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DRAFT
DO NOT QUOTE OR CITE
Revised May 2, 1994
Internal Review Draft
Chapter 9. Risk Characterization
NOTICE
THIS DOCUMENT IS A PRELIMINARY DRAFT. It has not been formally released by
the U.S. Environmental Protection Agency and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its technical accuracy and
policy implications.
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, D.C.
DRAFT--DO NOT QUOTE OR CITE
DISCLAIMER
This document is an interal draft for review purposes only and does not constitute
Agency policy. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
ii
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
2
Chapter 9
3
4
RISK CHARACTERIZATION OF DIOXIN AND RELATED COMPOUNDS
5
6
Introduction
7
Chlorinated dibenzo-p-dioxins and related compounds (commonly known
8
simply as dioxins) are environmental contaminants present in a variety of
9
environmental media. This class of compounds has caused great concern in the
10 general public as well as intense interest in the scientific community. Much of the
11 public concern revolves around the characterization of these compounds as among
12 the most potent "man-made" toxicants ever studied. Indeed, these compounds are
13 extremely potent in producing a variety of effects in experimental animals based on
14 traditional toxicology studies at levels hundreds or thousands of times lower than most
15 chemicals of environmental interest. In addition, human studies demonstrate that
16 exposure to dioxin and related compounds is associated with subtle biochemical and
17
biological changes and with chloracne, a serious skin condition associated with these
18 and similar organic chemicals Laboratory studies suggest the probability that
19 exposure to dioxin-like compounds may be associated with other serious health
20 effects including cancer. Human data are supportive of these concerns. Recent
21 laboratory studies have provided new insights into the mechanisms involved in the
22 impact of dioxins on various cells and tissues and, ultimately, on toxicity. Dioxins have
23 been demonstrated to be potent modulators of cellular growth and differentiation,
24 particularly in epithelial tissues. These data coupled with assumptions and inferences
25
regarding extrapolation from experimental animals to humans and from high doses to
26
low doses allow a characterization of dioxin hazards.
27
This chapter presents a risk characterization for dioxin and related compounds.
28
In the risk characterization, key findings pertinent to understanding the hazards and
29 risks of dioxin and related compounds are described and integrated. All of the
30
available information is considered in proposing hypotheses or in reaching
31
conclusions. The risk characterization is not meant to be an executive summary of the
32
extensive data base which has been analyzed in detail in preceeding chapters and in
1
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
the exposure document. Risk characterization requires a discussion of likely routes,
2
patterns and levels of exposure as well as aspects of hazard and dose response.
3
Information contained in the document entitled, Estimating Exposure to 2,3,7,8-
4
tetrachlorodibenzo-p-dioxin and Related Compounds (EPA,1994), hereafter referred
5
to as the Exposure Document, will be integrated with the health effects information on
6
this class of compounds found in previous chapters of this assessment. The risk
7
characterization contains an articulation of the strengths and weaknesses of the
8
available evidence, as well as clearly presenting assumptions made and inferences
9
used. Risk is characterized in both qualitative and quantitative terms, as appropriate.
10
Finally, overall conclusions regarding the health risks of dioxin and related
11
compounds are presented.
12
The process for development of this risk characterization of dioxin and related
13
compounds has been an open and participatory one. The health assessment and
14 exposure documents which provide the basis for this characterization have been
15 developed in collaboration with scientists from within and from outside of the Federal
16
government. Each of these has undergone extensive internal and external review
17
including review at a meeting of experts after a first draft was completed. Additional
18
input to this characterization comes from comments on those draft chapters as well as
19 from the panel of experts who met in September, 1992. This panel was asked to
20 provide their perspective on themes to be carried into the characterization and their
21
contributions are reflected here. Finally the characterization, as presented here,
22 represents review and comment by both those Federal scientists involved in
23
development of the health assessment and exposure chapters as well as
24
representatives of other Federal agencies. However, the views expressed in this
25
characterization are those of the collective authors and, as a draft undergoing public
26
comment and further external review, no Agency-level endorsement should be
27
inferred at this time.
28
Once fully peer reviewed and revised accordingly, this risk characterization is
29
meant to provide a balanced picture of the scientific findings of the health and
30
exposure assessments for use by risk managers in selecting risk management
31
options. As an integrated presentation of a complex data base, it is meant to answer
32
key questions concerning the science behind concerns for dioxins and should be
2
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
useful in developing strategies for risk communication.
2
3
CHEMICAL STRUCTURE AND PROPERTIES
4
Polychlorinated dibenzodioxins (PCDDs), polychlorinated dibenzofurans
5 (PCDFs), and polychlorinated biphenyls (PCBs) are chemically classified as
6 halogenated aromatic hydrocarbons (HAH). The chlorinated and brominated
7 dibenzodioxins and dibenzofurans are tricyclic aromatic compounds with similar
8 physical and chemical properties, and both classes are similar structurally. Certain of
9 the PCBs (the so-called co-planar or mono-ortho co-planar congeners) are also
10 structurally and conformationally similar. The most widely studied of these compounds
11
is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). This compound, often called simply
12 dioxin, represents the reference compound for this class of compounds. The structure
13
of TCDD and several related compounds is shown in Figure 9-1.
14
For purposes of this document, dioxin-like compounds are defined to include
15
the subset of this class of compounds which are generally agreed to produce dioxin-
16 like toxicity. These compounds are assigned individual Toxicity Equivalency Factor
17 (TEF) values as defined by international convention (EPA, 1989) in proportion to their
18 toxicity relative to TCDD. Results of in vitro and in vivo laboratory studies contribute to
19 the assignment of a relative toxicity value. All chlorinated dibenzodioxins (CDDs) and
20 chlorinated dibenzofurans (CDFs) with chlorines substituted in the 2,3,7, and 8
21 positions are assigned TEF values. Additionally, the analogous brominated dioxins
22 and furans (BDDs and BDFs) and certain polychlorinated biphenyls (PCBs) have
23 recently been identified as having dioxin-like toxicity and thus are also included in the
24 definition of dioxin-like compounds. Generally accepted TEF values are shown in
25 Table 9-1. A recent World Health Organization/International Program on Chemical
26
Safety meeting (Dec.,1993) held in the Netherlands considered the need to derive
27
internationally acceptable interim TEFs for the dioxin-like PCBs. Recommendations
28
arising from that meeting of experts suggest that in general only a few of the dioxin-like
29 PCBs are likely to be significant contributors to general population exposures to dioxin-
30
like compounds. It is estimated that these dioxin-like PCBs may be responsible for
31
approximately 1/4 to1/3 of the total toxicity equivalence associated with general
32
population environmental exposures to this class of related compounds. Both the
3
Fig. 9-1
Dioxin and Similar Compounds - Chemical Structure
o
CI
CI
CI
CI
CI
CI
CI
CI
O
O
2,3,7,8-Tetrachlorodibenzo-p-dioxin
2,3,7,8-Tetrachlorodibenzofuran
CI
O
CI
CI
CI
CI
CI
CI
CI
CI
o
O
CI
1,2,3,7,8-Pentaachlorodibenzo-p-dioxin
2,3,4,7,8-Pentachlorodibenzofuran
CI
CI
CI
CI
PAGE 3-a 3 a
CI
CI
CI
CI
CI
CI
CI
3,3',4,4',5-Pentachlorobiphenyl
3,3',4,4',5,5'-Hexachlorobiphenyl
Table 9-1. Toxicity Equivalency Factors (TEF) for CDDs and CDFs.
Compound
TEF
Mono-, Di-, and Tri-CDDs
0
2,3,7,8-TCDD
1
Other TCDDs
0
2,3,7,8-PeCDD
0.5
Other PeCDDs
0
2,3,7,8-HxCDD
0.1
Other HxCDDs
0
2,3,7,8-HpCDD
0.01
Other HpCDDs
0
OCDD
0.001
Mono-, Di-, and Tri-CDFs
0
2,3,7,8-TCDF
0.1
Other TCDFs
0
1,2,3,7,8-PeCDF
0.05
2,3,4,7,8-PeCDF
0.5
Other PeCDFs
0
2,3,7,8-HxCDF
0.1
Other HxCDFs
0
2,3,7,8-HpCDF
0.01
Other HpCDFs
0
OCDF
0.001
Source: EPA, 1989.
PAGE 3-6
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
refinement of the toxicity equivalence factors for dioxin-like PCB congeners (DeVito et
2
al., 1993) as well as a compilation and analysis of all of the available data on relative
3 toxicities of dioxin-like PCBs with respect to a number of endpoints (Ahlborg et al.,
4
1994) support these findings. Although these findings have been published recently,
5
additional review and data collection will be needed . The panel specifically
6 recommended that theseTEFs be used as intake values and urged caution in their use
7 with regard to toxicity equivalence in body burden measurements. In addition, the
8 panel urged investigation of companion TEFs for ecotoxicological use, based on data
9
from ecotoxicological studies.
10
There are 75 possible individual compounds comprising the CDDs depending
11
on the positioning of the chlorine(s) and 135 different CDFs. These are called
12
individual congeners. Likewise, there are 75 possible different positional congeners
13
of BDDs and 135 different congeners of BDFs (see Exposure Document, Table 2-1).
14
Only 7 of the 75 possible congeners of CDDs or of BDDs are thought to have dioxin-
15
like toxicity; these are ones with chlorine/bromine substitutions in, at least, the 2,3,7,8-
16 positions. Only 10 of the 135 possible congeners of CDFs or of BDFs are thought to
17 have dioxin-like toxicity; these also are ones with substitutions in the 2,3,7,8 -positions.
18
While this suggests 34 individual CDDs, CDFs, BDDs or BDFs with dioxin-like toxicity,
19 inclusion of the mixed chloro/bromo congeners substantially increases the number of
20 possible congeners with dioxin-like activity. There are 209 possible PCB congeners.
21
Only 13 of the 209 possible congeners are thought to have dioxin-like toxicity, these
22 are ones with 4 or more chlorines with just 1 or no substitutions in the ortho position.
23 These compounds are sometimes referred to as coplanar since they can assume a flat
24 configuration with rings in the same plane. Similarly configured polybrominated
25 biphenyls are likely to have similar properties although the data base on these
26 compounds with regard to dioxin-like activity has been less extensively evaluated.
27 Mixed chlorinated and brominated congeners will increase the number of compounds
28 considered dioxin-like. The physical/chemical properties of each congener vary
29 according to the degree and position of chlorine and/or bromine substitution.
30
In general these compounds have very low water solubility, high octanol-water
31
partition coefficients, low vapor pressure and tend to bioaccumulate. Volume II of the
32
Exposure Document presents congener specific values for water solubility, vapor
4
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
pressure, partition coefficients and photo quantum yields and discusses other physico-
2
chemical characteristics of the chlorinated dioxins and dibenzofurans. These physico-
3 chemical properties result in the environmental fate and transport discussed below.
4
Expanded discussions will be required in future documents to account for dioxin-like
5
PCB's and for brominated or mixed halogenated congeners.
6
7
ENVIRONMENTAL FATE
8
Despite a growing body of literature from laboratory, field, and monitoring
9 studies examining the environmental fate and environmental distribution of CDDs and
10 CDFs, the fate of these environmentally ubiquitous compounds is not yet fully
11 understood and the following represents our best understanding, based on available
12 data. In soil, sediment, the water column, and probably air, CDDs/CDFs are primarily
13 associated with particulate and organic matter because of their high lipophilicity and
14 low water solubility. They exhibit little potential for significant leaching or volatilization
15 once sorbed to particulate matter. The available evidence indicates that CDDs and
16 CDFs, particularly the tetra- and higher chlorinated congeners, are extremely stable
17 compounds under most environmental conditions, with environmental persistence
18 measured in decades. The only environmentally significant transformation process for
19 these congeners is believed to be photodegradation of chemicals not bound to
20 particles in the gaseous phase or at the soil- or water-air interface. Brominated
21 congeners are significantly more readily transformed by photodegradation.
22 CDDs/CDFs entering the atmosphere are removed either by photodegradation or by
23 dry or wet deposition. Burial in-place or erosion of soil to water bodies appears to be
24 the predominant fate of CDDs/CDFs sorbed to soil. CDDs/CDFs entering the water
25 column primarily undergo sedimentation and burial. The ultimate environmental sink
26
of CDDs/CDFs is believed to be aquatic sediments.
27
Little specific information exists on the environmental transport and fate of the
28
dioxin-like PCBs. However, the available information on the physical/chemical
29 properties of dioxin-like PCBs coupled with the body of information available on the
30
widespread occurrence and persistence of PCBs in the environment indicates that
31
these PCBs are likely to be associated primarily with soils and sediments, and to be
32
thermally and chemically stable. Soil erosion and sediment transport in water bodies
5
Draft . Do Not Quote or Cite - Draft
May 2, 1994
1
and emissions to the air (via volatilization, dust resuspension, or point source
2
emissions) followed by atmospheric transport and deposition are believed to be the
3
dominant transport mechanisms responsible for the widespread environmental
4 occurrence of PCBs. Photodegradation to less chlorinated congeners followed by
5
slow anaerobic and/or aerobic biodegradation is believed to be the principal path for
6 destruction of PCBs. Similar situations exist for the polybrominated biphenyls (PBBs).
7 Little information is available on the occurrence and fate of biphenyl congeners
8
containing both chlorine and bromine but their contribution to dioxin-like activity in the
9
environment is thought to be small.
10
11
SOURCES
12
The chlorinated and brominated dioxins and furans have never been
13
intentionally produced other than on a laboratory scale basis for use in chemical
14 analyses. Rather, they are generated as byproducts from various combustion and
15 chemical processes. PCBs were produced in relatively large quantities for use in such
16 commercial products as dielectrics, hydraulic fluids, plastics and paints. They are no
17 longer produced, but continue to be released to the environment through the use and
18 disposal of these products. A similar situation exists for the commercially produced
19
PBBs which were produced for a number of uses like flame retardants.
20
Dioxin-like compounds are released to the environment in a variety of ways and
21
in varying quantities depending upon the source. Studies of sediment cores in lakes
22
near industrial centers of the United States have shown that historical environmental
23 deposition of dioxins and furans was quite low until about 1920, peaked around 1980
24
and has declined thereafter. This trend suggests that the presence of dioxin-like
25 compounds in the environment has occured primarily as a result of industrial practices
26 and is likely to reflect changes in release over time. Although these compounds are
27
released from a variety of sources, the congener profiles of CDDs and CDFs found in
28 sediments have been linked to combustion sources (Hites, 1991). Three theories
29 have been suggested to explain formation of CDDs and CDFs during combustion: 1)
30
The CDDs and CDFs are present in the fuels or feed materials and pass through the
31
combustor intact; 2) precursor chemicals are present in the fuels or feed material and
32
undergo reactions catalyzed by particulates and other chemicals to form CDDs and
6
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
CDFs; and 3) the CDDs and CDFs are formed de novo from organic and inorganic
2
substrates bearing little resemblance in molecular structure.
3
The principal identified sources of environmental release of CDDs and CDFs
4
may be grouped into four major types:
5
Combustion and Incineration Sources: Dioxin-like compounds can be
6
generated and released to the environment from various combustion processes when
7
chlorine donor compounds are present. These sources can include incineration of
8
wastes such as municipal solid waste, sewage sludge, hospital and hazardous
9
wastes; metallurgical processes such as high temperature steel production, smelting
10
operations, and scrap metal recovery furnaces; and the burning of coal, wood,
11
petroleum products, and used tires for power/energy generation. Even cigarette
12
smoke, crematories, volcanoes, and forest fires have been shown to be minor
13 sources.
14
Chemical Manufacturing/Processing Sources: Dioxin-like compounds can be
15
formed as by-products from the manufacture of chlorine and such chlorinated
16
compounds as chlorinated phenols, PCBs, phenoxy herbicides, chlorinated benzenes,
17
chlorinated aliphatic compounds, chlorinated catalysts, and halogenated diphenyl
18 ethers. Although the manufacture of many chlorinated phenolic intermediates and
19
products, as well as PCBs, was terminated in the late 1970s in the United States,
20 production continued elsewhere around the world until 1990 and continued, limited
21
use and disposal of these compounds can result in releases of CDDs, CDFs, and
22
PCBs to the environment.
23
Industrial/Municipal Processes: Dioxin-like compounds can be formed
24
through the chlorination of naturally occurring phenolic compounds such as those
25 present in wood pulp. The formation of CDDs and CDFs resulting from the use of
26
chlorine bleaching processes in the manufacture of bleached pulp and paper has
27
resulted in the presence of CDDs and CDFs in paper products as well as in liquid and
28
solid wastes from this industry. Municipal sewage sludge has been found to
29
occasionally contain CDDs and CDFs.
