  Xx,q_ile there may be

  25        individuals in the population who might experience a higher cancer risk on the basis of genetic

  26        factors or other deten'ninams of cancer risk not accounted for in epidemiologic data or animal

  27        studies, the vast majority of the population is expected to have less risk per unit of exposm'e and

 28        some my have zero risk.  Based on these slope factor estimates (per pgTEQ/kgBW/day), average

  29        cun'ent background body bm'dens (5 ng/kgBW) which result from average intakes of

  30        approximately 3  pgTEQ/kgBW/day are in the range of 10'* to  10'2.  A very small percentage of

 

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     I         the population (<  1%) may experience risk which are 2-3 times higher than this if they are both

     '>        the most vulnerable and the most highly exposed (among the top 5%) based on dietary intake of

     3        dioxin and related compounds. This range of upper bound risk for the general population has

     4        increased all order of magnitude from the risk described at background exposure levels based on

     5        EPA's draft of this reassessment (10.t_ 10-3) (EPA, 1994).

     6                  Despite the use of the epidemiology data to describe an upper bound on cancer risk, the

     7        Peer Panel that met in Septeinber 1993 to review an earlier draft of the cancer epidemiology

     8        chapter suggested that the epidemiology data alone were still not adequate to implicate dioxin

     9        and related compounds as "known" human carcinogens but that the results from the human

    10        studies were largely consistent with observations from laboratory studies of dioxin-induced

   11         cancer and, therefore, should not be dismissed or ignored.  Other scientists, including those who

    12        attended the Peer Pa3'_eI meeting, felt either more or less strongly about the weight of the

    13        evidence from cancer epidemiology studies, representing the range of opinion that still exists on

    14        the interpretation of the these studies. Similar opinions were expressed in the comments

   15        documented in the SAB's report in  1995 (SAB,1995).  More recently, the kltemational Agency

    16        for Research on Cancer, in its re-evaluation of the cancer hazard of dioxin and related

    17        compounds (1997), found that the while the. epidemiologic data base for 2,3,7,8-TCDD was still

    18        "hmited," the overall weight of the evidence was sufficient to characterize 2,3,7,8-TCDD as a

    19        Category 1, "known" human carcinogen.  Other related members of the class of dioxin-like

   20        compounds were considered to have "inadequate" epidemiologic data to factor into hazard

   21        categorization.  A similar classification has been proposed within the context of the Department

   22        of Health and Human Services' Report on Carcinogens.  They too base their characterization on

   23        the broad base of human, arlimal and mode- of-action information in humans and animals which

   24        support this conclusior_. Therefore, given that 2,3,7,8-TCDD is contained in complex mixtures of

   25        dioxin and related compounds, and that the TEQ approach has been adopted as a reasonable

   26        approach to assessing risks of these complex mixtures, it is also reasonable to apply estimates of

   27        upper bound cancer potency derived fi'om epidemiology studies where 2,3,7,8-TCDD was

   28        associated with excess cancer risk to complex mixtures of dioxin and related compounds.

   29                  The current evidence suggests that both receptor binding and most early biochemical

   30        events such as enzyme reduction are likely to demonstrate low-dose lineariW.  The mechanistic

 

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     1         relationship of these early events to the complex process of carcinogenesis remains to be

     2        established.  If these findings imply low-dose linearity in biologically-based cancer models under

     3        development, then the probability of cancer risk will be linearly related to exposure to TCDD at

     4        low doses.  Until the mechanistic relationship between early cellular responses and the

     5        parameters m biologically based cancer models is better understood, the shape of the dose-

     6        response curve for cancer in tine below the range of observation can only be inferred with

     7        uncertainty.  Associations between exposure to dioxin and certain types of cancer have been

     8        noted in occupational cohorts with average body burdens of TCDD approximately 1- 3 orders of

     9        mag'nitude (10 - 1000 times) higher than average TCDD body burdens in the general population.

    10        The average body burden in these occupational cohorts level is within 1-2 orders of magnitude

    11         (10-100 times) of average background body burdens in the general population iii terms of TEQ

    12        (See Table 5-1 and Figure 5-1).  Thus, there is no need for large scale low dose extrapolations in

    13        order to evaluate background intakes and body burdens, and little, if any data to suggest large

    14        departures fi'om linearity in this somewhat narrow window between the lower end of the range of

    l 5        observation and the range of general population back_ound exposures.  Nonetheless, the

    16        relationsMp of apparent increases in cancer mortality in these worker populations to calculations

    17        of general population risk remains a source of uncertainty.

    18                  TCDD has been clearly shown to increase malignant tumor incidence in laboratory

    19        animals.  In addition, a number of studies analyzed in this reassessment demonstrate other

   20        biological effects of dioxin related to the process of carcinogenesis.  Initial attempts to construct

   21         a biologically-based model for certain dioxin effects as described in this reassessment will need

   22        to be continued and expanded to accommodate more of the available biology and to 'apply to a

   23        broader range of potential health effects associated with exposure to dioxin-like compounds.

   24

   25        Use a "Margin-of Exposure Approach" to Evaluate Risk for Noncancer and Some Cancer

   26        Endpoints.

   27                  The likelihood that noncancer effects may be occurring in the human population at

   28        environmental exposure levels is often evaluated using a "margin of exposure" (MOE) approach.

   29        The Agency has used this approach for a number of years in its assessment of the safety of

   30        pesticides.  This concept has also been incorporated into the revised Cancer Risk Assessment

 

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                     guidelines.  A MOE is calculated by dividing a "point of departure" for extrapolation purposes at

                     the Iow end of the range ofobser_'ation in human or animal studies (the human-equivalent animal

                     LOAEL, no observed adverse effect level (NOAEL), benchmark dose (BMD) or effective dose

                      (EDxx)) by the human exposure or body burden level of interest. Generally speaking, when

                     considering either background exposures or incremental exposures plus background, MOEs in

             ,        range of 100 -1000 are considered adequate to rule out the likelihood of significant effects

            ?        occurring in humans based on sensitive animal responses orresults from epidemiologic studies.

            8        The adequacy of the MOE to be protective of health takes into account the nature of the effect at

            9        the "point of departm'e," the slope of the dose response curve, the adequacy of the overall

           10        database, interindividual variability in the human population, as well as other factors.

