v>EPA
United States
Environmental Protection
Agency
Office of Research and
Development \
Washington DC' 20460
May 2000
External Review Draft
Chapter 9. Toxicity
Equivalence Factors
(TEF) for Dioxin and
Related Compounds
Review
Draft
(Do Not
Cite or
Quote)
Exposure and Human
Health Reassessment of
2,3,7,8-Tetrachlorodibenzo
p-Dioxin (TCDD) and
Related Compounds
Part II: Health Assessment for
2,3,7,8-Tetrachlorodibenzo-p-
dioxin (TCDD and Related
Compounds
Notice
This document is a preliminary draft. It has not been formally
released by EPA and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its
technical accuracy and policy implications.
-------
-------
DRAFT
DO NOT CITE OR QUOTE
NCEA-I-0836
May 2000
External Review Draft
www.epa.gov/ncea
Chapter 9. Toxicity Equivalence Factors (TEF) for Dioxin
and Related Compounds
Exposure and Human Health Reassessment
of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD)
and Related Compounds
Part II: Health Assessment for 2,3,7,8-Tetrachlorodibenzo-p-dioxin
(TCDD) and Related Compounds
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.
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC :
-------
DISCLAIMER
This document is a draft for review purposes only and does not constitute U.S.
Environmental Protection Agency policy. Mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
n
-------
CONTENTS—OVERVIEW
Parti:
Volume 1:
Volume 2:
Volume 3:
Volume 4:
Exposure and Human Health Reassessment
of 2,3,7,8-Tetrachlorodibenzo-/7-Dioxin (TCDD)
and Related Compounds
Estimating Exposure to Dioxin-Like Compounds (Draft Final)
(EPA/600/P-00/001 Ab, Ac, Ad) March 2000 ;
Executive Summary (EPA/600/P-00/001Aa) ŁMJJs^plinclHdedk tliis dfaft)*
Sources of Dioxin-Like Compounds in the United States (EPA/600/P-00/001Ab)
Chapters1 JiroughJ2 , '
Properties, Environmental Levels, and Background Exposures
(EPA/600/P-OO/OOlAc)
Chapters 1 through 6
Site-Specific Assessment Procedures (EPA/600/P-00/00 1 Ad)
Chapters 1 through 8
Addendum: Revisions since March are included as an addendum to Part I.
Part II:
Health Assessment for 2,3,7,8-TetrachIorodibenzo-/?-dioxin (TCDD) and
Related Compounds (Draft Final)
(EPA/600/P-00/001Ae) May 2000 ;
Chapter 1 . Disposition and Pharmacokinetics :
Chapter 2. Mechanism(s) of Actions
Chapter 3. Acute, Subchronic, and Chronic Toxicity
Chapter 4. Immunotoxicity
Chapter 5. Developmental and Reproductive Toxicity
Chapter 6. Carchiogenicity of TCDD in Animals
Chapter 7. Epidemiology/Human Data
Chapter 8. Dose-Response Modeling for 2,3,7,8-TCDD
(SAB Review Draft) \
Chapter 9. Toxicity Equivalence Factors (TEF) for Dioxin and Related Compounds
(External Review Draft) :
Part III:
Integrated Summary and Risk Characterization for
2,3,7,8-TetrachIorodibenzo-/?-Dioxin.(TCDD) and Related Compounds
(External^Review Draft) (EPA/600/P-00/001Ag) May 2000
-------
CONTENTS
9. TOXICITY EQUIVALENCE FACTORS (TEFs) FOR DIOXIN AND
RELATED COMPOUNDS 9_!
9.1. INTRODUCTION '.].......'.[ 9.1
9.2. HISTORICAL CONTEXT OF TEFs '..'.'.'.'.'.'.'.'.'.'.'. 9-1
9.2.1. TEFs for PCDDs and PCDFs , '.'.'.'.'.'.'.'.'.'. 9-1
9.2.2. TEFs for PCBs '.'.'.'.'.'.'.'.'. 9-3
9.2.3. The Most Recent Evaluation of TEFs for PCDDs, PCDFs, and PCBs 9-5
9.3. SPECIFIC ISSUES 9.8
9.3.1. Ah Receptor and Toxicity Factors 9_8
9.3.2. Ah Receptor Ligands 9_ 1 j
9.4. TOTAL TEQ AND THE ADDITIVITY CONCEPT ....'.'................... 9-16
9.4.1. Examination of Laboratory Mixtures of PCDDs and PCDFs 9-17
9.4.2. Examination of Commercial or Laboratory-Derived Mixtures of
PCDDs, PCDFs, and PCBs 9_20
9.4.3. Examination of Environmental Samples Containing PCDDs,
PCDFs, and/or PCBs 9-22
9.4.4. Nonadditive Interactions With Non-Dioxin-Like Chemicals 9-24
9.4.5. Examination of the TEF Methodology in Wildlife .., 9-26
9.4.6. Toxic Equivalency Functions „ 9-28
9.4.7. Endpoint and Dose-Specific TEFs 9-29
9.5. UNCERTAINTY ] 9_29
9.6. IMPLICATIONS FOR RISK ASSESSMENT 9-30
9.7. SUMMARY 9_31
REFERENCES FOR CHAPTER 9 9.35
IV
-------
LIST OF TABLES
9-1. Estimated relative toxicity of PCDD and PCDF isomers to 2,3,7,8-T4CDD 9-32
9-2. Toxic equivalency factors (TEFs) i 9.33
LIST OF FIGURES
9-1. Structures of polychlorinated dibenzo-p-dioxins, dibenzofurans and biphenyls 9-34
-------
-------
CHAPTER 9. TOXICITY EQUIVALENCE FACTORS (TEFs) FOR
DIOXIN AND RELATED COMPOUNDS
9.1. INTRODUCTION
Previous risk assessments of dioxin and dioxin-like chemicals from around the world
have employed the Toxic Equivalency Factor (TEF) methodology. This method is also used
throughout EPA's dioxin reassessment. This chapter has been added to the EPA's dioxin
reassessment effort to address questions raised by the Agency's Science Advisory Board (SAB)
in 1995. In its Report to the Administrator (U.S. EPA, 1995), the Committee said it "supports
EPA's use of Toxic Equivalencies for exposure analysis...." However, the SAB suggested that, as
the TEQ approach was a critical component of risk assessment for dioxin and related
compounds, the Agency should be explicit in its description of the history and application of the
process and go beyond reliance on the Agency's published reference documents on the subject
(U.S. EPA, 1987, 1989, 1991) to discuss issues raised in review and comment on this approach.
Significant additional literature is now available on the subject, and this chapter provides the
reader with a summary which is up-to-date through 1999. Future research will be needed to
address uncertainties inherent in the current approach. The WHO: has suggested that the TEQ
scheme be reevaluated every 5 years and that TEFs and their application to risk assessment be re-
analyzed to account for emerging scientific information (van den Berg et al., 1998).
9.2. HISTORICAL CONTEXT OF TEFs
A wide variety of polyhalogenated aromatic hydrocarbon (PHAH) compounds can be
detected as complex mixtures in both abiotic and biotic samples. Because of PHAHs' known
global environmental distribution and their toxicity to experimental animals (DeVito et al., 1995;
DeVito and Birnbaum, 1995; Grassman et al., 1998)(see Chapters 3-6 of this volume), to wildlife
(Giesy and Kannan, 1998; Ross, 2000), and to humans (IARC, 1997) (see also Chapter 7 of this
volume), hazard characterization and risk assessment activities have tended to focus on a subset
of polychlorinated dibenzo-p-dioxin (PCDDs), polychlorinated dibenzofurans (PCDFs), and
polychlorinated biphenyls (PCBs)(Figure 9-1). The subset of compounds known as "dioxin-like"
has been described and discussed in Chapter 1 of the dioxin reassessment. In this chapter, the
development of TEFs for these and other PHAHs is discussed.
9.2.1. TEFs for PCDDs and PCDFs
The first use of a TEF-like method was described by Eadon et al. (1986) as a means to
estimate potential health risks associated with a PCB transformer fire in Binghamton, NY. In
1983, the Ontario Ministry of the Environment produced a Scientific Criteria Document for
5/22/00
9-1 DRAFTr-DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
PCDDs and PCDFs which concluded, based on a review of available scientific information, that
dioxin and dibenzofurans were structurally similar compounds that shared a common cellular
! I
mechanism of action (activation of the Ah receptor [AhR]) and induced comparable biological
and toxic responses, and that the development of environmental standards for human health
concerns should be based on a "toxic equivalency" approach with 2,3,7,8-tetrachlorodibenzo-p-
dioxin (TCDD) as the prototype (OME, 1984). The final recommendation divided all
PCDD/PCDF congeners into their respective homologue groups and assigned to each group a
toxicity factor relative to TCDD (Table 9-1). These numerical factors could then be applied to
transform various concentrations of PCDDs and PCDFs into equivalent concentrations of
2,3,7,8-TCDD.
Following up on an initial risk assessment methodology designed to address the emission
of dioxins and furans from waste incinerators, EPA also concluded that TEFs were the best
available interim scientific policy for dealing with complex mixtures of these contaminants.
With the mandate to develop active research programs that would address *he limitations
inherent to this risk management technique, the Agency recommended TEFs for specific
congeners, rather than isomeric groups (Table 9-2; U.S. EPA, 1987). In an analogous fashion to
OME's approach, concentrations of PCDDs and PCDFs would be analytically determined, the
concentration of each congener would be multiplied by its respective TEF value, and all the
products would be summed to give a single 2,3,7,8-TCDD equivalent. This approach has been
described mathematically as:
Total Toxicity Equivalence (TEQ) = ^ C" * TEFn
n=l
Cn equals the concentration of the individual congener in the complex mixture under analysis.
TEFs were determined by inspection of the available congener-specific data and an assignment of
an "order of magnitude" estimate of relative toxicity when compared to 2,3,7,8-TCDD. In vitro
binding and in vitro and in vivo toxicity studies were considered in setting individual TEFs.
Scientific judgment and expert opinion formed the basis for these TEF values. External review
of the toxicity and pharmacokinetic data utilized by EPA in setting these TEFs supported the
basic approach as a "reasonable estimate" of the relative toxicity of PCDDs and PCDFs (Olson et
al, 1989).
A 3-year study conducted by the North Atlantic Treaty Organization Committee on the
Challenges of Modern Society (NATO/CCMS) also concluded that the TEF approach was the
best available interim measure for PCDD/PCDF risk assessment. On the basis of examination of
the available data dealing with exposure, hazard assessment, and analytical methodologies
related to dioxin and furans, an International Toxicity Equivalency Factor (I-TEF) scheme was
5/22/00
9-2
DRAFT—DO NOT CITE OR QUOTE
-------
presented (Table 9-2; NATO/CCMS, 1988). This review also concluded that "data strongly
support the role of the Ah receptor in mediating the biologic and toxic responses elicited by
2,3,7,8-TCDD and related PCDDs and PCDFs and provide the scientific basis for the
development of TEFs for this class of compounds." Various refinements to previous efforts
included selection of TEF values based more on in vivo toxicities, assigning TEF values to
octachlorodibenzo-p-dioxin and octachlorodibenzofuran, and removing any TEF values for all
non-2,3,7,8-substituted congeners. Although it was indicated that, theoretically, it may be
possible to detect nearly all of the 210 PCDD/DF isomers in the environment, seventeen 2,3,7,8-
substituted congeners were known to be preferentially retained and bioaccumulated. For
example, when fish or a variety of rodent species were exposed to a complex mixture of
PCDDs/PCDFs from incinerator fly ash, the 2,3,7,8-substituted bongeners, which were minor
components of the original mixture, predominated in the analysis of their tissues (Kuehl et al,
1986; van den Berg et al, 1994). In addition, when humans were exposed to a complex mixture
of more than 40 different PCDF congeners during the Oriental rice oil poisoning episodes, only
the 2,3,7,8-substituted congeners were detected in subsequent blood and adipose tissue analysis
(Ryan et al., 1990). EPA, which had participated in the NATO/CCMS exercise, officially
adopted the revised I-TEFs in 1989, with the caveat that this risk assessment approach remains
interim and continued revisions should be made (U.S. EPA, 1989; Kutz et al., 1990). The use of
the TEF model for risk assessment and risk management purposes has been formally adopted by
a number of countries (Canada, Germany, Italy, the Netherlands; Sweden, the United Kingdom,
U.S.A.) (Yrjanheiki, 1992), and as guidance by international organizations such as the
International Programme on Chemical Safety, WHO.
9.2.2. TEFs for PCBs
During the period of TEF development for PCDDs/PCDFs, a considerable body of
experimental evidence was also being generated regarding the structure-activity relationships
between the different polychlorinated biphenyl homologue classes (Safe, 1990, 1994). Following
the synthesis of analytical standards for all 209 theoretical PCB congeners by 1984, subsequent
analysis of a variety of commercial samples was able to identify all but 26 (Jones, 1988).
