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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
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1 documented. U.S.
workers in certain segments of the chemical industry, had elevated levels of
2 TCDD exposure, with some tissue measurements in the
thousands of ppt TCDD. There is no
3 clear evidence that elevated exposures are currently
occurring among U.S. workers. Documented
4 examples of past exposures for other groups include certain
Air Force personnel exposed to
5 Agent Orange during the Vietnam War and people exposed as a
result of industrial accidents in
6 Europe and Asia.
7 Consumption
of breast milk by nursing infants may lead to higher levels of exposure
8 compared to the intake of non-nursing
infants and dietary intakes later in life. A number of
9 studies have measured levels of the dioxin-like compounds
in human breast milk, yielding an
10 average of 35 ppt TEQDFP-WHO98. Based on a six month nursing scenario, the
average daily
11 intake for an infant is about 100 times higher than the
adult daily intake on a body weight basis:
12 the adult intake is 1 pg TEQDFP-WHO98/kg-d,
while the infant intake while breast feeding would
13 be 100 pg TEQDFP-WHO98/kg-d. The differences in body burden between
nursing infants and
14 adults are expected to be much less than the differences in
daily intake. On a mass basis, the
15 cumulative dose to the infant under this scenario is about 9% of the lifetime intake.
16 Consumption of unusually high amounts of fish,
meat, or dairy products containing
17 elevated levels of dioxins and
dioxin-like PCBs can lead to elevated exposures in comparison to
18 the general population. Most people eat some fish from multiple
sources, both fresh and salt
19
water The typical dioxin
concentrations in these fish and the typical rates of consumption are
20
included in the mean background calculation of exposure. People who
consume large quantities
21
of fish at typical
contamination levels may have elevated exposures since the concentration of
22
dioxin-like compounds in fish are generally higher than in other animal
food products. These
23
kinds of exposures are addressed within the estimates of variability of
background and are not
24
considered to result in highly exposed populations. If high-end consumers obtain their fish
from
25
areas where the concentration of dioxin-like chemicals in the fish is
elevated, they may constitute
26
a highly exposed subpopulation.
Although this scenario seems reasonable, no supporting data
27
could be found for such a highly exposed subpopulation in the U.S. One
study measuring dioxin-
28
like compounds in blood of sports fishers in the Great Lakes area showed
elevations over mean
29
background, but within the range of normal variability. Elevated CDD/CDF levels in human
30
blood have been measured in Baltic fishermen. Similarly elevated levels of coplanar PCBs have
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1 been measured in the blood of fishers on the north shore
of the Gulf of the St. Lawrence River
2 who consumed large amounts of seafood.
3 Similarly, high exposures to dioxin-like chemicals
as a result of consuming meat and
4 dairy products would only occur in situations where
individuals consume large quantities of
5 these foods and the level of these compounds is
elevated. Most people eat meat and
dairy
6 products
from multiple sources and, even if large quantities are consumed, they are not
likely to
7 have unusually high exposures. Individuals who raise their own livestock for basic subsistence
8 have the potential for higher exposures if local levels of
dioxin-like compounds are high. One
9 study in the U.S. showed elevated levels in chicken eggs
near a contaminated soil site. European
10 studies at several sites have shown elevated CDD/CDF levels
in milk and other animal products
I 1 near combustion sources.
12
13 4.6.
Environmental Trends (cross
reference: Part I, Volume III, Chapter 6)
14 Concentrations of CDD/CDFs and PCBs m the U.S.
environment were consistently Iow
15 prior to the 1930s.
Then, concentrations rose steadily until about 1970. At this time, the
trend
16 reversed and the concentrations have declined to the
present.
17 The most compelling supportive evidence of this
trend for the CDD/CDFs and PCBs
18 comes from dated sediment core studies. Sediment
concentrations in these studies are generally
19 assumed to be an indicator of the rate of atmospheric
deposition. CDD/CDF and PCB
20 concentrations in sediments began to increase around the
1930s, and continued to increase until
21 about 1970.
Decreases began in 1970 and have continued to the time of the most
recent sediment
22 samples (about 1990).
Sediment data from 20 U.S. lakes and rivers from seven separate research
23 efforts consistently support this trend. Additionally,
sediment studies in lakes located in several
24 European countries have shown similar trends.
25 It is reasonable to assume that sediment core
trends should be driven by a similar trend in
26 emissions to the environment. The period of increase generally matches the time when a variety
27 of industrial activities began rising and the period of
decline appears to correspond with growth
28 in pollution abatement.
Many of these abatement efforts should have resulted in decreases in
29 dioxin emissions, i.e. elimination of most open burning,
particulate controls on combustors,
30 phase out of leaded gas, and bans on PCBs, 2,4,5-T,
hexachlorophene, and restrictions on use of
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1
pentachlorophenol. Also, the
national source inventory of this assessment documented a
2
significant decline in emissions from the late 1980s to the mid-1990s.
Further evidence of a
3
decline in CDD/CDF levels in recent years is emerging from data,
primarily from Europe,
4
showing declines in foods and human tissues.
5 In addition to the congener-specific PCB data
discussed earlier, a wealth of data on total
6
PCBs and Aroclor mixtures exist which also supports these trends. It is reasonable to assume
7
that the trends for dioxin-like PCBs are similar to those for PCBs as a
class because the
8
predominant source of dioxin-like PCBs is the general production of PCBs
in Aroclor mixtures.
9
PCBs were intentionally manufactured in large quantities from 1929 until
production was banned
10
in the U.S. in 1977. U.S.
production peaked in 1970, with a volume of 39,000 metric tons.
11
Further support is derived from data showing declining levels of total
PCBs in Great Lakes
12
sediments and biota during the 1970s and 1980s. These studies indicate, however, that during
13
the 1990s the decline is slowing and may be leveling off.
14 Past human exposures to dioxins were most
likely higher than current estimates. This
is
15
supported by a study which applied a non-steady state pharmacokinetic
model to data on
16
background U.S. tissue levels of 2,3,7,8-TCDD from the 1970s and 80s.
Various possible intake
17
histories (pg/kg-day over time) were tested to see which best fit the
data. An assumption of a
18
constant dose over time resulted in a poor fit to the data. The "best fit" (statistically
derived) to
19
the data was found when the dose, like the sediment core trends, rose
through the 60s into the
20
70s, and declined to current levels.
Some additional support for this finding comes from a
21
limited study of preserved meat samples from several decades in the
twentieth century. One
22
sample, from before 1910, showed very low concentrations of dioxins and
coplanar PCBs.
23
Thirteen other samples, from the 1940s until the early 1980s,
consistently showed elevated levels
24 of all dioxin-like compounds as compared
to food surveys conducted during the 1990s.
25
26
5.0 DOSE-RESPONSE
CHARACTERIZATION
27
28
Previous sections of this integrated summary have focused on
characterizing the hazards of and
29
exposure to dioxin-like compounds. In order to bring these issues
together and provide an
30
adequate characterization of risk, the relationships of exposure to dose
and, ultimately, to
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1 response must be evaluated.
Key questions to be asked include:
1) What can be said about the
2 shape of the dose-response function in the observable range
and what does this imply about
3 dose-response in the range of environmental exposures? 2)
What is a reasonable limit (critical
4 dose or point of departure) at the lower end of the
observable range and what risk is associated
5 with this exposure?
In addition, one can address the issue of extrapolation beyond the range
of
6 the data in light of the answers to the above
questions. While extrapolation of risks
beyond the
7 range of observation in animals and/or humans is an
inherently uncertain enterprise, it is
8 recognized as an essential component of the risk assessment
process (NAS, 1983). The level of
9 uncertainty is dependent on the nature (amount and scope) of
the available data and on the
I0 validity of the models which have been used to characterize
dose-response. These form the bases
11
for scientific inference
regarding individual or population risk beyond the range of current
12 observation (NAS 1983, 1994)
13 In Part 2, Chapter 8, the body of literature
concerning dose-response relationships of
14 TCDD has been presented. This Chapter addresses the
important concept of selecting an
15 appropriate metric for cross-species scaling of dose and
presents the results of empirical
16 modeling for many of the available data sets on TCDD
exposures in humans and in animals.
17 Although not all human observations or animal experiments
are amenable to dose-response
18 modeling, over 200 data sets were evaluated for shape, and
an effective dose (ED) value
19 expressed as a percent response for the endpoint being
evaluated is presented e.g. ED01
equals an
20 effective dose for a 1% response. The analysis of dose-response relationships for TCDD,
21 considered within the context of toxicity equivalence,
mechanism of action and background
22 human exposures, helps to elucidate the common ground and
the boundaries of the science and
23 science policy components inherent in this risk
characterization for the broader family of dioxin-
24 like compounds, For
instance, the dose-response relationships provide a basis to infer a point of
25 departure for extrapolation for cancer and noncancer risk for
a complex mixture of dioxin-like
26 congeners given the assumption of toxicity equivalence as
discussed in Part 2, Chapter 9.
27 Similarly, these relationships provide insight into the
shape of the dose-response at the point of
28 departure which can help inform choices for extrapolation
models for both TCDD and total TEQ.
29 In evaluating the dose-response relationships for
TCDD as a basis for assessing this
30 family of compounds, both empirical dose response modeling
approaches as well as mode-of-
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I action based approaches have been developed and applied
(See Part2, Chapter 8; Portier et al,
2
1996). Empirical models have
advantages and disadvantages relative to more ambitious
3
mechanism-based models.
Empirical models provide a simple mathematical model that
4
adequately describes the pattern of response for a particular data set
and can also provide the
5
means for hypothesis testing and interpolation between data points. In
addition, they can provide
6
qualitative insights into underlying mechanisms. However, the major
disadvantage is their
7
inability to quantitatively link data sets in a mechanistically
meaningful manner. On the other
8
hand, mechanism-based modeling can be a powerful tool for understanding
and combining
9
information on complex biological systems. Use of a truly mechanism-based approach can, in
10
theory, enable more reliable and scientifically sound extrapolations to
lower doses and between
11
species. However, any scientific uncertainty about the mechanisms that
the models describe is
12
inevitably reflected in uncertainty about the predictions of the models.
13 Physiologically-based pharmacokinetic (PBPK) models
have been validated in the
14
observable response range for numerous compounds in both animals and
humans. The
15
development of PBPK models for disposition of TCDD in animals has
proceeded through
16
multiple levels of refinement, with newer models showing increasing
levels of complexity by
17
incorporating data for disposition of TCDD, its molecular actions with the
Ah receptor and other
18
proteins, as well as numerous physiological parameters (Part 2, Chapter
1). These have provided
19
insights into key determinants of TCDD disposition in treated animals.
The most complete PBPK
20
models give similar
predictions about TCDD tissue dose metrics.
The PBPK models have been
2l extended to generate predictions for early biochemical
consequences of tissue dosimetry of
22
TCDD such as induction of CYP1A1.
Nevertheless, extension of these models to more complex
23
responses are more uncertain at this time. Differences in interpretation of the mechanism of
24
action lead to varying estimates of dose-dependent behavior for similar
responses. The shape of
25
the dose-response curves governing extrapolation to low doses are
determined by these
26
hypotheses and assumptions. At
this time, the knowledge of the mechanism of action of dioxin
27
and receptor theory, and the available data of dose-response, do not
firmly establish a scientific
28
basis for replacing a linear procedure for estimating cancer
potency. Consideration of this same
29
information indicates that the use of different procedures to estimate
the risk of exposure for
30
cancer and noncancer endpoints may not be appropriate. Both the cancer and noncancer effects
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I
of dioxin appear to result from qualitatively similar modes-of-action of
action. Initial steps in the
2
process of toxicity are the same and many early events appear to be
shared. Thus, the inherent
3
potential for low dose significance of either 'type' of effect (cancer
or noncancer) should be
4
considered equal and evaluated accordingly. In the observable range
around 1% excess response,
5
the quantitative differences are relatively small. Below this response,
the different mechanisms
6
can diverge rapidly. The use of predicted biochemical responses as dose
metrics for toxic
7
responses is considered as a
potentially useful application of these models. However, greater
8
understanding of the linkages between these biochemical effects and
toxic responses is needed to
9
reduce the potentially large uncertainty associated with these
predictions.
10
5.1 Dose Metric(s)
11 One of the most difficult issues in risk
assessment is the determination of the dose metric
12
to use for animal-to-human extrapolations. To provide significant insight into differences in
13
sensitivity among species, the appropriate animal-to-human extrapolation
of tissue dose is
14
required. The most appropriate
dose metric should reflect both the magnitude and frequency of
15 exposure, and should be
clearly related to the toxic endpoint of concern by a well-defined
16
mechanism. This is, however, often difficult because human exposures
with observable
17
responses may be very. different from highly controlled exposures in
animal experiments. In
18
addition, comparable exposures may be followed by very different
pharmacokinetics (absorption,
19
distribution, metabolism and/or elimination) in animals and humans. Finally, the sequelae of
20
exposure in the form of a variety of responses related to age, organ-,
and species-sensitivity
21
complicate the choice of a common dose metric. Despite these complexities, relatively simple
22
default approaches including body surface or body weight scaling of
daily exposures have
23
often been recommended (EPA, 1992; EPA, 1996).
24 Given the data available on dioxin and related
compounds, dose can be expressed in a
25
multitude of metrics (Devito et al, 1995) such as daily intake
(ng/kg/d), current body burden
26
(ng/kg), average body burden over a given period of time, plasma
concentration, etc. Examples
27
of other dose metrics of relevance for TCDD and related compounds can be
found in the
28
literature including concentration of occupied Ah receptor (Jusko,
1995), induced CYP1A2
29
(Andersen et al, 1997; Kohn, 1993) and reduced epidermal growth factor
receptor (EGFR)
30
(Portier and Kohn, 1996). Considering the variety of endpoints seen with
TCDD, and expected
31
with other dioxin-like chemicals, in different species, it is unlikely
that a single dose metric will
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t be adequate for interspecies extrapolation for all of
these endpoints. The issue of an appropriate
2
dose metric for developmental effects considering the potential for a
narrow time window of
3
sensitivity, for instance, has been discussed in a number of places in
this document.
4
Furthermore, the use of different dose metrics with respect to the same
endpoint may lead to
5
widely diverse conclusions. This
latter point is discussed in more detail in Part 2, Chapter 8.
6
Nevertheless, it is possible to express dose in a form that allows for
comparison of responses for
7
selected endpoints and species.
This can be done by either choosing a given exposure and
8
comparing responses (e.g. unit risk level (URL)) or choosing a
particular response level and
9 comparing the associated
exposures (e.g. effective dose (ED)).