           11         Considering MOEs based on incremental exposures alone divided by the human exposure of

           12        interest, is not considered to give an accurate portrayal of the implications of that exposure unless

           13        background exposures are insignificant.  This is not the case for background exposures for dioxin

           14        and related compounds based on total TEQ,  The average levels of background intake and

           15        associated body burdens of dioxin-like compounds in ten'ns of TEQs in the general population

           16        would be well within a l_ctor of 100 of human-equivalent exposure levels associated with

           17        NOELS, LOAELs, BMDs, or ED0_ values in laboratory animals exposed to TCDD or TCDD

           l 8        equivalents. These estimates, although highly variable, suggest that  any choice of body burden,

           19        as a pomt-o!-departure, above 100 ng/kg would likely yield greater than 1% excess risk for some

          20        endpoint in humans.  Also, choosing ora point-6f-departure below 1 ng/kg would likely be an

          21         extrapolation below the range of these data and would likely represent a risk of less than 1%.

          22        Any choice in the middle range of lng/kg to 100 rig/kg, would be supported by the analyses,

          23        although the data provide the greatest support in the range of l 0ng/kg to 50 rig/kg.  In many

          24        cases, the MOE compared to background using these endpoints is a factor of 10 or less (See Part

          25        2, Chapter 8).  Because of the relatively high background compared to effect levels, the Agency

          26        is not recommending the derivation of a reference dose (RfD) for dioxin and related compounds.

          27        Although RfD's are often useful because they represent a health risk goal below which there is

          28        likely to be no appreciable risk ofnoncancer effects over a lifetime of exposure, their primm'y use

          29        is to evaluate increments of exposure from specific sources when background exposures are low

          30        and insignificant.  Any RfD that the Agency would recommend using the traditional approach for

 

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    1         setting an RfD is likely to be 2-3 orders of magnitude (100-1000) below current background

    2        intakes and body burdens.  Since exceeding the RfD is not a statement ofrisk, discussion of an

    3        RfD for an incremental exposure when the RfD has already been exceeded by average

    4        background exposures is meaningless.

    5                  When evaluating incremental exposures associated with specific sources, knowing the

    6        increment relative to background may help to urtderstand the impact of the incremental exposure.

    7        For instance, it would be misleading to suggest that an incremental exposure of

    8        0.001  pgTEQ/kg/day was below the RfD if"background" exposures were already at or above

    9        that level. On the other hand, as part of the total, the increment represents less than a 0.1%

   10        increase over average "background" and we estimate that individuals within the 50-95% range of

   11         exposure within the population may be 2-3 times (200%-300%) higher. This has led us to

   12        suggest that perhaps the best infonnation for a decision-maker to have is 1) a characterization of

   13        average "background" exposures; 2) a characterization of the percent increase over background

   14        of individuals or subpopulations of interest; and 3) a policy statement about when increases over

   15        average "background" become significant for the dec/sion.  This is not easy because one could

   16        argue that, given high "background", any addition, if it is widespread, is too much.  On the other

   17        )land, someone else could argue that a 10% increase in incremental exposure for a small

   18        population around a specific point source would be well within the general population exposures

   19        and would not constitute a disproportionate exposure or risk.  In this case, the strategy might be

  20        too bring average "background" exposures down and to focus on large incremental exposures or

  21        highly susceptible populations. This would be a strategy that would parallel the Agency's lead

  22        strategy.  Other parallel issues between dioxin-like compounds and lead are under discussion

  23        within the Agency.

  24                  ATSDR (1999) set a minimal risk level (MRL), which is defined similarly to the EPA's

  25         RtT), for dioxin and related compounds of 1.0 pgTEQ/kgBW/day.  Some of the data regarding

  26        lower bounds on the EDs0.'  from various noncancer effects call that MRL into question.   WHO

  27        (2000) has set a tolerable daily intake of 1-4 pgTEQ/kgBW/day and has indicated that, although

  28        current exposures in that range are "tolerable" (a risk mmmgement decision rather than a risk

  29        assessment), efforts should be made to ultimately reduce intake levels.  Findings in this

  30        reassessment appear to be supportive of that recommendation.

 

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      1

     2        Children's risk from exposure to dioxin and related compounds may be increased, but

     3        more data are needed to full)' address this issue.

     4                  The issue of children's risk fi'om exposure to dioxin-like compounds has been addressed

      5        in a number of sections tlzroughout this reassessment.  Data suggest a sensitivity of response in

      6        both humans and animals durirlg the developmental period, both prenatally and postnatally.

      7        However, data are limited.  Since evaluation of the impacts of early exposures on both children's

      8        health and health later in life is important to a complete characterization of risk, collection of

      9        additional data ill this area should be a high priority to reduce uncertainties in future risk

     10        assessments.

    11                  Data from the Dutch cohort of children exposed to PCBs and dioxin-like compounds

    12        suggest impacts of exposure to background levels of dioxin and related compounds prenatally

    13        and, perhaps, postnatally on neurobehavioral outcomes, thyroid function, and liver enzymes

    14        (AST and ALT).  While these effects can not be attributed solely to dioxin and related

    15        compounds, several associations suggest that these are, in fact, likely to be Ah-mediated effects.

     16        ,An investigation of background dioxin exposure and tooth development was done in Finnish

    17        children as a result of studies of dental effects in dioxin-exposed rats, mice, and nonhuman

     18        primates, and in PCB-exposed children.   The Finnish investigators examined enamel

     19        hypomineralization of permanent first molars in 6-7 year old children.  The length of time which

    20        infants breast fed was not significantly associated with either mineralization changes, or with

    21         TEQ levels in the breast milk.  However, when the levels and length of breast feeding were

    22        combined in an overall score, a statistically si_ificant association was observed ® = 0.3,

    23        p = 0.003, regression analysis).