30
Reservoir Sources: The persistent and hydrophobic nature of these
31
compounds cause them to accumulate in soils, sediments and organic matter and to
32
persist in waste disposal sites. The dioxin-like compounds in these "reservoirs" can be
7
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
redistributed by various processes such as dust or sediment resuspension and
2 transport. Such releases are not original sources in a global sense, but can be on a
3 local scale. For example, releases may occur naturally from sediments via
4 volatilization or via operations which disturb them such as dredging. Aerial deposition
5 and accumulation on leaves can lead to releases during forest fires or leaf composting
6 operations.
7
As awareness of these possible sources has grown in recent years, a number of
8
changes have occurred which should reduce the release rates. For example, releases
9 of dioxin-like compounds have been reduced due to the switch to unleaded
10 automobile fuels (and associated use of catalytic converters and reduction in
11 halogenated scavenger fuel additives), process changes at pulp and paper mills, new
12 emission standards and upgraded emission controls for incinerators, and reductions in
13 the manufacture of chlorinated phenolic intermediates and products.
14
Although dioxins in the environment may arise from a number of sources as
15 discussed above, the Exposure Document presents recent analyses of air emissions
16 of CDDs and CDFs for several European countries in terms of total toxic equivalents
17 (TEQs) based on international TEFs. These studies assume that emissions to air
18 make up the major portion of dioxins released to the environment. Estimates of total
19 release in these countries range from approximately 100-1000 g TEQ/year in West
20 Germany and 100-200g TEQ/year in Sweden to approximately 1000 and 4000 g TEQ/
21 year maximum emissions in the Netherlands and United Kingdom respectively.
22 Similar nationwide estimates for the U.S. have not been compiled prior to this
23 reassessment effort. The Exposure Document estimates an upper end on U.S.
24 emissions to be in the range of 14,000 g TEQ/year. Qualitatively speaking, major
25 contributors to this total include medical waste incinerators, municipal waste
26 incinerators, cement kilns,and industrial wood burning. Because of the limited number
27 of measurements and the large number of potential sources for each of these
28 emissions, total estimated emissions from these sources are considered highly
29 uncertain. Municipal waste incineration has a better data base of measurement data
30 than other air sources but emissions are highly variable among facilities so that the
31
overall estimate remains uncertain. Diesel-fueled vehicles, hazardous waste burning,
32
forest fires and metal smelting are more moderate contributors of dioxin-like
8
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
compounds but the magnitude of the contribution is also highly uncertain. Sewage
2
waste incineration and residential wood burning as well as a few minor processes
3
round out the current analysis and provide lower range estimates of medium to low
4
certainty. Although still other sources are recognized and releases to land and water
5
in addition to air are briefly mentioned in the Exposure Document, it is clear from this
6
exercise that additional measurement data will be needed to gain an adequate
7
appreciation for the nature and magnitude of major U.S. sources of CDD and CDF
8
emissions.
9
Several investigators have attempted to conduct "mass balance" checks on the
10
estimates of national dioxin releases to the environment. Basically, this procedure
11
involves comparing estimates of the emissions to estimates of aerial deposition. Such
12 studies in Sweden (Rappe, 1991) and Great Britain (Harrad and Jones, 1992) have
13
suggested that the deposition exceeds the emissions by about 10 fold. These studies
14
are acknowledged to be quite speculative due to the strong potential for inaccuracies
15
in emission and deposition estimates. In addition, the apparent discrepancies could.
16
be explained by long range transport from outside the country, resuspension and
17
deposition of reservoir sources or unidentified sources. Bearing these limitations in
18
mind, this procedure has been used in this reassessment to compare the estimated
19
emissions and deposition in the U.S.
20
Deposition measurements have been made at a number of locations in Europe
21
and two places in the US (See discussion of these studies in Volume II of the
22 Exposure Document). These limited data suggest that a deposition rate of 1 ng
23
TEQ/m2-yr is typical of remote areas and that 2-6 ng TEQ/m2-yr is more typical of
24
populated areas. Applying the values of 1 ng TEQ/m2-yr to Alaska and 2-6 ng TEQ/m2-
25 yr to the contiguous 48 states, the total U.S. deposition can be estimated as 20,000 to
26
50,000 g TEQ/yr. While this range is higher than the upper estimate of emissions for
27
the US (<14,000 g TEQ/yr) as presented in the Exposure Document, the upper
28
estimate may account for >30-70% of predicted deposition. As noted above,
29
interpreting such comparisons is highly speculative and supports the need to conduct
30
further emissions testing into all media and deposition measurement, if we are to
31
understand emisssions and deposition in terms of a mass balance.
32
9
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
Levels in the Environment and in Food
2
CDDs, CDFs and PCBs have been found throughout the world in practically all
3
media including air, soil, water, sediment, fish and shellfish, and other food products
4 such as meat and dairy products. The highest levels of these compounds are found in
5 soils, sediments, and biota; very low levels are found in water and air. The
6 widespread occurrence observed, particularly in industrialized countries, is not
7 unexpected considering the numerous sources that emit these compounds into the
8 atmosphere, and the overall resistance of these compounds to biotic and abiotic
9 transformation.
10
The average levels of these compounds found in the various media in North
11 America have been compiled in the Exposure Document. The levels shown for
12 environmental media and for food are based on few samples and must be considered
13 quite uncertain. However, they seem reasonably consistent with levels measured in a
14 number of studies in Western Europe and Canada. The consistency of these levels
15 across industrialized countries adds some confidence to the limited data from the U.S.
16 and provides some reassurance that these estimates are reasonable.
17
This assessment proposes the hypothesis that the primary mechanism by which
18
dioxin-like compounds enter the terrestrial food chain is via atmospheric deposition.
19 Dioxin and related compounds enter the atmosphere directly through air emissions or
20 indirectly, for example through volatilization from land or water or from re-suspension
21 of particles. Deposition can occur directly onto soil or onto plant surfaces. Soil
22 deposits can enter the food chain via direct ingestion (e.g. grazing animals, earth
23 worms, fur preening by burrowing animals). Dioxin-like compounds in soil can
24 become available to plants by volatilization and vapor absorption or particle
25 resuspension and adherence to plant surfaces. In addition, dioxin-like compounds in
26 soil can adsorb directly to underground portions of plants. Uptake from soil via the
27 roots into above ground portions of plants is thought to be insignificant.
28
Support for this air-to-food hypothesis is provided by Hites (1991) who
29 concluded that "background environmental levels of dioxin-like compounds are
30 caused by dioxin-like compounds entering the environment through the atmospheric
31 pathway." His conclusion was based on demonstrations that the congener profiles in
32 lake sediments could be linked to congener profiles of combustion sources. Further
10
Draft - Do Not Quote or Cite . Draft May 2, 1994
1
arguments supporting this hypothesis include: 1) numerous measurements show that
2
emissions occur from multiple sources and deposition occurs in most areas including
3
remote locations, 2) atmospheric transport and deposition is the only mechanism that
4
could explain the widespread distribution of these compounds in soil, and 3) other
5
mechanisms of uptake into food, for instance from direct contamination or through
6
packaging, are much less plausible. Direct uptake into food from soil or sediments is
7
possible and could be important for "local" exposures. These routes are less likely to
8
explain the general background level of dioxin and related compounds found in the
9
diet of the general population.
10
At present, it is unclear whether atmospheric deposition represents primarily
11
"new" contributions of dioxin and related compounds from all media reaching the
12
atmosphere or whether it is "old" dioxin and related compounds which persist and
13
recycle in the environment. Understanding the relationship between these two
14
scenarios will be particularly important in understanding the relative contributions of
15
individual point sources of these compounds to the food chain and assessing the
16
effectiveness of control strategies attempting to reduce the levels in food.
17
18
Background Exposure Levels
19
The term "background" exposure has been used throughout this reassessment
20
to describe exposure of the general population who are not exposed to identifiable
21
point sources. For the purposes of calculating background exposures to dioxin-like
22 compounds via dietary intake the upper-range background toxicity equivalent values
23 (TEQs) (i.e., those calculated using one-half the detection limit for the non-detects)
24 were used in the Exposure Document. The estimates are based on intake of dioxin-
25
like CDDs and CDFs and do not include estimates for dioxin-like PCBs or other dioxin-
26
like compounds. Inclusion of dioxin-like PCBs could raise these estimates by 35-50%.
27
A background exposure level of 119 pg TEQ/day for the U.S. was estimated. These
28
estimates are comparable to analogous estimates for European countries. These
29
include estimates for Germany, which range from 79 pg TEQ/day based on Furst, et al.
30
(1990) to 158 pg TEQ/day based on Furst, et al. (1991), 118-126 pg TEQ/day exposure
31
via numerous routes in the Netherlands (Theelen, 1991), and 140-290 pg TEQ/day for
32
the typical Canadian exposed mainly through food ingestion (Gilman and Newhook,
11
Draft - Do Not Quote or Cite . Draft
May 2, 1994
1
1991). It is generally concluded by these researchers that dietary intake is the primary
2
pathway of human exposure to CDDs and CDFs. These investigators among others
3
suggest that greater than 90 percent of human exposure occurs through the diet, with
4
foods from animal origins being the predominant sources.
5
This hypothesis remains to be validated. Although data are derived from
6
multiple studies from around the world, they represent limited numbers of samples.
7
Use of one-half of the detection level for non-detects is a reasonable but conservative
8
approach to estimating low levels in samples. For some data sets, use of zero values
9
for non-detects could result in significantly lower estimates. However, it is widely held
10
that such an approach would most likely underestimate true levels of exposure.
11
Similar estimates derived from different data sets, developed by different investigators
12
in several countries, strengthen the probability that this inference represents the true
13
picture for exposure of the general population in industrialized countries to dioxin and
14
related compounds.
15
Data on human tissue levels suggest that body burden levels among
16
industrialized nations are reasonably similar (Schecter, 1991). These data can also
17
be used to estimate background exposure through the use of pharmacokinetic models.
18
Using this approach, exposure levels to 2,3,7,8-TCDD in industrialized nations are
19
estimated to be about 20 to 40 pg/day ( .3-.6 pg TCDD/kg/day). This is generally
20
consistent with the estimates derived using diet based approaches to estimate total
21
TCDD intake.
22
The U.S. study of CDD/F body burdens contained in the National Human
23
Adipose Tissue Survey (NHATS) (EPA, 1991) analyzed for CDD/Fs in 48 human
24
tissue samples which were composited from 865 samples. These samples were
25
collected during 1987 from autopsied cadavers and surgical patients. While this was
26
an important study of chemical residues occuring in human fat , numerous technical
27
shortcomings of this study have been described. For instance, the sample
28
compositing prevents use of these data to examine the distribution of CDD/F levels in
29
tissue among individuals. However, it did allow conclusions in the following areas:
30
National Averages: The national averages for all TEQ congeners (but excluding
31
dioxin-like PCBs) were estimated and totaled to 28 pg TEQ/g lipid adjusted value (28
32
ppt).
12
Draft . Do Not Quote or Cite - Draft May 2, 1994
1
Age Effects: Tissue concentrations of CDD/Fs were found to increase with age.
2
.Geographic Effects: In general, the average CDD/F tissue concentrations
3 appeared fairly uniform geographically.
4 Race Effects: No significant difference in CDD/F tissue concentrations were found
5 on the basis of race.
6 Sex Effects: No significant difference in CDD/F tissue concentrations were found
7 between males and females.
8 -Temporal Trends: The 1987 survey showed decreases in tissue concentrations
9 relative to the 1982 survey for all congeners. However, it is not known whether these
10 declines were due to improvements in the analytical methods or actual reductions in
11 body burden levels. The percent reductions among individual congeners varied from
12 9 percent to 96 percent.
13
More recent data (Patterson et al., 1994) show similar trends with regard to
14 decreasing levels of dioxin-like PCBs in blood and fat. In addition, they showed a
15 wide variability of PCB congeners in human adipose tissue sample as compared to
16 concentrations of CDDs and CDFs which were less variable.
17
Inclusion of dioxin-like PCBs in TEQ calculations raises the average body
18 burden to 40-60 pg TEQ/g (40-60 ppt). Since available data from the two studies
19 discussed above do not provide a representative population sample, these
20 conclusions must be regarded as somewhat uncertain. Additional measurements will
21 be necessary to confirm these findings. Use of a protocol for sampling which allows
22 an evaluation of age adjusted population averages will be critical for understanding
23 the current body burden situation and evaluating impacts of future efforts to further
24 reduce exposures to this class of compounds.
25
Levels of dioxin-like compounds found in human tissue/blood appear similar in
26 Europe and North America. Schecter (1991) compared levels of dioxin-like
27 compounds found in blood among people from U.S. pooled samples (100 subjects)
28 and Germany (85 subjects). Although mean levels of individual congeners differed by
29 as much as a factor of two between the two populations, the total TEQ averaged 42 pg
30 TEQ/g (42 ppt) in the German subjects and was 41 pg TEQ/g ( 41 ppt) in the pooled
31
U.S. samples. These values do not include TEQs for PCBs.
32
New information on levels of dioxin-like compounds in human adipose tissue
13
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
and blood has recently been published (Patterson et al, 1994). This study reports
2
measurements of dioxin-like PCB congeners as well as CDD and CDF levels in
3 samples from 28 Atlanta residents. These measurements show that concentrations of
4 dioxin-like PCBs can be more than an order of magnitude higher than concentrations
5 of TCDD. Comparison with other published information suggests much higher levels
6 of non-dioxin-like congeners of PCBs and the possibility that concentrations of both
7 types of congeners will depend heavily upon previous human activities such as fish
8 consumption. These data are consistent with the previous statement that dioxin-like
9 PCBs may account for approximately 1/3 of the total TEQ in the general population.
10 Values in Patterson's study calculated TEQs for PCBs using the data of Safe (1990)
11 which were acknowledged by the author as being conservative and, based on more
12 recent data, are likely to overestimate the contribution of dioxin-like PCBs.
13
14 Highly Exposed Populations
15
Certain groups of people may have higher exposures to dioxin-like compounds
16 than the general population. This issue has been discussed previously in terms of
17 increased exposure due to dietary habits (See Exposure Document) or due to
18 occupational conditions or industrial accidents (See Chapter 7).
19
Consumption of breast milk by nursing infants may lead to higher levels of
20
exposure during the early postnatal period as compared to intake in the diet later in
21
life. Schecter et al. (1992) reports that a study of 42 U.S. women found an average of
22
16 pg TEQ /g (16 ppt), 3.3 ppt of which was 2,3,7,8-TCDD, in the lipid portion of breast
23 milk. A much larger study in Germany (n= 526) found an average of 29 pg TEQ / g (29
24 ppt ) in the lipid portion of breast milk. These estimates do not include a contribution
25 to total TEQ from dioxin-like PCBs. The level in human breast milk can be predicted
26 on the basis of the estimated dioxin intake by the mother. Such procedures are
27 presented in Volume II of the Exposure Document.
28
Using these procedures and assuming that an infant breast feeds for one year,
29 has an average weight during this period of 10 kg, ingests 0.8 kg/d of breast milk and
30 that the dioxin concentration in milk fat is 20 pg /g ( 20 ppt) of TEQ, the average daily
31
dose to the infant over this period is predicted to be about 60 pg TEQ/kg/d, not
32
including dioxin-like PCBs. This value is 10 to 20 times higher than the estimated
14
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May 2, 1994
1
range for background exposure to adults (i.e. 3-6 pg TEQ/kg/d). However, if a 70 yr
2
averaging time is used to obtain an added increment of lifetime daily dose, then the
3
increment of lifetime average daily dose is attributable to this nursing scenario is
4
estimated to be 0.8 pg of TEQ/kg/d. On a mass basis, the cumulative dose to the infant
5
under this scenario is about 210 ng compared to a lifetime background dose of about
6
1700 to 5100 ng (suggesting that 4 to 12 percent of the lifetime dose may occur as a
7
result of breast feeding for the first year of life). Traditionally, EPA has used the lifetime
8
average daily dose as the basis for evaluating incremental cancer risk and the
9 average daily dose (i.e., the daily exposure per unit body weight occurring during an
10 exposure event) as the more appropriate indicator of risk for certain noncancer
11 endpoints. The use of a lifetime average daily dose for high level, early exposures
12 may underestimate cancer risk if dose rate or perinatal sensitivity is important in the
13 ultimate carcinogenic outcome. The average daily dose approach may be particularly
14 important for the evaluation of non-cancer endpoints if exposure is occurring during
15
windows of sensitivity during prenatal and postnatal development.