10 As discussed above: dose can be expressed in a
number of ways. For TCDD and other
11
dioxin-like compounds, attention has focused on the consideration of
dose expressed as daily
12
intake (ng/kg/day), body burden (ng/kg), or area under the plasma
concentration versus time
13
curve (AUC) (Devito et al, 1995;
Aylward et al, 1996). While the AUC may
be a more precise
14 dose metric, the concept of physiological time (lifetime of
an animal) complicates the
15
extrapolation, as the appropriate scaling factor is uncertain for toxic
endpoints. Because body
16
burden incorporates differences between species in TCDD half-life (these
differences are large
17
between rodent species and humans (See Table 8.2)), this dose metric
appears to be the most
18
practical for this class of compounds (Devito et al, 1995). Average lifetime body burden is best
19
suited for steady-state conditions, with difficulties arising when this
dose metric is applied to
20
evaluation of acute exposures, such as those occurring in the 1976
accidental exposure of some
21
people living in Seveso, Italy (Bertazzi and di Domenico, 1994). In cases such as this, increased
22
body burden associated with the acute exposure event is expected to
decline (half-life for TCDD
23
is approximately 7 years) until it begins to approach a steady state
level associated with the much
24
smaller daily background intake. However, this issue of acute exposure
is not a major factor in
25
the current analyses. In
general, daily excursions in human exposure are relatively small and
26
have minor impact on average body burden. Instead, physiologically-based
pharmacokinetic
27
(PBPK) models suggest that human body burdens increase over time and
begin to approach
28
steady state after approximately 25 years with typical background
doses. Occupational
29
exposures represent the middle ground where daily excursions during the
working years can
30
significantly exceed daily background intakes for a number of years,
resulting in elevated body
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1 burdens. This is
illustrated in Table 5-1. Estimation of
the range and mean or median of
2 "attained" body burden in accidentally- or
occupationally exposed cohorts is presented and
3 compared to body burdens based on background
exposures. These data are presented
graphically
4 in Figure 5-1. As discussed earlier, using background of
total body burden (TEQDFP-WHO98) as a
5 point of
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100
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I comparison, these often- termed "highly exposed"
populations have maximum body burdens
2 which are relatively close to general population
backgrounds at the time. When compared
to
3 background body burdens of the late 1980s, many of the
median values and some of the mean
4 values fall within a range of one order of magnitude
(factor of 10) and all fall within a range of
5 two orders of magnitude (factor of 100). General population backgrounds at the time
are likely
6 to have been higher.
Since these are attained body burdens, measured at the time of the
Seveso
7 accident or back-calculated to the time of last known
elevated exposure, being compared to
8 background, average lifetime body burdens in these cohorts
will be even closer to lifetime
9 average background levels. This will be important if, as
demonstrated for some chronic effects in
10
animals and as assumed when relying on average body burden as a dose
metric, cancer and other
11 noncancer effects are a consequence of average tissue
levels over a lifetime. Body burdens
begin
12 to slowly decline soon after elevated exposure ceases. Some
data in humans and animals suggest
13 that elimination half-lives for dioxin and related
compounds may be dose dependent, with high
14 doses being eliminated more rapidly than lower doses. Nonetheless, the use of an approximately
15 7 year half-life of elimination presents a reasonable
approach for evaluating both back-calculated
16 and average lifetime levels since for most cohorts the
exposure is primarily to TCDD.
17 The ability to detect effects in epidemiologic
study is dependent on a sufficient difference
18 between control and
exposed populations. The relatively small difference (<10-100 fold)
19 between exposed and controls in these studies makes
exposure characterization in the studies a
20 particularly serious issue. This point also strengthens the
importance of measured blood or tissue
21 levels in the epidemiologic analyses, despite the
uncertainties associated with calculations
22 extending the distribution of measured values to the entire
cohort and assumptions involved in
23 back-calculations.
24 Characterization of the risk of exposure of
humans today remains focused on the levels
25 of exposure that occur in the general population, with
particular attention given to special
26 populations (See Part 1).
For evaluation of multiple endpoints and considering the large
27 differences in half-lives for TCDD across multiple species,
it is generally best to use body burden
28 rather than daily intake as the dose metric for comparison
unless data to the contrary is presented.
29 Further discussion of this point which provides the
rationale for this science-based policy choice
30 is presented in Part 2, Chapters 1 and 8.
31
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1 Calculations
of Effective Dose (ED)
2 Comparisons across multiple endpoints, multiple
species and multiple experimental
3 protocols are too complicated to be made on the basis of
the full dose-response curve. As
4 discussed above, comparisons of this sort can be made by
either choosing a given exposure and
5 comparing tine responses, or choosing a particular response
level and comparing the associated
6 exposures, in the analyses contained in Chapter 8 and
elsewhere in the reassessment,
7 comparison of responses are made using estimated exposures
associated with a given level of
8 excess response or risk.
To avoid large extrapolations, this common level of excess risk was
9 chosen such that for most studies, the estimated exposure is
in or near the range of the exposures
I0 seen in the studies being compared with extra weight given
to the human data. A common
11 metric for comparison is the effective dose or ED, which is
the exposure dose resulting in an
I2 excess response over background iii the studied
population. The USEPA has suggested
this
13 approach in calculating Benchmark Doses (BMD) (Allen et al,
1994) and in its proposed
14 approaches to quantifying cancer risk (EPA, 1996). While effective dose evaluation at the 10%
15 response level (ED_0 or lower bound on ED_0 (LED_0)) is
somewhat the norm, given the power of
16 most chronic toxicology studies to detect an effect, this
level is actually higher than those
17 typically observed in the exposed groups in studies of TCDD
impacts on humans. To illustrate,
18 h.tng cancer mortality has a background lifetime risk of
approximately 4% (smokers and
19 nonsmokers combined), so that even a relative risk of 2.0 (
2 times the background lifetime risk)
20 represents approximately a 4% increased lifetime risk. Based upon this observation and
21 recognizing that many of the .TCDD-induced endpoints
studied in the laboratory include 1%
22 effect levels in the experimental range, Chapter 8 presents
effect/ye doses of 1% or EDm. The
23 use of ED values
below 10% is consistent with the
Agency's guidance on the use of mode-of-
24 action in assessing risk, as described in the evaluation
framework discussed in Section 3.3., in
25 that the observed range for many "key events"
extends down to or near the 1% response level.
26 Deten'nining the dose at which key events for dioxin
toxicity begin to be seen in a heterogeneous
27 human population provides important information for
decisions regarding risk and safety.
28
29 5.2 Empirical Modeling of Individual Data Sets
30 As described in Chapter 8, empirical models have advantages
and disadvantages relative
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I
to more ambitious
mechanism-based models. Empirical
models provide a simple mathematical
2 model that adequately describes the pattern of response for
a particular data set and can also
3 provide the means roi' hypothesis testing and interpolation
between data points. In addition, they
4 can provide qualitative insights into underlying
mechanisms. However, the major disadvantage is
5 their inability to quantitatively link data sets in a
mechanistically meaningful manner. Data
6 available for several biochemical and toxicological effects
of TCDD, and on the mechanism of
7 action of this chemical, indicate that there is good
qualitative concordance between responses in
8 laborato_5' animals and humans (see Table 1). For example, human data on exposure zmd
cancer
9 response appear to be qualitatively consistent with
animal-based risk estimates derived from
10 carcinogenicity bioassays (See Part 2, Chapter 8). These
and other data presented tl'u'oughout this
11 reassessment would suggest that animal models are
generally an appropriate basis for estimating
12 human responses.
Nevertheless, there are clearly differences in exposures and responses
between
13 animals and humans, and recognition of these is essential
when using animal data to estimate
14 human risk. The
level of confidence ir! any prediction of human risk depends on the degree to
15 which the prediction is based on an accurate description of
these interspecies extrapolation
16 factors. See
Chapter 8 for a further discussion of this point.
17 Almost all data are consistent with the
hypothesis that the binding of the TCDD to the Ah
18 receptor is the first step in a series of biochemical,
cellular, and tissue changes that ultimately
19 lead to toxic responses observed in both experimental
animals and humans (See Part 2, Chapter
20 2). As such, an
analysis of dose-response data and models should use, whenever possible,
21 information on the quantitative relationships between
ligand (i.e. TCDD) concentration, receptor
22 occupancy, and biological response. However, it is clear that multiple
dose-response
23 relationships are possible when considering ligand-receptor
mediated events. For example,
24 dose-response relationships for relatively simple
responses, such as enzyme induction, may not
25 accurately predict dose-response relationships for complex
responses such as developmental
26 effects and cancer.
Cell- or tissue-specific factors may determine the quantitative
relationship
27 between receptor occupancy and the ultimate response. Indeed, for TCDD there is much
28 experimental data from studies using animal and human
tissues to indicate that this is the case.
29 This serves as a note of caution as empirical data on TCDD
are interpreted in the broader context
30 of complex exposures to mixtures of dioxin-like compounds
as well as to non-dioxin-like
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I
toxicants.
2 As for other chemical mechanisms where high
biological potency is directed ttu'ough the
3
specific and high affinity interaction between chemical and critical
cellular target, the
4
supposition of a response threshold roi' receptor-mediated effects is a
subject for scientific
5
debate. The basis of this
controversy has been recently sun'u'narized (Sewall and Lucier, 1996).
6 Based on classic receptor theory, the occupancy
assumption states that the magnitude of
7
biological response is
proportional to the occupancy of receptors by drug molecules. The
8
'typical' dose-response curve for such a receptor-mediated response is
sigmoidal when plotted on
9
a semilog graph or hyperbolic if plotted on a arithmetic plot. Implicit in this relationship is
10
low-dose linearity (0-10°_, fi'actional response) ti'u'ough the origin.
Although the law of mass
11
action predicts a single molecule of ligand can interact with a
receptor, thereby inducing a
12
response, it is also stated that there must be some dose that is so low
that receptor occupancy is
13
trivial and therefore no perceptible response is obtainable.
14 Therefore, the same receptor occupancy
assulnption of the classic receptor theow is
15
interpreted by different parties as support for and against the
existence of a threshold. It has been
16
stated that the occupancy assulnption cannot be accepted or rejected on
experinaental or
17
theoretical grounds (Goldstein et al,
1974). To detennine the
relevance of receptor interaction
18
for TCDD-mediated responses, one must consider (1) altematives as well
as limitations of the
19
occupancy theory; (2) molecular factors contributing to measured endpoints; (3) limitations of
20
experimental methods; (4) contribution of measured effect to a relevant
biological/toxic
21
endpoint; and (5) background exposure.
22 Throughout this reassessment, each of these
considerations has been explored within the
23
cul-rent context of the understanding of the mechanism of a action of
TCDD, of the
24
methods for analysis of dose-response for cancer and noncancer
endpoints, and of the available
25
data sets of TCDD dose and effect for several rodent species, and humans
that were
26
occupationally exposed to TCDD at levels exceeding the exposure of the
general population.
27
5.2.1 Cancer
28 As described in Section 2.2.2.4 above, TCDD has
been classified as a known human
29
carcinogen, mad is a carcinogen in all species and strains of laboratory
animals tested. The
30
epidemioiogical database for TCDD, described in detail in Part 2,
Chapter 7a, suggests that
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DRAFT-- DO NOT QUOTE OR CITE May 1, 2000
1
exposure may be associated with increases in all cancers combined, in
respiratory tumors, and,
2
perhaps, in soft-tissue sarcoma.
Although there are sufficient data in animal cancer studies to
3
model dose-response for a number of tumor sites, as with mm2y chemicals,
it is generally
4
difficult to find human data with sufficient information to model
dose-response relationships.
5
For TCDD, there exist tI'n'ee studies of human occupational exposure
which provide enough
6
information to perfon-n a quantitative dose-response analysis. These are
the NIOSH Study
7
(Fingerhut et al, 1991), the Hamburg Cohort Study (Manz et al, 1991),
and the BASF Cohort
8
Study (Zober et al, 1990). In Part 2, Chapter 8, simple empirical models
were applied to these
9
studies for which exposure-response data for TCDD are available in human
populations.
10
Modeling cancer in
humans uses slightly different approaches than those used in
11
modeling animal studies. The modeling approach used in the analysis of
the human
12
epidemiology data for all ca]cers comb/ned and lung cancer involves
applying estimated human
13
body burden to cancel' rcsponse, and estimating parameters in a linear
risk model for each data
14
set. A linear risk model was
used since the number of exposure groups available for analyses
15 was too small to
support more complicated models.
Because of this, evaluating the shape of the
16
dose-response data for the human studies was not done. Access to--th_ raw data may make it
17
possible to use more complicated mathematical forms which allow for the
evaluation of shape.
18
In the one case in which this has been done, the dose-response shape
suggested response which
19
was less than linear (dose raised to a power <1) (Becher et al,
1998). For these studies, there are
20
several assumptions and uncertainties involved in the modeling of the
data including
21
extrapolation of dosage, both in back-calculation and in elimination
ldnetics, and the type of
22
extrapolation model employed,
23 As described in Part 2, Chapter 8, the data used
in the analyses are from Aylward et al.
24
(1996) for the NIOSH study, Flesch-lanys et al. (1998) for the Hamburg
cohort, and Ott and
25
Zober (1996a;I996b) for the BASF cohort. The limited information
available from these studies
26
is in the form of standard mortality ratios (SMRs) and/or risk ratios by
exposure subgroups with
27
some estimate of cumulative subgroup exposures. Exposure subgroups were
defined either by
28
number of years of exposure to dioxin-yielding processes or by
extrapolated TCDD levels. No
29
study sampled TCDD blood serum levels for more than a fraction of their
cohort and these
30 samples were
generally taken decades after last known exposure. In each study, serum fat or
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i body fat levels of TCDD were back calculated using a
first-order model. The assumed
half-life
2 of TCDD used in the model varied from study to study. Aylward et al. used the average TCDD
3 levels of those sampled in an exposure subgroup to
represent the entire subgroup. Flesch-Janys et
4 al. and Ott and Zober performed additional calculations,
using regression procedures with data
5 on time spent at various occupational tasks to estimate
TCDD levels for all members of their
6 respective cohorts. They then divided the cohorts into
exposure groups based on the estimated
7 TCDD levels The information presented in the literature
cited above was used to calculate
8 estimated average TCDD dose levels in Chapter 8.
9 To provide EDc;_ estimates for comparison in
Chapter 8, Poisson regression (Breslow and
10 Day, 1987) was used
to fit a linear model to the data described above. Analysis of animal cancer
11 data suggests a mixture of linear and non-linear responses
with linear shape parameters
12 predominating; complex responses to TCDD, both cancer and
noncancer, are more often than
13 not nonlinear.
Besides the issue of use of a linear model, several other important
uncertainties
14 discussed in Chapter 8 are the representativeness and
precision of the dose estimates that were
15 used, the choice of half-life and whether it is dose
dependent, and potential interactions between
16 TCDD and smoking or other toxicants. Nevertheless, with these qualifications, it
is possible to
17 apply simple empirical models to studies in which exposure
data for TCDD are available in
18 human populations.
19 The analysis of these three epidemiological
studies of occupationally exposed individuals
20 suggest an effect of TCDD on all cancers, and hmg cm-tcers
in the adult human male. The EDs0_
21 based upon average excess body burden of TCDD ranged from
6 ngTCDD/kg to 161 ngTCDD/kg
22 in humans. The lower bounds on these doses (based on a
modeled 95% C.I.) range from 3.5
23 ngTCDD/kg to 77 n.gTCDD,q<g. For the effect of TCDD on
lung cancers, the only tumor site
24 increased in both rodents and humans, the human EDso_
ranged fi'om 24 ng/kg to 161 rig/kg.
25 The lower bounds on these doses (based on a modeled 95%
C.I.) range from 10.5 ngTCDD/kg
26 to 77 ngTCDD/kg. These estimates of EDs0_ are compared to
animal estimates later in this
27 discussion.
28 Both empirical and mechanistic models were used
to examine cancer dose-response in
29 animals. Port/er et
a1.(1984) used a simple multistage model of carcinogenesis with up to two
30 mutation stages affected by exposure to model the five
tumor types observed to be increased in
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1 the 2 year feed study of Kociba et al. (Sprague-Dawley
rats) (1978) and the eight tumor types
2 observed to be increased in the 2 year gavage cancer study
conducted by the National Toxicology
3 Program (Osbome-Mendel rats and B6C3F_ mice) (1982). The findings from this analysis which
4 examined cancer dose-response within the range of
observation are presented in Table 8.3.2.
5 which is reproduced with slight modifications as Table 5-2
below. All but one of the estimated
6 EDs0_ are above the lowest dose used in the experiment (approximately 1 ngTCDD/kg/day in
7 both studies) and are thus interpolations rather than
extrapolations. The exception, liver
cancer
8 in [emale rats from the Kociba study, is very near the
lowest dose used in this study and is only a
9 small extrapolation (from 1 ngTCDD/kg/day to 0.77
ngTCDD/kg/day). Steady-state body
10 burden calculations were also used to derive doses for
comparison across species. Absorption
11 was assumed to be 50% for the Kociba et al. study (feed
experiment) and 100% for the NTP
12 study (gavage
experiment). Also presented in Table
5-2 are the shapes of the dose-response
13 curves as detemlined by Portier et al (1984).