However, once released into the environment, PCBs are subject to a variety of photolysis and
biodegradation processes, to the extent that only 50-75 congeners are routinely detected in
higher trophic level species (van den Berg et al., 1995). Initial structure-activity relationship
studies revealed that those congeners substituted in only the meta and para positions were
approximate isostereomers of TCDD. Subsequent toxicological studies confirmed that these
non-ortho-substituted, "co-planar" PCBs (e.g., PCB 77, 81, 126, 169) did induce a variety of in
vitro and in vivo effects similar to TCDD (Leece et al., 1985). Maximum TCDD-like activity is
5/22/00
9-3
DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
obtained for PCBs when there are no ortho, two or more meta, and both para positions occupied
(Figure 9-1). Introduction of a single ortho substituent to the biphenyf. (mono-ortho "co-planars")
results in a diminishing, but not elimination, of TCDD-like activity and lexicological responses
resembling commercial mixtures of PCBs. The addition of a single ortho substituent also
increases the non-dioxin-like activity of the chemical. Several congeners from this group are
prevalent in both commercial PCBs and a -wide variety of environmental samples. Some of the
more persistent mono-ortho substituted PCBs (PCBs 105, 118, 156) can be found in human
serum and adipose samples at levels up to three orders of magnitude higher than the "co-planar"
PCBs, PCDDs and PCDFs (Patterson et al., 1994). In limited studies a third group of PCB
congeners, the di-ortho "non-co-planars," has exhibited only minor amounts of dioxin-like
activity (if any), usually 4-6 orders of magnitude less potent than TCDD (Safe, 1990). Recent
studies have demonstrated that some of the earlier methods of preparation of these di-ortho non-
co-planar PCBs had trace contaminants of PCDFs, which may account for the weak dioxin-like
activity of these chemicals (van der Kolk et al., 1992). In 1991, EPA convened a workshop to
consider TEFs for PCBs (Barnes et al., 1991). The consensus was that a small subset of the
PCBs displayed dioxin-like activity and met the criteria for inclusion in the TEF methodology.
Such proposals for the TEF methodology also seem to have utility in assessing risks to wildlife
(van den Berg et al., 1998; Giesey and Kannan, 1998; Ross, 2000).
PCBs are often classified into two categories: "dioxin-like" and "non-dioxin-like." The
dioxin-like PCBs bind to the AhR and produce dioxin-like effects in experimental animals. All
other PCBs then fall into the non-dioxin-like classification. Although the dioxin-like PCBs are
generally more potent at inducing biological effects, they constitute only a minor portion of the
mass of PCBs found in environmental and biological samples. The non-dioxin-like PCBs
account for a majority of the mass of the PCBs found in environmental and biological samples.
The use of the term non-dioxin-like PCBs is not necessarily useful. The PCBs not included in
the TEF scheme (i.e., the non-dioxin-like PCBs) are not a single class of chemicals and have
multiple toxicities with separate structure-activity relationships (Barnes et al., 1991). Not
enough congener-specific research has been performed to adequately characterize or classify
these chemicals. For example, the "neurotoxic" PCBs have been typically defined by structure-
activity relationships for decreasing dopamine concentrations or alterations in intracellular
calcium in cell culture (Shain et al., 1991; Kodavanti et al., 1996). However, few of these
congeners have been examined in vivo to determine the predictive ability of these in vitro
screens.
As part of the joint World Health Organization European Centre for Environmental
Health (WHO-ECEH) and the International Programme on Chemical Safety (IPCS) project to
harmonize TEF schemes for dioxin-like compounds, a database was generated consisting of all
5/22/00
9-4
DRAFT—DO NOT CITE OR QUOTE
-------
available relevant toxicological data for PCBs up to the end of 1993. Of almost 1,200 peer-
reviewed publications, 146 were selected and analyzed on the basis of the following criteria: at
least one PCS congener was investigated; TCDD or a reference co-planar PCB (77, 126, 169)
was used during the experiment or results were available from previous experiments (same
author, laboratory, experimental design); and the endpoint in question was affected by both the
reference compound and the PCB congener in question (i.e., dioxin-specific). TEFs were then
determined from a total of 60 articles/manuscripts on the basis of the reported results for 14
different biological/toxicological parameters. Following scientific consultation by 12 experts
from 8 different countries, interim TEF values were recommended for 13 dioxin-like PCBs
(Table 9-2), based on four inclusion criteria: (1) the compound should show structural similarity
to PCDDs and PCDFs; (2) it should bind to the Ah receptor; (3) it should induce dioxin-specific
biochemical and toxic responses; and (4) it should be persistent and accumulate in the food chain
(Ahlborg et al, 1994). Increased consideration was given to selection of a TEF value based on
repeat-dosing in vivo experiments, when available. '•
There is experimental evidence to suggest that a limited number of PCB congeners
classified as weak or non-AhR agonists could effect concentration-dependent nonadditive
interactions with dioxin-like compounds (Safe, 1990; 1994). Both antagonistic (Safe, 1990;
Morrissey et al., 1992; Smialowicz et al., 1997b) and synergistic (Safe, 1990; van Birgelen et al.,
1996a,b; van Birgelen et al., 1997) interactions between TCDD and PCBs have been observed in
experimental systems. These interactions usually occur at e?rtremely high doses of the PCBs that
are not environmentally relevant,, and thus the nonadditive interactions are thought not to
significantly detract from the TEF methodology (van den Berg et al., 1998; Birnbaum, 1999).
9.2.3. The Most Recent Evaluation of TEFs for PCDDs, PCDFs, and PCBs
An additional recommendation from the first WHO PCB TEF consultation was that the
current database should be expanded to include all relevant mfonrjation on PCDDs, PCDFs, and
other dioxin-like compounds that satisfied the four inclusion criteria. Prior to the second WHO-
ECEH consultation in 1997, various terminologies or definitions applicable to TEFs were
reviewed and standardized. Whereas previously the term TEF had been used to describe all
scientific endpoints used in comparison with TCDD, it was noted that a variety of experimental
parameters may not be considered "toxic," but are considered as biological/biochemical
responses, such as Ah receptor binding and alkoxyresorufin O-dealkylase induction. The
decision was that any experimental endpoint for which a numerical value of the relative potency
compared to TCDD had been generated from a single laboratory examining a single endpoint
would be known as a relative potency value, or REP. The term TEF would then be restricted to
describe an order-of-magnirude consensus estimate of the toxicity of a compound relative to the
5/22/00
9-5 DRAFTr-DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
toxicity of TCDD that is derived using careful scientific judgment of eill available data (van
Leeuwen., 1997; van den Berg et al., 1998).
At the second WHO-ECEH consultation in 1997, relative potency factors were calculated
based on the following methodology (van den Berg et al., 1998):
• Assigned as reported in the publication/manuscript (verified from available data).
• Calculated from the dose-response curves using linear interpolation of log doses
comparing the same effect levels with correction for different control levels.
• Calculated from ratios of low or no observed effect levels (LOELs, NOELs) and
effect concentration/dose 10%, 25% or 50% values (ED/EC10 25 50).
• Calculated from ratios of tumor promotion indexes or maximal enzyme induction
levels.
• Calculated from ratios of All receptor binding affinities (Kd).
Whereas the resulting range of in vitro/in vivo REP values for a particular congener may
span 3-4 orders of magnitude, final selection of a TEF value gave greater weight to REPs from
repeat-dose in vivo experiments (chronic > subchronic > subacute > acute). As with the PCB
TEF consultation, dioxin-specific endpoints were also given higher priority. A rounding-off
procedure (nearest 1 or 5) was also employed for final TEF selection (Table 9-2). It should be
noted that the TEF was rounded up or down depending on the chemical, the data, and scientific
judgment
Notable amendments to the previous NATO/WHO TEF schemes include:
• On the basis of new REPs from in vivo tumor promotion and enzyme induction, a
TEF of 1.0 was recommended for 1,2,3,7,8-PeCDD.
• Originally the TEF for OCDD was based on body burdens of the chemical
following subchronic exposures; a TEF based on administered dose is reduced to
0.0001.
• New in vivo enzyme induction potency and structural similarity with OCDD
support the TEF change to 0.0001 for OCDF.
• REPs from an in vivo subchronic toxicity study (enzyme induction, hepatic retinol
decreases) support reducing the TEF to 0.0001 for PCB 77.
• A TEF value of 0.0001 was assigned for PCB 81. Even though PCB 81 was not
assigned a TEF value at the 1993 WHO consultation because of lack of human
residue and experimental data, more recent data demonstrate similar qualitative
structural activity results compared to PCB 77.
5/22/00
9-6
DRAFT—DO NOT CITE OR QUOTE
-------
• Because of the lack of in vivo enzyme induction (CYP 1A1/A2) and reproductive
toxicity with structurally similar congeners (PCB 47 and PCS 153), the previous
interim TEF values for the di-ortho-substituted PCBs 170 and 180 were
withdrawn.
Although a number of uncertainties associated with the TEF concept have been identified
(nonadditive interactions with non-dioxin-like PCBs, natural ligands for the Ah receptor,
questionable low-dose linearity of REP responses), the 1997 WHO expert meeting decided that
an additive TEF model remained the most feasible risk assessment method for complex mixtures
of dioxin-like PHAHs.
The WHO working group acknowledged that there are a number of other classes of
chemicals that bind and activate the Ah receptor. The chemicals include, but are not limited to,
polyhalogenated naphthalenes, diphenyl ethers, fluorenes, biphenyl methanes, quaterphenyls, and
others. In addition, a number of brominated and chloro/bromo-substituteddioxin analogues of
the PCDDs and PCDFs have been demonstrated to cause dioxin-like effects. The WHO working
group concluded that "at present, insufficient environmental and toxicological data are available
to establish a TEF value for any of the above compounds" (van den Berg et al., 1998).
In January 1998, EPA and the U.S. Fish and Wildlife Service sponsored a meeting
entitled "Workshop on the Application of 2,3,7,8-TCDD Toxiciry Equivalency Factors to Fish
and Wildlife." The major objective of the workshop was to address uncertainties associated with
the use of the TEF methodology in ecological risk assessment. Twenty-one experts from
academia, government, industry, and environmental groups participated in the workshop. The
consensus of the workgroup was that while there are uncertainties in the TEF methodology, the
use of this method decreases the overall uncertainty in the risk assessment process. However,
quantifying the decrease in the uncertainty of a risk assessment using the TEF methodology
remains ambiguous, as does the exact uncertainty in the TEF methodology itself (U.S. EPA,
2000). :
This first section has outlined the process of assessing the relative potency of chemicals
and the assignment of a consensus TEF value. There are still many questions on the use of the
TEF method and the validity of some of the underlying assumptions. A detailed discussion and
review of the data supporting the development and use of the TEF method, as well as the data
relating to the issue of additivity, is included within the specific issues section that follows.
5/22/00
9-7
DEAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
9.3. SPECIFIC ISSUES
9.3.1. Ah Receptor and Toxicity Factors
Issues relating to the role of the Ah receptor as the common mediator of toxicity of
dioxin-like chemicals and the cross-species comparability of AhR structure and function
frequently arise when the TEF approach is discussed. Recent data relating to each of these issues
are discussed below.
The general basis for the TEF scheme is the observation that the AhR mediates most if
|
not all biological and toxic effects induced by dioxin-like chemicals (Safe, 1990; Okey et al.,
1994; Birnbaum, 1994; Hankinson, 1995). Binding to the receptor is necessary, but not
sufficient, to generate the wide variety of toxic effects caused by dioxin-like HAHs (Sewall and
Lucier, 1995; De Vito and Birnbaum, 1995) (for additional review references, see Chapter 2).
There are several lines of evidence that the Ah receptor is important in the toxicity of the dioxin-
like chemicals. A brief discussion of this evidence shall be presented in the following section.
Those wishing a more detailed discussion of this issue are referred to Chapter 2.
Initial studies on the toxicity of PAHs demonstrated that the sensitivity to these chemicals
varied by strain of mice and segregated with the Ah locus. The Ah locus was then found to
encode a receptor designated as the aryl hydrocarbon receptor or AhR. Sensitive strains of mice
expressed receptors with high binding affinity for these chemicals, while the resistant mice
expressed a receptor that poorly bound the PAHs. One of the best ligands for this receptor was
TCDD. Shortly after the discovery of the AhR, structure-activity relationship studies
demonstrated a concordance between binding affinity to the Ah receptor and toxic potency in
vivo in mice. Further support of the role of the Ah receptor hi the toxicity of dioxin-like
chemicals was demonstrated following the development of AhR knockout mice (Fernandez-
Salguero et al., 1995; Schmidt et al., 1996; Mimura et al., 1997; Lahvis and Bradfield, 1998).
Administration of TCDD at doses more than 10 times the LD50 of wild-type mice has not
produced any significant dioxin-like effects, either biochemical or toxicological, in the AhR
knockout mice (Fernandez-Salguero et al., 1996; Peters et al., 1999). These data as a whole
demonstrate that the binding to the AhR is the initial step in the toxicity of dioxin-like chemicals.
Although binding to the AhR initiates a cascade of molecular and cellular events leading
to toxicity, the exact mechanism of action of dioxin-like chemicals is not completely understood.
One difficulty in determining the mechanism is our limited understanding of the normal
physiological role of the AhR, which would aid in understanding of potential species differences
in response to dioxin-like chemicals. The available data indicate that the AhR does play an
important role in normal processes and that there are a number of similarities in the action of the
AhR between species. These data strengthen our confidence in species extrapolations with these
chemicals.