16
In addition, consumption of unusually high levels of fish or meat containing
17
elevated levels of dioxin and related compounds can lead to elevated blood levels in
18 comparison to the general population. Most people eat fish from multiple sources
19 where levels of dioxin-like compounds are likely to be low. Even if large quantities of
20 fish are consumed, they are not likely to have unusually high exposures. However,
21
individuals who fish regularly for purposes of basic subsistence are likely to obtain
22 their fish from a few sources and may have the potential for elevated exposures. Such
23 individuals may also consume large quantities of fish. Although average consumers
24
may eat a few fish meals a month (an average intake of 6.5 grams of fish a day), many
25 recreational anglers near large water bodies may consume, on average, 4 to 5 times
26 as much (approximately 30 grams per day); some individuals at the high end of the
27 consumption range may eat, on average, as much as140 grams per day. Certain
28 members of ethnic groups who are subsistence fishers may consume 2 to 3 times this
29 amount as an upper estimate. Svensson et al (1991) found elevated blood levels of
30
PCDDs and PCDFs in high fish consumers living near the Baltic Sea in Sweden. The
31
highest consumers, fishermen or workers in the fish industry, had blood level TEQs
32
that were approximately 3 times that of non-fish consumers (60 pg TEQ/g lipid versus
15
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
20 pg TEQ/g lipid). The difference in levels of dioxin-like compounds was particularly
2
apparent for the PCDFs. Dioxin-like PCBs were not accounted for in this study.
3
Studies are currently underway to examine fish consumption patterns in several
4
Native American groups. Recent results (Columbia River Intertribal Fish Commission,
5
1994) suggest that Native Americans living along the Columbia River may consume
6
an average of 30 grams of fish a day; some individuals consume much higher levels.
7
Studies are currently underway to determine levels of dioxin-like compounds in fish
8
from this region. No measurements of dioxin-like chemicals in the blood of these
9
Native American populations are currently available.
10
Dewailly et al. (1994) observed elevated levels of coplanar PCBs in the blood of
11
fishermen on the north shore of the Gulf of the St. Lawrence River who consume large
12
amounts of seafood. Coplanar PCB levels were 20 times higher among the 10 highly
13
exposed fishermen than among controls. This study also reported elevated levels of of
14
coplanar PCBs in the breast milk of Inuit women of Arctic Quebec. The principal
15
source of protein for the Inuit people is fish and sea mammal consumption.
16
The possibility of high exposures to dioxin-like chemicals as a result of
17
consuming meat and dairy products is most likely to occur in situations where
18
individuals consume large quantities of these foods from a locality where the level of
19
these compounds is elevated. Most people eat meat and dairy products from multiple
20
sources and, even if large quantities are consumed, are not likely to have unusually
21
high exposures. However, individuals who raise their own livestock for basic
22
subsistence have the potential for higher exposures if local levels of dioxin-like
23
compounds are high. Volume III of the Exposure Document presents methods for
24
evaluating this type of exposure scenario, but no studies were found in the literature to
25
demonstrate this potential based on measurements of dioxin-like chemicals from
26
source to livestock to humans.
27
Although the subpopulations discussed above have the potential for high
28
exposure to dioxin-like compounds, a careful evaluation of dietary habits is needed to
29
confirm this possibility. It would generally be inappropriate to compute the total intake
30
of dioxin-like compounds in a subpopulation by simply adding the dioxin intake from
31
highly consumed food to the general population intake level. The general population
32
background estimate assumes a typical pattern of food ingestion, whereas a
16
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1
subpopulation who has a high consumption rate of one particular food type is likely to
2
eat less of other food types. Ideally, the evaluation should be based on the entire diet
3
of the subpopulation and use case-specific values for food ingestion rates and
4
concentrations of dioxin-like compounds.
5
High blood levels of dioxin and related compounds based on high levels of
6
exposure have been documented for industrial exposures in segments of the chemical
7
industry and for industrial accidents. Health effects studies in human populations have
8
focused on these groups of highly exposed individuals. Results of these studies are
9
described in detail in Chapter 7. Other populations in proximity to industrial sites have
10
been evaluated for elevated blood levels of dioxin and related compounds. Higher
11
levels have been measured in a few situations.
12
13 DISPOSITION AND PHARMACOKINETICS
14
The disposition and pharmacokinetics of 2,3,7,8-TCDD and related compounds
15
have been investigated in several species and under various exposure conditions.
16
These data and models derived from them are critical in understanding the sequelae
17
of human exposure. Data related to disposition and pharmacokinetics of dioxin and
18
related compounds and efforts to develop models to further understand tissue
19
dosimetry are described in detail in Chapter 1 of the Health Assessment document.
20
The gastrointestinal, dermal and transpulmonary absorption of these
21
compounds represent potential routes for human uptake. Findings of studies in
22
experimental animals indicate that oral exposure to 2,3,7,8-TCDD in the diet or in an
23
oil vehicle results in the absorption of >50%, and often closer to 90%, of the
24 administered dose. Gastrointestinal absorption of related compounds is variable,
25
incomplete and congener specific. More soluble congeners, such as 2,3,7,8-TCDF,
26
are almost completely absorbed, while the extremely insoluble OCDD is very poorly
27
absorbed. In some cases, absorption has been found to be dose dependent, with
28
increased absorption occurring at lower doses (2,3,7,8-TBDD, OCDD). The limited
29
data base also suggests that there are no major interspecies differences in the
30
gastrointestinal absorption of these compounds among mammals. Limited data from a
31
single human volunteer suggests a high level (> 87%) of absorption of 2,3,7,8-TCDD
32
in corn oil from the gastrointestinal tract. Following absorption, a half life for
17
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
elimination was estimated to be 2120days.
2
Additional data also indicate the importance of the formulation or vehicle
3 containing the toxicant(s) on the relative bioavailability of 2,3,7,8-TCDD and related
4 compounds after exposure. For instance, rodent feeding studies indicate that the
5 bioavailability of 2,3,7,8-TCDD from soil varies between sites and 2,3,7,8-TCDD
6 content alone may not be indicative of potential human hazard from contaminated
7 environmental materials. Although data indicate that substantial absorption may occur
8 from contaminated soil, soil type and duration of contact may substantially affect the
9 absorption of 2,3,7,8-TCDD from soils obtained from different contaminated sites. This
10 uncertainty should be kept in mind as intake values are often used to estimate
11 potential risk from environmental samples.
12
In experiments measuring dermal absorption for 2,3,7,8-TCDD and several
13 CDFs, the percentage of administered dose absorbed decreased with increasing dose
14 while the amount absorbed (µg/kg) increased with dose. Results also suggest that the
15 majority of the compound remaining at the skin exposure site was associated with the
16 the outer skin layer ( the stratum corneum) and did not penetrate through to the dermis.
17 Together, these results on dermal absorption indicate that at lower doses (≤0.1
18 µmol/kg), a greater percent of this administered dose of 2,3,7,8-TCDD and three CDFs
19 was absorbed. Nonetheless, even following a low dose dermal application of 200
20 pmol (1 nmol/kg), the rate of absorption of 2,3,7,8-TCDD is still very slow (rate constant.
21 of 0.005 hour-1). Dermal exposure of humans to 2,3,7,8-TCDD and related
22 compounds usually occurs as a complex mixture of these contaminants in soil, oils or
23 other mixtures which would be expected to alter absorption. Available data suggest
24 that the dermal absorption of 2,3,7,8-TCDD depends on the formulation (vehicle or
25 adsorbent) containing the toxicant. Although no data are available to directly evaluate
26 human dermal absorption, the data available from in vitro and animal studies suggest
27 slow dermal absorption of these compounds which is likely to be dependent on the
28 vehicle or adsorbent containing the compounds and the duration of the contact.
29
The use of incineration as a means of solid and hazardous waste management
30
results in the emission of contaminated particles that may contain TCDD and related
31 compounds into the environment. Thus, exposure to TCDD and related compounds
32 may result from inhalation of contaminated fly ash, dust and soil. Systemic effects
18
Draft - Do Not Quote or Cite - Draft May 2, 1994
1 occur in animals after pulmonary exposure to TCDD, suggesting that transpulmonary
2 absorption of TCDD does occur. Further results suggest that the transpulmonary
3 absorption of 2,3,7,8-TCDD and 2,3,7,8-TBDD was similar to that observed following
4 oral exposure. These limited data provide evidence of efficient transpulmonary
5 absorption after intratracheal instillation in laboratory animals. No data from humans
6 or primates are available to address this issue. However, these data provide support
7 for the inference that efficient absorption will occur when particles containing dioxin
8 and related compounds are inhaled by humans.
9
Once absorbed into blood, 2,3,7,8-TCDD and related compounds readily
10 distribute to all organs. Tissue distribution within the first hour after exposure parallels
11 blood levels and reflects physiological parameters such as blood flow to a given tissue
12 and relative tissue size. There do not appear to be major species or strain differences
13 in the tissue distribution of 2,3,7,8-TCDD and 2,3,7,8-TCDFin mammals, with the liver
14 and adipose tissue being the primary disposition sites although human data to
15 address this issue are quite limited. The tissue distribution of the coplanar PCBs and
16 PBBs also appears to be similar to that of 2,3,7,8-TCDD and 2,3,7,8-TCDF based on
17 evaluation in experimental animals.
18
Multiple studies suggest that distribution of this class of compounds to internal
19 organs is likely to be dose dependent. At low doses in animal studies, adipose tissue
20 serves as the major depot; at high doses, a major fraction is sequestered in the liver.
21 The biochemical basis for this observation is under investigation. Induction of a
22 binding protein has been hypothesized to play a major role.
23
As discussed above, levels of 2,3,7,8-TCDD averaging 5-10 pg/g lipid (ppt)
24 have been reported for background populations. Sielken (1987) evaluated these data
25 and concluded that the levels of 2,3,7,8-TCDD in human adipose are log-normally
26 distributed and positively correlated with age. Among the observed U.S. background
27 levels of 2,3,7,8-TCDD in human adipose tissue, more than 10% were >12 pg/g (ppt).
28 Paired human serum and adipose tissue levels of 2,3,7,8-TCDD have been compared
29 by Patterson et al. (1988) and Kahn et al. (1988). Both laboratories reported a high
30 correlation between adipose tissue and serum 2,3,7,8-TCDD levels when the samples
31 were adjusted for total lipid content. This correlation indicates that serum 2,3,7,8-
32 TCDD provides a valid estimate of the 2,3,7,8-TCDD concentration in adipose tissue
19
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
under steady-state, low-dose conditions.
2
In a study of potentially heavily exposed Vietnam veterans, the Centers for
3
Disease Control (MMWR, 1988) reported on an Air Force study of Ranch Hand
4
veterans who were either herbicide loaders or herbicide specialists in Vietnam. The
5
herbicide, 2,4,5,T (Agent Orange) that was used in Viet Nam was contaminated with a
6
low percentage of 2,3,7,8-TCDD. The mean serum 2,3,7,8-TCDD levels of 147 Ranch
7
Hand personnel was 49 pg/g (ppt) in 1987, based on total lipid-weight, while the mean
8
serum level of the 49 controls was 5 pg/g (ppt). In addition, 79% of the Ranch Hand
9
personnel and 2% of the controls had 2,3,7,8-TCDD levels ≥10 pg/g (ppt). The
10 distribution of 2,3,7,8-TCDD levels in this phase of the Air Force health study indicates
11
that , while Ranch Hand veterans have higher lifetime exposures than controls, only a
12 small number of Ranch Hand personnel had unusually heavy 2,3,7,8-TCDD exposure.
13
This report also estimated the half-life of 2,3,7,8-TCDD in humans to be ~7 years on
14
the basis of 2,3,7,8-TCDD levels in serum samples taken in 1982 and 1987 from 36 of
15
the Ranch Hand personnel who had 2,3,7,8-TCDD levels >10 pg/g (ppt) in 1987.
16 Similar results were obtained by Kahn et al. (1988) who compared 2,3,7,8-TCDD
17
levels in blood and adipose tissue of Agent Orange-exposed Vietnam veterans and
18
matched controls. This study also examined moderately exposed Vietnam veterans
19
who handled herbicides regularly while in Vietnam. Although this study can
20
distinguish moderately exposed men from others, the data do not address the question
21
of identifying persons whose exposures are relatively low and who constitute the bulk
22
of the population, both military and civilian, who may have been exposed to greater
23
than background levels of 2,3,7,8-TCDD.
24
Although early in vivo and in vitro investigations were unable to detect the
25
metabolism of 2,3,7,8-TCDD, there is now evidence that a wide range of mammalian
26 and aquatic species are capable of slowly biotransforming 2,3,7,8-TCDD to polar
27
metabolites. Although metabolites of 2,3,7,8-TCDD have not been directly identified in
28 humans, recent analytic data from feces samples from an individual in a self-dosing
29 experiment suggests that humans can metabolize 2,3,7,8-TCDD (Wendling et al.,
30
1990). The metabolism of 2,3,7,8-TCDD and related compounds is required for
31
urinary and biliary elimination and therefore plays a major role in regulating the rate of
32
excretion of these compounds. Direct intestinal excretion of parent compound is
20
Draft . Do Not Quote or Cite - Draft May 2, 1994
1
another route for excretion of 2,3,7,8-TCDD and related compounds that is not
2
regulated by metabolism.
3
Structure-activity studies of 2,3,7,8-TCDD and related compounds support the
4
widely accepted principle that the parent compound is the active species, and the
5
relative lack of biological activity of readily excreted monohydroxylated metabolites of
6
2,3,7,8-TCDD and 3,3'4,4'-TCB suggests that metabolism is a detoxification process
7
necessary for the biliary and urinary excretion of these compounds. This concept has
8
also been generally applied to 2,3,7,8-TCDD-related compounds, although data are
9
lacking on the structure and toxicity of metabolites of other CDDs, BDDs, CDFs, BDFs,
10
PCBs and PBBs. It is still possible, however quite unlikely, that low levels of
11
unextractable and/or unidentified metabolites may contribute to one or more of the
12
toxic responses of 2,3,7,8-TCDD and related compounds.
13
Due to the lipophilic nature of dioxins and related compounds, lactation can
14
provide a mechanism for decreasing the body burden of these compounds in females.
15
This elimination of 2,3,7,8-TCDD through mother's milk can result in high exposure
16
levels in the infant, as discussed above. Since milk is highly absorbable, it would be
17
likely that this source would provide 2,3,7,8-TCDD and related compounds in a form
18
that is readily bioavailable to the nursing infant.
19
Physiologically-based pharmacokinetic (PB-PK) models have been developed
20
for 2,3,7,8-TCDD in mice, rats and humans. PB-PK models incorporate known or
21
estimated anatomical, physiological and physicochemical parameters to describe
22
quantitatively the disposition of a chemical in a given species. PB-PK models can
23 assist in the extrapolation of high-to-low dose kinetics within a species, estimating
24
exposures by different routes of administration, calculating effective doses and
25
extrapolating these values across species. These models are particularly important
26
given the limited empirical data on individual dioxin-like congeners.
27
Kedderis (1994) has recently reviewed biologically-based models of dioxin
28
pharmacokinetics. The early studies in rodents have recently been extended to
29
describe protein induction and tissue distribution data in the mouse (Leung et
30
al.,1990b) and rat (Leung et al., 1990a). Anderson and coworkers (Anderson et
31
al.,1993) refined the model to relate protein induction to interactions between dioxin,
32
the Ah receptor and DNA. This model also incorporated the concept of diffusion-
21
Draft - Do Not Quote or Cite . Draft
May 2, 1994
1
limited tissue distribution. The model described by Kedderis et al. (1993) for 2,3,7,8-
2
tetrabromodibenzo-p-dioxin (TBDD) extended the use of PBPK models to the
3
brominated congener of TCDD and designated the inducible cytochrome, CYP1A2, as
4
the dioxin binding protein in the liver. Kohn et al. (1993) used similar approaches to
5
describe tissue dosimetry of TCDD and additionally incorporated dioxin mediated
6
effects on growth factors. Other models have been proposed recently to describe
7
effects of TCDD on lipid metabolism (Roth et al., ,1993). An empirical dose dependent
8
model by Carrier (1991) related the varying fraction of the body burden of TCDD
9
associated with the liver in humans to the total body burden of TCDD. Kedderis (1994)
10 has suggested that, with our current understanding of the biologic determinants
11 driving the hepatic sequestration of dioxin, this empirical description may now be
12 interpreted in terms of a biologically-based model. The fact that the Carrier (1991)
13 model deals with a relatively large data base of human exposures to dioxin and
14
related compounds may facilitate predictions of human risk in terms of dosimetry as
15
well as biologic response.
16
Our uncertainty in the validity of predictions from PB-PK models is primarily
17
driven by the limited availability of congener and species-specific data that accurately
18
describe the dose- and time-dependent disposition of 2,3,7,8-TCDD and related
19
compounds. As additional data become available, particularly on the dose-dependent
20
disposition of these compounds, more accurate models can be developed. In
21
developing a suitable model in the human, it is also important to consider that the half-
22
life estimate of 7.1 years for 2,3,7,8-TCDD was based on two serum values taken
23
5 years apart, with the assumption of a single compartment, and assuming a first-order
24
elimination process (Pirkle et al., 1989). It is likely that the excretion of 2,3,7,8-TCDD
25
in humans is more complex, involving several compartments, tissue-specific binding
26
proteins and a continuous daily background exposure. Furthermore, changes in body
27
weight and body composition should also be considered in developing PB-PK models
28
for 2,3,7,8-TCDD and related compounds in humans.