    24                  In addition, effects have been seen where sig-nificantly elevated exposure occurred.  The

    25        incidents at Yusho and Yu-Cheng resulted in increased perinatal mortality and low birthweight in

    26        infants bom to women who had been exposed.  Rocker bottom heal was observed in Yusho

    27        infants, and functional abnon'nalities have been reported in Yu-Cheng children.  The similarity of

    28        effects observed in l'mman infants prenatally exposed to the complex mixture in Yusho and

    29        Yu-Cheng with those reported in adult monkeys exposed only to TCDD suggests that at least

    30        some of the effects children are due to the TCDD-Iike congeners in the contaminated rice oil

 

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        1         ingested by the mothers of these clzildren.  The similar responses include a clustering of effects in

        2        organs derived from the ectodermal germ layer, referred to as ectodermal dysplasia, including

        3        effects on the skin, nails, and Meibomian glands; developmental and psychomotor delay during

        4        developmental az_d cognitive tests.  Some investigators believe that, because all of these effects

        5        ill the Yusho and Yu-Cheng cohorts do not correlate with TEQ, some of the effects are

        6        exclusively due to nondioxin-tike PCBs or a combination of all the congeners.   Ill addition, on

        7        the basis of these data, it is still not clear to what extent there is an association between overt

        8        matemal toxicity and embryo/fetal toxicity in humans. FmXher studies in the offspring, as well as

        9        follow-up to the Seveso incident may shed further light on this issue.  In addition to chloracne

       10        and acute responses to TCDD exposure seen in Seveso children, elevated levels of serum GGT

       11         have been observed within a year after exposure in some of the more highly exposed Seveso

       12        children  Long-telTn pathologic consequences of elevated GGT have not been illustrated by

       13        excess mortality fi'om liver disorders or cancer, or in excess morbidity but further follow-up is

       14        needed.  It must be recognized that the absence of an effect thus far does not obviate the

       15        possibility that the enzyme levels may have increased concurrent to the exposure but declined

       16        after cessation.  The apparently transient elevations in- 'Ai]T'levels among the Seveso children

       17        suggest that hepatic enzyme levels other than GGT may react in this manner to 2,3,7,8-TCDD

       18        exposure.

       19                  Impacts on thyroid hormones provides an example of an effect of elevated, postnatal

       20        exposure to dioxin and related compounds.  Several studies of nursing infants suggest that

       21         ingestion of breast milk with a higher dioxin TEQ may alter thyroid function. Thyroid hormones

       22        play important roles in the developing nervous system in of all vertebrates species, including

       23        humans.  In fact, thyroid hormones are considered so important in development that in the U.S.

       24        all infants are tested for hypothyroidism shortly after birth.  Results from the studies mentioned

       25        above suggest a possible shift in the population distribution of thyroid hormone levels,

       26        particularly T4, and point out the need for collection of longitudinal data to assess the potential

       27        for long-term effects associated with developmental exposures. The exact processes accounting

       28        for these observations in humans are unknown, but when put in perspective of animal responses,

       29        tlne following might apply: dioxin increases the metabolism and excretion of thyroid hormone,

       30        mainly T4, in the liver.  Reduced T4 levels stimulate the pituitary to secrete more TSH, which

 

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       I         et_lances thyroid hormone production.  Early in the disruption process, the body can

       2        overcompensate for the loss ofT4, which may result in a small excess of circulating T4 in

       3         response to the increased TSH.  In animals, given higher doses of dioxin, the body is unable to

       4        maintain homeostasis, and TSH levels remain elevated and T4 levels decrease.

       5                  A large number of studies in animals have addressed the question of effects of dioxin-like

       6        chemicals after itl utero or lactational exposure.  These have included both single congener

       7         studics _tnd exposures to complex mixtures.  However, the vast majority of the data, are derived

       8         from studies of 2,3,7,8-TCDD, or single congeners (e.g. PCB 77) or commercial mixtures of

       9        PCBs.  Exposure patterns have included single doses to the dams as well as dosing on multiple

      10        days during gestation begim'fing as early as the first day of gestation.  These studies are discussed

      11         in detail in Part 2, Chapter 5.  The observed toxic effects include developmental  toxicity,

      12        neurobehavioral and neurochemical  alterations, endocrine effects and developmental

      13         immunotoxicity.  For instance, results of this body of work suggests that 2,3,7,8-TCDD clearly

      14        has the potential to produce alterations in male reproductive function (rats and hamsters) and

      15        male sexual behavior (rats) after prenatal exposure.  In addition, impacts on neuromotor and

      16        cognitive behavior as well as development of the immune system have been indicated in a

      17        number of studies.

      18                  No epidemiological data and limited animal data are available to address the question of

      19        the potential impact of exposure to dioxin-like compounds on childhood cancers or oil cancers of

      20        later life. Given the relative impact of nursing on body burdens (See the discussion of breast milk

      21         exposures and body burdens below), direct impacts of increased early post-natal exposure on the

      22        carcinogenic process is expected to be small.  This conclusion is based on the reasonable

      23         assumptions that cancer risk is a function of average lifetime body burden or that, because dioxin

      24        is a potent cancer promoter rather than a direct initiator of the cancer process, exposures later in

      25         life might be more important than those received earlier.  However, recent studies of Brown et al.

      26        (1998) suggest that prenatal exposure of rats to dioxin and related compounds may indirectly

      27        enhance their sensitivity as adults to chemical carcinogenesis from other chemical carcinogens.

      28        Further work is needed to evaluate this issue.

      29                  In addition to the potential vulnerability during development, the fetus, infants and

      30        children a'e exposed to dioxins through several routes.  The fetus is exposed in utero to levels of

 

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       I         dioxin and related compounds which reflect the body burden of the mother.  It is important to

       2        recognize that it is not the individual meals which a pre.ant woman eats during pregnancy

       3        which might affect development but the consequence of her exposure history, over her life, which

       4        has the greatest impact on her body burden.  Again, good nutrition which represents diet with

       5        appropriate levels of fat has consequences on dietary intake and consequent body burdens of

       6        dioxin and related compounds.  Nursing infants represent a special ease where, for a limited

       7        portion of their lives, ii,rants may have elevated exposures on a body weight basis when

       8        compared to non-nursing infants and adults (See discussion below).  In addition to breast milk

       9        exposures, intakes of CDD/Fs and dioxin-like PCBs are over tlwee times higher for a young child

      10        as compared to that of an adult, on a body weight basis.  Table 4-9 in Section 4 of this document

      11        describes the variability in average intake values as a function of age using age-specific food

      12        consumption rate and average food concentrations, as was done for adult intake estimates.

      13        However, like the case for the nursing infant discussed below, the differences in body burden

      14        between children and adults are expected to be much less thar_ the differences in daily intake.

      15        Assuming that body burden is the relevaa'n dose metric for most, if not all, effects, there is some

      16        assurance that these increased intake levels will have limited additional impact on risk as

      17        compared to overall lifetime exposure.