5/22/00
9-8
DRAFT—DO NOT CITE OR QUOTE
-------
There are several lines of evidence suggesting that the AhR is an important factor in
developmental and homeostatic processes. The AhR is a ligand-activated transcription factor
that is a member of the basic-helix-loop-helix-Per-Arnt-Sim (bHLH-PAS) superfamily. The
AhR is also a highly conserved protein that is present in all vertebrate classes examined,
including modern representatives of early vertebrates such as cartilaginous and jawless fish
(Hahn, 1998). In addition, an AhR homologue has been identified in C. elegans (Powell-
Coffman, 1998). The bHLH-PAS superfamily consists of a grooving list of at least 32 proteins
found in diverse organisms such as Drosophila, C. elegans, and humans. Many of these proteins
are transcription factors that require either hetero- or homodimerization for functionality. These
proteins regulate circadian rhythms (per and clock) and steroid receptor signaling (SRC-1, TIF2,
RAC3) and are involved hi sensing oxygen tension (Hif-1, EPAS-1/HLF) (Hahn, 1998). The
classification of the AhR as part of the bHLH-PAS superfamily and its evolutionary conservation
imply that this protein may play an important role in normal physiological function. It has been
proposed that understanding the function of the bHLH-PAS family of proteins and the
phylogenetic evolution of the AhR may lead to an understanding of the role of this protein in
normal processes (Hahn, 1998).
The process of development is a complex phenomenon that involves the specific
expression of numerous genes in a spatial and temporal pattern. The importance of a particular
gene in developmental biology is often inferred by its spatial and temporal expression during
development. The AhR is expressed in a tissue, cell, and temporal pattern during development
(Abbott et al., 1995). It is highly expressed in the neural epithelium, which forms the neural crest
(Abbott et al., 1995). The expression of the AhR during development suggests that this protein
has important physiological functions.
Further evidence of the role of the AhR in developmental processes is provided by the
development and study of AhR knockout mice. Three strains of AhR knockout mice have been
produced using a targeted disruption of the Ahr locus (Fernsindez-Salguero et al., 1995; Schmidt
et al., 1996; Mimura et al., 1998; Lahvis and Bradfield, 1998). The AhR -/- mice develop
numerous lesions with age (Fernandez-Salguero et al., 1995). Mortality begins to increase at
about 20 weeks, and by 13 months almost half of the mice either die or become moribund.
Cardiovascular alterations consisting of cardiomyopathy with hypertrophy and focal fibrosis,
hepatic vascular hypertrophy and mild fibrosis, gastric hyperplasia, T-cell deficiency in the
spleen, and dermal lesions are apparent in these mice and the incidence and severity increases
with age (Fernandez-Salguero et al., 1995). Although male and female AhR -/- mice are fertile,
the females have difficulty maintaining conceptus during pregnancy, surviving pregnancy and
lactation, and rearing pups to weaning (Abbott et al., 1999). It should be noted that the AhR
knockout mice are resistant to the toxic effects of TCDD. \
5/22/00
9-9
DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
Comparisons between the AhR of experimental animals (primarily rodents) and the
human AhR have revealed a number of similarities in terms of ligand and DNA binding
characteristics as well as biochemical functions. Tissue-specific patterns of expression of AhR
mRNA are similar hi rats, mice, and humans, with highest levels generally detected in lung, liver,
placenta, and thymus (Dolwick et al., 1993; Dohr et al., 1996). Nuclear AhR complexes isolated
from human and mouse hepatoma cells (Hep G2 and Hepa Iclc7, respectively) have similar
molecular weights. Although the human AhR was found to be more resistant to proteolytic
digestion by trypsin or chymotrypsin, the major breakdown products were similar between the
two species, and photolabeling analysis with TCDD suggested common features in the ligand
binding portion of the receptors (Wang et al., 1992).
Limited analysis has suggested the average human AhR exhibits a lower binding affinity
for various HAHs than "responsive" rodent strains. However, similar to a variety of
experimental animals, human populations demonstrate a wide variability in AhR binding affinity
(Micka et al., 1997). Recent determination of AhR binding affinity (Kd) toward TCDD in 86
human placenta samples showed a greater than twentyfold range in the binding affinity, and this
range encompasses binding affinities similar to those observed in sensitive and resistant mice
(Okey et al., 1997). Whereas the concentration of various ligands required to activate a human
AhR reporter gene construct was higher than required with rodent cell cultures, the actual rank
order of binding affinities was in agreement (Rowlands and Gustafsson., 1995). Although
comparisons have been made of the TCDD binding affinity to the AhR of different species,
caution should be used when applying this information to species sensitivity. For mice, the
sensitivity to the biochemical and toxicolpgical effects of TCDD can be correlated with the
relative binding affinity of the TCDD to the AhR hi different strains (Birnbaum et al., 1990;
Poland and Glover, 1990). However, the relative binding affinity of TCDD to the AhR across
species does not aid in the understanding of interspecies differences in the response or sensitivity
to TCDD (DeVito and Birnbaum, 1995).
The human AhR also demonstrates other slight differences when compared to the AhR
from experimental animal species. The molecular mass of the human AhR ligand-binding
subunit appears to be greater than the AhR subunit from certain TCDD "responsive" mouse
strains but similar to the receptor molecular mass for rats (Poland and Orlover, 1987). Currently
there has been no association established between differences in the molecular mass of the AhR
and sensitivity to a particular biochemical or toxicological response (Okey et al., 1994). The
non-liganded human AhR appears thermally more stable compared to AhR from various rodent
species, whereas the reverse situation exists with the liganded human AhR (Nakai and Bunce,
1995). Transformation of the ligand-bound human AhR receptor (isolated .from colon
adenocarcinoma cells) to the DNA-binding state, unlike rodent hepatic AhR, is temperature
5/22/00
9-10 DRAFT—DC) NOT CITE OR QUOTE
-------
dependent (Harper et al., 1992). However, in critical areas of receptor function such as ligand
recognition, transformation, and interaction with genomic response elements, the human AhR is
comparable to the AhR isolated from experimental animals. \
The bHLH structure of receptor proteins such as AhR ensures appropriate contact and
binding with DNA recognition sites. Amino acid sequence analysis between mouse and human
AhR shows an overall sequence homology of 72.5%, whereas the HLH domain shows 100%
amino acid concordance (Fujii-Kuriyama et al., 1995). In comparison, the deduced ammo acid
composition of the AhR from killifish was 78%-80%, similar to the amino acid sequence of
rodent and human AhR (Hahn and Karchner, 1995). Ligand-bound or transformed AhR from a
variety of mammalian species, including humans, all bind to a specific DNA sequence or "dioxin
response element" with similar affinities (Bank et al., 1992; Swanson and Bradfield, 1993).
The majority of scientific evidence to date supports the theory that binding to AhR is a
necessary first step prior to dioxin-like chemicals eliciting a response, as discussed in Chapter 2
of this volume. Current research has identified the AhR in a variety of human tissues and cells
that appear to function in a similar manner to the AhR from experimental animals, including fish,
birds, and mammals. When multiple endpoints are compared across several species, there exists'
a high degree of homogeneity in response and sensitivity to TCDD and related chemicals
(DeVito et al., 1995). Therefore, these data provide adequate support for the development of the
TEF methodology. However, these data also reflect the true complexity of intra- and interspecies
comparisons of biochemical and toxicological properties. Continued research into the variety of
additional cytoplasmic and nuclear proteins capable of interacting with the AhR signaling
pathway will ultimately lead to a better understanding of the observed species and strain
variability in the response to dioxin-like chemicals and may be useful in further refining TEFs.
9.3.2. Ah Receptor Ligands
A wide variety of structurally diverse anthropogenic and natural chemicals are capable of
interacting with the. AhR. These chemicals also have a broad range of potencies at inducing
dioxin-like effects in experimental systems. One of the major differences between the
anthropogenic chemicals included in the TEF methodology and the natural AhR ligands is their
pharmacokinetics. The anthropogenic chemicals included in the TEF methodology are persistent
and bioaccumulate in wildlife and humans. In contrast, most if not all of the natural AhR ligands
are rapidly metabolized and eliminated from biological systems. The following section will
examine the differences between the chemicals included in the TEF methodology and remaining
AhR ligands not included in this approach.
The synthetic compounds that bind to AhR include a number of different classes of
chemicals such as industrial chemicals (polyhalogenated biphenyls, halogenated napthalenes,
5/22/00
9-11 DRAFT—DO NOT CITE OR QUOTE
-------
1 polyhalogenated biphenyls, chlorinated paraffins, etc.), pesticides (hexachlorobenzene), and
2 contaminants (polyhalogenated dioxins and furans) associated with various manufacturing,
3 production, combustion, and waste disposal processes. In addition, pyrolysis of organic material
4 can produce a number of unsubstituted polycyclic aromatic hydrocarbons (PAHs) with moderate
5 to high affinity for AhR (Poland and Knudson, 1982; Nebert, 1989; Chaloupka et al., 1993).
6 Not all of the anthropogenic sources of dioxin-like chemicals are included in the TEF
7 methodology. Many of these chemicals, such as hexachlorobenzene and the brominated diphenyl
8 ethers, are only weakly dioxin-like and have significant toxicological effects that are not
9 mediated by the Ah receptor. For these chemicals, it is not clear that adding them to the TEF
10 methodology would decrease the uncertainty in the risk assessment process. For other classes of
11 chemicals, such as the chlorinated napthalenes, environmental concentrations and human
12 exposures are uncertain. Other anthropogenic chemicals such as the PAHs are not included
13 because of their short half-lives and relatively weak AhR activity.
14 Brominated dioxins, dibenzofurans, biphenyls, and napthalenes also induce dioxin-like
15 effects in experimental animals (Miller and Birnbaum, 1986; Zacherewski et al., 1988;
16 Birnbaum et al., 1991; Hornung et al., 1996; DeVito et al., 1997; Weber and Greim, 1997). The
17 brominated dioxins and dibenzofurans may be more or less potent than their chlorinated
18 orthologues, depending on the congener (Birnbaum et al., 1991; DeVito et al., 1997). The
19 sources of the brominated dioxin-like chemicals are not well characterized. Some of the
20 chemicals, such as the brominated biphenyls and naphthalenes, are synthesized and sold as
21 commercial flame retardants. Brominated dibenzofurans are produced, as byproducts of pyrolysis
22 of brominated flame retardants. There is some evidence of human exposure to brominated
23 dioxins and dibenzofurans from extruder operators (Ott and Zober, 1996). Polybrominated,
i I
24 polychlorinated, and mixed bromo and chloro dioxins and dibenzofurans have been found in soot
25 from textile processing plants (Sedlak et al., 1998). Although these chemicals have been found
26 in humans, these studies are limited to a small population and exposure to the general population
27 remains undetermined. Future examinations of the TEF methodology should include a more
28 detailed discussion of the of the brominated dioxins and dibenzofurans,
29 The evolutionary conservation of AhR and its biological function following activation by
30 dioxin-like chemicals have led to the hypothesis that there must be an endogenous or
31 physiological ligand(s) for this receptor. Presently, the endogenous ligand remains
32 irndeterrnined. However, efforts to discover the natural ligand have led to the discovery of a
33 number of naturally occurring AhR ligands. A number of naturally occurring chemicals present
34 in the diet are capable of binding to AhR and inducing some dioxin-like effects in experimental
35 animals (Bradfield and Bjeldanes, 1984,1987) and humans (Michnovicz and Bradlow, 1991;
5/22/00
9-12
DRAFT—DO NOT CITE OR QUOTE
-------
Sinha et al., 1994). The question of how the interaction of these chemicals relates to the toxicity
of those chemicals designated as dioxin-like has become the subject of much debate.
One class of naturally occurring chemicals that activate the AhR is the indole derivatives.
Indole derivatives, naturally present in a variety of cruciferous vegetables, are capable of
modulating the carcinogenicity of PAHs (Wattenberg and Loub, 1978). Indole-3-carbinol (I-3-C)
and 3,3'-diindolylmethane (DIM) are major secondary metabolites found in cruciferous
vegetables and induce both phase I and II metabolic enzymes (CYP1 A-dependent glutathione and
glucuronyl transferases, oxidoreductases) in experimental animals (Bradfield and Bjeldanes,
1984, 1987), human cell lines (Bjeldanes et al., 1991; Kleman et al., 1994), and humans
(Michnovich and Bradlow, 1990, 1991). Although both compounds induce CYP450 enzymes
under AhR transcriptional control, they exhibit relatively low binding affinity for the Ah receptor
(Gillner et al., 1985). Further investigation revealed that I-3-C is relatively unstable in the acidic
environment of the digestive tract and readily forms DIM. In turn, DIM can participate in acid
condensation reactions to form indolocarbazoles (ICZs) (Chen et al., 1995). ICZs can also be
produced by bacterial metabolism of the common dietary amino acid tryptophan. ICZs, in
particular indolo[3,2b]carbazole5 exhibit high binding affinity for the rodent AhR, approximately
equipotent to 2,3,7,8-tetrachlorodibenzofuran, and can induce CYP1A1 activity in cultured cells
(Bjeldanes et al., 1991; Gillner et al., 1993; Chen et al., 1995). ICZ and a methylated derivative
5,1 l-dimethylindolo[3,2b]carbazole (MICZ), are also capable of binding to and activating the '
AhR in human hepatoma cells (HepG2) (Kleman et al., 1994). With considerably lower efficacy,
I-3-C and DIM can partially displace TCDD from the AhR from human breast cancer cells
(T47D) (Chen et al., 1996). These results would suggest that this group of compounds may
represent a class of physiologically active AhR ligands derived from natural sources, which could
either mimic dioxin-like compounds in their action or act as competitors for AhR binding.