14
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109
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1,2000
1 this does not imply that a non-linear model such as the
quadratic or cubic would not fit these
2 data. In fact, it
is unlikely that in m'_y one case, a linear model or a quadratic model could be
3 rejected statistically for these cases. These studies had
only three experimental dose groups hence
4 these shape calculations are not based upon sufficient
doses to guarantee a consistent estimate;
5 they should be viewed with caution. The ED0_ steady state
body burdens range from a Iow value
6 of 14 ng/kg based upon the linear model associated with
liver tumors in female rats to as high as
7 1190 ng/kg based upon a cubic model associated with
thyroid follicular cell adenomas in female
8 rats. Lower bounds on the steady state body burdens in the
animals range from 10 ngTCDD/kg to
9 224 ng/kg. The corresponding estimates of daily intake
level at the ED0_ obtained from an
10 empirical linear model range fi'om 0.8 to 43 ngTCDD/kg body
weight/day depending on the
11 tumor site, species and sex of the animals investigated.
Lower confidence bounds on the
12 estimates of daily intake level at the ED0_ in the animals
range from 0.6 to 14 ngTCDD/kg body
13 weight/day. In
addition, using a mechanistic approach to modeling, Portier and Kohn (1996)
14 combined the biochemical response model of Kolm et al.
(1993) with a single initiated
15 phenotype two stage model of carcinogenesis to estimate
liver tumor incidence in female
16 Sprague-Daw!ey rats fi'om the two-year cancer bioassay of
Kociba et al. (197'8-57. By way of
17 comparison, the ED0t estimate obtained from this linear
mechanistic model of liver tm'nor
18 induction in female rats was 0.15 ngTCDD/kg body weight/day
based on intake, which is
19
equivalent to 2.7 ngTCDD/kg
steady state body burden. No lower bound on this modeled
20 estimate of steady state body burden ,,vas provided
21 As discussed in Part 2, Chapter 8, different dose
metrics can lead to widely diverse
22 conclusions. For
example, as described in Chapter 8, the ED0: intake for the animal tumor sites
23 presented above ranges [rom less than one to tens of
ng/kg/day m'_d the lowest dose with an
24 increased tumorigenic response (thyroid tun'mrs) in a rat
is 1.4 ng/kg/day (NTP, 1982). The daily
25 intake of TCDD in hun-tans is estimated to be 0.14 to .4
pgTCDD/kg/day. This implies that
26 humans are exposed to doses 3,500 to 10,000 times lower
thml the lowest tumorigenic dose in rat
27 thyroid. However,
1.4 ng&g/d in the rat leads to a steady state body burden of approximately
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1 25 ng/kg, assuming a half-life of TCDD of 23 days and
absorption from feed of 50%'. If the
2 body burden of TCDD in humans is approximately 5 ngTCDD/kg
lipid or 1.25 ng/kg body
3 Weight (assuming about 25% of body weight is lipid), humans are exposed to about 20 times
4 less TCDD than the minimal carcinogenic dose for the rat. If total TEQ is considered the
5 difference is even less, approaching only a factor of 2
difference. The difference between
these
6 two estimates is entirely due to the approximately 100
fold difference in the half-life between
7 humans and rats. At
least for this comparison, if cancer is a function of average levels in the
8 body, the most appropriate metric for comparison is the
average or steady state body burden
9 since the large differences in animal to human half-life
are accounted for.
10 Comparisons of human and animal EDs0_ from Part
2, Chapter 8 for cancer response on a
11 body burden basis show approximately equal potential for
the carcinogenic effects of TCDD. In
12 humans, restricting the analysis to log-linear models ill
Part 2, Chapter 8, resulted in cancer
13 EDso_ ranging fi'om 6 ng/kg to 161 ng/kt,. This was similar to the empirical modeling
estimates
14 from the animal studies, which ranged from 14 ng/kg to 1190
ng/kg (most estimates were in the
15 range from 14 to 500 ng&g). The lower bounds on the human body burdens at the EDso, (based
16 on a modeled 95% C.I.) range from 3.5 ngTCDD/kg to 77
ngTCDD/kg. Lower bounds On the
17 steady state body burdens m the animals range from 10
ngTCDD/kg to 224 ng/kg.. The estimate
18 for the single mechanism-based model presented earlier (2.7
ng/kg) was approximately 2 times
19 lower than the lower end of the range of human ED01
estimates and less than one times less than
20 the lower bound on the ED0, (LED0_). The same value was approximately 5 times
lower than the
21 lower end of the range of animal ED0, estimates and less
than 4 times less than the LEDm.
22 Using human and animal cancer EDs0_, their lower
bound estimates, and the value of 2.7
23 ngTCDD/kg from the single mechanism based model, slope
factors and comparable risk
24 estimates for a human background body burden of
approximately 5 ngTEQ/kg (20 ngTEQ/kg
25 lipid) can be calculated using the following equations:
26
I
steady state body burden (ng/kg) =( daily dose (nw'kg/day) *
(half-life)/'Ln(2)) (f) where fis the
fraction
absorbed from the exposure
route (umtless) and half-life is the half-life in days.
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1 Equation 5-1 Calculating Slope Factors from Body Burdens
at the ED0_
2
3 Slope Factor (per pgTEQ/kgBW/day) = Risk at ED0_ / Intake
(pgTEQ/kgBW/day) Associated
4 with Human Equivalent Steady State Body Burden at ED0_
5 where:
0 Risk at ED_: = 0.01; and
7 Intake (pgTEQ/kgBW/day) = [Body Burden at ED01 (ngTEQ/kg)*halflife
(days')] * f
8 Ln(2)
9 half life = 2593 days in humans and 25 days in rats (See
Table 8.1 in Part 2, Chapter 8)
10 f= fraction of dose absorbed; it is assumed to be 50% for
absorption from food (Kociba et al.,
11 1976) and 100%
for other routes.
12
13 Equation 5-2 Calculating Upper Bound on Excess Risk at
Human Background Body
14 Burden
15
16 Upper Bound on Excess Risk at Humm_ Background Body Burden
= ( Human-13aekground Body
17 Burden ( ng/kg))(risk at EDu_)/lower bound on Human
Equivalent Steady State Body Burden
18 (ng/kg) at EDoE
19 where:
20 Risk at ED0_ =0.01
21
22 Use of
these approaches reflects methodologies being developed within the context of
the
23 revised Cancer risk assessment guidelines. Slopes are estimated by a simple
proportional
24 method at the" point of departure" (LED0_) at the
Iow end of the range of experimental
25 observation. As
discussed below, these methods can be compared to previous approaches using
26 the hnearized multistage (LMS) procedure to determine if
the chosen approach has significantly
27 chm_ged the estimation of slope. The estimates of ED0_/LED0_ represent the human-equivalent
28 body burden for 1% excess cancer risk based on exposure to
TCDD and are assumed for
29 purposes of this analysis to be equal for TCDD equivalents
(total TEQ). This assumption is
30 based on the toxicity equivalence concept discussed
tlu-oughout this report and in detail on Part
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1
2, Chapter 9. All cancer slope factors can be compared to the Agency's
previous slope factor of
2
1.6 X 10.4
per
pgTCDD/kgBW/day (EPA, 1985).
3
4
Estimates of slope factors and risk at current background body burdens
based on human
5
data
6 Estimates of upper bound slope factors( pet'
pgTCDD/kgBW/day) calculated from the
7
human EDs0_ presented on Table 8.3.1 range fi'om 5.3 X 10'-_, if the
LED0, for all cancer deaths in
8
the Hamburg cohort is used, to 2.4 X I0'4
if the ED0_ for lung cancer deaths in the smaller BASF
9
cohort is used. ,4.11 of the
other slope factors for all cancer deaths or lung cancer deaths in the
10
thJ'e_: cohorts would fall within that range. LEDs0_ for all cancer
deaths span approximately an
11
order of magnitude and would generate slope factors in the range of 5 X
I04 to5 X10'4. Slightly
12
smaller slope factors are generated when LEDs0, roi' lung cancer are
used. The largest slope
13
factors based on LEDso_ come fi'om the Hamburg cohort (5.3 X 10.3 and
1.8 X 10.3 respectively
I4
for all cancer deaths and lung cancer deaths.) These estimates compare
well with the estimates of
15
risk associated with TCDD exposure in the Hamburg cohort published by
Becher (Becher etal.,
16
1998). The risk estimates of
Becher et al. derived from data on TCDD exposure to male workers
17
with a ten year latency and taking greater caution over other factors
affecting risk including
18
choice of model, latency, job category, dose metric and concmTent
exposures. These estimates
19
range from 1.3 X10'_ to 5.6 X 10'3 per pg TCDD,q;gBW/day. In this
analysis all excess cancers
20
are attributed to TCDD exposure despite significant levels of other
dioxin-like compounds in
21
blood measurements of this cohort (see Table 5~1). Although risk estimates using TCDD alone
22
in this cohort might suggest an
overestimate of risk, no evidence for this emerged from the
23
analysis and assuming that TCDD will still dominate total TEQ,
differences in slope factor
24
estimates are likely to be less than a factor of two and may not be
discemable. Taking into
25
account different sources of variation, Becher et al.(1998) suggest a
range of 10'3 to 10'_ for
26
additional lifetime cancer risk for a daily intake of 1 pgTCDD/kg BW/day. By inference, that
27
range couid also apply to total TEQ intake. As described in Section
4.4.2, current estimates of
28
intake in the U.S. are estimated to be approximately 1 pgTEQ/kg
BW/day. Using Equation 5-2,
29
the upper bound range of risks
estimated from current human body burdens of 5 ngTEQ/kgBW
30
(which equates to a serum level of 20 pg/g lipid (see Table 4.7)) based
on all cancer deaths in the
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1 tl'u'ee cohorts ranged from 1.4 X 10-2 tO
1.3 X 10'3; based on lung cancer deaths, the
upper bound
2 on the estimates of excess risk extended to 6 X10_. The range of these estimates provides
further
3 support for the perspective on risk provided by Becher et
al. Uncertainties associated with these
4 estimates fi'om human studies are discussed in Part 2,
Chapter 8 and in Becher et al. (1998).
5
6 Estimates of slope factors and risk at current background
body burdens based on animal
7 data
8 Upper bound slope factors ( per pgTCDD/kgBW/day)
for human cancer risk calculated
9 from lower bounds on EDs0t (LEDs0.) for the animal cancers
presented in Table 5-2 range from
10 1.9 X 10'" to
8.4 X10'5. This spans a range front
being 12 times gTeater than the previous upper
11 bound estimate on cancer slope ( 1.6 X10-4 (EPA, 1985)) to 2 times less. The largest slope
12 factor is derived from the same study as the 1985
estimate; that is, the slope factor derived from
13 the female liver cancer in the Kociba et a1.(1978) study
continues to give the largest slope factor.
14 In attempting these comparisons, two issues became
apparent. First, the body burden and the
15
intake at the ED0_ from
Pottier et al. (1994) does not result in the same slope factor as EPA
16 (1985). Despite the use of the same study results, a slope
factor of 1.8 X10'5 per
17 pgTCDD,'kgBW/day using the LMS approach. This is a factor of approximately 10 lower
than
18 the EPA (1985) estimate of the slope. The differences are attributable to the aims
of the
19 respective calculations at the time. Pottier et al. (1984)
calculated "virtually safe doses"
20 assuming that rodent and human doses scaled on a mg/kg
basis and he used the original tumor
21 counts fi-om the study.
EPA (1985), on the other hand used (BW)TM
to arrive at a human
22 equivalent dose and used the pathology results from a
re-read of the original Kociba study (
23 Albert, 1980). In
addition, rrm'lot counts were adjusted for early mortality in the study. The
24 factor to adjusi for (BW)3/4-scaling in the rat is
5.8. The correction for early mortality
can be
25 accounted for with a factor of 1.6 ( this is the ratio of
the intake values at the ED0_ with and
26 without the early mortality correction). If the Portier et al. slope factor (1.8
X10'Sper
27 pgTCDD/kgBW/day) is multiplied by these two factors, a
slope of 1.7 X10'4 per
28 pgTCDD/kgBW/day is calculated. This is equivalent to the EPA (1985) estimate of 1.6 X10'qper
29 pgTCDD/kgBW/day. Reconciling these issues is important to
assure appropriate comparisons of
30 slope factor estimates.
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1 More important is tine calculation of slope factor
estimates using current methods of analysis
2
which recognize the importance of the dose metric and the differences in
halfqife of dioxins in
3
the bodies of laboratmLy animals and humans (See part 2, Chapter 8 for
detailed discussion). The
4
major difference between the approaches used to calculate risks in the
mid-1980's (Portier et al,
5 1994; EPA, 1985)
and the current approach is the use of body burden as the dose metric for
6
animal to human dose equivalence.
All things being equal, the use of body burden accounts for
7
the approximately 100-fold
difference between half-lives of TCDD in humans and rats (2593
8
days versus 25 days (see Part 2, Table 8.1 )). Use of Equation 5-1 results in an estimated body
9
burden at the LEDo_ of 6.1
ng TEQ/Kg to be derived from the EPA (1985) Kociba tumor counts.
10
This compares favorably with the Portier estimate of 10 ng TEQ/Kg found
in Table 5-2. The
11 difference is entirely accounted for by the early deaths adjustment by EPA (1985). Use of these
12
body burdens at the LED0_ result in slope factor estimates of 1.9 X10'3
per pgTCDD/kgBW/day
13 and 3.1 X 10"
per pgTCDD/kgBW/day roi' the Chapter 8 re'id the newly derived body burden,
14
respectively. Again, the
difference is due solely to the adjustment for early mortality and EPA
15
believes this provides a better estimate of upper bound lifetime risk
than does the unadjusted.
16
EPA's new slope· factor (3.1 X
10'_ per pgTCDD/kgBW/day) is 19 times lower than the slope
17
factor from 1985.
18 A second issue with the modeling of the Kociba
data relates to the appropriate tumor
19
counts to use. As mentioned in
Section 2, Goodman and Sauer (1992) reported a second
20
re-evaluation of the female rat liver tumors in the Kociba study using
the latest pathology c_iteria
21
for such lesions. Results of this review are discussed in more detail in
Part 2, Chapter 6. The
22
review confin'ned only approximately one-third of the tumors of the
previous review (Albert,
23
1980). While this finding did
not change the determination of carcinogenic hazard since TCDD
24
induced tumors in multiple sites in this study, it does have an effect
on evaluation of
25
dose-response and on estimates of risk.. Since neither the original EPA
(1985) slope factor
26
estimate nor that of Portier et al. (1984) reflect this re-read, it is
important to factor these results
27
into the estimate of the ED0_ and slope factor. Using the LMS procedure which was used by
the
28
EPA in 1985 and the tumor counts as reported in Part 2, Chapter 6, Table
6.2, the revised slope
29
factor is reduced by approximately 3.6-fold to yield a slope factor of
4.4 X 10'5 per
30
pgTCDD/kgBW/day. However, since
the original estimates used a (BW)3/4-scaling, this n'mst be
115
i
DRAFT -- DO NOT QUOTE OR CITE May 1, 2000
1 adjusted to use body burden and obtain an appropriate
result. When dose is adjusted and
2 Equation 5-1 is
used, all L ED0_ of 22.2 ngTEQ/kg is derived and a slope factor of 8.3 X10'4
per
3 pgTCDD/kgBW/day.