      18

      19        Background Exposures to Dioxin and Related Compounds need to be Considered when

     20        Evaluating both Hazard and Risk.

     21                  The term "background" exposure has been used tI_oughout this reassessment to describe

     22        exposure of the general population, who are not exposed to readily identifiable point sources of

     23        dioxin-like compounds.  Adult daily intakes of CDD/CDFs and dioxin-like PCBs are estimated

     24        to average 45 and 25 pg TEQDFp-WHO98/day, respectively, for a total intake of 70 p_day

     25        TEQDF_,-vGq:Ouv   Daily intake is estimated by combining exposure media concentrations (food,

                                                                                                                   (

            26        soil, air) with contact rates (ingestion, inhalation),  Table 4-8 below summarizes the intake rates

     27        derived by this method.  The intake estimate is supported by an extensive database on food

     28        consumption rates and food data.  Phan'nacokinetic (PK) modeling provides further support for

     29        the intake estimates.  Current adult tissue levels reflect intakes from past exposure levels which

     30        are thought to be higher than current levels (see Trends Section 2.6).

 

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         1                  CDD/CDF and dioxin-like PCB intakes for the general population may extend to levels at

        2         least three times higher ti'mn the mean.  Variability in general population exposure is primarily

        3        the result in the differences in dietau choices that individuals make.  These are differences in

         4        both quantity and types of food consumed.  A diet which is disproportionately high in animal fats

         5        will result in an increased background exposure over the mean.  Data on variability of fat

         6        consumption indicate that the 95'h percenthe is about twice the mean and the 99th percenthe is

         7        approximately 3 times the meaa'_.  Additionally, a diet which substitutes meat sources that are low

         8         in dioxin (i e. beef, pork or poultry) with sources that are high in dioxin (i.e. fresh water fish)

         9        could result in exposures elevated over tltree times the mean.  This scenario may not represent a

        10        significant change in total animal fat consumption, even though it results in an increased dioxin

        11         exposure.  Intakes of CDD/Fs and dioxin-like PCBs are over tl'u'ee times higher for a young child

        12        as compared to that of an adult, on a body weight basis.  Using age-specific food consumption

        13        rate and average food concentrations, as was done above for adult intake estimates, Table 4-9

        14        describes the variability in average intake values as a function of age.

        15                  The average CDD/CDF tissue level for the general adult U.S. population appears to be

        16        declining and the best estimate of current (late 1990s) levels is 25 ppt (TEQDFp-WHO98, lipid

        17        basis).  The tissue samples collected in North America in the late 1980s and early 1990s showed

        18        an average TEQt)Fp WHO98  level of about 55 pg/g lipid.  This finding is supported by a number of

        19        studies which measured dioxin levels in adipose, blood and human milk, all conducted in North

       20        America.  The number of people in most of these studies, however, is relatively small and the

       21         participants were not statistically selected in ways that assure their representativeness of the

       22        general U.S. adult population.  One study, the 1987 National Human Adipose Tissue Survey

       23        (NHATS), involved over 800 individuals and provided broad geographic coverage, but did not

       24        address coplanar PCBs.  Similar tissue levels of these compounds have been measured in Europe

       25        and Japan during similar time periods.

       26                  Because dioxin levels in the environment have been declining since the 1970s (see trends

       27        discussion), it is reasonable to expect that levels in food, human intake and ultimately human

       28        tissue have also declined over this period.  The changes in tissue levels are likely to lag the

       29        decline seen in environmental levels and the changes in tissue levels cannot be assumed to occur

       30        proportionally with declines in enviromnental levels.  ATSDR (1999) summarized levels of

 

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        1        CDDs, CDFs and PCBs in human blood collected during the time period I995 to 1997.  The

       2        individuals smnpled were all U.S. residents with no known exposures to dioxin other than normal

       3        background.  The blood was collected from 400 individuals in seven different locations with an

       4        age range of 20 to 70 years.  All TEQ calculations were made assuming nondetects were equal to

       5        half the detection limit.  While these samples were not collected in a manner that can be

       6        considered statistically representative of the national population and lack wide geographic

       7        coverage, they are judged to provide a better indication of current tissue levels in the US than the

       8        earlier data (see Table 4-7 ). PCBs 105,  I18, and 156 are missing from the blood data for the

       9        comparison populations reported in the Calcasieu stud (ATSDR, 1999).  These congeners

      10        account for 62% of the total PCB TEQ estimated in the early 1990's.  Assuming that the missing

      I I         congeners fi-om the Calcasieu study data contribute the same proportion to the total PCB TEQ as

      12        in earlier data, they would increase our estimate of cun'ent body burdens by another 3.7 pgTEQ/g

      13        lipid  .tot a total PCB TEQ of' 5.9 pg/g lipid and a total DFP TEQ of 25 pg/g lipid.

      14                  Past background exposure of about 3 pgTEQ/kgBW/day leads to body burdens in the

      15        human population which currently average approximately 5ng/kg (20-30 pg TEQ/g lipid) when

      16        all dioxins, furans and PCBs are included and have been higher in the past.  DeVito et al. (1995)

      17        estimated that body burdens averaged 9-13 ng&g based on intake values of 4-6 pg TEQ/kg/day

      18        and blood levels of 40-60 pgTEQ/g lipid, based on data from the late 1980%.  If the general

      19        population were exposed to dioxins and related compounds at the current level of intake

      20        (approximately 1 pg TEQ/kg/day) for a lifetime, average steady state body burdens would be less

      21         than 2 ng/kg and blood levels would be 7-8 pg TEQ/g lipid.  These estimates are based on the

      22        assumption of 50% absorption of dioxin-like compounds from the diet,  Using the same

      23       . assumption used for intake values, high-end estimates of body burden of individuals in the

      24        general population (approximately the top 5% of the general population) may be greater than 2

      25        times higher than these average estimates. This is based on data for dietary fat consumption and

      26        the assumption that body burdens of dioxin and related compounds in the general population are

      27        associated with their fat consumption. The top 1% is likely to be 3 times higher based on their

      28        intake of fat.

      29                  Characterizing national background levels of dioxins in tissues is uncertain because the

      30        current data cannot be considered statistically representative of the general population.  It is also

 

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   1         complicated by the fact that tissue levels are a function of both age and birth year.  Because

   2        intake levels have var/ed over time, the accumulation of dioxins in a person who tumed 50 years

   3        old in 1990 is different than in a person who tumed 50 ill 2000.  Future studies should help

   4        address these uncertainties.  The National Health and Nutrition Examination Survey Ox!I'-IANrES)

   5        began a new national survey in 1999'which will measure dioxin blood levels in about 1700

   6        people per year (see http:www.cdc.gov/nchs/nhanes.htm).  The survey is conducted  at 15

   7        alii'ti:rent locations per year and is designed to select individuals statistically representative ofthe

   8        civilian U.S. population in terms of age, race and ethnicity.  These new data should provide a

   9        much better basis for estimating national background tissue levels and evaluating trends than the

  10        currently available data.