29
It is known that some exposure occurs to the developing fetus through placental
30
transfer of dioxin-like compounds in maternal blood via the placenta. In addition,
31
exposure is likely to increase in the early post-natal period through intake of mother's
32
milk containing dioxin-like compounds. Re-distribution of body burdens is likely to
22
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
occur with growth and development depending on relative intakes and changes in
2
body fat content. Fasting, aging and disease are all thought to alter steady state levels
3
of dioxin during life. These changes complicate standard pharmacokinetic models
4
and present the possibility for transient but potentially important increases in blood or
5 tissue levels of dioxin-like compounds during critical periods of development, growth
6 and aging. Additional data on both pharmacokinetics and pharmacodynamics in
7
relation to development and growth will be required to refine our perspectives on the
8
importance of these issues in evaluating dioxin hazards and risks.
9
10
MECHANISMS OF DIOXIN ACTION
11
Knowledge of the mechanisms of dioxin action may facilitate the risk
12
assessment process by imposing bounds upon the assumptions and models used to
13
describe possible responses to exposure to dioxin. In this document, current
14
knowledge of dioxin action has been reviewed, with emphasis on the contribution of
15
the specific cellular receptor for dioxin and related compounds, the Ah receptor, to the
16
mechanism. Other reviews referenced in Chapter 2 provide additional background on
17
the subject.
18
The remarkable potency of TCDD in eliciting its toxic effects suggested
19 the possible existence of a receptor for dioxin. Biochemical and genetic evidence
20 implicate the TCDD-receptor in the biological responses to dioxin-like compounds.
21
For example, studies of structure-activity relationships among congeners of TCDD
22 reveal a correlation between a compound's specific binding affinity and its potency in
23 eliciting biochemical responses, such as enzyme induction. Furthermore, inbred
24 mouse strains in which TCDD binds with lower affinity to the receptor exhibit
25 decreased sensitivity to dioxin's biological effects, such as thymic involution, cleft
26 palate formation and hepatic porphyria.
27
Electrophoretic studies to evaluate the properties of specific proteins from
28
inbred mouse strains reveal the existence of several forms of the TCDD-binding
29 protein. These observations imply the existence of multiple alleles at the Ah locus in
30
mice. The biochemical properties of the different forms of the Ah receptor remain to be
31
described. In particular, the extent to which the different receptor forms affect the
32
sensitivity to TCDD is not known.
23
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
Human cells contain an intracellular protein whose properties resemble those
2
of the Ah receptor in animals. Binding studies and hydrodynamic analyses have
3
identified an Ah receptor-like protein(s) in a variety of human tissues. Functional Ah
4 receptors have been found in many human tissues including lymphocytes, liver, lung,
5 and placenta. By analogy with the existence of multiple receptor forms in mice, it is
6 reasonable to anticipate that the human population will also be polymorphic with
7
respect to Ah receptor structure and function. Therefore, it is also reasonable to expect
8 humans to differ from one another in their susceptibilities to TCDD. The binding and
9 hydrodynamic properties of the Ah receptor differ relatively little across species and
10 tissues yet responses vary widely; it is difficult to, therefore, account for the diversity of
11 TCDD's biological effects by characteristics of the receptor alone.
12
The Ah receptor exists in cells as a complex of proteins. Upon binding of dioxin-
13
like compounds ( the "ligands" for this receptor), the Ah receptor dissociates from the
14
complex and interacts with a protein designated "Aryl hydrocarbon Receptor Nuclear
15 Transferase," or Arnt, forming a heterodimer (Hoffman, et al., 1991). Although originally
16 thought to participate in transfer of the dioxin-bound Ah receptor to the nucleus, more
17 recent studies suggest that Arnt is a nuclear protein that interacts with the liganded Ah
18 receptor to form a heteromeric, DNA-binding complex that can activate gene
19 transcription. Neither the ligand receptor nor Arnt exhibit substantial DNA-binding in
20 the absence of the other; the presence of both proteins is required to generate a
21 specific DNA-binding species and to activate the expression of specific genes. Both
22 the Ah receptor and Arnt belong to a class of transcription factors which function as
23 heterodimers and which contribute to the control of numerous genes (Kadesch, 1993).
24
By analogy to the multiple alleles that exist for the ligand-binding component of the Ah
25 receptor, it is reasonable to expect that the DNA-binding component of the receptor
26 will also exhibit polymorphisms and exist in multiple forms. In principle, such a
27 situation raises the possibility that different functional forms of the receptor complex
28 can exist, created by the association of receptor subunits in different combinations.
29 Such combinatorial diversity could contribute to the variety of biological responses
30
produced by TCDD.
31
The evidence to date implies that the Ah receptor participates in every
32
biological response to TCDD. A simplified diagram of this hypothesis is presented in
24
Draft - Do Not Quote or Cite . Draft May 2, 1994
1
Figure 9-2. This hypothesis predicts that TCDD will be found to activate the
2
transcription of other genes via a receptor- and enhancer-dependent mechanism
3 analogous to that described for the cytochrome P4501A1 (CYP1A1) gene. CYP1A1 is
4 one of a family of proteins involved in the activation and detoxification of both
5 endogenous and exogenous chemicals. Preliminary data from a number of
6 laboratories suggest that this is the case. For example, TCDD induces the expression
7 of the cytochrome P4501A2 gene, the glutathione S-transferase Ya subunit gene, an
8 aldehyde dehydrogenase gene, and a quinone reductase gene; in some cases,
9 induction is known to occur at the transcriptional level, to be Ah receptor-dependent,
10 and to involve a genomic regulatory element(s) analogous to that found upstream of
11 the CYP1A1 gene. In addition, recent observations suggest that, in human
12 keratinocytes, TCDD activates the transcription of plasminogen activator inhibitor-2
13 and interleukin-18, as well as other genes (Sutter et al., 1991). Recent data describe
14 the complete cDNA sequence of the mRNA of one of these genes as a new gene
15 subfamily of cytochrome P450 ( Sutter et al., 1994). The mechanism by which dioxin
16 activates the expression of these genes is currently unknown. For dioxin-responsive
17
genes other than CYP1A1, and especially for those genes that respond in tissue-
18 specific fashion, the presence of the receptor/enhancer system may not be sufficient
19 for dioxin action, and other, tissue-specific regulatory components may play a
20 dominant role in governing the response to TCDD. Thus, future research may reveal
21 the existence of additional positive or negative gene regulatory components that can
22 influence the response of the cell to TCDD.
23
Recent observations have suggested the presence of Ah-mediated changes in
24 phosphotyrosyl proteins following TCDD treatment. These changes may be due to
25 increased phosphorylation of preexisting proteins, increased synthesis of proteins that
26 are phosphorylated, decreased phosphatase activity or a combination of all three
27 mechanisms (DeVito et al., 1994). Protein tyrosine phosphorylation is known to play a
28 critical role in signal transduction and regulation of cellular events, such as entry into
29 the cell cycle. Changes in protein tyrosine phosphorylation following TCDD treatment
30 may indicate additional changes in signal transduction pathways which alone or in
31
combination with trancriptional alterations may result in altered cellular differentiation
32
or proliferation. Further research will be required to test this hypothesis and further
25
Draft . Do Not Quote or Cite - Draft May 2, 1994
1 elucidate the interactions among these regulatory processes.
2
Compensatory changes, which occur in response to TCDD's primary effects,
3 can complicate the analysis of dioxin action in intact animals. For example, TCDD can
4 produce changes in the levels of steroid hormones, peptide growth factors and/or their
5 cognate cellular receptors. In turn, such alterations have the potential to produce a
6 series of subsequent biological effects, which are not directly mediated by the Ah
7 receptor. Furthermore, the hormonal status of an animal appears to influence its
8 susceptibility to the hepatocarcinogenic effects of TCDD (Lucier et al., 1991).
9 Likewise, exposure to other chemicals can alter the developmental toxicity of TCDD
10 (Couture et al., 1990). Therefore, in some cases, TCDD may act in combination with
11 other chemicals to produce its biological effects. Such phenomena increase the
12 difficulty of analyzing dioxin action in intact animals and increase the complexity of risk
13 assessment, given that humans are routinely exposed to a wide variety of chemicals.
14
The fact that TCDD may induce a cascade of biochemical changes in the intact
15 animal raises the possibility that dioxin might produce a response such as cancer by
16 mechanisms that differ among tissues. For example, in one case, TCDD might activate
17 a gene(s) that is directly involved in tissue proliferation. In a second case, TCDD-
18 induced changes in hormone metabolism may lead to tissue proliferation secondary to
19 increased secretion of a trophic hormone. In a third case, TCDD-induced changes in
20 hormone receptors for growth factors or hormones may alter the sensitivity of a tissue
21 to proliferative stimuli. In a fourth case, TCDD-induced toxicity may lead to tissue
22 death, followed by regenerative proliferation. Thus, while this reassessment has
23 identified a number of hypothetical mechanisms for cancer induction by TCDD, there
24 remains considerable uncertainty about which mechanisms occur, with what levels of
25 sensitivity, and in which species.actually occurs, whether they would exhibit similar
26 sensitivities to TCDD, or whether they would occur in all animal species exposed to
27 the dioxin. Advances in knowledge regarding the role of such activities in dioxin
28 toxicity will facilitate the development of more definitive biologically-based models of
29 dioxin action.
30
Under some circumstances, exposure to TCDD elicits beneficial effects. For
31 example, TCDD can protect against the carcinogenic effects of polycyclic aromatic
32 hydrocarbons in mouse skin; this may reflect the induction of detoxifying enzymes by
26
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
dioxin (Cohen et al., 1979; DiGiovanni et al., 1980). In other situations, TCDD-induced
2 changes in hormone metabolism may alter the growth of hormone-dependent tumor
3 cells, producing a potential anti-carcinogenic effect (Spink et al., 1990). There is
4 considerable uncertainty about the magnitude and importance of these effects in
5 relation to both dose and response characteristics of dioxins in various species.
6 Nonetheless, these (and perhaps other) potentially beneficial effects of TCDD
7
complicate the risk assessment process for dioxin.
8
A substantial body of biochemical and genetic evidence indicates that
9 the Ah receptor mediates the biological effects of TCDD. This evidence implies that a
10 response to dioxin requires the formation of ligand-receptor complexes. TCDD-
11 receptor binding appears to obey the law of mass action and, therefore, depends upon
12 (1) the concentration of ligand in the target cell; (2) the concentration of receptor in the
13 target cell; and (3) the binding affinity of the ligand for the receptor. In principle, some
14 TCDD-receptor complexes will form even at very low levels of dioxin exposure.
15 However, in practice, at some finite concentration of TCDD, the formation of TCDD-
16 receptor complexes will be insufficient to elicit detectable effects. Furthermore,
17 biological events subsequent to TCDD-receptor binding may or may not exhibit a
18 linear response to dioxin. In some experimental systems with no direct relationship to
19 dioxin-induced responses, the induction of gene transcription appears to require a
20 threshold concentration of transcription factor(s) (Fiering et al., 1990). However, recent
21 studies in several laboratories have indicated no evidence of a threshold for relatively
22 simple responses to dioxin-like compounds such as CYP1A1 induction and others.
23 Further information will be required to determine if other responses to dioxin-like
24 compounds requiring gene transcription will also demonstrate low-dose linear
25 behavior.
26
While much of our understanding of TCDD impacts on genetic activity is
27 derived from studies on liver, studies of other tissues (e.g., skin, thymus) are likely to
28 reveal additional TCDD-responsive genes, which exhibit tissue-specific expression
29 (Sutter et al., 1991). Analyses of the mechanism of dioxin action in such systems
30
appear likely to reveal additional factors that influence the susceptibility of a particular
31 tissue to TCDD. In addition, studies of other TCDD-inducible genes, such as
32 glutathione S-transferase, quinone reductase, and aldehyde dehydrogenase, may
27
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
reveal whether differences in enhancer structure, receptor-enhancer interactions, or
2
promoter structure affect the responsiveness of the target gene to TCDD (Whitlock,
3 1990).
4
Further analyses of dioxin action may provide more insight into the mechanisms
5
by which TCDD and related compounds produce immunological effects, reproductive
6
and/or developmental effects or cancer, effects which are of particular public health
7
concern. A major challenge for the future will be the establishment of experimental
8
systems in which such complex biological phenomena are amenable to study at the
9
molecular level.
10
11
TOXIC EFFECTS OF DIOXIN
12
A.) General Comments
13
It is clear from the evaluation of the toxicologic literature that dioxin and related
14
compounds have the ability to produce a plethora of responses in animals and,
15
presumably, in humans (Table 9-2). Relatively few have been demonstrated to occur
16
in humans because of lack of knowledge regarding levels of dioxin exposure in the
17
general population, few comprehensive studies of more highly exposed populations,
18 the inherent insensitivity of epidemiologic studies, and the inability to rule out
19
confounding exposures. Evaluation of hazard and risk for dioxin and related
20
compounds must rely on a weight of the evidence approach in which all available data
21
are brought to bear on these issues. This often necessitates cross-species
22
extrapolation of effects.
23
The reliability of using animal data to estimate human hazard and risk has often
24
been questioned for this class of compounds. Although human data are limited,
25 evidence suggests that animal models are appropriate for estimating human risk if all
26
available data are considered. Humans have a fully functional Ah receptor and both in
27
vivo and in vitro studies demonstrate comparability of biochemical responses in
28
humans and animals. When comparing species and strains for their responses to
29
these compounds, a wide range of sensitivity to TCDD-induced toxicities has been
30
noted. Qualitatively speaking, however, almost every response can be produced in
31
every species if the appropriate dose is administered. Although outliers, i.e. species
32
which are either very sensitive or refractory, can be identified for a particular response,
28
Table 9-2. Effects of TCDD and Related Compounds in Different Animal Species
Guinea
Effect
Human
Monkey
Pig
Rat
Mouse
Hamster
Cow
Rabbit
Chicken
Fish
+
+
+
+
+
+
+
+
Presence of
+
+
8
AhR
+
+
+
+
+
+
+
Binding of
+
+
TCDD: AhR
Complex to
the DRE
(enhancer)
+
+
+
+
+
+
+
Enzyme
+
induction
+
+
+
+
+
+
+
+
Acute
lethality
+
+
+
+
+
+
Wasting
syndrome
Terato-
+
+
+
+
+
genesis/fetal
+/-
+
+
+
PAGE 28-a
toxicity,
mortality
+
+
+
Endocrine
+/-
effects
+
+
+
+
+
+
Immuno-
+/-
+
toxicity
+
+
+
+
Carcino-
+/-
genicity
+
+
Chlor-
+
+
acnogenic
effects
+ = observed.
+/- = observed to limited extent, or +/- results.
0 = not observed.
Table 9-2. (continued)
Guinea
Human
Pig
Rat
Mouse
1
Effect
Monkey
Hamster
Cow
Rabbit
Chicken
Fish
Porphyria
0
0
+
+
0
+
Hepato-
+
+/-
+
+
+/-
+
+
+
toxicity
Edema
+
0
0
+
+
+
+
Testicular
+
+
+
+
atrophy
Bone
+
+
+/-
+
marrow
hypoplasia
+ = observed.
+/- = observed to limited extent, or +/- results.
0 = not observed.
Page 28-6
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
no species is consistently sensitive or refractory for all effects. In addition, the majority
2
of species cluster in sensitivity for a given effect within approximately one order of
3
magnitude (factor of 10). Therefore, despite a range of sensitivities across species, it
4 is reasonable to assume that humans will not be refractory to all effects nor that they
5 will be as sensitive as the most sensitive responder for each effect. Humans are likely,
6 because of interindividual variability, which is greater than that found in individual
7 species of laboratory animals, to show a wide range of sensitivities for various dioxin-
8 induced toxicities. For purposes of the current assessment, therefore, unless there are
9 data to identify a particular species as being representative of humans for a particular
10 effect, average humans can be reasonably assumed to be of average sensitivity for
11 various effects, recognizing that individuals in the population might vary widely in their
12 sensitivity to individual effects. The uncertainty introduced by this assumption i.e. that,
13 on average, humans will respond as do average animal models for individual effects
14
of exposure to dioxin-like compounds and that an unknown range of variability exists
15 in the human population for individual effects, should be carefully considered as
16
results of this characterization are applied to individuals or specific subpopulations.