In addition to the plant-derived indoles, experimental animals consuming thermally
treated meat protein as well as humans fed cooked meat can exhibit induced CYP1A2 activity
(Degawa et al., 1989). High-temperature cooking (250°C, 22 minutes) of ground beef resulted in
the formation of a number of heterocyclic aromatic amines (HAAs) in part-per-billion levels,
which were thought to be responsible for the observed CYP1A2 induction in human volunteers
(Sinha et al., 1994). Mechanistic analysis of one particular HAA, 2-amino-3,8-
dimethylimidazo[4,5-flquinoxalme (MelQx), has shown that it is capable of both interacting with
the AhR and inducing CYP1A1/A2 activity in rats (Kleman and Gustafsson, 1996). These data
should be viewed cautiously because recent data indicate that CYP1A2 can be induced through
non-AhR mechanisms (Ryu et al., 1996). Because there are multiple pathways to induce
CYP1A2, the increase in CYP1A2 activity following exposure to complex mixtures, such as
cooked meat, does not necessarily indicate the presence of dioxin-like chemicals.
5/22/00 9-13 DRAFT—DO NOT CITE OR QUOTE
-------
1 Other diet-derived chemicals that can interact with the AhR include oxidized essential
2 amino acids. UV-oxidized tryptophan is capable of inducing CYP1A1 activity in mouse
3 hepatoma cells through an AhR-dependent mechanism (Sindhu et al., 1996). Rats exposed to
4 UV-oxidized tryptophan in vivo also exhibited induction of hepatic and pulmonary CYP1 Al
5 activity. Both hi vitro and in vivo enzyme induction were transient, with the oxidized tryptophan
6 possibly being metabolized by CYP1 Al (Sindhu et al., 1996). Tryptzinthrins, biosynthetic
7 compounds produced from the metabolism of tryptophan and anthranilic acid by yeast commonly
j, j
8 found hi food, are agonists for the rat AhR (Schrenk et al., 1997). Various tryptanthrins were
9 also capable of inducing CYP1 Al -related enzyme activity in mouse hepatoma cells with the
10 approximate efficacy of ICZ.
11 Recent studies have demonstrated that physiological chemicals can bind to the AhR.
12 Bilirubin was recently found to be capable of transforming the AhR from mouse hepatoma cells
13 into its DNA-binding state, resulting in CYP1 Al induction. Hemin and biliverdin can also be
14 metabolically converted to bilirubin, resulting in AhR-dependent gene activation (Sinai and
1
15 Bend, 1997). Despite these results, there is no clear evidence that these are the physiological
16 ligands for the AhR, nor is there evidence that these compounds can modulate the activity of
17 dioxin-like compounds or lead to dioxin-like toxic effects in humans or animals.
18 A number of "natural" or dietary compounds have been identified, which in certain in
19 vitro cases can function as AhR agonists with similar potency when compared to various
20 halogenated aromatics. It has been postulated that the endogenous ligands could be the major
21 contributors to the daily dose of TEQs, because of their higher estimated intakes (Safe, 1995).
22 Comparing the TEQ intake of natural or dietary AhR ligands to the halogenated aromatics, it has
23 been proposed that more than 90% of the TEQ is derived from the dietary or natural compounds
24 (Safe, 1995). The "natural" ligands tend to have short half-lives and do not accumulate. The
25 PCDDs/PCDFs and PCBs included in the TEF methodology clearly bioaccumulate. If
26 contributions to the total TEQ are estimated on steady-state body burdens of these chemicals
27 instead of daily intake, then TCDD and other PCDDs/PCDFs and PCBs contribute more than
28 90% of the total TEQ compared to the "natural" ligands (DeVito and Birnbaum, 1996). The
29 difference in the results of these analyses demonstrates our uncertainty of the relative potencies
30 and exposures to these natural AhR ligands.
31 When a comparison is attempted between the perceived relative risk from natural vs.
32 anthropogenic AhR agonists, a number of factors should be taken into consideration. The
33 toxicity of AhR ligands depends on several factors, including AhR binding affinity, biological
34 half-life, and exposure. The chemicals included in the TEF scheme are those that not only bind
35 to AhR but also bioaccumulate and have long biological half-lives in humans, typically on the
36 order of years. In contrast, the pharmacokinetics of the endogenous or natural group are not well
5/22/00
9-14
DRAFT—DO NOT CITE OR QUOTE
-------
studied, but these chemicals tend to be short-lived, with half-lives on the order of minutes to
hours. Although both PAHs and the halogenated aromatics bind to AhR and induce cytochrome
P450-related enzyme activities, only the latter group produces the additional dioxin-like
spectrum of toxicological responses. These toxicities are thought to be due to the persistent
exposures attributable to the long half-lives of these chemicals (Riddick et al., 1994).
Initial studies comparing the potency of indolo[3,2b]carbazole to TCDD demonstrate the
importance of the pharmacokinetic differences between these chemicals. For example, in Hepa-1
cells exposed for 4 hours, the relative potency of indolo[3,2b]carbazole compared to TCDD is
0.1 (Chen et al., 1995). If the relative potency is determined after 24 hours of exposure, the
potency of indolo[3,2b]carbazole drops 1,000-fold to 0.0001 (Chen et al., 1995). In addition, the
dioxin-like effects of low doses of indolo[3,2b]carbazole in Hepa-1 cells are transient. Similar
transient effects of other dietary-derived AhR ligands have also been reported (Xu and Bresnick,
1990; Berghard et al., 1992; Ridduck et al., 1994). These data demonstrate that the relative
potencies of these chemicals compared to TCDD are dependent upon the pharmacokinetic
properties of the chemicals and the experimental design used in the comparisons. These data
also demonstrate our uncertainty of the relative potency of the dietary-derived AhR ligands.
Though it is important to address these issues, the available data do not lend themselves to an
appropriate quantitative analysis of the issue.
One of the other limitations when comparing the relative exposures to dietary AhR
ligands and the anthropogenic AhR ligands is that few in vivo studies have examined the toxicity
of the dietary or natural AhR ligands. However, in utero exposure of rats to I-3-C resulted in a
number of reproduction-related abnormalities in male offspring, only some of which resemble
those induced by TCDD (Wilker et al., 1996). The relative in vivo potency of I-3-C in these
studies was approximately 0.000005 (Wilker et al., 1996). Although there are limited data on the
in vivo biochemical and toxicological effects of these ligands, the effects of mixtures of
anthropogenic and natural AhR ligands is lacking. There are some studies examining the
interactions of I-3-C and ICZ on the effects of TCDD in cell culture systems. However, it is
uncertain how to extrapolate these in vitro concentrations to present human in vivo exposures.
The limited data available do not adequately address the interactions between these chemicals.
Future in vivo studies are required in order to better understand the potential interactions between
these classes of AhR ligands. i
Another difficulty in comparing the natural AhR ligands to the dioxins is the multiple
effects induced by the natural AhR ligands. In vivo and in vitro studies of I-3-C indicate that it
induces a number of biochemical alterations that are not mediated through the AhR (Broadbent
and Broadbent, 1998). The activation of these additional pathways creates difficulties in making
direct comparisons with TCDD and related chemicals. Similarly, the PAHs also have non-AhR-
5/22/00
9-15 DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
mediated biochemical and toxicological effects that also complicate diiect comparisons with
TCDD and related dioxins. For example, interactions of TCDD with PAHs have demonstrated
both synergistic and antagonistic interactions (Silkworth et al., 1993).
Presently, there are several limitations in our understanding of the importance of naturally
occurring dioxin-like chemicals vs. the dioxin-like chemicals included in the TEF methodology.
First is the lack of data on the interactions between these classes of chemicals. Few if any
mixtures of natural AhR ligands and PCDDs or PCDFs examining a toxic response have been
published. Second, many of the natural AhR ligands have multiple mechanisms of action that
presently cannot be accounted for in the TEF methodology. For example, I-3-C has
anticarcinogenic properties in tumor promotion studies, and these effects may or may not be
mediated through AhR mechanisms (Manson et al., 1998). The lack of data and the role of non-
AhR mechanisms in the biological effects of these chemicals prohibit a definitive conclusion on
the role of natural vs. anthropogenic dioxins in human health risk assessment.
Although Safe has suggested that exposure to natural AhR ligands is 100 times that of
TCDD and other dioxin-like chemicals (Safe, 1995), the impact of the natural AhR ligands is
uncertain. Epidemiological studies suggest that human exposures to TCDD and related
chemicals are associated with adverse effects such as developmental impacts and cancer. In
many of these studies, the exposed populations have approximatley 100 times more TCDD
exposure than background populations (see Chapter 7). If the exposure to natural AhR ligands is
included in these comparisons, then the exposed populations should have only about 2 times
higher total TEQ exposures than the background population. It seems unlikely that
epidemiological studies could discriminate between such exposures. These data suggest that the
estimates of the contribution of the natural AhR ligands to the total TEQ exposure are
overestimated. In addition, regardless of the background human exposure to "natural" AhR
ligands, the margin of exposure to TCDD and related chemicals between the background
population and populations where effects are observed remains a concern.
9.4. TOTAL TEQ AND THE ADDITIVITY CONCEPT
The issue of the scientific defensibility of additivity in determining total TEQ has been
raised since the onset of the use of TEFs. Arguments regarding this approach include the
presence of competing agonists or antagonists in various complex mixtures from environmental
sources, interactions based on non-dioxin-like activities (inhibition or synergy), and the fact that
dose-response curves for various effects may not be parallel for all congeners assigned TEFs.
Although comparative pharmacokinetics have also been raised as an issue, this has generally
been accounted for by the heavier weight accorded to in vivo studies in the. assignment of TEFs.
Despite these concerns, empirical data support the use of the additivity concept, recognizing the
5/22/00
9-16 DRAFT—DO NOT CITE OR QUOTE
-------
imprecise nature of the TEFs per se. A substantial effort has been made to test the assumptions
of additivity and the ability of the TEF methodology to predict the effects of mixtures of dioxin-
like chemicals. These efforts have focused on environmental, commercial, and laboratory-
derived mixtures. In addition, endpoints examined ranged from biochemical alterations, such as
enzyme induction, to toxic responses such as tumor promotion, teratogenicity, and
immunotoxicity. A brief summary of some of the more important work is given and discussed in
the following section.
The TEF methodology has been examined by testing mixtures of chemicals containing
dioxins and sometimes other chemicals. These mixtures have either been combined and
produced in the laboratory or were actual environmental samples. Researchers have also used
different approaches in estimating the TCDD equivalents of the mixtures. Some researchers
have determined the REP of the components of the mixture in the same system in which the
mixture was tested and have used these REPs to estimate TCDD equivalents. These studies can
provide insight into the validity of the assumption of additivity of the TEF methodology. Other
researchers have used consensus TEF values to estimate the TCDD equivalents of the mixture. It
is not clear if these studies can be considered true tests of the additivity assumption. The
consensus TEF values have been described as conservative estimates of the relative potency of a
chemical in order to protect humans and wildlife. If the consensus TEF values are conservative
and protective, then they should overestimate the potency of mixtures tested in an experimental
system. In essence, using the consensus TEF values should generally overpredict the potency of
a mixture (and therefore underpredict the response) when compared to the equivalent
concentrations of TCDD in an experimental system. In the following discussion of the studies
examining the assumption of additivity, these differences in study design and their implications
for interpretation of the data must be considered.
9.4.1. Examination of Laboratory Mixtures of PCDDs and PiCDFs
Bock and colleagues evaluated the TEF methodology in several systems using both
individual congeners as well as laboratory-derived mixtures (Lipp et al., 1992; Schrenk et al.,
1991,1994). REPs or toxic equivalents or "TEs" (as designated by the authors) were determined
for 2,3,7,8-substituted PCDDs based on enzyme induction in human HepG2 cells, rat H4IIE
cells, and primary rat hepatocytes. The laboratory-defined mixtures, containing up to 49
chlorinated dibenzo-p-dioxins, were then examined in these same cell culture systems. The
TCDD equivalents of the mixtures were determined on the basis of the assumption of additivity
using the TEF methodology and the laboratory derived REPs or TEs as well as experimentally by
comparing the ECSOs of the mixtures with that of TCDD. According to the authors, in all three
systems the data demonstrated that the components of the mixture act in an additive manner
5/22/00
9-17 DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
(Lipp, 1991; Schrenk et al., 1991). For example, in the human HepG2 cells the EC50 of a
mixture of 49 different PCDDs was determined experimentally at 0.034 pg TEQ/plate, compared
to the calculated or predicted EC50 of 0.028 pg TEQ/plate. Interestingly, the TEF methodology
accurately predicted the effects of a mixture containing predominately OCDD, some heptaCDDs
and hexaCDDs, and no pentaCDDs or TCDD (Schrenck et al., 1991).
Bock and colleagues also tested a mixture of 49 PCDDs in a rat liver tumor promotion
study. In theses studies, rats received an estimated 2-200 ng TCDD/kg/d or 200-20,000 ng
mixture/kg/d. The doses of the mixture were equivalent to the TCDD doses using a TE of the
mixture of 0.01 based on enzyme induction in rat hepatocytes (Schrenk et al., 1991). A
comparison of the relative potency of the mixture was based on liver concentrations of the
• I
chemicals followed by TEQ calculations using the I-TEFs (NATO/CCMS, 1988). According to
the authors, in the low-dose region (2-20 ng TCDD/kg/d) the I-TEFs accurately predict the
enzyme-inducing activity of the mixture but tend to overestimate the potency of the mixture at
the higher doses (20-200 ng/kg/d). Also, according to the authors, the I-TEFs provide a rough
estimate of the tumor-promoting potency of the mixture but overestimate the mixture's potency .
However, the authors did not quantify or qualify the magnitude of the overestimation.