  11                  As described above, cun'ent intake levels from food sources m'e estimated in this

  12        Reassessment to be approximately 1  pgTEQ/KgBW/day.  Certain segments of the population

  13        may be exposed to additional increments of exposure by being in proximity to point sources or

  14        because of dietary practices.  These will be described below.

  I5

  16        Evaluation of Exposure of "Special" Populations and Developmental Stages is Critical to

  17        Risk Characterization.

  18                  As discussed above, background exposures to dioxin-like compounds may extend to

  19        levels at Ieast three times higher thart the mean.  This upper range is assumed to result from the

 20        non'nal variability of diet and human behaviors.  Exposures from local elevated sources or

 21         exposures resulting from unique diets would be in addition to this background variability.  Such

 22        elevated exposures may occur in small segments of the population such as individuals living near

 23        discrete local sources, or subsistence or recreational fishers.  Nursing infants represent a special

 24        case where, for a limited portion of their lives, these individuals may have elevated exposures on

 25        a body weight basis when compared to non-nursing infants and adults.  This exposure will be

 26        discussed in a separate section below.

 27                  Dioxin contamination incidents involving the commercial food supply have occurred in

 28        the U.S. and other countries.  For example, in the U.S., contaminated ball clay was used as an

 29        anticaking agent in soybean meal and resulted in elevated dioxin levels in some poultry and

 30        catfish. This incident involved less than 5% of the national poultry production and has since been

 

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                  eliminated.  Elevated dioxin levels have also been obse_'ed in a few beef and dairy animals

                  where the contamination was associated with contact with pentachlorophenol treated wood.

                  Evidence of this kind of elevated exposure was not detected in the national beef survey.

                  Consequently its occurrence is likely to be low, but it has not been determined. These incidents

         3        may have led to small increases in dioxin exposure to the general population.  However, it is

         5        unlikely that such incidents have led to disproportior_ate exposures to populations living near

        7        where these incidents have occulted, since, in the U.S., meat and dairy products are highly

        8        distributed on a national scale.  If contamination events were to occur in foods that are

        9        predominantly distributed on a local or regional scale, then such events could lead to highly

       10        exposed local populations

       I I                   Elevated exposures associated with the workplace or industrial accidents have also been

       12        documented.  U.S. workers in certain segments of the chemical industry had elevated levels of

       13        TCDD exposure, with some tissue measurements in the thousands ofppt TCDD.  There is no

       14        clear evidence that elevated exposures are currently occurring among U.S. workers. Documented

       15        examples of past exposures for other groups include certain Air Force personnel exposed to

       16        Agent Orange during the Vietnam War and people exposed as a result of industrial accidents in

       17        Europe and Asia.

       18                  Consumption of unusually high amounts of fish, meat, or dairy products containing

       19        elevated levels of dioxins and dioxin-like PCBs can lead to elevated exposures in comparison to

       20        the general population.   Most people eat some fish from multiple sources, both fresh, and salt

       21         water.  The typical dioxin concentrations in these fish and the typical rates of consumption are

       22        included in the mean background calculation of exposure.  People who consume large quantities

       23        of' fish at typical contamination levels may have elevated exposures since the concentration of

       24        dioxin-like compounds in fish are generally higher than in other animal food products. These

       25        kinds of exposures are addressed within the estimates of variability of background and are not

       26        considered to result in highly exposed populations.  If high-end consumers obtain their fish from

       27        areas where the concentration of dioxin-like chemicals in the fish is elevated, they may constitute

       28        a highly exposed subpopulation.  Although this scenario seems reasonable, no supporting data

       29        could be found for such a highly exposed subpopulation in the U.S. One study measuring dioxin-

       30        like compounds in blood of sports fishers in the Great Lakes area showed elevations over mean

 

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   I         background, but within the range of normal variability.  Elevated CDD/CDF levels in human

   2        blood have been measured in Baltic fishermen.  Similarly elevated levels ofcoplanar PCBs have

   3        been measured in the blood of fishers on the north shore of the Gulf of the St. Lawrence River

   4        who consume large amounts of seafood.

   5                  Similm'ly, high exposures to dioxin-Iike chemicals as a result of consuming meat and

   6        dairy products would only occur in situations where individuals consume large quantities of

   7        these foods and the level of these compounds is elevated.  Most people eat meat and dairy

   8        products fi'om multiple sources and, even if large quantities are consumed, they are not likely, to

   9        have unusually high exposures.  Individuals who raise their own livestock for basic subsistence

  10        have the potential for higher exposures if local levels of dioxin-like compounds are high.  One

  11         study in the U.S. showed elevated levels in chicken eggs near a contaminated soil site.  European

  12        studies at several sites have shown elevated CDD/CDF levels in milk and other animal products

  13        near combustion sources.

  14                  In smnmary, in addition to general population exposure, some individuals or groups of

  15        individuals may also be exposed to dioxin-like compounds from discrete sources or pathways

  I6        locally within their enviromnent.  Examples of these "special" exposures include: contamination

  17        incidents, occupational exposures, direct or indirect exposure to local populations from discrete

  18        sources, or exposures to subsistence or recreational fishers.

  19

  20        Breast feeding infants have higher intakes of dioxin and related compounds for a short but

  21         developmentally important part of th eir life, but benefits of breast feeding are widely

  22        recognized to outweigh the risks.

  23                  A number of studies have measured levels of the dioxin-like compounds in human breast

  24        milk, yielding an average of 35 ppi TEQD_F,-WHOw  Several studies around the world suggest

  25        that these levels are declining.  Whether the perceived decline is based on declines in current

  26        intake levels in women whose breast milk levels would otherwise have been higher, or whether it

  27        reflects younger women :vith lower historical exposures entering the breast feeding population,

  28        or both, is not known.  Additional studies which take age, reproductive history and dioxin intake

 29        levels into account are needed.

