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1
2
3
4
5
6
7
8
9 Exposure
and Health Assessment for
l0 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
l l and Related Compounds
12
13
PART 3
14 Integrated Summary and Risk
Characterization for
15
2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
16 and Related
Compounds
17
18
19
20
2I NOTICE
22
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
28
29
30
31 National Center
for Environmental Assessment
32 Office of Research and
Development
33 U.S. Environmental Protection Agency
34 Washington, D.C.
35
36
37
TABLE of CONTENTS
4
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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
23
24
25
26
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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)
12
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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
8
9
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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.
I0
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.
23
24
25
26
27
28 Table 1-1. The TEF Scheme for I-TEQDF*
29
13
13
14
15
16
17
18
19
20
21
22
23
24
25
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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
20
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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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22
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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.,
24
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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
25
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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
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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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DRAFT -- DO NOT QUOTE OR CITE May 1, 2000
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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DRAFT -- DO NOT QUOTE OR
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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
51
DRAFT -- DO NOT QUOTE OR CITE May 1, 2000
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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DRAFT-- DO NOT QUOTE OR CITE May 1, 2000
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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DIL4.FT -- DO NOT QUOTE OR CITE May 1, 2000
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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DRAFT-- DO NOT QUOTE OR CITE May 1, 2000
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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2000
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
90
DRAFT -- DO NOT QUOTE OR
CITE May 1, 2000
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