In the studies by Schrenk and colleagues, the TEQs were based on tissue dose, not
administered dose. Recent studies by DeVito et al. (1997b, 2000) indicate that the REP for
dioxin-Iike chemicals can differ when determined based on administered or tissue dose. The
higher chlorinated dioxins tend to accumulate in hepatic tissue to a greater extent than does
TCDD, and their REPs tend to decrease when estimated based on tissue dose (DeVito et al.,
1997b, 2000). Because the I-TEFs are based on an administered dose, they may not predict the
response when the TEQ dose is expressed as liver concentration. If the TEQ dose in the data by
Schrenk et al. (1994) is compared on an administered dose, then the dose-response relationship
for increases in relative volume of preneoplastic ATPase-deficient hepatic foci (% of liver) are
comparable between TCDD and the mixture, indicating that additive TEFs provided an
approximation of the tumor-promoting ability of a complex mixture of PCDDs (Schrenck et al.,
1994).
In responsive mouse strains, induction of cleft palate and hydronephrosis by TCDD
occurs at doses between 3 and 90 u.g TCDD/kg (Nagao et aL, 1993; Weber et al., 1985;
Birnbaum et al., 1985,1987,1991). Several groups have examined the assumption of additivity
using teratogenic effects of dioxins as an endpoint. Birnbaum and colleagues examined TEF
methodology using mouse teratogenicity as an endpoint (Weber et al., 1985; Birnbaum et al.,
1985,1987,1991). REPs were derived for 2,3,7,8-TCDF, 2,3,4,7,8-PeCDF, 1,2,3,7,8-PeCDF,
and 1,2,3,4,7,8-HxCDF (Weber et al., 1984, 1985; Birnbaum et al., 1987).. Analysis of the dose-
response for these chemicals, based on administered dose, demonstrated parallel slopes.
5/22/00
9-18 DRAFT—DO NOT CITE OR QUOTE
-------
According to the authors, dose-response analysis of two mixtures containing either TCDD and
2,3,7,8-TCDF or 2,3,4,7,8-PeCDF and 1,2,3,4,7,8-HxCDF demonstrated strict additivity
(Birnbaum et al., 1987; Weber et al., 1985).
Nagao et al. (1993) also examined the TEF methodology using teratogenicity in mice as
an endpoint. Mice were exposed to a single dose of TCDD (5-90 (ig/kg) or a mixture of PCDDs,
or one of two different mixtures of PCDFs. The mixtures contained no detectable TCDD. The I-
TEFs were used to determine the TEQ of the mixtures. According to the authors, the I-TEFs
predicted the potency of the PCDD mixture, and the dose-response relationship was consistent
with the assumption of additivity. The I-TEFs overestimated the potency of the PCDF mixtures
by two- or fourfold. All three mixtures contained significant concentrations of non 2,3,7,8-
chloro-substituted PCDDs and PCDFs in addition to the dioxin-like chemicals present.' In the
studies by Birnbaum and colleagues (Weber et al, 1985; Bimbaum et al, 1985, 1987, 1991) and
Nagao et al. (1993) examining the assumption of additivity using teratogenicity as an endpoint,
the TEF methodology proves useful in estimating the effects of these mixtures.
Rozman and colleagues have examined the assumption of additivity of PCDDs in both
acute and subchronic studies. In acute studies, TCDD (20-60 ng/kg), 1,2,3,7,8-PCDD (100-300
fxg/kg), 1,2,3,4,7,8-HxCDD (700-1,400 ng/kg), and 1,2,3,4,6,7,8-HpCDD (3,000-8,000 u.g/kg)
were administered to male rats, and REP values were determined: for lethality. A mixture of all
four chemicals was then prepared and dose-response studies were performed with the mixture at
doses that would produce 20%, 50%, and 80% mortality. The mixture studies demonstrated
strict additivity of these four chemicals for biochemical and toxicological effects (Stahl et al,
1992; Weber et al, 1992a,b). Following the acute studies, Viluksela et al. (1998a,b) prepared a
mixture of these chemicals and estimated the TEQ based on the REPs from the acute studies. A
loading/maintenance dose regimen was used for 90 days and the animals were followed for an
additional 90 days. According to the authors, the assumption of additivity predicted the response
of the mixture for lethality, wasting, hemorrhage, and anemia, as well as numerous biochemical
alterations such as induction of hepatic EROD activity and decreases in hepatic
phosphenolpyruvate carboxykinase and hepatic tryptophan 2,3-dioxygenase (Viluksela et al,
1997, 1998). Increases in serum tryptophan concentrations and decreases in serum thyroxine
concentrations were also predicted by the TEF methodology (Viluksila et al, 1998a).
Rozman and colleagues followed up these initial studies by examining the assumption of
additivity of the effects of PCDDs as endocrine disrupters (Gao et al, 1999). Ovulation is a
complex physiological phenomenon that requires the coordinated interaction of numerous
endocrine hormones. In a rat model, ovulation can be inhibited by TCDD at doses between 2 to
32 fig/kg (Gao et al, 1999). Dose-response analysis of TCCD, 1,2,3,7,8-PeCDD, and
1,2,3,4,7,8-HxCDD demonstrate that the slopes are parallel and the REPs are 0.2 and 0.04,
5/22/00
9-19 DRAFT—DO NOT CITE OR QUOTE
-------
1 respectively. According to the authors, the dose response for a mixture of these chemicals, in
2 which the components were at equally potent concentrations, further demonstrated the response
3 additivity of mixtures of PCDDs and the predictive ability of the TEF methodology (Gao et al,
4 1999).
5 The research on the interactions between mixtures of PCDDs and PCDFs has taken two
6 approaches. The first is to derive REP vahaes in the same system in which the mixtures shall be
7 tested. These studies confirm that the assumption of additivity can predict the response of
8 mixtures of PCDDs and PCDFs. A second approach is to use the I-TEFs to assess the potency of
9 a mixture. These studies tend to indicate that the I-TEFs overestimate the potency of a mixture
10 by factors of two to four. Recently, the WHO TEFs have been described as "order of magnitude"
11 estimates of the potency of dioxin-like chemicals. However, the studies using consensus TEFs
12 demonstrate that for mixtures of PCDDs and PCDFs, the TEF methodology will predict within a
13 half-order of magnitude or less (Schrenck et al., 1994; Nagao et al., 1993). In either case, the
14 TEF methodology accurately predicts the responses of experimentally defined mixtures of
15 PCDDs and PCDFs.
16
17 9.4.2. Examination of Commercial or Laboratory-Derived Mixtures of PCDDs, PCDFs,
18 and PCBs
19 Commercial mixtures of PCBs elicit a broad spectrum of biological and toxicological
20 responses in both experimental animals and humans. Some of the observed effects resemble
21 those induced by dioxin and furans (enzyme induction, immunotoxicity, teratogenicity, endocrine
22 alterations, etc.). Attempts to expand the TEF approach to risk assessment of PCBs have
23 investigated the ability of both commercial PCBs and individual congeners, selected on the basis
24 of structure-activity relationships, to induce dioxin-like effects and to interact with TCDD. One
25 of the first studies to examine the interactions of individual PCB congeners with TCDD used
26 mouse teratogencity as an endpoint (Birnbaum et al., 1985, 1987). A rnono-ortho PCB
27 (2,3,455,3',41-HxPCB or PCB 156) at doses of 20 mg/kg or higher (Binibaum, 1991) induced
28 hydronephrosis and cleft palate in mice. When mice were co-exposed to PCB 156 and 3.0 (ig
29 TCDD/kg the interactions resulted in strict additivity.
30 The interaction of TCDD with dioxin-like PCBs has been examined by van Birgelen et al.
31 (1994a,b) in subchronic rat feeding studies. Concentrations of PCB 126 in the diet between 7
32 and 180 ppb induced several dioxin-like effects, including CYP1A1 induction, thymic atrophy,
33 liver enlargement, and decreases in hepatic retinol concentrations, body weight gains, and plasma
34 thyroxine concentrations. The REP for PCB 126 was estimated by the authors at between 0.01
35 and 0.1 (van Birgelen et al., 1994a). Co-exposure to PCB 126 and TCDD (0.4 or 5.0 ppb) in the
i
36 diet demonstrated additivity for all responses except induction of CYP1A2 and decreases in
5/22/00
9-20
DRAFT—DO NOT CITE OR QUOTE
-------
hepatic retinol, where antagonism occurred at the highest doses of PCB 126 and TCDD tested.
These nonadditive interactions were not observed at more environmentally relevant exposures,
according to the author. In a similar study design, PCB 156 alsoinduced dioxin-like effects with
a REP estimated between 0.00004 and 0.001 (van Birgelen et al., 1994b). Similar to the
interactions between PCB 126 and TCDD, additive interactions were observed in animals
receiving mixtures of PCB 156 and TCDD in the low-dose region for all responses examined.
However, at the highest exposures of PCB 156 and TCDD, the authors reported slight
antagonistic interactions for decreases in hepatic retinol (van Birgelen et al., 1994b). For both
PCB 126 and PCB 156, antagonistic interactions were observed with TCDD only at exposures
that produced maximal CYP1 Al induction. The authors concluded that the antagonistic
interactions are unlikely to occur at relevant human exposures.
In a series of studies examining the TEF methodology, TCDD (1.5-150 ng/kg/d),
1,2,3,7,8-PeCDD; 2,3,7,8-TCDF; 1,2,3,7,8-PeCDF; 2,3,4,7,8-PeCDF; OCDF; the co-planar
PCBs 77, 126, and 169; and the mono-ortho substituted PCBs 105, 118, and 156 were
administered to mice 5 days/week for 13 weeks. REPs were determined for EROD induction, a
marker for CYP1 Al, in liver, lung, and skin; ACOH activity, a marker for CYP1A2, in liver;
and hepatic porphyrins (DeVito et al., 1997a; 2000; van Birgelenet al., 1996c). These data '
demonstrate that the dose-response curves for the PCDDs and PCDFs were parallel (DeVito et
al., 1997a). Dose-response curves for some of the enzyme induction data for the individual
PCBs displayed evidence of non-parallelism in the high-dose region (DeVito et al., 2000). A
laboratory-derived mixture of these chemicals with congener mass ratios resembling those in
food was administered to mice and rats, and indicated that despite the evidence of non-
parallelism for the PCBs, the assumption of additivity predicted the potency of the mixture for
enzyme induction, immunotoxicity, and decreases in hepatic retinoids (Birnbaum and DeVito,
1995; van Birgelen et al., 1996; 1997; DeVito et al., 1997; Smialowicz et al., 1996). In addition,
the REPs estimated in mice also predicted the response of the mixture in rats for enzyme
induction and decreases in hepatic retinyl palmitate concentrations (van Birgelen et al., 1997d;
Ross et al., 1997; DeVito et al., 1997b). These studies indicate that not only do the REPs for
enzyme induction in mice predict other responses, such as immunotoxicity and decreases in
hepatic retinyl palmitate, they also can be used to predict responses of mixtures in another
species.
The commercial PCB mixtures induce a variety of dioxin-like effects. Rats exposed to
commercial Aroclors and observed for 2 weeks exhibited dose-dependent induction of hepatic
CYP1A activity (EROD) but no thymic atrophy (Harris et al., 1993). Using REP values derived
for EROD induction in rats, the TEF methodology provided good agreement with experimental
estimates of the ED50 for enzyme induction. However, use of the conservative TEF values of
5/22/00
9-21 DRAFT—DO NOT CITE OR QUOTE
-------
1 Safe (1990) overestimated the potency of the Aroclor mixutres (Harris et al., 1993). In contrast,
2 similar studies examining immunotoxiciry as an endpoint demonstrate that both experimentally
3 derived REP values and the conservative TEF values of Safe (1990) overestimate the potency of
4 the Aroclor mixtures by a factor of 1.2 - 22 (Harper et al., 1995). These data demonstrate that
5 there are nonadditive interactions between dioxin-like chemicals and the non-dioxin-like PCBs
6 and that these interactions are response specific and most likely are not due to AhR antagonism.
7 In in vitro systems, using H4IIe cells and rat hepatocytes, Schmitz et al. (1995, 1996)
8 examined the assumption of additivity for individual congeners as well as commercial mixtures.
9 After deriving REP values for enzyme induction, the authors concluded that a laboratory mixture
10 of PCBs 77, 105, 118, 126, 156, and 169 demonstrated perfect additive behavior in these cell line
11 systems (Schmitz et al., 1995). However, when the mixture was combined with a tenfold surplus
12 of a mixture containing non-dioxin-like PCBs (PCB 28, 52, 101, 138, 153 and 180), the mixture
13 demonstrated an approximate threefold higher TEQ than predicted. The authors concluded that a
14 moderate synergistic interaction is responsible for the increased enzyme-inducing potency of the
15 mixture containing dioxins and non-dioxin-like PCBs. Further studies by Schmitz et al. (1996)
16 also demonstrated a slight synergistic deviation (less than threefold) from strict additivity when
17 the calculated TEQ based on chemical analysis of Aroclor 1254 and Clophen A50 was compared
18 to the CYP1 A-induction TEQ derived in an established rat hepatoma cell line (H4IIE) (Schmitz
19 etal., 1996).
20 Researchers have evaluated the applicability of the TEF methodology to mixtures
21 containing dioxin-like PCBs by examining the interactions of binary mixtures, laboratory-derived
22 mixtures, or commercial mixtures of PCBs. The studies examining the binary mixtures or
23 laboratory-derived mixtures have demonstrated that the assumption of additivity provides good
24 estimates of the potency of a mixture of PCBs and other dioxin-like chemicals. In contrast,
25 studies using commercial mixtures of PCBs suggest that the assumption of additivity may be
26 endpoint specific, and that both synergistic and antagonistic interactions may occur for some
27 mixtures of dioxins and PCBs for certain endpoints. A more detailed examination of these issues
28 follows hi the section on nonadditive interactions with non-dioxin-like chemicals.