  30                  Based on a six month nursing scenario, the average daily intake for an infant is about 100

 

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           1         times higher than the adult daily intake on a body weight basis: the adult intake is 1 pg TEQDrp-

           2        WHOg_/kg-d, while the infant intake while breast feeding would be 100 pg TEQDFp-WHOgg/kg-d.

           3        The differences in body burden between nursing infants and adults are expected to be much less

           4        than the differences in daily intake, and rapidly approach steady state at close to adult levels.

            5        This is because of rapid growth of the infant and an in:.reasing total amount of body fat. Because

            6        of rapid equilibration throughout the body, tissue doses will better reflect doses to target tissues

            7        than will intake.  This is important if windows of vulnerability occur during this period of

            8        development.  On a mass basis, however, the cumulative dose to the infant under this scenario is

            9        about 9% of the lifetime retake.

          I0                  The American Academy of Pediatrics (1997) has made a compelling argument for the

          1 l         diverse advantages of breast-feeding and the use of human milk for infant feeding to infants,

          12        mother, families and society.  These include health, nutritional, inununologic, developmental,

          13        psychological, social, economic, and environmental benefits.  Breast milk is the point of

          14        comparison for all infant food and the breast-fed infant is the reference for evaluation of all

          15        altemative feeding methods.  In addition, increasing the rates of breast-feeding initiation is a

          16        national health objective and one of the goals of the U.S. Govemment's Healthy People 2010.

          17        The World Health Organization (1988) maintained that the evidence did notsupport an alteration

          18        of WHO recommendations which promote and support breast feeding.  A more recent

          19        consultation in 1998 (WHO, 2000) reiterated these conclusions.  While it is important that the

          20        recommendations of these groups continue to be re-evaluated in light of emerging scientific

          21         infon'nation, the Agency does not believe that finding contained in this report provides a

          22        scientific basis for initiating such a re-evaluation.  This conclusion is based on the fact that

          23        stronger data have been presented that body burden, not intake, is the best dose metric; that many

          24        of the noncancer effects, particularly those seen in children, are more strongly associated with

          25        pre-natal exposure and the mother's body burden rather than post-natal exposures and breast milk

          26        levels; and that dioxin-like compounds are strong promoters of carcinogenicity, a mode-of-

          27        action that depends on late stage impacts rather than early stage impacts on the carcinogenic

          28        process.

          29

          30

 

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            Many Dioxin sources have been identified and emissions to the environment are being

             reduced.

                      Currer_t emissions of CDDs/CDFs/PCBs to the U.S. environment result principally from

             anthropogenic activities.  Evidence which supports this fnding include: matches in time of rise

   3        of environmental levels with time when general industrial activity began rising rapidly (see trend

   6        discussion in Section 4.6), lack of any identified large natural sources and observations of higher

   7        CDD/CDF/PCB body burdens in industrialized vs. less industrialized countries (see discussion

   8        on human tissue levels in Section 4.4).

   9                  The principal identified sources of environmental release may be grouped into five major

  10        types: Combustion and Incineration Sources; Chemical Manufacturing/Processing Sources;

  11         Industrial/Municipal Processes; Biological and Photochemical Processes; and Reservoir Sources.

  12        Development of release estimates is difficult because only a few facilities in most industrial

  13        sectors have been tested for CDD/CDF emissions.  Thus an extrapolation is needed to estimate

  14        national emissions.  The extrapolation method involves deriving an estimate of emissions per

  15         unit of activity at the tested facilities and multiplying this by the total activity level in the

  16        untested facilities.  In order to convey the level of uncertainty in both the measure of activity and

  l 7        the emission factor, EPA developed a qualitative confidence rating scheme.  The confidence

  18        rating scheme, presented in Section 4, Table 4-1, uses qualitative criteria to assign a high,

  19        medium, or low confidence rating to the emission factor re'Id activity level for those source

  20        categories for which emission estimates can be reliably quantified.  The Dioxin Reassessment

  21         has produced an Inventory of source releases for the U.S. (Table 4-2).  The Inventory was

  22        developed by considering all sources identified in the published literature and numerous

 23        individual emissions test reports. The Inventory is limited to sources whose releases can be

  24        reliably quantified (i.e. those with confidence ratings orA, B or C as defined above).  Also, it is

 25         limited to sources with releases that are created essentially simultaneously with formation.  This

 26        means that the reservoir sources are not included.  The Inventory presents the environmental

 27        releases in terms of two reference years: I987 and 1995.  EPA's best estimates ofreleases of

 28        CDD/CDFs to air, water and land from reasonably quantifiable sources were approximately

 29        2,800 gram (g) (l.3 pounds) TEQo_-WH09_ in 1995 versus 13,500 g (6 pounds) TEQr_F-WHOos

 30        in 1987. The decrease in estimated releases of CDD/CDFs between 1987 and 1995

 

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            (approximately 80%) was due primarily to reductions in air emissions from municipal and

            medical waste incinerators.  The environmental releases of CDD/Fs in the U.S. occur from a

            wide variety of sources, but are dominated by releases to the air from combustion sources.

            Insufficient data are available to comprehensively estimate point source releases of dioxin-like

            compounds to water.  Sound estimates of releases to water are only available for chlorine

            bleached pulp and paper mills and manufacture of ethylene dichlor/de/vinyl chloride monomer.

            The contribution of dioxin-like compounds to waterways from nonpoint source reservoirs is

    I        Iii<ely to be greater than the contributions from point sources.  Current data are only sufficient to

   :_        support preliminat'y estimates of nonpoint source contributions of dioxin-like compounds to

   0        water (i.e., urban storm water runoffand rural soil erosion).  These estimates suggest that, on a

   I         nationwide basis, total nonpoint releases are significantly larger than point source releases.