17
B.) Chloracne
18
Chloracne and associated dermatologic changes are widely recognized
19
responses to TCDD and other dioxin-like compounds in humans. Chloracne is a
20 severe acne-like condition which develops within months of first exposure to high
21 levels of dioxin. For many individuals, the condition disappears after discontinuation
22 of exposure, despite serum levels of dioxin in the thousands of parts per trillion; for
23 others, it may remain for many years. The duration of persistent chloracne is on the
24 order of 25 years although cases of chloracne persisting over 40 years have been
25 noted. There are very little human data from which to determine definitively the doses
26 at which chloracne is likely to occur. Data from occupational studies suggest that
27 persistent chloracne is more often associated with exposures of high intensity, for long
28 duration and commencing at an early age. Acute exposures, or chronic lower level
29 exposures, if resulting in chloracne, have generally resulted in a condition which
30 resolves itself in a matter of months to a few years. Details of cloracnegenic response
31
in occupationally-exposed humans are described in detail in Chapter 7 of the Health
32
Assessment document
29
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
Induction of chloracne in humans after exposure to dioxin and related
2
compounds is supported by studies in laboratory animals. Rabbits, monkeys and
3
hairless mice have all proved useful in investigating this response. In addition, cellular
4
systems provide a research tool in elucidating the chloracne response at the cellular
5
level. Keratinocytes, the principal cell type in the epidermis, have been used as an in
6
vitro model for studies of TCDD-induced hyperkeratosis, a feature of chloracne, in
7
human- and animal-derived cell cultures. The response in these systems is
8
analogous to the hyperkeratinization observed in vivo as a part of chloracne.
9
There is little doubt that chloracne is a human condition often attributable to
10
exposure to dioxin and related compounds. The specific risk factors associated with
11
this response are still obscure. Recognition of chloracne has been associated with
12
high level exposure to these compounds, and as such may represent a biomarker of
13
exposure. Because of the wide variability of the chloracnegenic response in humans
14
and its varied persistence, however, the absence of chloracne is not a reliable
15
indicator of low exposure to dioxin and related compounds.
16
17
C.) Carcinogenicity
18
Since the last EPA review of the human data base relating to the
19
carcinogenicity of TCDD and related compounds in 1988, several new followup
20
mortality studies have been completed. Among the most important of these are a
21
study of 5,172 workers by Fingerhut et al. (1991), a study with 1,583 workers by Manz
22
et al. (1991), a smaller study of 247 workers by Zober et al. (1990), and a study of over
23
18,000 workers by Saracci et al. (1991). Although uncertainty remains in interpreting
24 these studies because not all potential confounders have been ruled out and
25 coincident exposures to other carcinogens is likely, all provide support for an
26 association between exposure to dioxin and related compounds and increased cancer
27
mortality. With the exception of the study by Saracci et al. (1991), these studies have
28
some exposure information that permits an assessment of dose response. These data
29
have in fact served as the basis for fitting the additive and multiplicative risk models in
30
Chapter 8. In addition, more limited results have been presented recently on the
31
Seveso cohort (Bertazzi et al., 1993) and on women exposed to chlorophenoxy
32
herbicides, chlorophenols and dioxins (Kogevinas et al., 1993). While these two
30
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
studies have methodologic short-comings which are described in Chapter 7, they
2 provide findings, particularly for exposure to women, which warrant additional follow-
3 up.
4
While the data base from epidemiologic studies remains controversial, it is the
5 view of this reassessment that this body of evidence support the laboratory data
6 indicating that TCDD probably increases cancer mortality of several types. Although
7 not all confounders were ruled out, positive associations between surrogates of dioxin
8 exposure, either occupational or proximity to a known source combined with some
9 information on body burden, and cancer have been reported. These data alone
10 suggest a role for dioxin exposure to contribute to a carcinogenic response but do not
11 confirm a causal relationship between exposure to dioxin and increased cancer
12 incidence. Available human studies alone cannot demonstrate whether a cause and
13 effect relationship between dioxin exposure and increased incidence of cancer exists.
14 Therefore, evaluation of cancer hazard in humans must include an evaluation of all of
15 the available animal and in vitro data as well as the data from exposed human
16 populations. The Peer Panel that met in September, 1993, to review an earlier draft of
17 the cancer epidemiology chapter suggested that the epidemiology data alone were
18 still not adequate to implicate dioxin and related compounds as "known" human
19 carcinogens but that the results from the human studies were largely consistent with
20 observations from laboratory studies of dioxin-induced cancer and, therefore, should
21 not be dismissed or ignored. Other scientists, including those who attended the Peer
22 Panel meeting felt either more or less strongly about the weight of the evidence from
23 epidemiology studies, representing the range of opinion that still exists on the
24 interpretation of the cancer epidemiology studies.
25
Many of the earlier epidemiological studies that suggested an association with
26 soft tissue sarcoma (STS) were criticized for a variety of reasons. Nonetheless, the
27 incidence of soft tissue sarcoma is elevated in several of the recent studies, supporting
28 the findings from previous studies. The fact that similar results were obtained in
29 independent studies of differing design and evaluating populations exposed to dioxin-
30 like compounds under varying conditions, along with the rarity of this tumor type,
31 weighs in favor of a consistent and real association. On the other hand, arguments
32 regarding selection bias, differential exposure misclassification, confounding, and
31
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
chance in each individual study have been presented in the scientific literature which
2
increase uncertainty around this association. In addition, excess respiratory cancer
3
was noted by Fingerhut, Zober, and Manz. These results are also supported by
4 observations subsequent to the Japanese rice oil poisoning accident where exposure
5 to PCDFs and PCBs occurred. Again, while smoking as a confounder can not be
6 totally eliminated as a potential explanation of these results, analyses conducted to
7
date suggest that smoking is not likely to explain the entire increase in lung cancer.
8 The question of multiple confounders, such as exposure to asbestos and other
9 chemicals, in addition to smoking has not been entirely ruled out and must be
10
considered as potentially adding to the observed increases. Although increases of
11
cancer at other sites (e.g. non-Hodgkin's lymphoma, stomach cancer) have been
12
reported, the data for an association with exposure to dioxin-like chemicals is less
13 compelling. What emerges from an analysis of the epidemiology data is a view of
14
dioxin-like compounds as potentially multi-site carcinogens in more highly exposed
15 human populationsthat have been studied, consisting primarily of men. There are
16
currently very few data for xposed women and children. Although uncertainty in this
17
view remains, the cancer findings are generally consistent with results from studies of
18 laboratory animals, and appears to be plausable given what is known about
19
mechanisms of dioxin action.
20
While both past and more recent human studies have focused on males, there
21
are some, limited data suggesting responses in females. Because both laboratory
22
animal data and mechanistic inferences suggest that males and females may respond
23 differently to dioxin-like activity, further data will be needed to address this question.
24
An extensive data base on the carcinogenicity of dioxin and related compounds
25
in laboratory studies exists and is described in detail in Chapter 6. There is adequate
26 evidence that 2,3,7,8-TCDD is a carcinogen in laboratory animals based on long-term
27 bioassays conducted in both sexes of rats and mice. All studies have produced
28 positive results, leading to the conclusions that TCDD is a multistage carcinogen
29 increasing the incidence of tumors at sites distant from the site of treatment and at
30
doses well below the maximum tolerated dose (MTD). Since this issue was last
31
reviewed by the Agency in 1988, TCDD has been shown to be a carcinogen in
32
hamsters, which are relatively resistant to the lethal effects of TCDD. Recent data have
32
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
also shown TCDD to be a liver carcinogen in the small fish, Medaka (Johnson, et al.,
2
1992). Few attempts have been made to demonstrate the carcinogenicity of other
3
dioxin-like compounds. Other than a mixture of two isomers of
4
hexachlorodibenzodioxin (HCDDs) which produced liver tumors in both sexes of rats
5 and mice (NTP, 1980), the more highly chlorinated CDDs and CDFs have not been
6 studied in long-term animal cancer bioassays. However, it is generally recognized
7 that these compounds bioaccumulate and exhibit toxicities similar to TCDD and are,
8 therefore, also likely to be carcinogens (EPA Science Advisory Board, 1989).
9
In addition to the demonstration of TCDD as a complete carcinogen in long term
10 cancer bioassays, a number of dioxin-like PCDDs and PCDFs, as well as several
11
PCBs, have also been demonstrated to be tumor promoters in two stage (initiation-
12
promotion) protocols in rodent liver and skin. In addition, recent data have
13 demonstrated the ability of TCDD to neoplastically transform immortalized human cells
14
in culture at very low concentrations of TCDD. While dioxin and related compounds
15 are not generally considered to be "genotoxic" in traditional terms, both empirical data
16 and the results of modeling efforts suggest that they may be functioning indirectly to
17
produce irreversible genetic changes in exposed cells. All of these data add
18 substantially to the weight of the evidence that dioxin and related compounds are
19 likely to possess carcinogenic potential in humans, at least under some
20 circumstances.
21
Despite the relatively large number of bioassays on TCDD, the study of Kociba
22
et al. (1978) and those of the NTP (1982), because of their multiple dose groups and
23 large dose range, continue to be the focus of additional review. Sauer (1990) re-
24 evaluated the female rat liver tumors in the Kociba study using the latest pathology
25 criteria for such lesions. The review confirmed only approximately one-third of the
26 tumors of the previous review (Squire, 1980). While this finding has little impact on the
27
question of carcinogenic hazard, since TCDD induced tumors in multiple sites in this
28 study, it does have an effect on evaluation of dose-response and on estimates of risk
29 at low doses. These issues will be discussed in a later section of this chapter.
30
One of the more interesting findings in the Kociba bioassay was reduced tumor
31
incidences of the pituitary, uterus, mammary gland, pancreas and adrenals. These
32
findings, coupled with the sex specificity of the TCDD-induced liver tumors in rats
33
Draft - Do Not Quote or Cite . Draft
May 2, 1994
1
emphasize that the carcinogenic actions of TCDD involve a complex interaction of
2
hormonal factors. Moreover, it is hypothesized that cell-specific factors modulate
3
TCDD/hormone actions relevant to cancer. The findings of reduced tumor incidence in
4
certain tissues suggest that dioxin exposure may be exerting an anti-carcinogenic
5 effect under certain circumstances or in certain tissues. The complex interplay
6 between dioxin and hormones in terms of both carcinogenic and anti-carcinogenic
7
responses will continue to a be matter of hypothesis until such data to address these
8 issues are obtained.
9
10
D.) Reproductive and Developmental Effects
11
The potential for dioxins and related compounds to cause reproductive
12
and developmental toxicity in animals has been recognized for many years and the
13
data base regarding these effects is analyzed in Chapter 5. Recent laboratory studies
14
have suggested that altered development may be among the most sensitive TCDD
15 endpoints in laboratory animal systems. Although the discussion of these effects is
16
divided into developmental toxicity and male and female reproductive toxicity, it is
17
important to recognize the interrelatedness of developmental and reproductive events
18 at all levels of biological complexity. For example, effects of TCDD on circulating
19
levels of sex hormones and/or on responsiveness to sex hormones in laboratory
20 animals or humans may be translated into reproductive dysfunction if exposure occurs
21
in adulthood as well as abnormal development if exposure occurs perinatally.
22 Likewise, even though organ structure and growth are considered separate
23 manifestations in developmental toxicity that are associated with perinatal exposure to
24
TCDD in laboratory animals, the development of an organ in all biological systems is
25 dependent on normal growth processes and inhibiting prenatal growth can
26
significantly disrupt the structural integrity of an organ system.
27
In the current data base, developmental toxicity endpoints are observed at
28
lower TCDD exposure levels than are endpoints of male and female reproductive
29 toxicity in a number of animal systems. The lowest effective TCDD egg burden for
30 causing developmental toxicity in fish and birds and the lowest effective maternal
31
TCDD body burden for producing a wide range of developmental responses in
32
mammals are summarized in Chapter 5. Of particular interest to the risk assessment
34
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
process is the fact that a wide variety of developmental events, crossing three
2
vertebrate classes and several species within each class, can be perturbed,
3
suggesting that dioxin has the potential to disrupt a large number of critical
4
developmental events at specific developmental stages. Not only can these changes
5
lead to increases in embryo/fetal mortality, but they can disrupt organ system structure
6
and irreversibly impair organ function.
7
Since developmental toxicity following exposure to TCDD-like congeners
8
occurs in fish, birds, and mammals, it is likely to occur at some level in humans. It is
9
not currently possible to state exactly how or at what levels humans in the population
10
will respond with adverse impacts on development or reproductive function. Data
11
analyzed in Chapter 5 and Chapter 7 suggest, however, that adverse effects may be
12
occuring at levels lower than originally thought to represent "no observed adverse
13
effect levels." Related effects in human infants exposed to a complex mixture of PCBs,
14
CDFs and PCQs in the Yusho and Yu-Cheng poisoning episodes were probably
15
caused by the combined exposure to those PCB and CDF congeners that are Ah
16
receptor agonists. Similarity of the effects observed in human infants perinatally
17
exposed to this complex mixture, with those reported in adult monkeys exposed only to
18
TCDD, increases the probability of at least some of the effects in the Yusho and Yu-
19
Cheng children being due to the TCDD-like congeners in the contaminated rice oil
20
ingested by the mothers of these children. Most significant is a clustering of effects in
21
organs derived from the ectodermal germ layer, a syndrome referred to as ectodermal
22
dysplasia. Included in this syndrome are effects on the skin, nails, and meibomian
23
glands that occur in both adult monkeys exposed to TCDD and in Yusho and Yu-
24
Cheng infants exposed, transplacentally to PCB, CDF and PCQ contaminated rice oils.
25
In addition, accelerated tooth eruption has been reported both in human infants
26
affected by the Yusho and Yu-Cheng exposures and in neonatal mice exposed to
27
TCDD. Yu-Cheng children exposed to PCB, CDF and PCQ contaminated rice oil
28
transplacentally have also exhibited developmental and psychomotor delay during
29
developmental and cognitive tests. Monkeys perinatally exposed to TCDD are also
30
affected by a deficit in cognitive function. The concept that the ectodermal dysplasia
31
syndrome in Yusho and Yu-Cheng infants may be caused by the combination of PCB
32
and CDF congeners in the rice oil that are Ah receptor agonists, but are less potent
35
Draft . Do Not Quote or Cite - Draft
May 2, 1994
1
than TCDD, is consistent with structure activity results for various developmental
2
endpoints in different species of fish, birds, and mammals.
3
In mammals, postnatal functional alterations involving learning behavior and
4
the developing reproductive system appear to be the developmental events most
5
sensitive to perinatal dioxin exposure. The developing immune system may also be
6
highly sensitive. Alterations in structural endpoints and diminished prenatal viability
7
and growth begin to predominate at maternal TCDD body burdens and/or daily TCDD
8
doses during gestation that are above 100 ng/kg in virtually every species tested.
9
These doses of TCDD are not maternally toxic. Higher dose levels can be
10
demonstrated to result in prenatal mortality. A general finding in fish, bird, and
11
mammalian species is that the embryo or fetus is more sensitive to TCDD-induced
12
mortality than the adult. Thus, the timing of TCDD exposure during the life history of an
13
animal can greatly influence its susceptibility to overt dioxin toxicity.
14
With respect to male and female reproductive endpoints, there are clear effects
15
following dioxin exposure of the adult animal. Such reproductive effects generally
16
occur at TCDD body burdens that are higher than those required to cause the more
17
sensitive developmental endpoints. For example, TCDD exposure of the adult male
18
rodent causes reduced testis and accessory sex organ weights, abnormal testis
19
structure, decreased spermatogenesis, reduced fertility, decreased testicular
20
testosterone synthesis, reduced plasma androgen concentrations, and altered
21
regulation of pituitary LH secretion. However, in laboratory animal studies, these
22
effects are detectable only at TCDD exposure levels that are overtly toxic to the animal.
23
In the more limited studies focusing on female reproduction, the primary effects include
24
decreased fertility, inability to maintain pregnancy, and in the rat, decreased litter size.
25
Signs of ovarian dysfunction and alterations in hormone levels have also been
26
reported.
27
Exposure of female mice and rats to TCDD has an antiestrogenic effect. The
28
dose of TCDD required to produce this response is generally higher than that needed
29
to cause the most sensitive signs of developmental toxicity in these species. More
30
specifically, hydronephrosis and cleft palate in mice and reductions in
31
spermatogenesis in rats occur at maternal doses of TCDD which are far less than
32
those needed to exert a demonstrable antiestrogenic effect when adult female mice
36
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
and rats are exposed to dioxin. The precise mechanism of TCDD's antiestrogenic
2
effect is not fully understood. It may be caused by both a decrease in available
3
estrogen receptor number and/or by an increase in cytochrome P-4501A-mediated
4
estrogen metabolism within the target cell.
5
These studies indicate that while there is variability between species in the
6
profile of developmental responses elicited by TCDD, essentially all dioxin-like PCB,
7
CDD, and CDF congeners that have Ah receptor affinity and intrinsic activity produce
8
the same pattern of developmental effects within a given vertebrate species if a
9
sufficiently high dose of the congener is given. Data to support these conclusions
10
regarding reproductive and developmental hazards of dioxin and related compounds
11
continue to accumulate, but the weight of the evidence is still a subject of much
12
scientific debate.