29
30 9.4.3. Examination of Environmental Samples Containing PCDDs, PCDFs, and/or PCBs
31 One of the first tests of the TEF methodology examined soot from a transformer fire in
32 Binghamton, NY (Eadon et al., 1986). Benzene extracts of soot from a PCB transformer fire
33 which contained a complex mixture of PCDDs, PCDFs, PCBs, and polychlorinated
34 biphenylenes were administered to guinea pigs as single oral doses, and LD50 values were
: . i
35 compared to TCDD. Relative potency values for the PCDDs and PCDFs based on guinea pig
36 LD50 values were used to estimate the TCDD equivalents of the mixture. Eadon and co-workers
5/22/00
9-22
DRAFT—DO NOT CITE OR QUOTE
-------
exposed guinea pigs to either TCDD alone or the soot and determined their LD50s. With these
relative potency values, the soot extract had a TCDD equivalent concentration of 22 ppm.
Comparison of the LDSOs for TCDD and the soot led to a TCDD equivalent of 58 ppm for the
mixture. Other endpoints examined included alterations in thymus weight, body weight, serum
enzymes, and hepatotoxicity. Experimentally the TCDD equivalents of the soot varied from 2 to
58 ppm. The authors concluded that because the benzene extract of the soot contained hundreds
of chemicals including PCDDs, PCDFs, and PCBs, the difference between the calculated TEQ of
22 ppm and the experimentally derived TEQs between 2 and 58 seems minimal. (Note: the
initial analytical TEQ value of soot [22 ppm] was calculated on the basis of guinea pig LD50
values of the respective components; using the current recommended TEF scheme [van den Berg
et al, 1998], the "calculated" TCDD TEQ would be approximately 17 ppm.)
Shortly after the studies on the Binghamton transformer fire soot, investigators applied the
TEF methodology to the leachate from Love Canal, NY. The organic phase of the leachate
consisted of more than 100 different organic compounds including PCDDs and PCDFs. The
leachate did not contain PCBs or PAHs. The authors estimated the TEQ of the mixture on the
basis of REP values for teratogenicity (cleft palate and hydronephrosis in mice) for the PCDDs
and PCDFs present in the leachate. The authors state that the leachate contained the equivalent
of 3 [ig TCDD/g and that more than 95% of the TEQ was contributed by TCDD. There were two
other PCDFs present in the leachate, and then- contribution to the total TEQ was approximately
5% (Silkworth et al., 1989). When the TEQ of the mixture was based on dose-response analysis
of the mixture compared to TCDD, the leachate was estimated to contain between 6.6 and 10.5
u.g TCDD/g (Silkworth et al., 1989). The authors concluded, there was a good agreement
between the experimental TCDD equivalents (6.6-10.5 ng TCDD/g) and the analytical TEQs (3
jig TCDD/g). In addition, these studies illustrate that the non-AhR components of the leachate
did not interfere with receptor-mediated teratogenicity (Silkworth et al., 1989). Additional
investigations have shown that the same complex mixture of non-AhR agonists slightly
potentiated TCDD-induced thymic atrophy and immunosuppression (plaque-forming cells/spleen
response) while decreasing the hepatic CYP1 A-inducing ability of the TCDD component
(Silkworth et al., 1993).
The assumption of additivity was also examined using a PCDD/PCDF mixture extracted
from fly ash from a municipal waste incinerator (Suter-Hofrnann and Schlatter, 1989). As a
purification step, rabbits were the fed organic extracts from the fly ash. After 10 days the livers
were removed and analyzed for PCDDs and PCDFs. The rabbit livers contained predominately
2,3,7,8-substituted PCDDs/PCDFs. Based on the chemical analysis of the liver, pulverized liver
lyophilisate was added to the standard rat diet. This diet was fed to rats for 13 weeks and body
weights and terminal thymus weights were recorded. The authors concluded that the mixture of
5/22/00
9-23 DRAFT—DO NOT CITE OR QUOTE
-------
1 PCDDs and PCDFs produced equivalent toxicities as TCDD, and the assumption of additivity
2 was confirmed.
3
4 9.4.4. Nonadditive Interactions With Non-Dioxin-Like Chemicals
5 For a number of toxicological responses, there appears to be evidence for nonadditive
6 interactions in defined dose ranges by both commercial Aroclors and major congeners with little
7 if any AhR agonist activity (i.e., PCB 153). Both commercial Aroclors and a PCB mixture
8 comprised of major congeners found in human breast milk were shown to antagonize the
9 immunotoxic effects of TCDD in mice (Biegel et al., 1989; Davis and Safe, 1989; Harper et al.,
10 1995). When immunotoxicity-derived TEF values for a variety of PCB congeners were used in
11 an additive manner to estimate TCDD TEQs for commercial Aroclors, in comparison to the
12 experimental TEQs, they were approximately predictive for Aroclor 1254 and 1260 (Harper et
13 al., 1995). However, the TEF approach tended to overestimate the immunotoxicity of Aroclors
14 1242 and 1248, suggesting some antagonism.
15 Typical responses to TCDD exposure in rodents include CYP1 enzyme induction and
: I ;
16 thymic atrophy. Rats consuming a diet containing 5 ppb TCDD for 13 weeks exhibited a 33-fold
17 increase in hepatic CYP1A activity (EROD) and a greater than 50% reduction in relative thymus
18 weight. Addition of PCB 153 to the diet at concentrations up to 100 ppm had no significant
19 effect on either response (van der Kolk et al., 1992). Mice dosed simultaneously with TCDD and
20 up to a 106-fold molar excess of PCB 153 (1 nmol/kg.vs. 1 mmol/kg) exhibited no significant
21 dose-dependent alteration in hepatic CYP1A1/A2 protein compared to the TCDD dose group
22 alone (De Jongh et al., 1995). There was, however, an approximate twofold increase in hepatic
23 EROD activity in the highest combined PCB 153 :TCDD dose group. Subsequent tissue analysis
24 revealed that the increase in EROD activity was probably related to PCB 153 increasing hepatic
25 TCDD concentrations. The same PCB congener at high doses (358 mg/kg) is able to almost
26 completely inhibit TCDD-induced suppression of the plaque-forming cell (PFC) response toward
27 sheep red blood cells in male C57BL/6J mice (Biegel et al., 1989; Smialowicz et al., 1997).
28 However, as PCB 153 displays negligible AhR binding affinity, the exact mechanism(s) behind
29 these interactions is unknown. Recently, it has been shown that PCB 153 at high doses (greater
30 than 100 mg/kg) actually enhances the PFC response in female B6C3F1 mice, thereby raising the
31 "control" set point. When combined doses of TCDD and PCB 153 are then compared to the
32' elevated PCB 153 response, an immunosuppressive effect is observed (Smialowicz et al., 1997).
33 The relevance of this functional antagonism is uncertain, as the doses required to inhibit the
34 TCDD-like effects are at least 100 mg/kg of PCB 153. These doses of PCB 153 seem unlikely to
35 occur in human populations except under extreme conditions.
5/22/00
9-24
DRAFT—DO NOT CITE OR QUOTE
-------
1
Commercial PCBs and various PCB congeners have been .shown to potentiate or
antagonize the teratogenicity of TCDD depending upon the dose, ranges and response examined
(Biegel et al., 1989; Morrissey et al., 1992). Treatment of developing chicken embryos with
TCDD and dioxin-like PCBs induces a characteristic series of responses, including embryo
lethality and a variety of embryo malformations/deformities. Combined exposure of chicken
embryos to both PCB 126 and PCB 153 (2 fig/kg and 25-50 mg/kg, respectively) resulted in
protection from PCB 126-induced embryo malformations, edema, and liver lesions, but not
mortality (Zhao et al., 1997). In mice, doses of 125 mg PCB 153/kg or higher inhibit the
induction of cleft palate by TCDD (Biegel et al., 1989; Morrissey et al., 1992). The induction of
hydronephrosis by TCDD was slightly antagonized by PCB 153, but only at doses of 500 mg/kg
or higher. Once again, the environmental relevance of exposures of 100 mg/kg of PCB 153 or
higher remains quite speculative, and nonadditive interactions are not expected at environmental
exposures.
Nonadditive interactions have also been observed in rodents exposed to both TCDD and
mixtures of various PCB congeners for hepatic porphyrin accumulation and alterations in
circulating levels of thyroid hormones. A strong synergistic response was seen with hepatic
porphyrin accumulation in female rats following the combined dietary exposure to TCDD and
PCB 153 (van Birgelen, 1996a). The mechanism accounting for the interaction was thought to
be a combination of both AhR-dependent (CYP1A2 induction) and AhR-independent (6-
aminolevulinic acid synthetase [ALAS] induction) events. Additionally, subchronic exposure of
mice to a mixture of PCDDs, PCDFs, and dioxin-like PCBs in a ratio derived from common
foods also resulted in a highly synergistic response, when compared to an equivalent dose of
TCDD alone, for both hepatic porphyrin accumulation and urinary porphyrin excretion (van
Birgelen et al., 1996b). PCB 153, although not porphyrinogenic alone, when added to the
mixture further enhanced the synergistic response of hepatic porphyrin accumulation. Non-AhR-
mediated induction of ALAS activity by both the dioxin-like mono ortho-substituted PCBs in the
mixture and by PCB 153 was hypothesized to partially explain the synergism.
Decreases in thyroid hormone levels have been observed in both experimental animals and
humans following exposure to both dioxin-like and non-dioxin-like compounds (Nagayama et
al., 1998; Koopman-Esseboom et al., 1997). It is currently thought that multiple mechanisms,
including induction of specific isozymes of hepatic UDP-glucoronyl transferase (UDPGT) and
binding to thyroid hormone transport proteins (thyroid binding globulin, transthryetin) could be
involved. Exposure of female rats to a food-related mixture of PCDDs, PCDFs, and dioxin-like
PCBs for 90 days resulted in an approximately 85% decrease in decrease in plasma levels of
thyroxine. In contrast, the TCDD equivalent dose produced no effect on serum thyroxine (van
Birgelen et al., 1997). Increased induction of several isoforms of UDPGT by the HAH mixture
5/22/00
9-25 DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
as compared to TCDD was thought to only partially explain the observed response with
thyroxine levels.
Several studies examining the interactions of dioxins and non-dioxins for rat liver tumor
promotion and additive and nonadditive interactions have been reported. Synergistic interactions
for tumor promotion have been observed for combinations of PCB 77 and PCB 52 (2,2',5,5T-
tetrachJorbiphenyl) in rat liver (Sargent et al., 1992). Eager et al. (1995) reported greater than
additive interactions of PCBs 126 and 153 in a rat liver tumor promotion model. The assumption
of additivity was examined in a laboratory-derived mixture of PCDDs, PCDFs, and PCBs in a rat
liver tumor promotion model (van der Plas et al., 1999). The mixture contained TCDD,
1,2,3,7,8-PeCDD, 2,3,4,7,8-PeCDF, and PCBs 126, 118, and 156. In addition, a dose-response
study was performed using the mixture with PCB 153 added, van der Plas and colleagues
concluded that the TEF methodology predicted the tumor-promoting potency of the mixture quite
well, within a factor of two (van der Plas et al., 1999).
The interactions of dioxins with non-dioxin-like chemicals results in additive and
nonadditive responses. The antagonistic interactions, while endpoint specific, appear to occur at
dose levels that greatly exceed most human exposures and should not affect the overall use of the
TEF methodology. One of the difficulties in addressing the nonadditive interactions is
understanding the mechanism behind these interactions. For the greater than additive
interactions for induction of porphyria and decreases in serum thyroxine, there are hypotheses
that may explain these effects. The mechanism of the antagonistic interactions of non-dioxin-
like PCBs and TCDD on immunotoxicity and teratogenicity in mice is uncertain. For other
responses, such as developmental reproductive toxicity, the interactions of PCDDs, PCDFs, and
PCBs have not been examined. In addition, it has also been suggested that antagonism of Ah
receptor-mediated events may be species specific. For example, addition of PCB 52, a congener
commonly found in biotic samples, inhibited the TCDD-induced expression of a reporter gene
under the regulatory control of the Ah receptor hi mouse and rat cells, but not in guinea pig or
human hepatoma cells (Aarts et al., 1995). Our limited understanding of the interactions
between dioxins and non-dioxins for a variety of responses requires further research before their
impact on the TEF methodology can be fully understood.
9.4.5. Examination of the TEF Methodology in Wildlife
Many wildlife species also exhibit toxic effects associated with exposure to halogenated
aromatic hydrocarbons. Early life stage (ELS) or sac fry mortality in fish, characterized by
edema, structural malformations, and growth reduction prior to fry mortality can be induced in
trout species following exposure to dioxin-like PCDDs, PCDFs, and PCBs-(Walker and
Peterson, 1991). Binary combinations of a variety of PCDDs, PCDFs, and both dioxin and non-
5/22/00
9-26 DRAFT—DO NOT CITE OR QUOTE
-------
dioxin-like PCB congeners injected into fertilized trout eggs were also capable of inducing ELS
mortality, with the majority of interactions between the congeners described as strictly additive
(Zabel et al., 1995). When a synthetic complex mixture of PCDDs, PCDFs, and PCBs, in
congener ratios that approximated Great Lakes fish residues, was tested in the ELS mortality
assay, the lethal potency observed for the mixture, compared to TCDD, deviated less than
twofold from an additivity approach (Walker et al., 1996). Recently, the TCDD TEQ of an
environmental complex mixture of PCDDs, PCDFs, and PCBs extracted from lake trout and
applied to the ELS bioassay could also be predicted by an additivity approach (Tillitt and Wright,
1997). These results suggest that additional halogenated aromatic compounds, including non-
dioxin-like PCBs, present in fish do not significantly detract from an additivity response for this
AhR-mediated event.