  12        Other releases to water bodies that camot be quantified on the basis of existing data include

  13        effluents from POTWs and most industrial/commercial sources.  Based on the available

  14        information, the inventory includes only a limited set of activities that result in direct

  15        environmental releases to land.  The only releases to land quantified in the inventory are land

  16        application of sewage sludge and pulp and paper mill wastewater sludges.  Not included in the

  17        Inventory's definition of an environmental release is _he'T-d_posal of sludges and ash into

  18        approved landfills.  While this inventory is the most comprehensive and well-documented in the

  19        world, the inventory is likely to underestimate total releases.  The magnitude of the

 20        underestimate is unl_lown but it is unlikely that non-combustion sources today ,other than

 21        reservoir sources, play a dominant role in human exposure.  In terms of 1995 releases from

 22        reasonably quantifiable sources, this document estimates releases of 2800 g WHO98TEQDF for

 23        contemporary formation sources and 2900 g WHO0_TEQDF for reservoir sources.  In addition,

 24        there remains a number of unquantifiable and poorly quantified sources which are described in

 25        Section 4.

 26                  As described above, combustion appears to be the most significant process of formation

 27        of CDDs/CDDFs today.  Important factors which can affect the rate of dioxin formation include

 28        the overall combustion efficiency, post combustion flue gas temperatures and residence  times,

 29        and the availability of surface catalytic sites to support dioxin synthesis.  Although chlorine is an

 30        essential component for the formation of CDD/'Fs in combustion systems, the empirical  evidence

 

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   1         indicates that for commercial scale incinerators, chlorine levels in feed are not the dominant

   2        controlling factor for rates of CDD/F stack emissions.  The conclusion that chlorine in feed is not

  3        a strong determinant of dioxin emissions applies to the overall population of commercial scale

   4        combustors.  For any individual commercial scale combustor, circumstances may exist in which

   5        changes in chlorine content of feed could affect dioxin emissions.  For uncontrolled combustion,

   6        such as open burning of house-hold waste, chlohne content of wastes may play a more

   7        significant role in affecting levels of dioxin emissions than observed in commercial scale

   8        combustors.

   9                  No significant release of newly fom_ed dioxin-like PCBs is occulting in the U.S.  Unlike

  10        CDD/CDFs, PCBs were intentionally manufactured in the U.S. in large quantities from 1929

  11         unlil production was banned in  1977.  Although it has been demonstrated that small quantities of

  12        coplanar PCBs can be produced during waste combustion,  no strong evidence exists that the

  13        dioxin-like PCBs make a significant contribution to TEQ releases during combustion. The

  14        occurrences of dioxin-like PCBs in the U.S. environment most likely reflects past releases

  15        associated with PCB production, use and disposal.  Further support of this finding is based on

  16        observations of reductions since 1980s in PCBs in Great Lakes sediment and other areas.

  17                  It is unlikely that the emission rates of CDD/CDFs fi'om known sources correlate

  18        proportionally with general population exposures.  Although the emissions inventory shows the

  19        relative contribution of various sources to total emissions, it cannot be assumed that these

  20        sources make the same relative contributions to human exposure. It is quite possible that the

  21         major sources of dioxin in food (see discussion in Section 2.6 indicating that the diet is the

  '_'_        dominant exposure pathway for humans) may not be those sources that represent the largest

  23         fractions of total emissions in the U.S.  The geographic locations of sources relative to the areas

  24         fi'om which much off the beef, pork, milk, and fish come, is important to consider.  That is, much

  25        of the agricultural areas which produce dieta_ animal fats are not located near or directly down

  26        wind of the major sources of dioxin and related compounds.

 27                  The contribution of reservoir sources to human exposure may be significant.  Several

  28         factors support this finding.  First, human exposure to the dioxin-like PCBs is thought to be

  29        derived almost completely fi'om reservoir sources.  Since one third of general population TEQ

 30        exposure is due to PCBs, at least one third of the overall risk from dioxin-Iike compounds comes

 

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     from reservoir sources.  Second, CDD/CDF releases from soil via soil erosion and nmoffto

            waterways appear to be greater than releases to water fi.om the primary sources included in the

             inventory.  CDD/CDFs in waterways can bioaccumulate in fish leading to human exposure via

             consumption of fish which makes up about one third of the total general population CDD/CDF

             TEQ exposure.  This suggests that a significant portion of the CDD/CDF TEQ exposure could be

             due to releases fi'om the soil reservoir.  Finally, soil reservoirs could have vapor and particulate

             releases wlfich deposit on plants and entel' the terrestrial food chain.  The magnitude of this

    ;        contribution, however, is unknown.

                       TI'ds assessment adopts the hypothesis that the primary mechanism by which dioxin-like

   0        compounds enter the terrestrial food chain is via atmospheric deposition.  Dioxin and related

    I         compounds enter the atmosphere directly through air emissions or indirectly, for example,

  12        tl_ough volatilization from land or water or from re-suspension of particles.  Once introduced

  13        into the enviro_m'_ent, dioxin-like compounds are widely distributed in the enviromnent as a

  14        result of a number of physical and biological processes.  The dioxin-like compounds are

  15        essentially insoluble in water, generally classified as semi-volathe and tend to bioaccumulate in

  16        animals,  Some evidence has shown that these compounds can degrade in the enviromnent, but in

  17        general they are considered very persistent and relatively immobile in. soils and sediments. These

  18        compounds are transported through the atmosphere as vapors or attached to air-bome particulates

  19        and can be deposited on soils, plants, or other surfaces (by wet or dry deposition).  The dioxin-

  20        like compounds enter water bodies primarily via direct deposition from the atmosphere, or by

  21         surface runoff and erosion.  From soils, these compounds can reenter the atmosphere either as

  22        resuspended soil particles or as vapors.  In water, they can be resuspended into the water colunm

  23        from sediments, volatilized out of the surface waters into the atmosphere or become buried in

  24        deeper sediments.  Immobile sedim,ents appear to serve as permanent sinks for the dioxin-like

  25        compounds.  Though not always considered an environmental compartment, these compom_ds

  26        are also found in anthropogenic materials (such as pentachlorophenol) and have the potential to

  27        be released from these materials into the broader environment.

  28                  The two primary pathways for the dioxin-like compounds to enter the ecological food

  29        chains and human diet are: air-to-plant-to-animal and water/sediment-to-fish.  Vegetation

  30        receives these compounds via atmospheric deposition in the vapor and particle phases.  The

 

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                   compounds are retained on plant surfaces and bioaccumulated in the fatty tissues of animals that

                   feed on these plants.  Vapor phase transfers onto vegetation have been experimentally shown to

                   dominate the air-to-plant pathway for the dioxin-like compounds, particularly for the lower

                   chlorinated congeners.  In the aquatic food chain, dioxins enter water systems via direct

                   discharge or deposition m'_d n.moff from watersheds.  Fish accumulate these compounds through

                   their direct contact with water, suspended particles, bottom sediments and through the

                   consumption of aquatic organisms.  Although these two pathways are thought to normally

                   dominate contribution to the commercial food supply, others can also be important.  Elevated

                   dioxin levels ill cattle resulting fi'om animal contact with pentacholorophenol treated wood have

          )        been documented by the USDA.  Animal feed contamination episodes have led to elevations of

          I         dioxins in poultu in the United States, mill< in Germany, and meat/dairy products in Belgium.