13
14
E.) Immunotoxicity
15
Concern over the potential toxic effects of chemicals on the immune
16
system arises from the critical role that the immune system plays in maintaining health.
17
It is well recognized that suppressed immunological function can result in increased
18
incidence and severity of infectious diseases as well as some types of cancer.
19
Conversely, the inappropriate enhancement of immune function or the generation of
20
misdirected immune responses can precipitate or exacerbate the development of
21
allergic and autoimmune diseases. Thus, suppression as well as enhancement of
22
immune function are considered to represent potential immunotoxic effects of
23
chemicals.
24
Extensive evidence has accumulated over the past 20 years to demonstrate that
25
the immune system is a target for toxicity of TCDD and structurally related compounds,
26
including PCDDs,PCDFs,PCBs, and PBBs. This evidence is described in detail in
27
Chapter 4. The evidence has derived from numerous studies in various animal
28
species, primarily rodents, but also guinea pigs, rabbits, monkeys, marmosets, and
29
cattle. Epidemiological studies also provide some evidence for the immunotoxicity of
30
HAH in humans. In animal studies, relatively high doses of HAH produce lymphoid
31
tissue depletion, except in the thymus where cellular depletion occurs at lower doses.
32
Alterations in specific immune effector functions and increased susceptibility to
37
Draft - Do Not Quote or Cite - Draft May 2, 1994
1 infectious disease have been identified at doses of TCDD well below those which
2 cause lymphoid tissue depletion. Both cell-mediated and humoral immune responses
3 are suppressed following TCDD exposure, suggesting that there are multiple cellular
4 targets within the immune system that are altered by TCDD. Evidence also suggests
5 that the immune system is indirectly targeted by TCDD-induced changes in
6 nonlymphoid tissues. In addition, in parallel with increased understanding of the
7 cellular and molecular mechanisms involved in immunity, studies on TCDD are
8 beginning to establish biochemical and molecular mechanisms of TCDD
9 immunotoxicity.
10
The ability of an animal to resist and/or control viral, bacterial, parasitic, and
11 neoplastic diseases is determined by both nonspecific and specific immunological
12 functions. Decreased functional activity in any immunological compartment may result
13 in increased susceptibility to infectious and neoplastic diseases. In terms of risk
14 assessment, host resistance is often accorded the "bottom line" in terms of relevant
15 immunotoxic endpoints. Animal host resistance models that mimic human disease are
16 available and have been used to assess the effect of TCDD on altered host resistance.
17 Results from host resistance studies provide evidence that exposure to TCDD results
18 in increased susceptibility to bacterial, viral, parasitic, and neoplastic disease. These
19 effects are observed at relatively low doses and likely result from TCDD-induced
20 suppression of immunological function. The specific immunological functions
21 targeted by TCDD in each of the host resistance models remain to be fully defined.
22
The difficulty in demonstrating consistent, direct effects of TCDD in vitro on
23 lymphocytes, the dependence of those effects on serum components, and the
24 requirement for high concentrations of TCDD are all consistent with the potential for an
25 indirect mechanism of TCDD on the immune system. One potentially important
26 indirect mechanism is via effects on the endocrine system. Several endocrine
27 hormones have been shown to regulate immune responses, including glucocorticoids,
28 sex steroids, thyroxine, growth hormone, and prolactin. Importantly, TCDD and other
29 related compounds have been shown to alter the activity of all of these hormones.
30
It is important to consider, however, that if an acute exposure to TCDD even
31 temporarily raises the TCDD body burden at the time when an immune response is
32 initiated, there may be a risk of adverse impacts even though the total body burden
38
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
may indicate a relatively low average TCDD level. Furthermore, since TCDD alters the
2
normal differentiation of immune system cells, the human embryo may be very
3
susceptible to long-term impairment of immune function from in utero effects of TCDD
4
on developing immune tissue. There are currently no data to directly support this
5
hypothesis. Concern arises as a consequence of inferences derived from an
6
understanding of dioxin action and observations in humans and laboratory animals.
7
F.) Other Effects
8
A number of other effects of dioxin and related compounds have been
9
discussed in some detail throughout the chapters in this assessment. While they serve
10
to illustrate the wide range of effects produced by this class of compounds, some may
11
be specific to the species in which they are measured and may have limited relevence
12
to the human situation. On the other hand, they may be indicative of the fundamental
13
level at which dioxin produces its biological impact and may represent a continuum of
14
response expected from these fundamental changes. While all may not be adverse
15
effects (some may be adaptive and of neutral consequence, and some may be
16
beneficial), several effects have been noted in human studies or in primates which
17
deserve special mention:
18
Circulating Reproductive Hormones
19
Two cross-sectional epidemiologic studies have detected an association
20
between levels of reproductive hormones and exposure to TCDD. Decreased
21
testosterone levels were detected in two of the three studies where testosterone was
22
evaluated and luteinizing hormone (LH) was increased in one of the two studies
23
evaluating that endpoint. Animal data are available to support the plausability of these
24 findings. The mechanism(s) responsible for this effect are largely unknown but
25
changes in receptor level or function, and hormone metabolism and homeostasis
26
need to be investigated. If these data continue to hold up in future observations, their
27
clinical significance will need to be further evaluated. Follow-up studies are currently
28
underway.
29
Diabetes and Fasting Serum Glucose Levels
30
Epidemiologic evidence has been presented to suggest an increased risk of
31
diabetes and for an elevated prevalence of abnormal fasting serum glucose levels
32
with dioxin exposure. Three studies found that individuals with elevated serum levels
39
Draft . Do Not Quote or Cite - Draft May 2, 1994
1
of TCDD had a slight but statistically significant or borderline significant increased risk
2 for developing diabetes or having elevated fasting serum glucose. There are virtually
3
no animal data to corroborate these finding although some data have indicated effects
4 of TCDD on glucose metabolism. While the findings of a greater prevalence of
5 elevated fasting glucose may presage the development of diabetes, in the NIOSH
6 study of chemical workers, the traditional risk factors for diabetes (age, body mass
7 index or weight, and family history of diabetes) appear substantially more influential
8 than TCDD exposure in the development of the disease.
9
Enzyme induction - - One of the best characterized effects of exposure to
10 dioxin-like compounds is the induction of cytochrome P-450 1A1 (CYP1A1). CYP1A1
11 is one of a family of proteins involved in the activation and detoxification of both
12 endogenous and exogenous chemicals. Dioxin also increases the activity of a
13 number of other enzymes involved in biotransformation reactions. Increased activity of
14 these enzymes has been implicated mechanistically in the toxic responses seen in
15 animals in response to dioxin-like compounds. For example, it has been hypothesized
16 that increáses in UDP-glucuronyltransferases, which metabolize thyroxine, may lead
17 indirectly to increased Thyroid Stimulating Hormone (TSH) synthesis by the pituitary
18 and subsequent hyperplastic and hypertrophic responses by the thyroid. There is
19 speculation that such prolonged stimulation may lead to the thyroid tumors seen in
20 both rats and mice exposed to TCDD. Therefore, while changes in enzyme activity in
21 response to dioxin and related compounds may result in detoxification of certain
22 chemicals, examples exist in experimental animals of changed metabolism leading
23 directly or indirectly to adverse effects, some as severe as cancer. Data to confirm this
24 effect of dioxin and related compounds in humans are not available.
25
Gamma glutymyl transferase (GGT) activity GGT is one of the many
26 hepatic enzymes that are measured in humans to evaluate liver toxicity. Of these, GGT
27 is the only hepatic enzyme found in a number of human studies to be chronically
28 elevated in adults exposed to high levels of TCDD. The consistency of the findings in a
29 number of studies suggests that the finding may reflect a true effect of exposure but for
30 which the clinical significance is unclear. Long term, pathologic consequences of
31 elevated GGT have not been illustrated by excess mortality from liver disorders or
32 cancer or in excess morbidity in the available cross-sectional studies. There are few
40
Draft - Do Not Quote or Cite - Draft May 2, 1994
1 animal data to support these findings.
2
Endometriosis
3
Endometriosis is a serious disorder of the female reproductive system which is
4 of unknown etiology and a major cause of infertility in women. The prevalence of
5 endometriosis in the general population is unknown but is estimated to be 10% among
6 reproductive-age women, indicating that endometriosis may be present in 6.6 million
7 women in the U.S. (Wheeler, 1992). Recent studies have determined that chronic
8 exposure to TCDD increases the risk of endometriosis in rhesus monkeys (Rier et
9 al,1993). The severity of the disease was dependent on the dose given. Previous
10 work has described an association between endometriosis in rhesus monkeys and
11 exposure to polychlorobiphenyl (PCB) compounds (Campbell et al, 1985). Additional
12 studies are underway which may confirm these observations in rhesus monkeys and
13 studies are planned to evaluate women exposed at Seveso for any correlation
14 between dioxin body burden and incidence or severity of endometriosis. Further
15 evaluation of this important health endpoint awaits reports from these studies.
16
17 DOSE-RESPONSE CONSIDERATIONS
18
The current efforts to evaluate the risks of dioxin and related compounds have
19 focussed on the understanding of the biological basis of response as well as
20 evaluation of the weight of the empirical observations on inferences regarding hazard
21 and risk. Previous sections have discussed the relationship of binding of this class of
22 compounds to a specific receptor and subsequent events. It is generally accepted that
23 all well-studied responses to dioxin appear to be mediated by receptor binding. This
24 situation is not unlike the signal transduction pathways which have been described for
25 hormone action, particularly exemplified by the well studied family of steroid
26 hormones, although the dioxin receptor does not belong to the steroid receptor family.
27
As with the steroid hormones, the earliest events in the biochemical signal
28 transduction process are likely to be linearly related to ligand concentration. The fact
29 that much of the biological activity of this class of compounds follows the rank order of
30 binding affinity of the congeners to the Ah-receptor supports the concept that these
31 earliest steps play a determining role in the probability that later responses will occur.
32 This does not suggest that a simple proportional relationship between receptor
41
Draft . Do Not Quote or Cite . Draft May 2, 1994
1
binding and biological response can explain the diversity of biological responses
2 described for dioxin and related compounds. It is likely that differences in response
3
will be due to tissue and cell-specific factors that modulate the qualitative relationship
4 between receptor binding, or more precisely, occupancy and response. It is expected
5 that there may be markedly different dose response relationships for different effects of
6 dioxin depending on the respective roles of modulating activities. Coordinated
7 biological responses, such as TCDD-mediated increases in cell proliferation, likely
8 involve other cellular factors and hormone systems. This means that the dose-
9 response for relatively simple sequelae of the early binding events such as
10 cytochrome (CYP1A1) induction may not accurately predict dose-response
11 relationships for more complex responses such as cancer. Much additional
12 knowledge will be required before we can more accurately predict these complex
13
dose-response relationships.
14
Development of biologically-based dose response models for dioxin and
15 related compounds as a part of this reassessment has led to considerable and
16 valuable insights regarding both mechanisms of dioxin action and dose response
17 relationships for dioxin effects. These are described in some detail in Chapter 8:
18 These efforts have not resulted in an alternative model to replace the linearized
19 multistage (LMS) procedure for estimating cancer potency or the uncertainty factor
20 approach for estimating levels below which non-cancer effects are not likely to occur.
21 These efforts have, however, provided additional perspectives on these traditional
22 methods and have provided a biological-based rationale for what had been primarily
23 statistical approaches. The development of models allows for an iterative process of
24 data development and hypothesis testing. These efforts will result in incorporation of
25 more of the available biological data into models to predict human risk at low
26 increments of exposure.
27
Table 9-2 summarizes estimated body burdens and effect levels for a variety of
28 species, including the low observed effect levels (LOELs) for some of the more
29 sensitive indicators of biological response induced by dioxin and related compounds.
30
Important assumptions used in deriving these values are included as part of this Table.
31
It is particularly important to note that the estimated body burdens associated with
32
several of these doses are quite low relative to background body burdens in the
42
Draft - Do Not Quote or Cite . Draft May 2, 1994
1
general human population. The implications of this observation will be discussed later
2 in this chapter. [Note to reviewers: This Table will be modified in the
3 external review draft to make it easier to understand. It is included in its
4 entirety here for your comment.]
5
Comparison of recent cancer modeling efforts using rodent data with the LMS
6 procedure show no compelling arguments for use of alternative slope factors to
7 estimate upper bounds on potential human cancer risk. All of these methods, when
8 incorporating data from the most recent pathology re-evaluation of the Kociba rat
9 study, result in upper bound estimates of a one in a million (10-6) risk specific dose of
10 approximately .01 pg TEQ/kg bw/day and an upper bound unit risk estimate of
11 approximately 1 X 10-4 per pg/kg bw/day. Analysis of human data from several
12 epidemiology studies yield similar, but slightly higher, estimates, although lack of
13 sufficient knowledge regarding human hazard, exposure and potential confounders
14 makes these estimates highly uncertain. Modeling efforts have indicated the
15 sensitivity of certain model parameters to choice of data sets and/or assumptions.
16 Particularly with regard to the slope of the response for surrogate markers of low dose
17 response such as enzyme induction or indirect mutagenic activity, estimates of cancer
18 risk are highly dependent on these assumptions and could predict very different,
19 generally lower, risks if other parameters are shown to be more appropriate.
20
An additional consideration regarding the evaluation of dose-response for
21 dioxin and related compounds involves the ubiquity of background exposure to these
22 compounds. Body burdens of these compounds have been discussed previously in
23 several parts of this assessment. In all studies, both in laboratory animals and in
24 humans, incremental exposures are being added onto an existing body burden which
25 is present at birth and appears to increase with age. This background is often
26 insignificant from the standpoint of added dose in experimental studies or for highly
27 exposed human cohorts. On the other hand, it has real implications relative to the
28 detectability of response at low incremental exposures and may have implications for
29 the use of models which assume additivity to ongoing processes which may have
30 been stimulated by background levels. Modeling estimates suggest that, if dioxin and
31 related compounds are adding to human cancer burden, current background
32 exposures may result in upper bound population cancer risk estimates in the range of
43
Table 9-3
1
ESTIMATED BODY BURDENS OF EXPERIMENTAL ANIMALS AND HUMANS EXPOSED
TO LOW EFFECT LEVELS OF 2,3,7,8-TCDD.
EXPERIMENTAL
EFFECT
SPECIES
DOSE
BODY BURDEN
REF/note
CHLORACNE
HUMANS
36-3,000 ng/kg
1,2/a
CHLORACNE
MONKEY
1,000n g/kg
1,000 ug/kg
3/b
CHLORACNE
RABBITS
4 ng/kg
5d/wk/4wk
220 ng/kg
4/C
CHLORACNE
MICE
5,000 ug/kg
17,000 ng/kg
5/d
3d/wk/2wk
DECREASE
TESTOSTERONE
HUMANS
13 ng/kg
6/e
DECREASE
TESTOSTERONE
RATS
12,500 ng/kg
10,200 ng/kg
7/£
sac day 7
ALTERED
GLUCOSE
TOLERANCE
HUMANS
110 ng/kg
8/g
ALTERED
GLUCOSE
TOLERANCE
HUMANS
14 ng/kg
9/h
DECREASE
GLUCOSE
UPTAKE
GUINEA
30 ng/kg
ADIPOCYTES
PIGS
sac day 1
30 ng/kg
10/i
DECREASE
SERUM
100 ng/kg/d
GLUCOSE
RATS
30 days
1,900 ng/kg
11/j
DECREASE
BIRTH
HUMANS
Mother body
WEIGHT
burden
1,400 ng/kg
12/k
1,400 ng/kg
DECREASE
GROWTH
HUMANS
47 ng/kg
13/1
DECREASE
GROWTH
RATS
125 ng/kg/d
maternal dose
gd day 6-15
1,250 ng/kg
14/m
2
EXPERIMENTAL
EFFECT
SPECIES
DOSE
BODY BURDEN
REF/note
DECREASE
GROWTH
RATS
400 ng/kg
maternal dose 400 ng/kg
15/m
gd 15
ALTERED
LYMPHOCYTE
RHESUS
25 ppt in diet
SUBSETS
MONKEYS
for 4 years
270 ng/kg
16/n
ALTERED
MARMOSETS
LYMPHOCYTE
0.3 ng/kg/wk
17/0
SUBSETS
for 24 weeks 6-8 ng/kg
1.5 ng/kg/wk
for 12 weeks
ENHANCED
VIRAL
MICE
10 ng/kg
SUSCEPTIBILITY
sac day 7
7 ng/kg
18/p
ENDOMETRIOSIS
MONKEYS
5 ppt in diet
4 years
27 ng/kg
19/n
DECREASED
64 ng/kg
SPERM
maternal dose 64 ng/kg
20/m
COUNT
RATS
gd 15
CANCER
HUMANS
100-7,000 ng/kg 21,22,23/q
CANCER
HAMSTERS
100 ug/kg
6 doses
Loom
24/I
(600 ug/kg
total dose)
CANCER
RATS
100 ng/kg/d
for 2 years
1,400 ng/kg
25/s
TUMOR
PROMOTION
RATS
125 ng/kg/d
30 weeks
24,000 ng/kg
26/t
3
EXPERIMENTAL
EFFECT
SPECIES
DOSE
BODY BURDEN
REF/note
CANCER
MICE
7.5 ng/kg/wk
SKIN TUMOR
for 20 wks
1,100 ng/kg
27/u
PROMOTION
dermal exposure
DOWN REGULATION
OF EGFR IN
HUMANS
1,400 ng/kg
12/k
PLACENTA
(MAXIMAL
EFFECT)
DOWN REGULATION
OF EGFR IN
RATS
125 ng/kg/d
LIVER
30 weeks
24,000 ng/kg
28/t
(MAXIMAL
EFFECT)
INCREASE
HUMANS
1,400 ng/kg
12/k
IN
PLACENTAL
CYP1A1
(MAXIMAL
EFFECT)
INCREASE
RATS
125 ng/kg/d
LIVER
30 weeks
24,000 ng/kg
29/t
CYP1A1
(MAXIMAL
EFFECT)
ENZYME
INDUCTION
RATS
1 ng/kg
1 ng/kg
30/v
CYP1A1
single dose
(LOEL)
sac 24 hr
ENZYME
1.5 ng/kg/d
INDUCTION
5 d/wk 13 wk
23 ng/kg
31/w
CYP1A1/1A2
MICE
(LOEL)
BACKGROUND
HUMAN
60 TEQ ppt
9 ng/kg
X
in serum
BACKGROUND
MOUSE
4 ng/kg
y
4
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exposure of male rats to 2,3,7,8-tetrachlorodibenzo-p-dioxin.