There are also numerous studies that have examined the effects of environmental mixtures
in marine mammals and avian species (Ross, 2000; Giesy and Kannan, 1998; Ross et al., 1996;
Shipp et al., 1998a,b; Resrum et al., 1998; Summer et al., 1996a,b). Ross and colleagues
examined captive harbor seals fed herring from either the Atlantic Ocean (low levels of
PCDDs/PCDFs/PCBs) or the Baltic Sea (high levels of PCDDs/PCDFs/PCBs). The seals fed
herring from the Baltic Sea displayed immunotoxic responses including impaired natural killer
cell activity and antibody responses to specific antigens. These effects were correlated with the
TEQ concentrations in the herring. Using mink as a model, Aulerich, Bursian, and colleagues
have also examined the TEF methodology. Minks were fed diets containing carp from Saginaw
Bay to provide exposures of 0.25, 0.5, or 1 ppm PCB in the diet. In a series of reports, the
authors demonstrated that the diet induced dioxin-like effects ranging from enzyme induction to
reproductive and developmental effects, and that these effects were correlated with the dietary
intake of TEQs (Giesy and Kannan, 1998). Similar studies in White Leghorn hens also
demonstrated that the TEQ approach provided accurate estimates Of the potency of the mixtures
(Summer et al., 1996).
In summary, current experimental evidence suggests that for PCDDs, PCDFs, co-planar
dioxin-like PCBs, and strictly AhR-mediated events, the concept of TEF additivity adequately
estimates the dioxin-like toxicity of either synthetic mixtures or environmental extracts, despite
the variations in relative contributions of each congener. Addition of the more prevalent mono-
and di-ortho-substituted PCBs to a mixture, at least in the case of environmental extracts and
wildlife, does not seem to significantly detract from this assumption of additivity. Interactions
other than additivity (antagonism, synergism) have been observed with a variety of effects
(teratogenicity, irnmunotoxicity, hepatic porphyrin accumulation, thyroid hormone metabolism)
in both binary combinations and complex synthetic mixtures of dioxin and partial or non-Ah
receptor agonists (commercial PCBs, PCB 153). However, it appears that at these high-dose
5/22/00 9-27 DRAFT—DO NOT CITE OR QUOTE
-------
1 exposures, multiple mechanisms of action not under the direct control of the Ah receptor are
2 responsible for these nonadditive effects.
3 Additional research efforts should focus on complex mixtures common to both
4 environmental and human samples and the interactions observed with biological and
'I
5 lexicological events known to be under Ah receptor control. In the interim, the additive
6 approach with TEFs derived by scientific consensus of all available data appears to offer a good
7 estimation of the dioxin-like toxicity potential of complex mixtures, keeping in mind that other
8 effects may be elicited by non-dioxin-like components of the mixture.
9
10 9.4.6. Toxic Equivalency Functions
11 The TEF methodology has been described as an "interim" methodology. Since this
12 interim method has been applied, there have been few proposed alternatives published. One
13 recent proposal suggests that the TEF value be replaced by a toxic equivalency function
14 (Putzrath, 1997). It has been proposed that the REPs for PCDDs/PCDFs are better described by
15 a function as compared to a factor or single-point estimate (Putzrath, 1996). Recent studies have
16 examined this possibility for a series of PCDDs/PCDFs and PCBs (DeVito et al., 1997; DeVito
17 et al., 2000). For the PCDDs/PCDFs, the data indicate that the REPs estimated from enzyme
18 induction data hi mice are best described by a factor and not a function. For some of the PCBs
19 examined, a function fit better, but the change in the REP was within a factor of two to five for
20 most of the four enzymatic responses examined (DeVito et al., 2000). In addition, the dose
21 dependency was observed only at the high-dose and not hi the low-dose region (DeVito et al.,
22 2000).
23 Even though these studies suggest that a TE function may be useful, there are numerous
24 difficulties in applying this method. If the REPs are really functions and not factors, there must
25 be a mechanistic basis for these differences, and these mechanisms would most likely be
26 response specific and perhaps species specific. This would then require that for all critical
27 responses, every chemical considered in the TEF methodology would, have to be examined.
28 Once again, it is highly unlikely that 2-year bioassays and multigenerational studies will be
29 performed on all the TEF congeners in the foreseeable future. The use of a TEF function
30 requires extensive data sets that are not available and are unlikely to be collected.
31 There are instances where exposures to PCBs are the major problem. The TEF
32 methodology provides risk assessors with a useful tool to estimate potential dioxin-related health
33 risks associated with these exposures. Typically, the congener makeup of environmental
34 exposures to PCBs does not resemble the congener profile of any of the commercial mixtures
35 produced. Because the environmental mixtures do not resemble the commercial mixtures, it is
36 not clear that using total PCB concentrations and comparing them to any of the commercial
5/22/00
9-28
DRAFT—DO NOT CITE OR QUOTE
-------
mixtures provides an accurate assessment of the potential risks. However, the use of the TEF
methodology allows for the estimation of the risk associated with the dioxin-like effects of the
mixture and may provide a more accurate assessment of the risk in conjunction with the use of
total PCBs. The Agency has recently published an application of this approach to the evaluation
of PCS carcinogenicity (U.S. EPA, 1996, Cogliano, 1998)
9.4.7. Endpoint and Dose-Specific TEFs
It is often suggested that species, endpoint, and dose-specific TEFs may be required for the
TEF concept to provide accurate estimates of risk. Although these proposals are interesting,
specific TEFs would require a much more complete data set than is available at this time. One
reason the TEF methodology was developed was because these data are not available, and it was
unlikely that all relevant chemicals would be tested for all responses in all species, including
humans. For example, it is extremely unlikely that 2-year bioassays for carcinogenesis or multi-
generational studies will be performed on all chemicals included in the TEF methodology. Even
though there are significant data demonstrating that a number of chemicals produce dioxin-like
toxic effects, clearly the data set is not complete. For this reason, WHO recommends revisiting
these values every 5 years.
9.5. UNCERTAINTY
TEFs are presented as point estimates, in spite of the feet that variability hi supporting
experimental data can range several orders of magnitude for a particular congener. It has been
proposed that some of this variability can be attributed to differences in exposure regimens, test
species, or purity of the test compound; however, the reasons for much of this variability have
not been adequately examined experimentally and remain unknown. Because of the multiple
methods of deriving the REP values for a particular chemical, it is difficult to estimate the
variability or uncertainty of a TEF point estimate. Consequently, the TEQ approach as currently
practiced does not provide for a quantitative description of the uncertainty for individual TEF
values, nor has any proposed method for incorporating quantitative uncertainty descriptors into
TEFs received general support or endorsement from the scientific community. Suggestions have
been made to use meta-analytic approaches or Monte Carlo techniques, however (Finley et al.,
1999), and these approaches are only as good as the data available. Given the incompleteness of
the available database, it seems unlikely that these approaches would provide much useful insight
at this tune.
Qualitative statements of confidence are embodied in the discussions associated with the
establishment and revision of TEFs. These qualitative judgments, when examined in the context
of a specific risk assessment, can provide valuable insight into the overall uncertainty of some
5/22/00
9-29 DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
TEQ estimates. For example, using WHO TEFs (van den Berg et al., 1998) to look at
background exposure from a typical U.S. diet, it is clear that only a limited number of congeners
significantly contributed to the total TEQ. More than 60% of the TEQvmojg associated with
background dietary exposure (1 pg/kg/d) comes from only four congeners: 2,3,7,8-TCDD (8% ),
1,3,7,8-PCDD (21.5% ), 2,3,4,7,8-PeCDF (10.7%), and PCB 126 (21%) (EPA Exposure Volume
III). The variability of the REP values found in the literature for these congeners is much lower
than for congeners that are minor contributors to background TEQ. The confidence in the TEFs
for major congener constituents of background exposure (or other exposure with a similar
; |, i
congener profile) has consistently been determined empirically to be within a factor of 2-3, but it
is unlikely that the estimated TEQ overestimates the "true" TEQ by more than a factor of five.
Additionally, for exposures in the background range it is unlikely that non-dioxin-like PCBs
significantly affect the uncertainty of TEQ estimates based upon the earlier discussions of
additivity. The uncertainty in TEQ estimates is only one component of the overall uncertainty in
a dioxin risk assessment. The TEQ uncertainty only addresses the co:nfidences associated in
ascribing 2,3,7,8-TCDD equivalents to a mixture. It does not address the uncertainly associated
with quantitatively linking health effects to 2,3,7,8-TCDD exposure, or the uncertainties
associated with exposure estimates themselves.
9.6. IMPLICATIONS FOR RISK ASSESSMENT
The TEF methodology provides a mechanism to estimate potential health or ecological
effects of exposure to a complex mixture of dioxin-like chemicals. However, the TEF method
must be used with an understanding of its limitations. This methodology estimates the dioxin-
like effects of a mixture by assuming dose-additivity and describes the mixture in terms of an
equivalent mass of 2,3,7,8-TCDD. Although the mixture may have the toxicological potential of
2,3,7,8-TCDD it should not be assumed for exposure purposes to have the same environmental
fate as 2,3,7,8-TCDD. The environmental fate of the mixture is still the product of the
environmental fate of each of its constituent congeners. Different congeners have different
physical properties such as vapor pressure, practical vapor partition, water octanol coefficient,
photolysis rate, binding affinity to organic mater, water solubility, etc; Consequently, both the
absolute concentration of a mixture in an environmental medium and 'the relative concentration
of congeners making up an emission will change as the release moves through the environment.
For some situations, treating emission as equivalent to exposure, which assumes that modeling
fate and exposure can be reasonably accomplished by treating a mixture as if it were all 2,3,7,8,-
TCDD, is a useful but uncertain assumption. However, for many risk assessments the
differences in fate and transport of different congeners must be taken into consideration and TEQ
must be calculated at the point of exposure if more accurate assessments are to be achieved.
5/22/00
9-30
DRAFT—E)O NOT CITE OR QUOTE
-------
Similarly, many dioxin releases are associated with the release of non-dioxin-like compounds
such as pesticides, metals, and non-dioxin-like PHAHs, and their risk potential may also need to
be assessed in addition to dioxin-related risk.
9.7. SUMMARY
AhR mediates the biochemical and lexicological actions of dioxin-like chemicals and
provides the scientific basis for the TEF/TEQ methodology. In its 20-year history, this approach
has evolved, and decision criteria supporting the scientific judgment and expert opinion used in
assigning TEFs have become more transparent. Numerous countries and several international
organizations have evaluated and adopted this approach to evaluating complex mixtures of
dioxin and related compounds. It has become the accepted interim methodology, although the
need for research to explore alternative approaches is widely endorsed. Although this method
has been described as a "conservative, order of magnitude estimate" of the TCDD dose,
experimental studies examining both environmental mixtures and laboratory-defined mixtures
indicate that the method provides a greater degree of accuracy and may not be as conservative as
described. Clearly, basing risk on TCDD alone or assuming all chemicals are as potent as TCDD
is inappropriate on the basis of available data. Although uncertainties in the TEF methodology
have been identified, one must examine this method in the broader context of the need to
evaluate the public health impact of complex mixtures of persistent bioaccumulative chemicals.
The TEF methodology decreases the overall uncertainties in the risk assessment process (U.S.
EPA, 1999); however, this decrease cannot be quantified. One of the limitations of the TEF
methodology in risk assessment is that the risk from non-dioxin-like chemicals is not evaluated.
Future research should focus on the development of methods that will allow risks to be predicted
when multiple mechanisms are present from a variety of contaminants.
5/22/00
9-31
DRAFT—DO NOT CITE OR QUOTE
-------
Table 9-1. Estimated relative toxicity of PCDD and PCDF isomers to 2,3,7,8-
Isomer groups
DD
M,CDD
D2CDD
T3CDD
T
-------
Table 9-2. Toxic equivalency factors (TEFs)
Congener
EPA/87 a
NATO/89 b
WHO/94 c
WHO/97d
PCDDs
2,3,7,8-TCDD
1,2,3,7,8-PeCDD
1,2,3,4,7,8-HxCDD
1,2,3,7,8,9-HxCDD
1,2,3,6,7,8-HxCDD
1,2,3,4,6,7,8-HpCDD
1,2,3,4,6,7,8,9-OCDD
1
0.5
0.04
0.04
0.04
0.001
0
1
0.5
0.1
0.1
0.1
0.1
0.001
1
1
0.1
0.1
0.1
0.01
0.0001
PCDFs
2,3,7,8-TCDF
1,2,3,7,8-PeCDF
2,3,4,7,8-PeCDF
1,2,3,4,7,8-HxCDF
1,2,3,7,8,9-HxCDF
1,2,3,6,7,8-HxCDF
2,3,4,6,7,8-HxCDF
1,2,3,4,6,7,8-HpCDF
1,2,3,4,7,8,9-HpCDF
1,2,3,4,6,7,8,9-OCDF
0.1
0.1
0.1
0.01
0.01
0.01
0.01
0.001
0.001
0
0.1
0.05
0.5
0.1
0.1
0.1
0.1
0.01
0.01
0.001
0.1
0.05
0.5
0.1
0.1
0.1
0.1
0.01
0.01
0.0001
PCBs
IUPAC # Structure
77 3,3',4,4'-TCB
81 3,4,4',5-TCB
105 2,3,3',4,4'-PeCB
114 2,3,4,4',5-PeCB
118 2,3'A4',5-PeCB
123 2',3,4,4',5-PeCB
126 3,3',4,4',5-PeCB
156 2,3,3',4,4',5-HxCB
157 2,333',454'35'-HxCB
167 253',4,4',555I-HxCB
169 3,3'A4',5,5I-HxCB
170 2,2',3,3'A4135-HpCB
180 2,2'53,454'55,5'-HpCB
189 23353l,454',555'-HpCB
0.0005
0.0001
0,0005
0.0001
0.0001
0.1
0.0005
0.0005
0.00001
0.01
0.0001
0.00001
0.0001
0.0001
0.0001
0.0001
0.0005
0.0001
0.0001
0.1
0.0005
0.0005
0.00001
0.01
0.0001
a U.S. EPA, 1987.
bNATO/CCMS, 1989.
c Alhlborg et al., 1994.
d van Leeuwen, 1997.