         ,2                  Deposition can occur directly onto soil or onto plant surfaces.  At present, it is unclear

        13        whether atmospheric deposition represents primarily current contributions of dioxin and related

        14        compounds fi'om all media reaching the atmosphere or whether it is past emissions of dioxin and

        15        related compounds which persist and recycle in the environment.  Understanding the relationship

        16        between these two scenarios will be particularly important in understanding the relative

        17        contributions of individual point sources of these compounds to the food chain and assessing the

        18        eftbctiveness of control strategies focused on either current or past emissions of dioxins in

        19        attempting to reduce the levels in food.

        20                  As discussed in Section 4.3, estimates for background levels of dioxin-like compounds in

        21         environmental media are based on a variety of studies conducted at different locations in North

        22        Ame_'ica.  Ofthe studies available for this compilation, only those conducted in locations

        23        representing "background"  were selected.  The amount and representativeness of the data varies,

        24        but in general these data lack the statistical basis to establish true national means.  The

        25        environmental media concentrations were consistent among the various studies, mid were

        26        consistent with similar studies in Westem Europe.  These data are the best available for

        27        comparing site specific values to national backm'ound levels.  Because of the limited number of

        28        locations exan_ined, however, it is not known if these ranges adequately capture the full national

        29        variability, if sigalificant regional variability exists making national means of limited utility, or if

        30        elevated levels above this range could still be the result of background contanlination processes.

 

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            As new data are collected these ranges are likely to be expanded and refined.  The limited data on

            dioxin-like PCBs ia envirorm'_ental media are sammarized in the document (Partl, Volume III,

             Chapter 4), but were not judged adequate for estimating background levels.

                      Concentrations of CDD/CDFs and PCBs in the U.S. envirolunent were consistently low

             prior to the 1930s.  Then, concentrations rose steadily until about 1970.  At this time, the trend

             reversed and the concentrations have declined to the present.  The most compelling supportive

             evidence of this trend for the CDD/CDFs aa_d PCBs comes from dated sediment core studies.

             Sediment concentrations in these studies are generally assumed to be an indicator of the rate of

    ,_        atmospheric deposition.  CDD/CDF and PCB concentrations in sediments began to increase

   O        around the 1930s, and continued to increase until about 1970.  Decreases began in 1970 and have

    1         continued to the time of the most recent sediment samples (about 1990).  Sediment data from 20

  12        U.S. lakes and rivers from seven separate research efforts consistently support this trend.

  13        Additionallyl sediment studies in lakes located in several European countries have shown similar

  14        trends.

  15                  It is reasonable to assume that sediment core trends should be driven by a similar trend in

  1O        emissions to the envirorunent.  The period of increase generally matches the time when a variety

  17        of industrial activities began r/sing and the period of decline appears to correspond with _-owth

  18        in pollution abatement.  Many of these abatement efforts should have resulted in decreases in

  19        dioxin emissions, i.e. elimination of most open burning, particulate controls on combustors,

  20        phase out of leaded gas, and bans on PCBs, 2,4,5-T, hexachlorophene, and restrictions on use of

  21        pentachlorophenol.  Also, the national source inventory of this assessment documented a

  22        significant decline in emissions from the late 1980s to the mid-1990s.  Further evidence cfa

  23        decline in CDD/CDF levels in recent years is emerging from data, primarily from Europe,

  24        showing declines in foods and human tissues.

  25                  In addition, as to the congener-specific PCB data discussed earlier, a wealth of data on

  26        total PCBs and Aroclor mixtures exist which also supports these trends.  It is reasonable to

  27        assume that the trends for dioxin-like PCBs are similar to those for PCBs as a class because the

  28        predominant source of dioxin-like PCBs is the general production of PCBs in Aroclor mixtures.

  29        PCBs were intentionally marmfactm'ed in large quantities from 1929 until production was banned

  30        in the U.S. in 1977.  U.S. production peaked in 1970, with a volume of 39,000 metric tons.

 

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                  Further support is derived from data showing declining levels of total PCBs in Great Lakes

          :        sediments and biota during the 1970s and 1980s.  These studies indicate, however, that during

                   the 1990s the decline is slowing and may be leveling off.

         4                  Because dioxin-like chemicals are persistent and accumulate in biological tissues,

         5        particularly in animals, the major route of human exposure is through ingestion of foods

         6        containing minute quantities (part per trillion or ppi levels) of dioxin-like compounds.  This

         7        results in wide-spread, low-level exposure of the general population to dioxin-like compounds.

         8        The issue of general population, background exposm'e is discussed above.

         9

        10        Risk Characterization Summary Statement

        11                  Based on all of the data reviewed in this reassessment and scientific inference, a picture

        12        emerges of TCDD and related compounds as potent toxicants in animals with the potential to

        13        produce a spectrum of effects.  Some of these effects may be occurring in humans at general

        14        population background levels and  may be resulting in adverse impacts on human health.  The

        15        potency and fundamental  level at which these compounds act on biological systems is analogous

        leo        to several well studied l'lormones.  Dioxin and related compounds have the ability to alter the

        17        pattern of growth and differentiation ora number of cellular targets by initiating a series of

        18        biochemical and biological events resulting in the potential for a spectrum of cancer and non-

        19        cancer responses in animals and humans.  Despite this potential, there is currently no clear

        20        indication of increased disease in the general population attributable to dioxin-like compounds.

        21         The lack of a clear indication of disease in the general population should not be considered

        22        strong evidence for no effect of exposure to dioxin-like compounds,  Rather lack of a clear

        23        indication of disease may be a result of the inability of our cmTent data and scientific tools to

        24        directly detect effects at these levels of human exposure.  Several factors suggest a need to

        25        further evaluate the impact of these chemicals on humans at or near current background  levels.

        26        These are: the weight of the evidence on exposm'e and effects; an apparently low margin-of-

        27        exposure/hr noncancer effects; and potential for significant risks to some portion of the  general

        28        population and additivity to background processes related to carcinogenicity in the case  of

        29        incremental exposures above background.

        30

 

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                                                                                                                  May 1, 2000

 

     DRAFT-- DO NOT QUOTE OR CITE

 

 

      7.0 REFERENCES

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

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