This represents EPA's most current upper bound estimate of human
4
cancer risk based on animal
data. It is 5.2 times larger than the
slope factor calculated in
5 EPA(1985). This
number reflects the increase in slope factor based on use of the body burden
6 dose metric (19 times greater) and the use of the Goodman
and Sauer (I 992) pathology (3.6
7
times less).
8
9 Estimates of slope factors and risk at current background
body burdens based on a
10 mechanistic model
1 l As
discussed above, Portier and Kohn (1996) combined the biochemical response
model
12 of Kohn et al. (1993) with a single initiated phenotype two
stage model of carcinogenesis to
13 estimate Iix,et' tumor incidence in female Sprague-Dawley
rats from the Kociba et al. (1978)
14 bioassay. The model
is described in more detail in Part 2, Chapter 8. This model adequately fit
15 the tumor data, although it overestimated the observed
tumor response at the lowest dose in the
16 Kociba study. The
shape of the dose-response curve was approximately linear and the eg it-_d
17 ED0_ value for this model was 1.3 ng/kg/day. The corresponding body burden giving a 1%
18
increased effect was 2.7
ng/kg. The model authors believe that
the use of cYP1A2 as a dose
19 metric for the first mutation rate is consistent with its
role as the major TCDD-inducible estradiol
20 hydrolase in liver and with it hypothesized role in the
production of estrogen metabolites leading
21 to increased oxidative DNA damage and increased mutation
(Yager and Liehr, 1996; Hayes et al.,
22 1996; Dan.nan et al., 1986;Roy et al., 1992). Although no lower bound estimate of the ED0t
is
23 calculated, a maximum likelihood estimate of the slope
factor can be calculated. It is 7.1 X
10.3
24 per pgTCDD&gBW/day. This estimate represents an example
of the type of modeling, based on
25 key events in a mode-of-action for carcinogenesis, which is
consistent with future directions in
26 dose-response modeling described in EPA's revised draft
Cancer Risk Assessment Guidelines
27 (EPA, 1999). While
a number of uncertainties remain regarding structure and parameters of the
28 model, the slope estimate is consistent with those derived
from humans and animals. More
29 detain on this model can be found in Part2, Chapter 8.
30
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1 5.2.2 Non Cancer Endpoints
2 At this point, sufficient data are not available
to model noncancer endpoints in humans.
3 Many studies are available to estimate ED0_ values for
noncancer endpoints in animals. However,
4
there are a number of
difficulties and uncertainties that should be considered when comparing the
5 same or different endpoints across species. Some of these include differences in
sensitivity of
6 endpoints, times of exposure, exposure routes, species and
strains, use of multiple or single
7 doses, and variability between studies even for the same
response. The estimated EDs0_ may be
8 influenced by experimental design, suggesting that caution
should be used in comparing values
9 fi'om different designs. In addition, caution should be
used when comparing studies that
10 extrapolate EDs0_ outside the experimental range. Furthermore, it may be difficult to compare
11 values across endpoints.
For example, the hm-nan health risk for a 1% change ofbody weight
12 may not be equivalent to a 1% change in enzyme
activity. Finally, background exposures
are not
13 often considered in these calculations simply because they
were not known. The latter
14 consideration is particularly important since the inclusion
of these may alter the shape of the
15 dose-response curve, possibly increasing the shape
parameter so that the responses would
16 demonstrate more threshold-like effects.[Chris- This needs
to be explained more fully.]
17 Nevertheless, given these considerations several
general trends were observed arid
18 discussed m Part 2, Chapter 8. The lowest EDs0_ tended to be for biochemical effects, followed
19 by hepatic responses, immune responses arid responses in
tissue weight. An analysis of shape
20 parm'neters implies that many dose-response curves are
consistent with linearity over the range of
21 doses tested. This
analysis does not imply that the cuD'es would be linear outside this range of
22 doses but it does inform the choices for
extrapolation. This is particularly
true when body
23 burdens or exposures at the lower end of the observed range
are close to body burdens or
24 exposures of interest for humans, wkich is the case with
dioxin-like chemicals.
25 Overall, shape parameter data suggest that biochemical
responses to TCDD are more likely to be
26 linear within the experimental dose range, while the more
complex responses are more likely to
27 assume a nonlinear shape. However, a large number (greater
than 40%) of the more complex
28 responses have shape paralneters that are more consistent
with linearity than non-linearity.
29 The tissue weight changes seen for animals (using only data sets with good or moderate
30 empirical fits to the model) yielded a median ED0_ at
average body burdens of 510 ng/kg in the
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DRAFT-- DO NOT QUOTE OR CITE May 1,2000
1
multidose studies (range; 11 to
28000 ng/kg) and a median EDm of 160 ng/kg ( range 0.0001 to
2
9700 ng/kg)in the single dose studies. Toxicity endpoints from the
single dose studies resulted in
3
a median value at average body burdens of 4300 ng/kg (range 1.3 to 1,000,000 ng/kg). For
4
tissue weight changes, 43% of the dose response curves exhibited linear
response. In contrast, the
5
toxicity endpoints fi'om the single dose studies exhibited predominantly
non-linear responses
6
(80%). All multi-dose studies demonstrated a greater degree of linear response
(41%) than did
7
single dose studies (37%), especially for tissue weight changes m_d
toxicity endpoints. (50%
8
linear for multidose versus 34% for single dose). In general, it is not
possible to dissociate the
9
differences between ca'_cer and non-cancer dose response as being due to
differences in endpoint
10
response or simply due to differences ill the length of dosing and
exposure. Also, a greater
11
percentage of the non-cancer EDs0_ were extrapolations below the lower
range of the data (42%)
12
than was the case for the cancer endpoints (8% in animals and no
extrapolations in humans).
13
14
5.3 Mode-of-Action Based Dose-Response Modeling
15 As described in Chapter 8, mechanism-based
modeling can be a powerful tool for
16
understanding and combining infom_ation on complex biological
systems. Use of a truly,
17
mechanism-based approach can, in theory, enable reliable and scientifically
sound extrapolations
18
to lower doses and between species. However, any scientific uncertainty
about the mechanisms
19
that the models describe is inevitably reflected in uncertainty about
the predictions of the models.
20
The assumptions and uncertainties involved in the mechanistic modeling
described in Chapter 8
21
are discussed at length in that Chapter and in cited publications.
22 The development and continued refinement of
physiologically based pharmacokinetic
23
(PBPK) models of the tissue dosimetry of dioxin have provided important
infom_ation
24
conceming the relationships between administered does and dose to tissue
compartments (section
25
8.2). Aspects of these models have been validated in the observable
response range for multiple
26
tissue compartments, species, and class of chemical. These models will continue to provide
27
important new information for future revisions of this health assessment
document. Such
28
information will likely include improved estimates of tissue dose for
liver and other organs
29
where toxicity has been observed, improved estimates of tissue dose(s)
in humans, and improved
30
estimates of tissue dose for dioxin related compounds.
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1 As a part of this reassessment, the development
of biologically based dose-response
2 (pharmacodynamic) models for dioxin and related compounds
has lead to considerable and
3 valuable insights regardingboth mechanisms of dioxin action
and dose-response relationships for
4 dioxin effects.
These efforts, described in some detail in Chapter 8, have provided
additional
5 perspectives on traditional methods such as the linearized
multistage procedure for estimating
6 cancer potency or tine uncertainty factor approach for
estimating levels below which noncancer
7 effects are not likely to occur. These methods have also provided a biologically based rationale
8 for what had been primarily statistical approaches. The development of models like those in
9 Chapter 8 allows for an iterative process of data
development, hypotheses testing and model
10 development.
11
12 5.4 Summary Dose-Response Characterization
13 All humans tested contain detectable body burdens
of TCDD and other dioxin-like
14 compounds that are likely to act tl'u'ough the same
mode-of-action. It is possible that any
15 additional exposure above cun'ent background body burdens
will be additive to ongoing
16 responses. The
magnitude ofthe additional response will be a function of the toxicitY
17 equivalence of the incremental exposure. This observation, the relatively small
margin of
18 exposure fol' "key events", and the high
percentage of observed linear responses suggests that a
19 proportional model should be used when extrapolating beyond
the range of the experimental
20 data. Short of
extrapolating to estimate risk in the face of uncertainties described above, a
simple
21 margin of exposure approach may be useful to
decision-makers when discussing risk
22 management goals.
However, this decision would have to be based upon a policy choice since
23 this analysis does not strongly support either choice.
24 Because human data for cancer dose-response
analysis were available and because of a
25 strong desire to stay within the range of responses
estimated by these data, the risk chosen for
26 deten'nining a point of departure was the 1% excess
risk. Doses and exposures associated
with
27 this risk (the EDs0t ) were estimated from the available
data using both mechanistic and empirical
28 models. Comparisons
were made on the basis ofbody burdens to account for differences in
29 half-life across the numerous species studied.
30 In humans, restricting the analysis to
log-linear models resulted in cancer EDs0_ ranging
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1 from 6 ng/kg to 161 ng/kg. This ,,,,,as similar to the estimates, from empirical modeling, fi'om
the
2 animal studies which ranged from 14 ng,q(g to 1190
ng/kg (most estimates were in the range
3 from 14 to 500 ng/kg), and 2.7 ng/kg for the single
mechanism-based model. Lower bounds on
4
these ED0_ estimates were used to calculate upper bound slope factors
and risk estimates for
5 average background body burdens. These estimates are presented above. Upper bound slope
6
factors allow the calculation of the probability of cancer risk for the
highly vulnerable in the
7
population (estimated to be the top 5% or greater. V_qfile there may be individuals in the
S
population who might experience a higher cancer risk on the basis of
genetic factors or other
9 determinants of
cancer risk not accounted for in epidemiologic data or animal studies, the vast
10
majority of the population is expected to have less risk per unit of
exposure and some my have
11 zero risk. Based
on these slope factor estimates (per pgTEQ/kgBW/day), average current
12
background body burdens (5 ng/kgBW) which result from average intakes of
approximately 3
13
pgTEQ/kgBW/day are in the range of 10'3 to 10-2. A very small percentage of the population
14
(< 1%) may' experience risk which are 2-3 times higher than this if
they are both the most
15
vulnerable and the most highly exposed (among the top 5%) based on
dietary intake of dioxin
16
and related compounds. This range of upper bound risk for the general
population has increased
............
17
an order of magnitude from the risk described at background exposure
levels based on EPA's
18
draft of this reassessme[lt (10'4-
10'3) (EPA, 1994).
] 9 Estimates for non-cancer endpoints showed much
greater variability, ranging over 10
20
orders of magnitude. In general,
the rloncancer endpoints displayed lower EDs0_ for short-term
21 exposures versus longer tem-_ exposures, and for simple
biochemical endpoints versus more
22
complex endpoints such as tissue
weight changes or toxicity. In
addition, the noncancer
23
endpoints generally displayed higher estimated EDs0_ than the cancer
endpoints, with most
24
estimates ranging from 100 ng/kg
to 100,000 ng/'kg. The mechanism-based models for
25
noncancer endpoints gave a lower range of EDs0_ (0.17 to 105 ng/kg).
While most of these
26
estimates were based upon a single model the estimate from the hepatic
zonal induction model
27
gave an ED0_ for CYP1A2
induction of 51 ng/kg and hence was within the same range.
28
These estimates, although highly variable, suggest that any choice of body burden, as a
29
point-of-departure, above 100 ng/kg would likely yield greater than 1%
excess risk for some
30
endpoint in humans. Also,
choosing of a point-of-departure below 1 ng/kg would likely be an
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1 extrapolation below the range of these data and would
likely represent a risk of less than 1%.
2
Any choice in the middle range of l ngfkg to 100 ng/kg, would be supported by the analyses,
3 although the data provide the greatest support in the range
of 10rig/kg to 50 n_d'kg.
4
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1 6.0 RISK
CHAILatCTERIZATION
2
3 Characterizing risks from dioxin and related
compounds requires the integration of
4
. complex data sets and the
use of science-based inferences regarding hazard, mode-of-action,
5 dose-response and exposure. It also requires consideration of incremental exposures in the
6 context of'an existing background exposure which is, for
the most part, independent of local
7 sources and dominated by exposure tl-u-ough the food
supply. Finally, this characterization
must
8 consider risks to special populations and developmental
stages (subsistence fishers, children,
9 etc.) as well as the general population. It is important that this characterization
convey the
10 current understanding of the scientific community regm'ding
these issues, highlight uncertainties
11 in this understanding, and speci_ where assumptions or
inferences have been used in the absence
12 of data. Although
characterization of risk is inherently a scientific exercise, by its nature, it
must
13 go beyond empirical observations and draw conclusions in
areas which are untested. In some
14 cases, these conclusions are, in fact, untestable given the
current capabilities in mlalytical
15 chemistry, toxicology and epidemiology. This situation should not detract from our
confidence
16 in a well
structured and documented characterization of risk but should se_'e to confirm
the
17 importance of considering risk assessment as an iterative
process which benefits from evolving
18 methods and data collection.
19
20 Dioxin and related compounds can produce a wide variety of
effects in animals and might
21 produce many of the same effects in humans.
22 There is adequate evidence based on all available
information discussed in Parts 1 and 2
23 of this reassessment, as well as that discussed in this
Integ-rated Sun'unary, to support the
24 inference that humans are likely to respond with a broad
spectrum of effects from exposure to
25 dioxin and related compounds. These effects will likely range from biochemical changes at or
26 near background levels of exposure to adverse effects with
increasing severity as body burdens
27 increase above background levels. Enzyme induction, chm'lges in hormone levels and indicators
28 of altered cellular function seen in humans and laboratory
animals represent examples of effects
29
of unlcnown clinical
significance but which may be early indicators of toxic response. Induction
30 of activating/metabolizing enzymes at or near background
levels, for instance, may be adaptive,
31 and in some cases, beneficial, or may be considered
adverse. Induction may lead to more
rapid
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DILAFT -- DO NOT QUOTE OR
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1
metabolism and elimination of potentially toxic compounds, or may lead
to increases in reactive
2
intermediates and may potentiate toxic effects. Demonstration of examples of both of these
3
situations is available in the published literature and events of this
type formed the basis for a
4
biologically based model which is discussed in Section 5. Subtle
effects, such as the impacts on
5
neurobehav_ora- outcomes, thyroid function, and liver enzymes (AST and
ALT) seen in the
6
Dutch chl!dren exposed to background levels of dioxin and related
compounds, or changes in
7
circulating reproductive hormones in men exposed to TCDD, illustrate the
types of responses
8
that support the finding of arguably adverse effects at or near
background body burdens. Clearly
9
adverse effects including, perhaps, cancer may not be detectable until exposures contribute to
10
body burdens which exceed back_ound by one or two orders of magnitude
(10 or 100 times).
11
The mechanistic relationships of biochemical and cellular changes seen
at or near background
12
body burden levels to production of adverse effects detectible at higher
levels remains uncertain
13
but data are accumulating to suggest mode-of-action hypotheses for
further testing.
14 It is well known that individual species vary in
their sensitivity to any particular dioxin
15
effect. However, the evidence
available to date indicates that humans most likely fall in the
16
middle of the range of sensitivity for individual effects among animals
rather than at either
17
extreme. In other words,
evaluation of the available data suggests that humm'm, in general, are
la
_cithc_ extrcnx:ly sensitive nor insensitive to the individual effects
of dioxin-like compounds.
19
Human data provide direct or indirect support for evaluation of likely
effect levels for several of
20
the endpoints discussed in the reassessment although the influence of
variability among hm'nans
21
remains difficult to assess.
Discussions have highlighted certain prominent, biologically
22
significant effects of TCDD and related compounds. In TCDD-exposed men, subtle changes in
23
biochemistry and physiology such as enzyme induction, altered levels of
circulating reproductive
24
hormones, or reduced glucose tolerar_ce and, perhaps, diabetes, have
been detected in a limited
25
number of epidemiologic studies.