III Effects on spermatogenesis and reproductive capability.
6
Toxicol. Appl. Pharmacol. 114:108.
21 Fingerhut, M.A., Halpern, Q.E., Narlow, B.S., Piacetelli,
L.A., Honchar, P.A., Seeney, M.H., greife, A.L., Dill, P.A.,
Steenland, K., and Suruda, A.J. (1991). Cancer mortality in
workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin N.
Engl. J. Med. 342:212.
22 Bertazzi, P.A., Pesatori, A.C., Consonni, D., Tironi, A.,
Landi, M.T., and Zocchetti, C. (1993). Cancer incidence in a
population accidentally exposed to 2,3,7, 8-tetrachlorodibenzo-
p-dioxin. Epidemiology 4:398.
23 Rao, M.S., Subbaro, V., Prasad, J.D., and Scarpelli, D.C.
(1988). Carcinogenicity of 2,3,7,8-tetrachlorodibenzo-p-
dioxin in the Syrian hamster. Carcinogenesis. 9 (9) 1677-
1679.
24 Kociba, R.J., Keyes, D.G., Beyer, J.E., Carreon, R.M., Wade,
C.E., Dittenber, D.A., Kalnins, R.P., Frauson, L.E., Park,
C.N., Barnard, S.D., Hummel, R.A., and Humiston, C.G. (1978)
Results of a two-year chronic toxicity and oncogenicity study
of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rats. Toxicol.
Appl. Pharmacol. 46:279.
25 Maronpot, R.R., Foley, J.F., Takahashi, K., Goldsworthy, T.,
Clark, G., Tritscher, A., Portier, C., and Lucier, G. (1993).
Dose response for TCDD promotion of hepatocarinogenesis in
rats initiated with DEN: histologic, biochemical and cell
proliferation endpoints. Environ. Health Perspect. 101:634-
642.
26
Poland, A., Palen, D. and Glover, E. (1982). Tumor promotion
by TCDD in skin of HRS/J hairless mice. Nature 300:271.
27 Sewall, C.
28 Tritscher, A.M., goldstein, J.A., Portier, C.J., McCoy, Z.,
Clark, G.C., and Lucier, G.W., (1992). Dose response
relationships for chronic exposure to 2,3,7,8-
tetrachlorodibenzo-p-dioxin in a rat tumor promotion model:
Quantification and immunolocalization of CYP1A1 and CYP1A2 in
the liver. Cancer Res. 52:3436-3442.
29 Van den Heuvel, J.P., Clark, G.C., Kohn, M.C., Tritscher,
A.M., Greenlee, W.F., Lucier, G.W., and Bell, D.A. (1994).
Dioxin-responsive genes: Examination of dose-response
relationships using quantitative reverse transcriptase-
7
polymerase chain reaction. Cancer Res. 54:62-68.
30 DeVito, M.J., Ma, X., Babish, J.G., Menache, M., and Birnbaum,
L.S. (1994). Dose-response relationships in mice following
subchronic exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin:
CYP1A1, CYP1A2, estrogen receptor and protein tyrosine
phosphorylation. 124:82-90.
31 Ganong W.F. (1982). Review of Medical Physiology. 11th
edition, Lange Medical Publications, Los Altos CA
32 Diliberto, J.J., Akube, P.I., Luebke, R.W., and Birnbaum, L.S.
(Submitted) Dose-respone relationships of tissue distribution
and induction of CYP1A1 and CYP1A2 enzymatic actitivities
following acute exposure to 2,3,7, 8-tetrachlrodibenzo-p-dioxin
(TCDD) in mice.
33 Rose, J.Q., Ramsey, J.C., Wentzler, T.H., Hummel, R.A., and
Gehring, P.J. (1976). The fate of 2,3,7,8-
tetrachlorodibenzo-p-dioxin following single and repeated oral
doses to the rat. Toxicol. Appl. Pharmacol. 36:209.
34 Beck, H., Dross, A., and Mathar, W. (1994) PCDD and PCDF
exposure and levels in humans in Germany. Environ. Health
Perspect. 102: suppl 1:173-185.
8
FOOTNOTES
a
The two values presented from this data are from the persons
with chloracne who had the lowest exposure (2) and the average
level of persons with chloracne from Yu Cheng (1) Estimates of
body burden for the average Yu Cheng pateint with chlroacne were
determined by authors (1). In the patient with the lowest value,
adipose tissue levels at the time of exposure are estimated by the
authors (2) assuming a half-life for TCDD of 5.8 years and are
expressed as pg/g of lipid. Body burdens are estimated from serum
levels at time of exposure (2) assuming that all TCDD in the body
is equally distributed in the lipid of the body. The average
worker is assumed to be a male weighing 70 kg with 15% of the
weight as lipid (31).
b
Animal administered lug/kg TCDD and it is assumed that
essentially no TCDD was eliminated when the animal developed a
chloracnegenic response.
C
Assumes the same rate of elimination as the rat and that the
animals weights 2.5 kg throughout the experiment.
d
Assumes a half-life of 11 days and an average weight of the
animal at 25 grams.
From reference (6) in which workers with levels of TCDD of 76
ppt e in serum or higher had lower testosterone levels. Also assumed
that the background TEQ was 60 ppt so that the total serum TEQ was
136 ppt (lipid adjusted). Average worker was male weighing 70 kg
with 15% body fat.
f sacrificed 7 days after dosing. Assumes a half-life of 23.4 days
Animals received single exposure of 12.5 ug/kg (LOAEL) and
and body burden corrected for elimination.
g
Same assumptions in e except average serum levels in affected
workers is 640 ppt.
h
From Ranch Hand study (8), assumes that high exposed group
(>33 ppt) had a background of 60 TEQ ppt. Thus this group had at
least 93 TEQ ppt. Assumes average ranch hand patient was male
weighing 70 kg with 15% body fat.
hours i after dose. Assumes that no TCDD was eliminated at this
Guinea pigs received 0.03 ug TCDD/kg ip. and sacrificed 24
time.
Animals were treated with 0.1 ug/kg/day for 30 days and
j assumes half-life of TCDD in the rat is 23.4 days.
k
According to the author (12), there is a decrease in birth
weights of children born from these patients and that the epidermal
9
growth factor receptor (EGFR) and CYP1A1 are maximally affected in
these patients. Body burdens determined based on levels of
2,3,4,7,8-pentachloro-dibenzofurar (TEF = 0.1) and 1,2,3,4,7,8-
hexachlorodibenzofuran in placenta tissue. Assumes placenta is 1%
lipid (34) and that women have a fat content of 21% of body weight
(31).
1
Body burdens estimated from serum levels presented by authors
(6). The authors (6) published that the average body weight for
the children was 30 kg with 25% of the weight as body fat. All the
dioxins are assumed to be equally distributed in the body fat.
m
Assumes pups exposed to an equal dose of TCDD as are the dams
on a weight basis and that the pups do not eliminate any of the
TCDD.
n
Assumes a single first-order elimination rate constant and a
a half-life for the whole body elimination of 400 days (3) and a
gastrointestinal absorption of 86% (33).
Assuming a single first-order elinination rate constatne and
O a half-life of 6-8 wks. Body Burdens calculated by authors (17).
Body burden determined in these animals (32). Approximately
P 70% of the body burden remains at 7 days after dosing.
Estimated highest body burden at time of last exposure.
q calculations based on measured TCDD levels in serum (lipid
adjusted) and assuming a first-order elimination kinetics and a
half-life for eliniationof 7.1 years. Also assumes a vody weight
of 70 kg and 22% body fat. Calculations for estimated serum
concentrations at last time of exposure performed by authors, not
adjusted for background levels.
Animals administered 100 ug/kg 6 times over a 4 week period.
Assumes r a half-life of 23.4 days and that animals are sacrificed at
10 months after the first dose.
Assumes a single first-order elimination rate constant and a
half-life S for the whole body elimination of 23.7 days (33) and a
gastrointestinal tract absorption of 86% (33).
t
Liver levels measured in study at approximately 300 ppb (lipid
adjusted). Also assumes animal is 10% body fat by weight.
u
Assumes an elimination rate of 11 days and a body weight of 20
grams.
V
Animals received a single dose and were sacrificed 24 hours
later. Assumes no TCDD eliminated at this time.
3
Animals received 1.5 ng/kg/d 5d/wk for 13 wk. Animals
10
sacrificed 3 days after last dose. Hepatic, dermal and pulmonary
EROD activity induced at this dose. Tissue levels measured in
liver, skin and fat. Assumes that this is the LOEL and that 100%
of the dose is in liver, skin and fat.
X
Assumes a background TEQ of 60 ppt for dioxins, dibenzofurans
and PCBs. Also assumes a body weight of 70 kg with 15% body fat.
y
Data from DeVito and Birnbaum. TEQ for TCDD, 1,2,3,7,8-PCDD;
2,3,7,8-TCDF; 1,2,3,7,8-PCDF; 2,3,4,7,8-PCDF; and OCDF in 150 day
old female B6C3F1 mice. Chemicals were determined in liver, fat
and skin of these animals. Assumes that 100% of the body burden is
in liver, fat, and skin.
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May 2, 1994
1
one in ten thousand (10-4) to 1 in a thousand (10-3) attributable to exposure to dioxin
2
and related compounds. Actual risk for individuals in the population is likely to be less
3
and, for some, may even be zero.
4
Background levels also complicate the evaluation of "No Observed or Low
5
Observed Adverse Effect Levels" (NOAELs or LOAELs). Incremental exposures must
6
be considered in light of existing body burdens in determining whether increased
7
probability of effects having biological thresholds are likely. The concept that an
8
incremental exposure is below an experimental threshold is moot unless the
9
combined background and incremental exposure or dose are below the threshold
10
level. This has important consequences for the assessment of compounds like dioxin
11
where certain effects can be detected at or near equivalent human background body
12
burden levels.
13
14
KEY ASSUMPTIONS AND INFERENCES
15
One of the primary functions of the risk characterization is to present key
16
assumptions and inferences which are used to reach conclusions in the absence of
17
definitive information. Not all scientists may agree with the use of these specific
18
assumptions and inferences. The degree to which thereis disagreement will have
19
profound effects on the acceptance of this analysis. While many of these assumptions
20
and inferences are discussed in previous sections, it is important that they be
21
recognized in order to put our overall conclusions in a proper perspective. Some of
22
the key assumptions and inferences are:
23
The limited information on sources, fate and transport in the environment
24
provide a reasonable basis for predicting human exposure. While data are limited
25
and, therefore, uncertain, information from a variety of studies in industrialized
26
countries coupled with our detailed knowledge of physico-chemical properties for this
27
class of compounds allows reasonable assumptions to be made regarding relative
28
ranking of sources with regard to their contribution to environmental loading, the
29
persistence of this class of compounds under specific environmental conditions and
30
the likelihood that the chemical will be transferred from the environment to biological
31
systems. Nonetheless, these are assumptions which are arguable and which will be
32
refined as more data become available.
44
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May 2, 1994
1
The air to food hypothesis is plausible and is supported by enough data to
2
warrant its use in the absence of more complete information. The air-to-food
3
hypothesis is founded on data evaluating deposition, environmental transport,
4
bioaccumulation and consumption patterns. It is supported by studies from Europe and
5
Canada. While individual measurement data are still quite limited, the consistency of
6
the evidence supporting the validity of the hypothesis is compelling. The hypothesis
7
has been accepted by a large segment of the knowlegable scientific community.
8
Because airborne dioxin may come from direct releases to air or from re-cycling of
9
dioxin-like compounds released into various environmental media from a number of
10
sources, this hypothesis provides a perspective on how dioxin-like compounds move
11
through the environment to humans but does not allow attribution of exposure to
12
particular sources.
13
Toxicity equivalence (TEQs) is a valid, interim method for assessing exposure
14
to a complex mixture of dioxin and related compounds and predicting likely health
15
outcomes. The EPA and the international scientific community have agreed that the
16
use of toxicity factors (TEFs) to predict relative toxicities of mixtures of this class of
17
compounds has an adequate empirical basis, is thoeretically sound, and, in the
18
absence of more complete data sets on the toxicity of individual members of this class,
19
is a useful procedure. This is not to say that the use of TEFs is a certain procedure.
20
Since 1986 when the first Agency-wide consensus on the use of TEFs was published,
21
additional refinements to the data bases and to the use of TEFs have occurred.
22
Published revisions in accord with international agreement appeared in 1989. In the
23
course of this reassessment, critical data were collected and agreement was reached
24
regarding the contribution of dioxin-like PCBs to overall TEQs. Additional validation of
25
the TEQ concept in predicting effects of this class of compounds on wildlife species
26
lends further support to the use of this approach. It must be recognized that this
27
relatively simple, additive approach does not take into account interactions between
28
dioxin-like compounds and other chemical exposures. These interactions may result
29
in either an overestimate or an underestimate of likely effects of the complex mixture.
30
Use of one-half the non-detect level for estimating low levels of exposure is a
31
reasonable but conservative approach to evaluating limited blood and tissue level
32
data. For some data sets, use of zero values for non-detects could result in
45
Draft - Do Not Quote or Cite - Draft May 2, 1994
1 significantly lower estimates. However, it is widely held that such an approach would
2 most likely underestimate true levels of exposure. Similar estimates derived from
3 different data sets, developed by different investigators in several countries, strengthen
4 the probability that this inference represents the true picture for exposure of the
5 general population in industrialized countries to dioxin and related compounds.
6
The limited data available from studies of levels of dioxin and related
7 compounds in humans provides an adequate basis to infer general population body
8 burdens. Although there are still limited measurements of general population body
9 burdens, the data provide a consistent picture of background body burdens for
10 industrialized countries. While additional data will help to refine the range of general
11 population body burdens as a function of location, human activity, age and the like,
12 there are adequate data to estimate current body burdens in the general population for
13 the purposes of this assessment. If estimates were to change with new data, it is not
14 likely that we would be far offand it is highly unlikely that these estimates would
15 represent a sensitive parameter in estimating margins-of-exposure within an order of
16 magnitude.
17
Laboratory animal studies provide useful information in evaluating potential
18 human responses to dioxin and related compounds. Based on our knowledge of the
19 biochemical and biological similarities between laboratory animals and humans, our
20 understanding of some of the fundamental impacts of this class of compounds on
21 biological systems, and comparable responses from animal and human studies both
22 in vitro and in vivo, our decision to use laboratory animal data to contribute to weight-
23 of- the-evidence conclusions on human hazard and risk is reasonable. Humans do
24 not appear to be an outlier for dioxin effects, that is, they do not, on average, appear to
25 be either refractory to or exquisitly sensitive to the effects of dioxin-like compounds.
26 While positive human data is preferable for ascribing hazard or risk, the lack of
27 adequate human data to demonstrate causality for many suspected dioxin effects is
28 assumed not to negate the findings from laboratory animal and in vitro studies.