These findings, coupled with knowledge derived from animal
26
experiments, suggest the potential for adverse impacts on human
metabolism, and developmental
27
and/or reproductive biology, and, perhaps, other effects in the range of
current human exposures.
28
These biochemical, cellular, and
organ-level endpoints have been shown to be affected by
29
TCDD, but specific data on these endpoints do not generally exist for
other congeners. Despite
30
this lack of congener specific data, there is reason to infer that these
effects may occur for all
31
dioxin-like compounds, based on the concept of
toxicity equivalence.
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1 In this volume, dioxin and related compounds are
characterized as carcinogenic,
2 developmental, reproductive, immunological and
endocrinological hazards. The deduction
that
3 humans are likely to respond with noncancer effects from
exposure to dioxin-like compounds is
4 based on the fundamental level at which these compounds
impact cellular regulation and the
5 broad range of species which have proven to respond with
adverse effects. Since, for example,
6 developmental toxicity following exposure to TCDD-iike
congeners occurs in fish, birds, and
7 mammals, it is likely to occur at some level in
humans. It is not currently possible to
state
8 exactly how or at
what levels individuals will respond with specific adverse impacts on
9 development or reproductive function, but analysis of the
Dutch cohort data and laboratory
10 animal studies
suggest that some effects may occur at or near background levels. Fortunately,
11 there have been few human cohorts identified with TCDD
exposures high enough to raise body
12 burdens significantly over background levels (See Table 5-
1 and Figure 5-1 in Section 5), and
13 when these cohorts have been examined, relatively few
clinically significant effects were
14 detected. 'Fine
lack of exposure gradients, adequate human information and the focus of most
15 currently available epidemiologic studies on
occupationally, TCDD-exposed adult males makes
16 evaluation of the inference, that noncancer effects
associated with exposure to dioxin-like
17 compounds may be occurring, difficult. It is important to
note, however, that when exposures to
18 very high levels of dioxin-like compounds have been
studied, such as in the Yusho and Yu- ·
19 Cheng cohorts, a spectrLtm of adverse effects have been
detected in men, women and children.
2[) Some ha_ c argued that to deduce that a spectrum of
noncancer effects will occur in humans in
21 the absence of better human data overstates the science;
most scientists involved in the
22 reassessment as authors and reviewers have indicated that
such inference is reasonable given the
23 weight-of-the-evidence fi'om available data. As presented, this logical conclusion represents a
24 testable hypothesis which may be evaluated by further data
collection. The EPA, its Federal
25 colleagues and others in tine general scientific community
are continuing to fill critical data gaps
26 which will reduce our uncertainty regarding both hazard and
risk characterization for dioxin and
27 related compounds.
28
29 Dioxin and related compounds are structurally related and
elicit their effects through a
30 common mode-of-action.
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1 The scientific community has identified and
described a series of common biological
2
steps that are necessary for most if not all of the observed effects of
dioxin and related
3
compounds in vertebrates including humans. Binding of dioxin-like compounds to a cellular
4
protein called the "Ah receptor" represents the first step
in a series of events attributable to
5
exposure to dioxin-)ike compounds including biochemical, cellular and
tissue-level changes in
6
normal biological processes.
Binding to the .Kh receptor appears to be necessary, for all well-
7
studied effects of dioxin but is not sufficient, in and ofitself, to
elicit these responses. There
8
remains some uncertainty as to whether every dioxin response is Ah
receptor-mediated. Sensitive
9
biological tools like the aryl hydrocarbon receptor deficient (Ahr'/-)
mice indicate a small residual
10
of effects to exposure to TCDD which does not allow us to rule out
altemative pathways which
11 are receptor-independent. The well documented effects
elicited by exposure of animals, and in,
12
some cases, humans, to 2,3,7,8-TCDD are shared by other chemicals which
have a similar
13
structure and Ah receptor binding characteristics. In the last five years, significant data has
14
accumulated which supports the concept of toxicity equivalence, a
concept which is at the heart
15
of risk assessment for the complex mixtures of dioxin and related
compounds which are
16
encountered in the envirom'nent. These data have been analyzed and
summarized in Part 2,
17
Chapter 9. This Chapter has been added to the EPA's dioxin re-assessment
effort to address
18
questions raised by the Agency's Science Advisory Board (SAB) in
1995. The SAB suggested
19
that, since the TEQ approach was a critical component of risk assessment
for dioxin and related
20
compounds, the Agency should be explicit in its description of the
history and application of the
21 process and go beyond reliance on the Agency's published
reference documents on the subject
22
(EPA 1987, 1989).
23
24
EPA and The Intemational Scientific Community have Adopted Toxicity
Equivalence of
25
Dio×in and Related Compounds as Prudent Science Policy.
26 Dioxin and related compounds always exist in
nature as complex mixtures. As
discussed
27
in the Exposure Document, these complex mixtures can be characterized
through analytic
28
methods to determine concentrations of individual congeners. Dioxin and related compounds
29
can be quantified and biological activity of the mixture can be
estimated using relative potency
30
values and an assumption of dose additivity. Such an approach has evolved over time to form
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DRAFT-- DO NOT QUOTE OR
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1
the basis for the use of"toxicitv equivalence" (TEQ) in risk assessment
for this group of
2
compounds. While such an approach is dependent on critical assumptions
and scientific
3
judgement, it has been characterized as a "useful, interim"
way to deal with the complex mixture
4
problem and has been accepted by numerous countries and several
intemational organizations.
5
Altemative approaches, including the assumption that all congeners carry
the toxicity
6
equivalence of 2,3,7,8-TCDD, or that all congeners other than
2,3,7,8-TCDD can be ignored,
7
have been generally rejected as inadequate for risk assessment purposes.
8 Significant additional literature is now
available on the subject of toxicity equivalence of
9
dioxin and related compounds and Chapter 9 provides the reader with a
summary which is up to
10
date through 1999. A recent
intemational evaluation of all of the available data ( van den Berg et
11
al., t 998) has re-affirmed the TEQ approach and has provided the
scientific community with the
12
latest values for toxicity equivalence factors (TEFs) for PCDDs, PCDFs
and dioxin-like PCBs.
13
Consequently, we can infer with greater confidence that humans will
respond to the cumulative
14
exposure of Ah receptor-mediated chemicals. The position taken in this
Reassessment is that
15
these 1998 TEFs should be adopted for use by the Agency. Future research will be needed to
16
address remaining uncertainties inherent in the current approach. The WHO has suggested that
17
the TEQ scheme be re-evaluated on a periodic basis and that TEFs and
their application to risk
18
assessment be re-analysed to account for emerging scientific
information.
19
20
Complex Mixtures of Dioxin and Related Compounds are Highly Potent,
"Likely"
21
Carcinogens.
22 With regard to carcinogenicity, a weight-of-the-evidence evaluation suggests
that
23
mixtures of dioxin and related compounds (CDDs, CDFs, and dioxin-like
PCBs) are strong
24
cancer promoters, weak direct or indirect initiators and likely to
present a cancer hazard to
25
humans. Since dioxin and related
compounds always occur in the environment and in humans as
26
complex mixtures of individual congeners, it is appropriate that the
characterization apply to the
27
mixture. According to the
Agency's revised draft Cancer Guidelines, the descriptor, likely, is
28
appropriate when the available tumor effects and other key data are
adequate to demonstrate
29
carcinogenic potential to humans.
Adequate data are recognized to span a wide range. The data
30
for complex mixtures of dioxin and related compounds represents a case
which, according to the
31
draft Guidelines, would approach the strong evidence end of the adequate
data spectrmn.
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i Ep_demiologic obse_'ations of an association between
exposure and cancer responses (TCDD),
2
unequivocal positive
responses in both sexes, multiple species and different routes in lifetime
3
bioassays or initiation-promotion protocols or other shorter-term in
vivo systems e.g. transgenic
4
models CI'CDD plus numerous PCDDs, PCDFs, dioxin-like PCBs), and
mechmfistic or mode-of
5
action data that are assumed to be relevant to human carcinogenicity
(PCDDs, PCDFs, dioxin-
6
like PCBs) all support the description ofcomplex mixtures of dioxin and
related compounds as
7
likely human carcinogens.
8 g,,_ile the data base from cancer epidemiologic
studies remains controversial, it is the
9
view of this reassessment that this body of evidence is supported by the
laboratory data
10
indicating that TCDD probably increases cancer mortality of several
types. Although not all
11
confounders were ruled out ill any one study, positive associations
between surrogates of dioxin
12
exposure, cithcr length of occupational exposure or proximity to a known
source combined with
13
some information based on measured blood levels, and cancer have been
reported. These data
14
suggest a role for dioxin exposure to contribute to a carcinogenic
response but do not confirm a
15
causal relationship between exposure to dioxin and increased cancer
incidence, Available human
16
studies alone cannot demonstrate whether a cause and effect relationship
between dioxin
17
exposure and increased incidence of cancer exists. Therefore, evaluation of cancer hazard in
18
humans must include an evaluation of all of the available animal and in
vitro data as well as the
19
data fi'om exposed human populations.
20 As discussed earlier in Section 2.2.1.4, under
EPA's current approach, individual
21
congeners can also be characterized as to their carcinogenic
hazard. TCDD is best characterized
22
as "carcinogenic to humans."
This means that, based on the weight of all of the evidence
23
(human, animal, mode-of-action), TCDD meets the criteria that allows EPA
and the scientific
24
community to accept a causal relationship between TCDD exposure and
cancer hazard. The
25
guidance suggests that "carcinogenic to humans" is an
appropriate descriptor of human
26
carcinogenic potential when there is an absence of conclusive
epidemiologic evidence to clearly
27
establish a cause and effect relationship between human exposure and
cancer, but there is
28
compelIing carcinogenicity in animals and mechanistic information in
animals and humans
29
den-tonstrating similar modes of carcmogerfic action· The "carcinogenic to humans"
descriptor is
30
suggested for TCDD since all of the following conditions are met:
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1
2 there is evidence from occupational epidemiologic
studies for an association between
3 TCDD exposure and increases in cancer at all
sites combined, in lung cancer, and,
4 perhaps, at other sites, but the data are
insufficient on their own to demonstrate a causal
5 association;
6
7 there is extensive carcinogenicity ill both sexes
of multiple species at multiple sites;
8
9
- there is general
agreement that the mode ofTCDD's carcinogenicity is Ah receptor
10 dependent and proceeds tkrough modification of
the action of a number of receptor and
11 honnone systems involved in cell growth and
differentiation such as the epidermal
12 growth factor receptor and estrogen receptor; and
13
14 key events such as equivalent body burdens in
animals mad in human populations
15 expressing an association between exposure to
TCDD and cancer, mid the determination
16 of active
,ah receptor and dioxin responsive elements in the general human population.
17 There is no reason to believe that these events
would not occur in the occupational
18 cohorts studied.
19
20
Other individual
dioxin-like compounds are characterized as "likely" human carcinogens
21
primarily because of the lack of epidemiological evidence associated
with their carcinogenicity
22
although the inference based on toxicity equivalence is strong that they
would behave in humans
23
as TCDD does. Other factors,
such as the lack of congener specific ckronic bioassays also
24
support this characterization.
For each congener, the degree of certainty is dependent on the
25
available congener specific data and its consistency With the
generalized mode-of-action which
26
underpins toxicity equivalence for TCDD and related compounds. Based on this logic, complex
27
enviroru'nental mixtures of TCDD and dioxin-like compounds should be
characterized as "likely"
28
carcinogens, with the degree of certainty of the characterization being
dependent on the
29
constituents of the n-fixture, when known. For instance, the hazard potential, although "likely,"
30
would be characterized differently for a mixture whose TEQ was dominated
by OCDD as
31
compared to one which was dominated by pentaCDF.
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While uncertainties remain
regarding quantitative estimates of upper bound cancel' risk
From dioxin and related
compounds, efforts of this reassessment to bring more data into the
evaluation of cancer potency have
resulted irl evaluation of the slope of the dose response curve
at the low end of the observed
range (using the LED00 using a simple proportional (linear) model
and a calculation of both upper
bound risk and margm of exposure (MOE) based on human
equivalent background exposures
and associated body burdens. Evaluation
of shape parameters
(which is used to estimate degree
of linearity or non-linearity of dose-response within the range
of observation) for biochemical
effects indicates that many of these biochemical effects can be
9
hypothesized to be to key events in a generalized, dioxin mode-of action
model. These analyses
0
do not argtte for significant departures from linearity below a
calculated ED0_ for endpoints
11
potentially related to cancer response, for at least one to two orders
ofmagnitude lower exposure.
12
Risk estimates for intakes associated with background body burdens or
incremental exposures
13
based on this slope factor represent a plausible upper bound on risk
based on the evaluation of
14
animal and human data. The slope
factors based on the most sensitive cancer responses, both
15
animal and human, calculated in Section 5, fall in a range of 5 X 10'3
to 5 X 10'4 per
16
pgTEQ/kgBW/day. The ranges of
estimates of upper bound cancer potency calculated from the
17
human and animal data analyzed in Part 2, Chapter8 overlap and the range
above is bounded on
18
the upper end by the estimate of slope fi'om the Hamburg cohort
epidemiology study and on the
19
lower end by' the estimate fi'om the re-analyzed Kociba study. Consequently, the Agency,
20
although fully recognizing this range and the public health conservative
nature of the slope
21
factors that make up the range, suggests the use of 5 X I0-3 per
pgTEQ/kgBW/day as an
22
estimator of upper bound cancer risk for both background intakes and
incremental intakes above
23
background. Slope factors allow
the calculation of the probability ofcancer risk for the highly'
24
vulnerable in the population (estimated to be the top 5% or
zeater). Xx,q_ile there may be
25
individuals in the population who might experience a higher cancer risk
on the basis of genetic
26
factors or other deten'ninams of cancer risk not accounted for in
epidemiologic data or animal
27
studies, the vast majority of the population is expected to have less
risk per unit of exposm'e and
28
some my have zero risk. Based on
these slope factor estimates (per pgTEQ/kgBW/day), average
29
cun'ent background body bm'dens (5 ng/kgBW) which result from average
intakes of
30
approximately 3 pgTEQ/kgBW/day are
in the range of 10'* to
10'2. A very small percentage of
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I
the population (< 1%) may
experience risk which are 2-3 times higher than this if they are both
'> the most vulnerable and the most highly exposed (among the
top 5%) based on dietary intake of
3
dioxin and related compounds. This range of upper bound risk for the
general population has
4
increased all order of magnitude from the risk described at background
exposure levels based on
5
EPA's draft of this reassessment (10.t_ 10-3) (EPA, 1994).
6
Despite the use of the epidemiology data to describe an upper bound on
cancer risk, the
7
Peer Panel that met in Septeinber 1993 to review an earlier draft of the
cancer epidemiology
8
chapter suggested that the epidemiology data alone were still not
adequate to implicate dioxin
9
and related compounds as "known" human carcinogens but that
the results from the human
10
studies were largely consistent with observations from laboratory studies
of dioxin-induced
11
cancer and, therefore, should not be dismissed or ignored. Other scientists, including those who
12
attended the Peer Pa3'_eI meeting, felt either more or less strongly
about the weight of the
13
evidence from cancer epidemiology studies, representing the range of
opinion that still exists on
14
the interpretation of the these studies. Similar opinions were expressed
in the comments
15
documented in the SAB's report in
1995 (SAB,1995). More recently,
the kltemational Agency
16
for Research on Cancer, in its re-evaluation of the cancer hazard of
dioxin and related
17
compounds (1997), found that the while the. epidemiologic data base for
2,3,7,8-TCDD was still
18
"hmited," the overall weight of the evidence was sufficient to
characterize 2,3,7,8-TCDD as a
19
Category 1, "known" human carcinogen. Other related members of the class of
dioxin-like
20
compounds were considered to have "inadequate" epidemiologic
data to factor into hazard
21
categorization. A similar
classification has been proposed within the context of the Department
22
of Health and Human Services' Report on Carcinogens. They too base their characterization on
23
the broad base of human, arlimal and mode- of-action information in
humans and animals which
24
support this conclusior_. Therefore, given that 2,3,7,8-TCDD is
contained in complex mixtures of
25
dioxin and related compounds, and that the TEQ approach has been adopted
as a reasonable
26
approach to assessing risks of these complex mixtures, it is also
reasonable to apply estimates of
27
upper bound cancer potency derived fi'om epidemiology studies where
2,3,7,8-TCDD was
28
associated with excess cancer risk to complex mixtures of dioxin and
related compounds.