29 Although some scientists may disagree, in our estimation, the data base on dioxin and
30 related compounds is one of the most comprehensive among all environmental
31 chemicals. The fundamental understanding of mechanisms of dioxin action provides a
32 unifying theory for the mechanisms for observed effects in laboratory animals and
46
Draft - Do Not Quote or Cite - Draft
May 2, 1994
1
humans, and for using a weight-of-the-evidence approach considering all relevent
2
data to infer the human health impacts of dioxin and related compounds.
3
Observations of effects from exposure to dioxin and related compounds in
4
humans and other animals suggest that fundamental changes in cellular biochemistry
5
and biology may be related to frankly adverse effects which can be more readily
6
observed at higher levels of exposure. Observations described in this assessment
7
suggest a continuum of response to exposure to dioxin-like chemicals. This
8
continuum provides a basis for inferring a relationship between some early events
9
which are not necessarily considered to be adverse effects with later events which are
10
adverse effects. Considerable uncertainty remains in inferring how these events are
11
related, although we know more about how dioxin-like compounds may elicit effects
12
than we know about the mechanisms of action for most chemicals. This inference may
13
be the most contentious of all and it is likely that a wide range of opinion will be
14
provided by the scientific community regarding the relationship of these mechanistic
15
observations and prediction of potential for adverse effects in exposed humans.
16
17 OVERALL CONCLUSIONS REGARDING THE IMPACT OF DIOXIN AND
18
RELATED COMPOUNDS ON HUMAN HEALTH
19
20
Dioxin exposure from multiple sources may result in a number of
21
biochemical and biological effects in both humans and other animals,
22
many of which are considered adverse or toxic effects, and some of
23
which occur at very low levels of exposure. A large variety of sources of dioxin
24
have been identified and others may exist. Because dioxin-like chemicals are
25
persistent and accumulate in biological tissues, particularly in animals, the major route
26
of human exposure is through ingestion of foods containing minute quantities of dioxin-
27
like compounds. This results in wide-spread exposure of the general population of
28
industrialized countries to dioxin-like compounds. Certain sub-populations may be
29
exposed to additional increments of exposure by being in proximity to point sources or
30
because of dietary practices. Some of the effects of dioxin and related compounds
31
have been observed in laboratory animals and humans at or near levels to which
32
people in the general population are exposed. Other effects are detectable only in
47
Draft - Do Not Quote or Cite - Draft May 2, 1994
1
highly exposed populations, and there may or may not be a likelihood of response in
2
individuals experiencing lower levels of exposure. Evaluation of effects in this health
3
assessment document are based on the concept that lipid adjusted serum levels
4
approximate the body burden of dioxin and related compounds, and that there will be
5
a dose-response relationship between effects and body burden. Adverse effects
6
associated with temporary increases in dioxin blood levels based on short term high
7
level exposures, such as those that might occur in an industrial accident scenario or
8
infrequent contact with highly contaminated environmental media, may be dependent
9
on exposure coinciding with a window of sensitivity of biological processes. It is
10
reasonable to assume that developing organisms may be particularly sensitive to
11
adverse impacts from fluctuations in exposure levels. Such exposures may also lead
12
to higher tissue levels over the long term because of the long half-life for elimination of
13
dioxin and related compounds.
14
The scientific community has identified and described a common
15
initiating mechanism that may account for most if not all of the observed
16 effects in vertebrates including humans. This mechanism involves binding of
17 dioxin-like compounds to a cellular receptor called the "Ah receptor." Binding to the
18
Ah receptor appears to be necessary for all well-studied effects of dioxin but is not
19
sufficient to elicit these responses. Receptor binding represents the first step in a
20 cascade of events attributable to exposure to dioxin-like compounds including
21
biochemical, cellular and tissue-level changes in normal biological processes. The
22 effects elicited by exposure to 2,3,7,8-TCDD are shared by other chemicals which
23 have a similar structure and Ah receptor binding characteristics. Consequently, the
24 biological system responds to the cumulative exposure of Ah receptor-mediated
25 chemicals rather than to the exposure to any single dioxin-like compound. The
26 concept of toxicity equivalence within this class of compounds and the use of toxicity
27 equivalence factors (TEFs) is widely accepted by the scientific community. While
28 some uncertainty remains with regard to the additivity of complex mixtures of these
29 compounds and with the impacts of co-exposure to non-dioxin-like compounds, the
30 use of this approach is consistent with the Agency's guidance on the evaluation of
31
complex mixtures in the absence of data on the impact of the actual mixture. This
32
approach to the evaluation of dioxin and related compounds represents one of the
48
Draft - Do Not Quote or Cite . Draft May 2, 1994
1 best studied and most widely accepted applications of this guidance although
2 additional validation studies to reduce uncertainty would be welcome.
3
There is adequate evidence from studies in human populations as
4 well as in laboratory animals and from ancillary experimental data to
5 support the inference that humans are likely to respond with a plethora of
6 effects from exposure to dioxin and related compounds. These effects will
7 likely range from adaptive changes at or near background levels of exposure which
8 may be adverse or may be beneficial, to adverse effects with increasing severity as
9 exposure increases above background levels. Induction of activating/metabolizing
10 enzymes, for instance, can lead to increases in reactive intermediates and may
11 potentiate toxic effects, or may lead to more rapid metabolism and elimination of
12 potentially toxic compounds. Demonstration of examples of both of these situations is
13 available in the published literature. The mechanistic relationships of biochemical and
14 cellular changes seen at very low levels of exposure to production of adverse effects
15 detectible at higher levels remains uncertain and controversial.
16
Individual species vary in their sensitivity to any particular dioxin effect.
17 However, the evidence available to date indicates that humans most likely fall in the
18 middle of the range of sensitivity for individual effects among animals rather than at
19 either extreme. In other words, evaluation of the available data suggest that humans,
20 in general, are neither extremely sensitive nor refractory to the individual effects of
21 dioxin-like compounds. Human data provide direct or indirect support for evaluation of
22 likely effect levels for several of the endpoints discussed in previous sections, although
23 the influence of variability among humans remains difficult to assess. Discussions in
24 previous chapters have highlighted certain prominent biologically significant effects of
25 TCDD and related compounds to emphasize some of the more sensitive indicators of
26 toxicity in animals and also, potentially, in humans. These endpoints have been
27 shown to be affected by TCDD, but specific data on the endpoints of concern do not
28 generally exist for other congeners. Concern for these effects based on the concept of
29 toxicity equivalence remains, however, for all dioxin-like compounds.
30
In humans, subtle changes in enzyme activity indicating liver
31 changes, in levels of circulating reproductive hormones in males, in
32 reduced glucose tolerance, and in cellular changes related to immune
49
Dr ft - Do Not Quote or Cite - Dr ft M y 2, 1994
1 function suggest the potential for adverse impacts on human metabolism,
2 reproductive biology and immune competence at or within one order of
3 magnitude of average background body burden levels. Average human
4 daily intakes of dioxin and related compounds, including the dioxin-like PCBs, are in
5 the range of 3-6 pg TEQ/ kg BW/day. This results in average body burdens estimated
6 to be in the range of 30-60 pg TEQ/g lipid(30-60 ppt) or 5-10 ng/kg body weight. The
7 effects described above are seen at or just several fold above these average levels.
8 Since exposures within the general population are thought to be log-normally
9 distributed, individuals at the high end of the general population range may be
10 experiencing some of these effects. Some more highly exposed members of the
11 population may be at risk for a number of adverse effects including developmental
12 toxicity, reduced reproductive capacity in males based on decreased sperm counts,
13 higher probability of experiencing endometriosis in women, reduced ability to
14 withstand an immunological challenge and others. This inference is supported by
15 observations in animals, by some human information from highly exposed cohorts and
16 by scientific inference. Fortunately, there have been few human cohorts identified with
17 exposures in the high end of this range. While the lack of adequate human
18 information and the insensitivity of epidemiologic studies makes validation of these
19 inferences difficult, they are not unreasonable given the weight-of-the-evidence from
20 available data. They represent testable hypotheses which may be strengthened by
21 further data collection.
22
The background levels in humans described above would be well within a
23 factor of 100 of levels representing low observed adverse effect levels (LOAELs) in
24 laboratory animals. For several of the effects noted in animals, a "margin of exposure"
25 (MOE) of less than an order of magnitude, based on intake levels or body burdens, is
26 likely to exist. A MOE is calculated by dividing the human-equivalent animal LOAEL or
27 no observed adverse effect level (NOAEL) with the human exposure level. The original
28 basis for MOE calculations was the observation that exposure in the range of 1-10 ng
29 TEQ/kg/day represented a no observed adverse effect level (NOAEL) for a sensitive
30 non-cancer endpoint and, therefore, that an intake of up to 10 pg TEQ/kg/day might
31
represent an adequate MOE for all other non-cancer effects. Recent data suggest that
32 "high end" average exposures in the general population are likely to approach this
50
Dr It - Do Not Quote or Cite . Dr It M y 2, 1994
1 intake level and that several effects, both subtle and frank, can be demonstrated to
2 occur in animals at intake values significantly lower than 1-10 ng TEQ/kg/day. It is,
3 therefore, highly unlikely that a margin of exposure (MOE) of 100 or more currently
4 exists for these effects at background intake levels, at least for some members of the
5 human population. We need to continue to monitor trends in human body burden for
6 dioxin and related compounds. If levels are declining, the relationship of background
7 body burdens to observed effect levels in animal and human studies will need to be re-
8 evaluated.
9
The USEPA has frequently defined a reference dose (RfD) for toxic chemicals to
10 represent a scientific estimate of the dose below which no appreciable risk of non-
11 cancer effects is likely to occur following chronic exposures. In the case of dioxin and
12 related compounds, calculation of an RfD based on human and animal data and
13 including standard uncertainty factors to account for species differences and sensitive
14 subpopulations would result in reference intake levels on the order of 10-100 times
15 below the current estimates of daily intake in the general population. For most
16 compounds where RfDs are applied, background exposures are generally low, are not
17 persistent and are not taken into account. Dioxin and related compounds presents an
18 excellent example of a case where background levels in the general population are
19 likely to have significance for evaluation of the relative impact of incremental
20 exposures associated with a specific source. Since RfDs refer to the total chronic dose
21 level, the use of the RfD in evaluating incremental exposures in the face of a
22 background intake exceeding the RfD would be inappropriate.
23
In addition to the concern for various non-cancer health endpoints discussed
24 above, the potential immunotoxicity of dioxin and related compounds represent a
25 special situation. Impairment of the immune system can be considered an adverse
26 outcome in its own right, being responsible for induced pathologies. At the same time,
27 immunotoxicity can function as a modulator of the disease process. The immune
28 system functions to protect against both pathogenic challenge and continued growth of
29 malignant cells. Alterations in the ability of the immune system to perform these
30 primary functions would result in either the promotion of the pathogenic process or the
31 progression of cancer. While it is relatively simple to determine experimentally the
32 effects of TCDD on the ability of the immune system to respond to a variety of specific
51
Draft - Do Not Quot or Cite - Draft M y 2, 1994
1
antigens or immunogens in laboratory animals, it is much more difficult to establish the
2
effects on longer term immune surveillance and the effect of dioxin-like compounds on
3
the immune system of humans.
4
Nonetheless, it has been clearly established that TCDD is immunotoxic and that
5
it can impair normal immune function in laboratory animals and that it is likely to do so
6
in humans as well. Although it is difficult to identify the cell type that is primarily
7
affected in each of the testing protocols, it is clear that several animal species are
8
sensitive to the immunotoxic effects of TCDD at single doses below 1 ug/kg. Although
9 it is possible that humans may be less sensitive than animal models to dioxin
10 immunotoxicity, there are currently limited data to evaluate the impact of immunotoxic
11
responses to dioxin and related compounds in humans.
12
With regard to carcinogenicity, a weight-of-the-evidence
13 evaluation suggests that dioxin and related compounds are likely to
14 present a cancer hazard to humans. This hazard is likely by oral and inhalation
15 routes of exposure and is less likely, although possible, by the dermal route of
16 exposure based on bioavailability and uptake studies. As daily doses through these
17 routes and subsequent body burdens approach those seen in occupational studies,
18 the uncertainty of the hazard characterization is reduced. While the epidemiological
19 data alone are not yet deemed sufficient to characterize the cancer hazard of this class
20 of compounds as being "known," the unequivocal evidence in animal studies,
21 inferences drawn from mechanistic data and the suggestive evidence of recent
22 epidemiology studies support the characterization of dioxin and related compounds as
23 likely cancer hazards. Extent of cancer risk will depend on such parameters as route
24 and level of exposure, overall body burden, dose to target tissues, and hormonal
25 status.
26
The current evidence suggests that both receptor binding and early biochemical
27
events such as enzyme induction are likely to demonstrate low-dose linearity. The
28 relationship of these early events to the complex process of carcinogenesis remains to
29 be determined. If these findings imply low-dose linearity of biologically based cancer
30 models under development, then probability of cancer risk will be linearly related to
31
exposure to TCDD at low doses. However, until the relationship between early
32
cellular responses and the parameters in biologically based cancer models is better
52
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May 2, 1994
1
understood, the shape of the dose-response curve for cancer in the low-dose region
2
can only be inferred with uncertainty. However, given that background exposures to
3 dioxin are ubiquitous and associations between exposure to dioxin-like compounds
4 and certain types of cancer have been noted at body burdens within 1-2 orders of
5 magnitude (10-100 times) of average background body burdens, there is no need for
6 large scale low dose extrapolations. However, since human data to support this
7 conclusion remain limited and based on individuals who were highly exposed for
8
some time in their life, the relationship of apparent increases in cancer mortality in
9 these populations to calculations of general population risk remains uncertain.
10
The fact that dioxin-like compounds are ubiquitous in the environment may
11
have further implications for low-dose risk assessment. Specifically, humans are
12 currently exposed to background levels of dioxin-like compounds on the order of 3-6
13 pg TEQ/kg bw/day, including dioxin-like PCBs. This is more than 500-fold higher than
14
the EPA's 1985 risk-specific dose associated with an upper bound one in a million
15 (1x10-6) risk of 0.006 pg TEQ/kg bw/day and 75-150-fold higher than revised risk
16 specific dose estimates presented in Chapter 8 of this reassessment. For populations
17 who are more highly exposed based on proximity to specific sources or specific
18 human activity patterns such as consumption of higher amounts of foods containing
19 average or higher levels of dioxin-like compounds, the additive background model of
20 Crump et al. (1986) implies that the addition of an incremental dose to an existing
21
background exposure would support the assumption of linearity within the exposure
22 range, particularly if that background exposure is within 1-2 orders of magnitude (10-
23 100 times) of the range of observation of purported dioxin-induced tumors in highly
24
exposed humans.
25
TCDD has been clearly shown to increase malignant tumor incidence in
26 laboratory animals. In addition, a number of studies have been conducted which
27 elucidate other biological effects of dioxin. These studies have been used to develop
28 biologically-based models of the pharmacokinetics of dioxin, of binding to the Ah
29 receptor and of induction of various proteins. In addition, bioassay data on TCDD
30
reported by Kociba have been analyzed using the two-stage clonal expansion model
31
of carcinogenesis. There is evidence to suggest that hormonal factors may be
32 involved in TCDD carcinogenesis. The role of such factors warrants additional study.
53
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1
Ideally, a biologically-based model for cancer induction by TCDD should explicitly
2
consider hormonal influences. Initial attempts to construct a biologically-based model
3 for certain dioxin effects as a part of this re-assessment will need to be continued and
4 expanded to accommodate more of the available biology and to apply to a broader
5 range of potential health effects associated with exposure to dioxin-like compounds.
6
Based on all of the the data reviewed in this reassessment, a
7 picture emerges of TCDD and related compounds as potent toxicants
8 producing a wide range of effects at very low levels when compared with
9 other environmental contaminants. The fundamental level at which these
10 compounds act on biological systems is analogous to several well studied hormones.
11 Dioxin and related compounds have the ability to alter the pattern of growth and
12 differentiation of a number of cellular targets by initiating a cascade of biochemical and
13 biological events resulting in a wide range of responses. While not all of these
14 responses are adverse, and some may even be beneficial, the weight of the evidence
15 suggests concern for the impact of these chemicals on humans at or near current
16 background levels. Additional, incremental exposures occurring as a result of
17 proximity to a point source of release or specific human activity patterns, such as
18 consumption of high levels of more highly contaminated foods, should be evaluated
19 relative to background levels and the impact of the incremental exposure on both
20 transient and steady-state body burdens. This situation is somewhat akin to the
21
scientific approach taken for evaluating and characterizing lead exposure in children.
22 This approach has been useful in providing public health-based advise to decision-
23 makers faced with difficult regulatory choices.
24
25
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PHOTOCOPY
PRESERVATION
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