29 The current evidence suggests that both receptor
binding and most early biochemical
30
events such as enzyme reduction are likely to demonstrate low-dose
lineariW. The mechanistic
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1
relationship of these early events to the complex process of
carcinogenesis remains to be
2
established. If these findings
imply low-dose linearity in biologically-based cancer models under
3
development, then the probability of cancer risk will be linearly
related to exposure to TCDD at
4
low doses. Until the mechanistic
relationship between early cellular responses and the
5
parameters m biologically based cancer models is better understood, the
shape of the dose-
6
response curve for cancer in tine below the range of observation can
only be inferred with
7
uncertainty. Associations
between exposure to dioxin and certain types of cancer have been
8
noted in occupational cohorts with average body burdens of TCDD
approximately 1- 3 orders of
9
mag'nitude (10 - 1000 times) higher than average TCDD body burdens in
the general population.
10
The average body burden in these occupational cohorts level is within
1-2 orders of magnitude
11
(10-100 times) of average background body burdens in the general
population iii terms of TEQ
12
(See Table 5-1 and Figure 5-1).
Thus, there is no need for large scale low dose extrapolations in
13
order to evaluate background intakes and body burdens, and little, if
any data to suggest large
14
departures fi'om linearity in this somewhat narrow window between the
lower end of the range of
l 5
observation and the range of general population back_ound
exposures. Nonetheless, the
16
relationsMp of apparent increases in cancer mortality in these worker
populations to calculations
17
of general population risk remains a source of uncertainty.
18 TCDD has been clearly shown to increase malignant
tumor incidence in laboratory
19
animals. In addition, a number
of studies analyzed in this reassessment demonstrate other
20
biological effects of dioxin related to the process of
carcinogenesis. Initial attempts to
construct
21
a biologically-based model for certain dioxin effects as described in
this reassessment will need
22
to be continued and expanded to accommodate more of the available
biology and to 'apply to a
23
broader range of potential health effects associated with exposure to
dioxin-like compounds.
24
25
Use a "Margin-of Exposure Approach" to Evaluate Risk for
Noncancer and Some Cancer
26
Endpoints.
27 The likelihood that noncancer effects may be
occurring in the human population at
28
environmental exposure levels is often evaluated using a "margin of
exposure" (MOE) approach.
29
The Agency has used this approach for a number of years in its
assessment of the safety of
30
pesticides. This concept has
also been incorporated into the revised Cancer Risk Assessment
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guidelines. A
MOE is calculated by dividing a "point of departure" for
extrapolation purposes at
the Iow end of the range
ofobser_'ation in human or animal studies (the human-equivalent animal
LOAEL, no observed
adverse effect level (NOAEL), benchmark dose (BMD) or effective dose
(EDxx)) by the human
exposure or body burden level of interest. Generally speaking, when
considering either
background exposures or incremental exposures plus background, MOEs in
, range of 100 -1000 are considered adequate to rule out the
likelihood of significant effects
? occurring in humans based on sensitive animal responses orresults from epidemiologic studies.
8 The adequacy of the MOE to be protective of health takes
into account the nature of the effect at
9 the "point of departm'e," the slope of the dose
response curve, the adequacy of the overall
10 database,
interindividual variability in the human population, as well as other factors.
11 Considering MOEs based on incremental exposures alone
divided by the human exposure of
12 interest, is not considered to give an accurate portrayal
of the implications of that exposure unless
13 background exposures are insignificant. This is not the case for background
exposures for dioxin
14 and related compounds based on total TEQ, The average levels of background intake and
15 associated body burdens of dioxin-like compounds in ten'ns
of TEQs in the general population
16 would be well within a l_ctor of 100 of human-equivalent
exposure levels associated with
17 NOELS, LOAELs, BMDs, or ED0_ values in laboratory animals
exposed to TCDD or TCDD
l 8 equivalents. These estimates, although highly variable,
suggest that any choice of body burden,
19 as a pomt-o!-departure, above 100 ng/kg would likely yield
greater than 1% excess risk for some
20 endpoint in humans.
Also, choosing ora point-6f-departure below 1 ng/kg would likely be an
21 extrapolation below the range of these data and would
likely represent a risk of less than 1%.
22 Any choice in the middle range of lng/kg to 100 rig/kg,
would be supported by the analyses,
23 although the data provide the greatest support in the range
of l 0ng/kg to 50 rig/kg. In many
24 cases, the MOE compared to background using these endpoints
is a factor of 10 or less (See Part
25 2, Chapter 8).
Because of the relatively high background compared to effect levels, the
Agency
26 is not recommending the derivation of a reference dose
(RfD) for dioxin and related compounds.
27 Although RfD's are often useful because they represent a
health risk goal below which there is
28 likely to be no
appreciable risk ofnoncancer effects over a lifetime of exposure, their primm'y
use
29 is to evaluate increments of exposure from specific sources
when background exposures are low
30 and insignificant.
Any RfD that the Agency would recommend using the traditional approach
for
I32
DRAFT-- DO NOT QUOTE OR CITE May I, 2000
1
setting an RfD is likely to be 2-3 orders of magnitude (100-1000) below
current background
2
intakes and body burdens. Since
exceeding the RfD is not a statement ofrisk, discussion of an
3 RfD for an incremental exposure when the RfD has already
been exceeded by average
4
background exposures is meaningless.
5 When evaluating incremental exposures associated
with specific sources, knowing the
6
increment relative to
background may help to urtderstand the impact of the incremental exposure.
7
For instance, it would be misleading to suggest that an incremental
exposure of
8
0.001 pgTEQ/kg/day was below the
RfD if"background" exposures were already at or above
9
that level. On the other hand, as part of the total, the increment
represents less than a 0.1%
10
increase over average "background" and we estimate that
individuals within the 50-95% range of
11 exposure within
the population may be 2-3 times (200%-300%) higher. This has led us to
12
suggest that perhaps the best infonnation for a decision-maker to have
is 1) a characterization of
13
average "background" exposures; 2) a characterization of the
percent increase over background
14
of individuals or subpopulations of interest; and 3) a policy statement
about when increases over
15
average "background" become significant for the dec/sion. This is not easy because one could
16
argue that, given high "background", any addition, if it is
widespread, is too much. On the other
17
)land, someone else could argue that a 10% increase in incremental
exposure for a small
18
population around a specific point source would be well within the
general population exposures
19
and would not constitute a disproportionate exposure or risk. In this case, the strategy might be
20
too bring average "background" exposures down and to focus on
large incremental exposures or
21
highly susceptible populations. This would be a strategy that would
parallel the Agency's lead
22
strategy. Other parallel issues
between dioxin-like compounds and lead are under discussion
23
within the Agency.
24 ATSDR (1999) set a minimal risk level (MRL),
which is defined similarly to the EPA's
25
RtT), for dioxin and related compounds of 1.0 pgTEQ/kgBW/day. Some of the data regarding
26
lower bounds on the
EDs0.' from various noncancer effects
call that MRL into question. WHO
27
(2000) has set a tolerable daily intake of 1-4 pgTEQ/kgBW/day and has
indicated that, although
28
current exposures in that range are "tolerable" (a risk
mmmgement decision rather than a risk
29
assessment), efforts should be made to ultimately reduce intake
levels. Findings in this
30
reassessment appear to be supportive of that recommendation.
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DRAFT-- DO NOT QUOTE OR
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1
2
Children's risk from exposure to dioxin and related compounds may be
increased, but
3
more data are needed to full)' address this issue.
4 The issue of children's risk fi'om exposure to
dioxin-like compounds has been addressed
5
in a number of sections tlzroughout this reassessment. Data suggest a sensitivity of response in
6
both humans and animals durirlg the developmental period, both
prenatally and postnatally.
7
However, data are limited. Since
evaluation of the impacts of early exposures on both children's
8
health and health later in life is important to a complete
characterization of risk, collection of
9
additional data ill this area should be a high priority to reduce
uncertainties in future risk
10
assessments.
11 Data from the Dutch cohort of children exposed to
PCBs and dioxin-like compounds
12
suggest impacts of exposure to background levels of dioxin and related
compounds prenatally
13
and, perhaps, postnatally on neurobehavioral outcomes, thyroid function,
and liver enzymes
14
(AST and ALT). While these
effects can not be attributed solely to dioxin and related
15
compounds, several associations suggest that these are, in fact, likely
to be Ah-mediated effects.
16
,An investigation of background dioxin exposure and tooth development
was done in Finnish
17
children as a result of studies of dental effects in dioxin-exposed
rats, mice, and nonhuman
18
primates, and in PCB-exposed
children. The Finnish investigators
examined enamel
19
hypomineralization of permanent first molars in 6-7 year old
children. The length of time which
20
infants breast fed was not significantly associated with either
mineralization changes, or with
21
TEQ levels in the breast milk.
However, when the levels and length of breast feeding were
22
combined in an overall score, a statistically si_ificant association was
observed ® = 0.3,
23
p = 0.003, regression analysis).
24 In addition, effects have been seen where
sig-nificantly elevated exposure occurred.
The
25
incidents at Yusho and Yu-Cheng resulted in increased perinatal
mortality and low birthweight in
26
infants bom to women who had been exposed. Rocker bottom heal was observed in Yusho
27
infants, and functional abnon'nalities have been reported in Yu-Cheng
children. The similarity of
28
effects observed in l'mman infants prenatally exposed to the complex
mixture in Yusho and
29
Yu-Cheng with those reported in adult monkeys exposed only to TCDD
suggests that at least
30
some of the effects children are due to the TCDD-Iike congeners in the
contaminated rice oil
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1 ingested by the mothers of these clzildren. The similar responses include a clustering
of effects in
2 organs derived from the ectodermal germ layer, referred to
as ectodermal dysplasia, including
3 effects on the skin, nails, and Meibomian glands;
developmental and psychomotor delay during
4 developmental az_d cognitive tests. Some investigators believe that, because all
of these effects
5 ill the Yusho and Yu-Cheng cohorts do not correlate with TEQ,
some of the effects are
6 exclusively due to nondioxin-tike PCBs or a combination of
all the congeners. Ill addition, on
7 the basis of these data, it is still not clear to what
extent there is an association between overt
8 matemal toxicity and embryo/fetal toxicity in humans.
FmXher studies in the offspring, as well as
9 follow-up to the Seveso incident may shed further light on
this issue. In addition to chloracne
10 and acute responses to TCDD exposure seen in Seveso
children, elevated levels of serum GGT
11 have been observed within a year after exposure in some of
the more highly exposed Seveso
12 children Long-telTn
pathologic consequences of elevated GGT have not been illustrated by
13 excess mortality fi'om liver disorders or cancer, or in
excess morbidity but further follow-up is
14 needed. It must be
recognized that the absence of an effect thus far does not obviate the
15 possibility that the enzyme levels may have increased
concurrent to the exposure but declined
16 after cessation.
The apparently transient elevations in- 'Ai]T'levels among the Seveso
children
17
suggest that hepatic enzyme
levels other than GGT may react in this manner to 2,3,7,8-TCDD
18 exposure.
19 Impacts on thyroid hormones provides an example
of an effect of elevated, postnatal
20 exposure to dioxin and related compounds. Several studies of nursing infants suggest
that
21 ingestion of breast milk with a higher dioxin TEQ may
alter thyroid function. Thyroid hormones
22 play important roles in the developing nervous system in of
all vertebrates species, including
23 humans. In fact,
thyroid hormones are considered so important in development that in the U.S.
24 all infants are tested for hypothyroidism shortly after
birth. Results from the studies
mentioned
25 above suggest a possible shift in the population
distribution of thyroid hormone levels,
26 particularly T4, and point out the need for collection of
longitudinal data to assess the potential
27 for long-term effects associated with developmental
exposures. The exact processes accounting
28 for these observations in humans are unknown, but when put
in perspective of animal responses,
29 tlne following might apply: dioxin increases the metabolism
and excretion of thyroid hormone,
30 mainly T4, in the liver.
Reduced T4 levels stimulate the pituitary to secrete more TSH, which
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I et_lances thyroid hormone production. Early in the disruption process, the body
can
2 overcompensate for the loss ofT4, which may result in a
small excess of circulating T4 in
3 response to the increased TSH. In animals, given higher doses of dioxin, the body is unable to
4 maintain homeostasis, and TSH levels remain elevated and T4
levels decrease.
5 A large number of studies in animals have
addressed the question of effects of dioxin-like
6 chemicals after itl utero or lactational exposure. These have included both single congener
7 studics _tnd exposures to complex mixtures. However, the vast majority of the data, are
derived
8 from studies of 2,3,7,8-TCDD, or single congeners (e.g.
PCB 77) or commercial mixtures of
9 PCBs. Exposure
patterns have included single doses to the dams as well as dosing on multiple
10 days during gestation begim'fing as early as the first day
of gestation. These studies are
discussed
11 in detail in Part 2, Chapter 5. The observed toxic effects include developmental toxicity,
12 neurobehavioral and neurochemical alterations, endocrine effects and developmental
13 immunotoxicity.
For instance, results of this body of work suggests that 2,3,7,8-TCDD
clearly
14 has the potential to produce alterations in male
reproductive function (rats and hamsters) and
15 male sexual behavior (rats) after prenatal exposure. In addition, impacts on neuromotor and
16 cognitive behavior as well as development of the immune
system have been indicated in a
17 number of studies.
18 No epidemiological data and limited animal data
are available to address the question of
19 the potential impact of exposure to dioxin-like compounds
on childhood cancers or oil cancers of
20 later life. Given the relative impact of nursing on body
burdens (See the discussion of breast milk
21 exposures and body burdens below), direct impacts of
increased early post-natal exposure on the
22 carcinogenic process is expected to be small. This conclusion is based on the reasonable
23 assumptions that cancer risk is a function of average
lifetime body burden or that, because dioxin
24 is a potent cancer promoter rather than a direct initiator
of the cancer process, exposures later in
25 life might be more important than those received
earlier. However, recent studies of
Brown et al.
26 (1998) suggest that prenatal exposure of rats to dioxin and
related compounds may indirectly
27 enhance their sensitivity as adults to chemical
carcinogenesis from other chemical carcinogens.
28 Further work is needed to evaluate this issue.
29 In addition to the potential vulnerability during
development, the fetus, infants and
30 children a'e exposed to dioxins through several
routes. The fetus is exposed in utero
to levels of
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I
dioxin and related compounds
which reflect the body burden of the mother.
It is important to
2 recognize that it is not the individual meals which a
pre.ant woman eats during pregnancy
3 which might affect development but the consequence of her
exposure history, over her life, which
4 has the greatest impact on her body burden. Again, good nutrition which represents diet
with
5 appropriate levels of fat has consequences on dietary
intake and consequent body burdens of
6 dioxin and related compounds. Nursing infants represent a special ease where, for a limited
7 portion of their lives, ii,rants may have elevated
exposures on a body weight basis when
8 compared to non-nursing infants and adults (See discussion
below). In addition to breast milk
9 exposures, intakes of CDD/Fs and dioxin-like PCBs are over
tlwee times higher for a young child
10 as compared to that of an adult, on a body weight
basis. Table 4-9 in Section 4 of this
document
11 describes the variability in average intake values as a
function of age using age-specific food
12 consumption rate and average food concentrations, as was
done for adult intake estimates.
13 However, like the case for the nursing infant discussed
below, the differences in body burden
14 between children and adults are expected to be much less
thar_ the differences in daily intake.
15
Assuming that body burden
is the relevaa'n dose metric for most, if not all, effects, there is some
16 assurance that these increased intake levels will have
limited additional impact on risk as
17 compared to overall lifetime exposure.
18
19 Background Exposures to Dioxin and Related Compounds need
to be Considered when
20
Evaluating both Hazard and Risk.
21 The term "background" exposure has been
used tI_oughout this reassessment to describe
22
exposure of the general population, who are not exposed to readily
identifiable point sources of
23
dioxin-like compounds. Adult
daily intakes of CDD/CDFs and dioxin-like PCBs are estimated
24
to average 45 and 25 pg TEQDFp-WHO98/day, respectively, for a total
intake of 70 p_day
25
TEQDF_,-vGq:Ouv Daily intake is
estimated by combining exposure media concentrations (food,
(
26 soil, air) with contact rates
(ingestion, inhalation), Table 4-8
below summarizes the intake rates
27
derived by this method. The
intake estimate is supported by an extensive database on food
28
consumption rates and food data.
Phan'nacokinetic (PK) modeling provides further support for
29
the intake estimates. Current
adult tissue levels reflect intakes from past exposure levels which
30 are thought to be
higher than current levels (see Trends Section 2.6).
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May 1, 2000
1 CDD/CDF and dioxin-like PCB intakes for the
general population may extend to levels at
2 least three times higher ti'mn the mean. Variability in general population exposure
is primarily
3 the result in the differences in dietau choices that
individuals make. These are differences
in
4 both quantity and types of food consumed. A diet which is disproportionately high in
animal fats
5 will result in an increased background exposure over the
mean. Data on variability of fat
6 consumption indicate that the 95'h percenthe is about twice
the mean and the 99th percenthe is
7 approximately 3 times the meaa'_. Additionally, a diet which substitutes meat sources that are low
8 in dioxin (i e. beef, pork or poultry) with sources that
are high in dioxin (i.e. fresh water fish)
9 could result in exposures elevated over tltree times the mean. This scenario may not represent a
10 significant change in total animal fat consumption, even
though it results in an increased dioxin
11 exposure. Intakes
of CDD/Fs and dioxin-like PCBs are over tl'u'ee times higher for a young child
12 as compared to that of an adult, on a body weight
basis. Using age-specific food
consumption
13 rate and average food concentrations, as was done above for
adult intake estimates, Table 4-9
14 describes the
variability in average intake values as a function of age.
15 The average CDD/CDF tissue level for the general
adult U.S. population appears to be
16 declining and the best estimate of current (late 1990s)
levels is 25 ppt (TEQDFp-WHO98, lipid
17 basis). The tissue
samples collected in North America in the late 1980s and early 1990s showed
18 an average TEQt)Fp WHO98 level of about 55
pg/g lipid. This finding is supported
by a number of
19 studies which measured dioxin levels in adipose, blood and
human milk, all conducted in North
20 America. The number
of people in most of these studies, however, is relatively small and the
21
participants were not
statistically selected in ways that assure their representativeness of the
22 general U.S. adult population. One study, the 1987 National Human Adipose Tissue Survey
23 (NHATS), involved over 800 individuals and provided broad
geographic coverage, but did not
24 address coplanar PCBs.
Similar tissue levels of these compounds have been measured in Europe
25 and Japan during similar time periods.
26 Because dioxin levels in the environment have
been declining since the 1970s (see trends
27 discussion), it is reasonable to expect that levels in
food, human intake and ultimately human
28 tissue have also declined over this period. The changes in tissue levels are likely to
lag the
29 decline seen in environmental levels and the changes in
tissue levels cannot be assumed to occur
30 proportionally with declines in enviromnental levels. ATSDR (1999) summarized levels of
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1 CDDs, CDFs and PCBs in human blood collected during the
time period I995 to 1997. The
2 individuals smnpled were all U.S. residents with no known
exposures to dioxin other than normal
3 background. The
blood was collected from 400 individuals in seven different locations with an
4 age range of 20 to 70 years. All TEQ calculations were made assuming nondetects were equal to
5 half the detection limit.
While these samples were not collected in a manner that can be
6 considered
statistically representative of the national population and lack wide
geographic
7 coverage, they are judged to provide a better indication of
current tissue levels in the US than the
8 earlier data (see Table 4-7 ). PCBs 105, I18, and 156 are missing from the blood data
for the
9 comparison populations reported in the Calcasieu stud
(ATSDR, 1999). These congeners
10 account for 62% of the total PCB TEQ estimated in the early
1990's. Assuming that the missing
I I congeners fi-om the Calcasieu study data contribute the
same proportion to the total PCB TEQ as
12 in earlier data, they would increase our estimate of
cun'ent body burdens by another 3.7 pgTEQ/g
13 lipid .tot a total
PCB TEQ of' 5.9 pg/g lipid and a total DFP TEQ of 25 pg/g lipid.
14 Past background exposure of about 3
pgTEQ/kgBW/day leads to body burdens in the
15 human population which currently average approximately
5ng/kg (20-30 pg TEQ/g lipid) when
16 all dioxins, furans and PCBs are included and have been
higher in the past. DeVito et al.
(1995)
17 estimated that body burdens averaged 9-13 ng&g based on
intake values of 4-6 pg TEQ/kg/day
18 and blood levels of 40-60 pgTEQ/g lipid, based on data from
the late 1980%. If the general
19 population were exposed to dioxins and related compounds at
the current level of intake
20 (approximately 1 pg TEQ/kg/day) for a lifetime, average
steady state body burdens would be less
21 than 2 ng/kg and blood levels would be 7-8 pg TEQ/g
lipid. These estimates are based on the
22 assumption of 50% absorption of dioxin-like compounds from
the diet, Using the same
23
. assumption used for intake values, high-end estimates of body burden
of individuals in the
24 general population (approximately the top 5% of the general
population) may be greater than 2
25 times higher than these average estimates. This is based on
data for dietary fat consumption and
26 the assumption that body burdens of dioxin and related
compounds in the general population are
27 associated with their fat consumption. The top 1% is likely
to be 3 times higher based on their
28 intake of fat.
29 Characterizing national background levels of
dioxins in tissues is uncertain because the
30 current data cannot be considered statistically
representative of the general population.
It is also
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1
complicated by the fact that tissue levels are a function of both age
and birth year. Because
2
intake levels have var/ed over time, the accumulation of dioxins in a
person who tumed 50 years
3
old in 1990 is different than in a person who tumed 50 ill 2000. Future studies should help
4
address these uncertainties. The
National Health and Nutrition Examination Survey Ox!I'-IANrES)
5
began a new national survey in 1999'which will measure dioxin blood
levels in about 1700
6
people per year (see http:www.cdc.gov/nchs/nhanes.htm). The survey is conducted at 15
7
alii'ti:rent locations per year and is designed to select individuals
statistically representative ofthe
8
civilian U.S. population in terms of age, race and ethnicity. These new data should provide a
9
much better basis for estimating national background tissue levels and
evaluating trends than the
10
currently available data.
11 As described above, cun'ent intake levels from
food sources m'e estimated in this
12
Reassessment to be approximately 1
pgTEQ/KgBW/day. Certain segments
of the population
13
may be exposed to additional increments of exposure by being in
proximity to point sources or
14
because of dietary practices.
These will be described below.
I5
16
Evaluation of Exposure of "Special" Populations and
Developmental Stages is Critical to
17
Risk Characterization.
18 As discussed above, background exposures to
dioxin-like compounds may extend to
19
levels at Ieast three times higher thart the mean. This upper range is assumed to result from
the
20
non'nal variability of diet and human behaviors. Exposures from local elevated sources or
21
exposures resulting from unique diets would be in addition to this
background variability. Such
22
elevated exposures may occur in small segments of the population such as
individuals living near
23
discrete local sources, or subsistence or recreational fishers. Nursing infants represent a special
24
case where, for a limited portion of their lives, these individuals may
have elevated exposures on
25
a body weight basis when compared to non-nursing infants and
adults. This exposure will be
26
discussed in a separate section below.
27 Dioxin contamination incidents involving the commercial
food supply have occurred in
28
the U.S. and other countries.
For example, in the U.S., contaminated ball clay was used as an
29
anticaking agent in soybean meal and resulted in elevated dioxin levels
in some poultry and
30 catfish. This incident involved less than
5% of the national poultry production and has since been
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eliminated. Elevated dioxin levels have also been
obse_'ed in a few beef and dairy animals
where the contamination was
associated with contact with pentachlorophenol treated wood.
Evidence of this kind of
elevated exposure was not detected in the national beef survey.
Consequently its occurrence
is likely to be low, but it has not been determined. These incidents
3 may have led to small increases in dioxin exposure to the
general population. However, it is
5 unlikely that such incidents have led to disproportior_ate
exposures to populations living near
7 where these incidents have occulted, since, in the U.S.,
meat and dairy products are highly
8 distributed on a national scale. If contamination events were to occur in foods that are
9 predominantly distributed on a local or regional scale,
then such events could lead to highly
10 exposed local populations
I I Elevated exposures associated with the workplace
or industrial accidents have also been
12 documented. U.S.
workers in certain segments of the chemical industry had elevated levels of
13 TCDD exposure, with some tissue measurements in the
thousands ofppt TCDD. There is no
14 clear evidence that elevated exposures are currently
occurring among U.S. workers. Documented
15
examples of past exposures for
other groups include certain Air Force personnel exposed to
16 Agent Orange during the Vietnam War and people exposed as a
result of industrial accidents in
17 Europe and Asia.
18 Consumption of unusually high amounts of fish, meat, or
dairy products containing
19 elevated levels of dioxins and dioxin-like PCBs can lead to
elevated exposures in comparison to
20 the general population.
Most people eat some fish from multiple sources, both fresh, and salt
21 water. The typical
dioxin concentrations in these fish and the typical rates of consumption are
22 included in the mean background calculation of exposure. People who consume large quantities
23 of' fish at typical contamination levels may have elevated
exposures since the concentration of
24 dioxin-like compounds in fish are generally higher than in
other animal food products. These
25 kinds of exposures are addressed within the estimates of
variability of background and are not
26 considered to result in highly exposed populations. If high-end consumers obtain their fish from
27 areas where the concentration of dioxin-like chemicals in
the fish is elevated, they may constitute
28 a highly exposed subpopulation. Although this scenario seems reasonable, no supporting data
29 could be found for such a highly exposed subpopulation in
the U.S. One study measuring dioxin-
30 like compounds in blood of sports fishers in the Great
Lakes area showed elevations over mean
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DRAFT-- DO NOT QUOTE OR CITE May 1, 2000
I
background, but within the range of normal variability. Elevated CDD/CDF levels in human
2
blood have been measured in Baltic fishermen. Similarly elevated levels ofcoplanar PCBs have
3
been measured in the blood of fishers on the north shore of the Gulf of
the St. Lawrence River
4
who consume large amounts of seafood.
5 Similm'ly, high exposures to dioxin-Iike
chemicals as a result of consuming meat and
6
dairy products would only occur in situations where individuals consume
large quantities of
7
these foods and the level of these compounds is elevated. Most people eat meat and dairy
8
products fi'om multiple sources and, even if large quantities are
consumed, they are not likely, to
9
have unusually high exposures.
Individuals who raise their own livestock for basic subsistence
10
have the potential for higher exposures if local levels of dioxin-like
compounds are high. One
11
study in the U.S. showed elevated levels in chicken eggs near a
contaminated soil site. European
12
studies at several sites have shown elevated CDD/CDF levels in milk and
other animal products
13
near combustion sources.
14 In smnmary, in addition to general population
exposure, some individuals or groups of
15
individuals may also be exposed to dioxin-like compounds from discrete
sources or pathways
I6
locally within their enviromnent.
Examples of these "special" exposures include: contamination
17
incidents, occupational exposures, direct or indirect exposure to local
populations from discrete
18
sources, or exposures to subsistence or recreational fishers.
19
20
Breast feeding infants have higher intakes of dioxin and related
compounds for a short but
21
developmentally important part of th eir life, but benefits of breast
feeding are widely
22
recognized to outweigh the risks.
23 A number of studies have measured levels of the
dioxin-like compounds in human breast
24
milk, yielding an average of 35 ppi TEQD_F,-WHOw Several studies around the world suggest
25
that these levels are declining.
Whether the perceived decline is based on declines in current
26
intake levels in women whose breast milk levels would otherwise have
been higher, or whether it
27
reflects younger women :vith lower historical exposures entering the
breast feeding population,
28
or both, is not known.
Additional studies which take age, reproductive history and dioxin
intake
29
levels into account are needed.
30 Based on a six month nursing scenario, the
average daily intake for an infant is about 100
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DRAFT-- DO NOT QUOTE OR CITE May 1,2000
1 times higher than the adult daily intake on a body weight
basis: the adult intake is 1 pg TEQDrp-
2 WHOg_/kg-d, while the infant intake while breast feeding
would be 100 pg TEQDFp-WHOgg/kg-d.
3 The differences in body burden between nursing infants and
adults are expected to be much less
4 than the differences in daily intake, and rapidly approach
steady state at close to adult levels.
5 This is because of rapid growth of the infant and an
in:.reasing total amount of body fat. Because
6 of rapid equilibration throughout the body, tissue doses
will better reflect doses to target tissues
7 than will intake.
This is important if windows of vulnerability occur during this period
of
8 development. On a
mass basis, however, the cumulative dose to the infant under this scenario is
9 about 9% of the lifetime retake.
I0 The American Academy of Pediatrics (1997) has
made a compelling argument for the
1 l diverse advantages of breast-feeding and the use of human
milk for infant feeding to infants,
12 mother, families and society. These include health, nutritional, inununologic, developmental,
13 psychological, social, economic, and environmental
benefits. Breast milk is the point of
14 comparison for all infant food and the breast-fed infant is
the reference for evaluation of all
15 altemative feeding methods. In addition, increasing the rates of breast-feeding initiation is
a
16 national health objective and one of the goals of the U.S.
Govemment's Healthy People 2010.
17 The World Health Organization (1988) maintained that the
evidence did notsupport an alteration
18 of WHO recommendations which promote and support breast
feeding. A more recent
19 consultation in 1998 (WHO, 2000) reiterated these
conclusions. While it is important that
the
20 recommendations of these groups continue to be re-evaluated
in light of emerging scientific
21 infon'nation, the Agency does not believe that finding
contained in this report provides a
22 scientific basis for initiating such a re-evaluation. This conclusion is based on the fact that
23 stronger data have been presented that body burden, not
intake, is the best dose metric; that many
24 of the noncancer effects, particularly those seen in
children, are more strongly associated with
25 pre-natal exposure and the mother's body burden rather than
post-natal exposures and breast milk
26 levels; and that dioxin-like compounds are strong promoters
of carcinogenicity, a mode-of-
27 action
that depends on late stage impacts rather than early stage impacts on the
carcinogenic
28 process.
29
30
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DRAFT -- DO NOT QUOTE OR CITE May 1, 2000
Many Dioxin sources have been
identified and emissions to the environment are being
reduced.
Currer_t emissions of
CDDs/CDFs/PCBs to the U.S. environment result principally from
anthropogenic activities. Evidence which supports this fnding include:
matches in time of rise
3
of environmental levels with time when general industrial activity began
rising rapidly (see trend
6
discussion in Section 4.6), lack of any identified large natural sources
and observations of higher
7
CDD/CDF/PCB body burdens in industrialized vs. less industrialized
countries (see discussion
8
on human tissue levels in Section 4.4).
9 The principal identified sources of environmental
release may be grouped into five major
10
types: Combustion and Incineration Sources; Chemical
Manufacturing/Processing Sources;
11
Industrial/Municipal Processes; Biological and Photochemical Processes;
and Reservoir Sources.
12
Development of release estimates is difficult because only a few
facilities in most industrial
13
sectors have been tested for CDD/CDF emissions. Thus an extrapolation is needed to estimate
14
national emissions. The
extrapolation method involves deriving an estimate of emissions per
15
unit of activity at the tested facilities and multiplying this by the
total activity level in the
16
untested facilities. In order to
convey the level of uncertainty in both the measure of activity and
l 7
the emission factor, EPA developed a qualitative confidence rating
scheme. The confidence
18
rating scheme, presented in Section 4, Table 4-1, uses qualitative
criteria to assign a high,
19
medium, or low confidence rating to the emission factor re'Id activity
level for those source
20
categories for which emission estimates can be reliably quantified. The Dioxin Reassessment
21
has produced an Inventory of source releases for the U.S. (Table
4-2). The Inventory was
22
developed by considering all sources identified in the published
literature and numerous
23
individual emissions test reports. The Inventory is limited to sources
whose releases can be
24
reliably quantified (i.e. those with confidence ratings orA, B or C as
defined above). Also, it is
25
limited to sources with releases that are created essentially
simultaneously with formation. This
26
means that the reservoir sources are not included. The Inventory presents the environmental
27
releases in terms of two reference years: I987 and 1995. EPA's best estimates ofreleases of
28
CDD/CDFs to air, water and land from reasonably quantifiable sources
were approximately
29
2,800 gram (g) (l.3 pounds) TEQo_-WH09_ in 1995 versus 13,500 g (6
pounds) TEQr_F-WHOos
30 in 1987. The decrease in estimated
releases of CDD/CDFs between 1987 and 1995
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(approximately 80%) was due
primarily to reductions in air emissions from municipal and
medical waste incinerators. The environmental releases of CDD/Fs in the
U.S. occur from a
wide variety of sources, but are
dominated by releases to the air from combustion sources.
Insufficient data are available to
comprehensively estimate point source releases of dioxin-like
compounds to water. Sound estimates of releases to water are only
available for chlorine
bleached pulp and paper mills and
manufacture of ethylene dichlor/de/vinyl chloride monomer.
The contribution of dioxin-like
compounds to waterways from nonpoint source reservoirs is
I
Iii<ely to be greater than the contributions from point sources. Current data are only sufficient to
:_
support preliminat'y estimates of nonpoint source contributions of
dioxin-like compounds to
0
water (i.e., urban storm water runoffand rural soil erosion). These estimates suggest that, on a
I
nationwide basis, total nonpoint releases are significantly larger than
point source releases.
12
Other releases to water bodies that camot be quantified on the basis of
existing data include
13
effluents from POTWs and most industrial/commercial sources. Based on the available
14
information, the inventory includes only a limited set of activities
that result in direct
15
environmental releases to land. The
only releases to land quantified in the inventory are land
16
application of sewage sludge and pulp and paper mill wastewater
sludges. Not included in the
17
Inventory's definition of an environmental release is _he'T-d_posal of
sludges and ash into
18
approved landfills. While this
inventory is the most comprehensive and well-documented in the
19 world, the inventory is likely to underestimate total
releases. The magnitude of the
20
underestimate is unl_lown but it is unlikely that non-combustion sources
today ,other than
21
reservoir sources, play a dominant role in human exposure. In terms of 1995 releases from
22
reasonably quantifiable sources, this document estimates releases of
2800 g WHO98TEQDF for
23
contemporary formation sources and 2900 g WHO0_TEQDF for reservoir
sources. In addition,
24
there remains a number of unquantifiable and poorly quantified sources
which are described in
25
Section 4.
26
As described above,
combustion appears to be the most significant process of formation
27
of CDDs/CDDFs today. Important
factors which can affect the rate of dioxin formation include
28
the overall combustion efficiency, post combustion flue gas temperatures
and residence times,
29
and the availability of surface catalytic sites to support dioxin
synthesis. Although chlorine is an
30
essential component for the formation of CDD/'Fs in combustion systems,
the empirical evidence
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DRAFT -- DO NOT QUOTE OR CITE May 1, 2000
1
indicates that for commercial scale incinerators, chlorine levels in
feed are not the dominant
2
controlling factor for rates of CDD/F stack emissions. The conclusion that chlorine in feed is not
3
a strong determinant of dioxin emissions applies to the overall
population of commercial scale
4
combustors. For any individual
commercial scale combustor, circumstances may exist in which
5
changes in chlorine content of feed could affect dioxin emissions. For uncontrolled combustion,
6
such as open burning of house-hold waste, chlohne content of wastes may
play a more
7
significant role in affecting levels of dioxin emissions than observed
in commercial scale
8
combustors.
9 No significant release of newly fom_ed
dioxin-like PCBs is occulting in the U.S.
Unlike
10
CDD/CDFs, PCBs were intentionally manufactured in the U.S. in large
quantities from 1929
11
unlil production was banned in
1977. Although it has been
demonstrated that small quantities of
12
coplanar PCBs can be produced during waste combustion, no strong evidence exists that the
13
dioxin-like PCBs make a significant contribution to TEQ releases during
combustion. The
14
occurrences of dioxin-like
PCBs in the U.S. environment most likely reflects past releases
15
associated with PCB production, use and disposal. Further support of this finding is based on
16
observations of reductions since 1980s in PCBs in Great Lakes sediment
and other areas.
17 It is unlikely that the emission rates of
CDD/CDFs fi'om known sources correlate
18
proportionally with general population exposures. Although the emissions inventory shows the
19 relative
contribution of various sources to total emissions, it cannot be assumed that
these
20
sources make the same relative contributions to human exposure. It is
quite possible that the
21
major sources of dioxin in food (see discussion in Section 2.6
indicating that the diet is the
'_'_
dominant exposure pathway for humans) may not be those sources that
represent the largest
23
fractions of total emissions in the U.S. The geographic locations of sources relative to the areas
24
fi'om which much off the beef, pork, milk, and fish come, is important
to consider. That is, much
25
of the agricultural areas which produce dieta_ animal fats are not
located near or directly down
26
wind of the major sources of dioxin and related compounds.
27 The contribution of reservoir sources to human
exposure may be significant. Several
28
factors support this finding.
First, human exposure to the dioxin-like PCBs is thought to be
29
derived almost completely fi'om reservoir sources. Since one third of general population TEQ
30
exposure is due to PCBs, at least one third of the overall risk from
dioxin-Iike compounds comes
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from reservoir sources. Second, CDD/CDF
releases from soil via soil erosion and nmoffto
waterways appear to be greater
than releases to water fi.om the primary sources included in the
inventory. CDD/CDFs in waterways can bioaccumulate in
fish leading to human exposure via
consumption of fish which makes
up about one third of the total general population CDD/CDF
TEQ exposure. This suggests that a significant portion of
the CDD/CDF TEQ exposure could be
due to releases fi'om the soil
reservoir. Finally, soil reservoirs
could have vapor and particulate
releases wlfich deposit on plants
and entel' the terrestrial food chain.
The magnitude of this
;
contribution, however, is unknown.
TI'ds assessment adopts the hypothesis that the primary
mechanism by which dioxin-like
0
compounds enter the terrestrial food chain is via atmospheric
deposition. Dioxin and related
I
compounds enter the atmosphere directly through air emissions or
indirectly, for example,
12
tl_ough volatilization from land or water or from re-suspension of
particles. Once introduced
13
into the enviro_m'_ent, dioxin-like compounds are widely distributed in
the enviromnent as a
14
result of a number of physical and biological processes. The dioxin-like compounds are
15
essentially insoluble in water, generally classified as semi-volathe and
tend to bioaccumulate in
16
animals, Some evidence has shown
that these compounds can degrade in the enviromnent, but in
17
general they are considered very persistent and relatively immobile in.
soils and sediments. These
18
compounds are transported through the atmosphere as vapors or attached
to air-bome particulates
19
and can be deposited on soils, plants, or other surfaces (by wet or dry
deposition). The dioxin-
20
like compounds enter water bodies primarily via direct deposition from
the atmosphere, or by
21
surface runoff and erosion. From
soils, these compounds can reenter the atmosphere either as
22
resuspended soil particles or as vapors. In water, they can be resuspended into the water colunm
23
from sediments, volatilized out of the surface waters into the
atmosphere or become buried in
24
deeper sediments. Immobile
sedim,ents appear to serve as permanent sinks for the dioxin-like
25
compounds. Though not always
considered an environmental compartment, these compom_ds
26
are also found in anthropogenic materials (such as pentachlorophenol)
and have the potential to
27
be released from these materials into the broader environment.
28 The two primary pathways for the dioxin-like compounds
to enter the ecological food
29
chains and human diet are: air-to-plant-to-animal and
water/sediment-to-fish. Vegetation
30
receives these compounds via atmospheric deposition in the vapor and
particle phases. The
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compounds are retained on
plant surfaces and bioaccumulated in the fatty tissues of animals that
feed on these plants. Vapor phase transfers onto vegetation have
been experimentally shown to
dominate the air-to-plant
pathway for the dioxin-like compounds, particularly for the lower
chlorinated congeners. In the aquatic food chain, dioxins enter
water systems via direct
discharge or deposition
m'_d n.moff from watersheds. Fish
accumulate these compounds through
their direct contact with water, suspended particles,
bottom sediments and through the
consumption of aquatic
organisms. Although these two pathways
are thought to normally
dominate contribution to
the commercial food supply, others can also be important. Elevated
dioxin levels ill cattle
resulting fi'om animal contact with pentacholorophenol treated wood have
) been documented by the USDA. Animal feed contamination episodes have led to elevations of
I dioxins in poultu in the United States, mill< in
Germany, and meat/dairy products in Belgium.
,2 Deposition can occur directly onto soil or onto
plant surfaces. At present, it is unclear
13 whether atmospheric deposition represents primarily current
contributions of dioxin and related
14 compounds fi'om all media reaching the atmosphere or
whether it is past emissions of dioxin and
15 related compounds which persist and recycle in the
environment. Understanding the
relationship
16 between these two scenarios will be particularly important
in understanding the relative
17 contributions of individual point sources of these
compounds to the food chain and assessing the
18 eftbctiveness of control strategies focused on either
current or past emissions of dioxins in
19 attempting to reduce the levels in food.
20 As
discussed in Section 4.3, estimates for background levels of dioxin-like
compounds in
21 environmental media are based on a variety of studies
conducted at different locations in North
22 Ame_'ica. Ofthe
studies available for this compilation, only those conducted in locations
23 representing "background" were selected. The amount and representativeness of the data varies,
24 but in general these data lack the statistical basis to
establish true national means. The
25 environmental media concentrations were consistent among
the various studies, mid were
26 consistent with similar studies in Westem Europe. These data are the best available for
27 comparing site
specific values to national backm'ound levels.
Because of the limited number of
28 locations exan_ined, however, it is not known if these
ranges adequately capture the full national
29 variability, if sigalificant regional variability exists
making national means of limited utility, or if
30 elevated levels above this range could still be the result
of background contanlination processes.
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As new data are collected these
ranges are likely to be expanded and refined.
The limited data on
dioxin-like PCBs ia envirorm'_ental media are sammarized in
the document (Partl, Volume III,
Chapter 4), but were not judged
adequate for estimating background levels.
Concentrations of
CDD/CDFs and PCBs in the U.S. envirolunent were consistently low
prior to the 1930s. Then, concentrations rose steadily until
about 1970. At this time, the trend
reversed and the concentrations
have declined to the present. The most
compelling supportive
evidence of this trend for the
CDD/CDFs aa_d PCBs comes from dated sediment core studies.
Sediment concentrations in these
studies are generally assumed to be an indicator of the rate of
,_
atmospheric deposition. CDD/CDF
and PCB concentrations in sediments began to increase
O
around the 1930s, and continued to increase until about 1970. Decreases began in 1970 and have
1
continued to the time of the most recent sediment samples (about
1990). Sediment data from 20
12
U.S. lakes and rivers from seven separate research efforts consistently
support this trend.
13
Additionallyl sediment studies in lakes located in several European
countries have shown similar
14
trends.
15 It is
reasonable to assume that sediment core trends should be driven by a similar
trend in
1O
emissions to the envirorunent.
The period of increase generally matches the time when a variety
17
of industrial activities began r/sing and the period of decline appears
to correspond with _-owth
18
in pollution abatement. Many of
these abatement efforts should have resulted in decreases in
19
dioxin emissions, i.e. elimination of most open burning, particulate
controls on combustors,
20
phase out of leaded gas, and bans on PCBs, 2,4,5-T, hexachlorophene, and
restrictions on use of
21
pentachlorophenol. Also, the
national source inventory of this assessment documented a
22
significant decline in emissions from the late 1980s to the
mid-1990s. Further evidence cfa
23
decline in CDD/CDF levels in recent years is emerging from data,
primarily from Europe,
24
showing declines in foods and human tissues.
25
In addition, as to the
congener-specific PCB data discussed earlier, a wealth of data on
26
total PCBs and Aroclor mixtures exist which also supports these
trends. It is reasonable to
27
assume that the trends for dioxin-like PCBs are similar to those for
PCBs as a class because the
28
predominant source of dioxin-like PCBs is the general production of PCBs
in Aroclor mixtures.
29
PCBs were intentionally marmfactm'ed in large quantities from 1929 until
production was banned
30
in the U.S. in 1977. U.S.
production peaked in 1970, with a volume of 39,000 metric tons.
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Further support is derived
from data showing declining levels of total PCBs in Great Lakes
: sediments and biota during the 1970s and 1980s. These studies indicate, however, that during
the 1990s the decline is
slowing and may be leveling off.
4 Because dioxin-like chemicals are persistent and
accumulate in biological tissues,
5 particularly in animals, the major route of human exposure
is through ingestion of foods
6 containing minute quantities (part per trillion or ppi
levels) of dioxin-like compounds. This
7 results in wide-spread, low-level exposure of the general
population to dioxin-like compounds.
8 The issue of general population, background exposm'e is
discussed above.
9
10 Risk Characterization Summary Statement
11 Based on all of the data reviewed in this
reassessment and scientific inference, a picture
12 emerges of TCDD and related compounds as potent toxicants
in animals with the potential to
13 produce a spectrum of effects. Some of these effects may be occurring in humans at general
14 population background levels and may be resulting in adverse impacts on human health. The
15 potency and fundamental
level at which these compounds act on biological systems is analogous
leo to several well studied l'lormones. Dioxin and related compounds have the
ability to alter the
17 pattern of growth and differentiation ora number of
cellular targets by initiating a series of
18 biochemical and biological events resulting in the
potential for a spectrum of cancer and non-
19 cancer responses in animals and humans. Despite this potential, there is currently
no clear
20 indication of increased disease in the general population
attributable to dioxin-like compounds.
21 The lack of a clear indication of disease in the general
population should not be considered
22 strong evidence for no effect of exposure to dioxin-like
compounds, Rather lack of a clear
23 indication of disease may be a result of the inability of
our cmTent data and scientific tools to
24 directly detect effects at these levels of human
exposure. Several factors suggest a
need to
25 further evaluate the impact of these chemicals on humans at
or near current background levels.
26 These are: the weight of the evidence on exposm'e and
effects; an apparently low margin-of-
27 exposure/hr noncancer effects; and potential for
significant risks to some portion of the
general
28 population and additivity to background processes related
to carcinogenicity in the case of
29 incremental exposures above background.
30
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7.0 REFERENCES
151