5/22/00
9-33
DRAFT—DO NOT CITE OR QUOTE
-------
TCDD (2,3,7,8)
TCDF (2,3,7,8)
Cl
Cl
2,3,3',4,4'-PeCB
Cl Cl
2,2',4,4I,5,5'-HCB
Figure 9-1. Structures of polychlorinated dibenzo-p-dioxins, diberizofurans and biphenyls.
The prototype chemical 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD[2,357581), and example of a
dioxin-hke dibenzofuran 2,3,7,8-tetrachlorodibenzfuran (TCDF[2,3,7,8J), a mono-ortho dioxin-
like PCB, 253,3'34,4l-pentachlorobiphenyl (2,3,3',4,4'-PeCB), a dioxin-like co-planar PCS,
3,3',4,4',5-pentachlorobiphenyl (3,3',4,4',5-PeCB) and an example of a non-dioxin-like di-ortho
substituted PCB9 252',4,4',5,5'-hexachlorobiphenyl (2,2',4,4'5555I-HCB).
5/22/00
9-34
DRAFT—DO NOT CITE OR QUOTE
-------
REFERENCES FOR CHAPTER 9
Aarts, JMMJG; Denison, MS; Cox, MA; et al. (1995) Species-specific antagonism of Ah receptor action by 2,2',5,5'-
tetrachloro- and 2,2',3,3',4,4'-hexachlorobiphenyl. Eur J Pharmacol- Environ Toxicol Pharmacol Sect 293(4):463-
474. ' ;
Abbott, BD; Bimbaum, LS; Perdew, GH. (1995) Developmental expression of two members of a new class of
transcription factors: I. Expression of aryl hydrocarbon receptor in the C57BL/6N mouse embryo. Devel Dyn
204(2):133-143.
Abbott, BD; Schmid, JE; Pitt, JA; et al. (1999) Adverse reproductive outcomes in the transgenic Ah
receptor-deficient mouse. Toxicol Appl Pharmacol 155(1):62-70.
Ahlborg, UG; Brouwer, A; Fingerhut,, MA; et al. (1992) Impact of polychlorinated dibenzo-p-dioxins,
dibenzofurans, and biphenyls on human and environmental health, with special emphasis on application of the toxic
equivalency factor concept. Eur J Pharmacol 228(4): 1 79- 1 99.
Ahlborg, U; Becking, GC; Bimbaum, LS; et al. (1994) Toxic equivalency factors for dioxin-like PCBs: report on a
WHO-ECEH and IPCS consultation, Dec. 1993. Chemosphere 28(6): 1049- 1067.
Eager, Y; Hemming, H; Flodstrom, S; et al. (1995) Interaction of 3,4,5s3',4'-pentachlorobiphenyl and
2,4,5,2',4',5'-hexachlorobiphenyI in promotion of altered hepatic foci in rats. Pharmacol Toxicol 77(2): 149- 154.
Bank, PA; Yao, EF; Phelps, CL; et al. (1992) Species-specific binding of transformed Ah receptor to a dioxin
responsive transcriptional enhancer. Eur J Pharmacol 228(2-3):85-94.
Barnes, D; Alford-Stevens, A; Birnbaum, L; et al. (1991) Toxicity equivalency factors for PCBs? Qual Assur
Berghard, A; Gradin, K; Toftgard, R. (1992) The stability of dioxin-receptor ligands influences cytochrome
P450IA1 expression in human keratinocytes. Carcinogenesis 13(4):651-655.
Biegel, L; Harris, M; Davis, D; et al. (1989) 2,2',4,4',5,5'- hexachlorobiphenyl as a
2,3,7,8-tetrachlorodibenzo-p-dioxin antagonist in C57BL/6J mice. Toxicol Appl Pharmacol 97(3):561-571.
Birnbaum, LS. (1994) The mechanism of dioxin toxicity: relationship to risk assessment. Environ Health Perspect
102(Suppl9):157-167.
Birnbaum, LS. (1999) TEFs: a practical approach to a real-world problem. Hum Ecol Risk Assess 5:13-23.
Birnbaum, LS; DeVito, MJ. (1995) Use of toxic equivalency factors for risk assessment for dioxins and related
compounds. Toxicology 105:391-401.
Birnbaum, LS; Weber, H; Harris, MW; et al. (1985) Toxic interaction of specific polychlorinated biphenyls and
2,3,7,8-tetrachlorodibenzo-p-dioxin: increased incidence of cleft palate in mice. Toxicol Appl Pharmacol 77:292-
302.
Birnbaum, LS; Harris, MW; Crawford, DD; et al. (1987) Teratogenic effects of polychlorinated dibenzofurans in
combination in C57BL/6N mice. Toxicol Appl Pharmacol 91:246-255.
Birnbaum LS; McDonald, MM; Blair, PC; et al. (1990) Differential toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin
(TCDD) in C57BL/6J mice congenic at the Ah locus. Fundam Appl Toxicol 15(1): 186-200.
Birnbaum, LS; Morrissey, RE; Harris, MW. (1991) Teratogenic effects of 2,3,7,8-tetrabromodibenzo-p-dioxin and
three polybrominated dibenzofurans in C57BL/6N mice. Toxicol Appl Pharmacol 107:141-152.
5/22/00
9-35
DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
55
Bjeldanes LF; Kim, JY; Grose, KR; et al. (1991) Aromatic hydrocarbon responsiveness-receptor agonists generated
o ^o/o,0!^^ " Vitr° and *"• Viv°: ''''"Pa"80115 with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Proc Natl Acad
bci JJ8(2I):9543-9547.
Bradfield, CA; Bjeldanes, LF. (1984) Effect of dietary indole-3-carbinol on intestinal and hepatic monooxygenase
glutathione S-transferase and epoxide hydrolase activities in the rat. Food Chem Toxicol 22(12):977-982.
Bradfield, CA; Bjeldanes, LF. (1987) Structure-activity relationships of dietary indoles: a proposed mechanism of
action as modifiers of xenobiotic metabolism. J Toxicol Environ Health 21(3):31 1-323.
Broadbent, TA; Broadbent, HS. (1998) 1. The chemistry and pharmacology of indole-3-carbinol(indole-3-methanol)
and 3-(methoxymethyl)indole. [Part II]. Curr Med Chem 5(6):469-491.
Chaloupka, K; Harper, N; Krishnan, V; et al. (1993) Synergistic activity of polynudear aromatic hydrocarbon
mixtures as aryl hydrocarbon (Ah) receptor agonists. Chem-Biol Inter 89:141-158.
Chen, YH; Riby, J; Srivastava, P; et al. (1995) Regulation of CYP1A1 by indolo[3,2-b]carbazole in murine
hepatoma cells. J Biol Chem 270 (38):22548-22555.
Chen, I; Safe, S; Bjeldanes, L. (1996) Indole-3-carbinol and diindolylmethane as aryl hydrocarbon (Ah) receptor
agonsits and antagonists in T47D human breast cancer cells. Biochem Pharmacol 51:1069-1076.
Cogliano, VJ. (1998) Assessing the cancer risk from environmental PCBs. Environ Health Perspect 106(6,):3 17-323.
Davis, D; Safe, S. (1989) Dose-response immunotoxicities of commercial polychlorinated biphenyls (PCBs) and
their interaction with 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicol Lett 48:35-43.
Degawa, M; Tanimura, S; Agatsuma, T; et al. (1989) Hepatocarcinogenic heterocyclic aromatic amines that induce
cytochrome P-448 isozymes, mainly cytochrome P-448H (P-450IA2), responsible for mutagenic activation of the
carcinogens in rat liver. Carcinogenesis 10(6):1 1 19-1 122.
De Jongh, J; DeVito, M; Nieboer, R; et al. (1995) Induction of cytochrome P450 isoenzymes after toxicokinetic
interactions between 2,3,7,8- tetrachlorodibenzo-p-dioxin and 2,2',4,4',5,5'-hexachlorobiphenyl in the liver of the
mouse. Fundam Appl Toxicol 25(2):264-270.
llS *123 BimbaUm> LS' (1"5:) Dioxins: model chemicals for assessing receptor-mediated toxicity. Toxicology
DeVito, MJ; Bimbaum, LS. (1996) The use of body burdens vs. daily dose in comparisons of endo- and exodioxins
and in assessing human health risks. Organohalogen Compounds 29:424-429.
DeVito MJ; Birnbaum, LS; Farland, WH; et al. (1995) Comparisons of estimated human body burdens of dioxinlike
chemicals and TCDD body burdens in experimentally exposed animals. Environ Health Perspect 103(9):820-831.
DeVito, MJ; Ross, DG; van Birgelen, APJM; et al. (1997a) The effects of mixtures of PCDDs, PCDFs and PCBs on
hepatic retinyl palmitate concentrations in mice. Organohalogen Compounds 34:49-54.
DeVito, MJ; Diliberto, JJ; Ross, DG; et al. (1997b) Dose-response relationships for polyhalogenated dioxin and
dibenzofurans following subchronic treatment in mice: I. Cyplal and Cypla2 enzyme activity in liver, lung and
skin. Toxicol Appl Pharmacol 147:267-280.
DeVito, MJ; Ross, DG; Dupuy, AE, Jr; et al. (1998) Dose-response relationships for disposition and hepatic
sequestration of polyhalogenated dibenzo-p-dioxins, dibenzofurans, and biphenyls following subchronic treatment in
mice. Toxicol Sci 46(2):223-234.
5/22/00
9-36 DRAFT—DO NOT CITE OR QUOTE
-------
DeVito, MJ; Diliberto, JJ; Ross, DG; et al. (2000) Dose-response relationships for induction of cyplal and cypla2
enzyme activity in liver, lung and skin in female mice following subchronic exposure to poychlorinated biphenyls
Toxicol Sci (2000).
Dohr, O; Li, W; Donat, S; et al. (1996) Aryl hydrocarbon receptor mRNA levels in different tissues of 2,3,7,8,-
tetrachlorodibenzo-p-dioxin-responsive and nonresponsive mice. Adv Exp Mol Biol 387:447-459.
Dolwick, KM; Schmidt, JV; Carver, LA; et al. (1993) Cloning and expression of a human Ah receptor cDNA Mol
Pharmacol44(5):911-917.
Eadon, G; Kaminsky, L; Silkworm, J; et al. (1986) Calculation of 2,3,7,8-TCDD equivalent concentrations of
complex environmental contaminant mixtures. Environ Health Perspect 70:221-227.
Fernandez-Salguero, P; Pineau, T; Hilbert, DM; et al. (1995) Immune system impairment and hepatic fibrosis in
mice lacking the dioxin-binding Ah receptor. Science 268:722-726.
Fernandez-Salguero, PM; Hilbert, DM; Rudikoff, S; et al. (1996) Aryl-hydrocarbon receptor-deficient mice are
resistant to 2,3,7,8-tetrachlorodibenzo- p-dioxin-induced toxicity. Toxicol Appl Pharmacol 140:173-179.
Finley, B; Kirman, C; Scott, P. (1999) Derivation of probabilistic distributions for the W.H.O. mammalian toxic
equivalency factors. Organo Halogen 42:225-228.
Fujii-Kuriyama, Y; Ema, M; Mimura; J; et al. (1995) Polymorphic forms of the Ah receptor and induction of the
CYP1A1 gene. Pharmacogenetics 5:8149-153.
Gao, X; Son, DS; Terranova, PF; et al. (1999) Toxic equivalency factors of pblychlorinated dibenzo-p-dioxins in an
ovulation model: validation of the toxic equivalency concept for one aspect of endocrine disruption. Toxicol Appl
Pharmacol 157(2): 107-116.
Giesy, JP; Kannan, K. (1998) Dioxin-like and non-dioxin-like toxic effects of polychlorinated biphenyls (PCBs):
implications for risk assessment. Crit Rev Toxicol 28(6):511-569.
Gillner, M.; Bergman, J; Cambillau, C; et al. (1985) Interactions of indoles with specific binding sites for 2,3,7,8,-
tetrachlorodibenzo-p-dioxin in rat liver. Mol Pharmacol 28:357-363.
Gillner, M; Bergman, J; Cambillau, C; et al. (1993) Interactions of indolo[3,2-b]carbazoles and related polycylic
aromatic hydrocarbons with specific binding sites for 2,3,7,8,-tetrachlorodibenzo-p-dioxin in rat liver. Mol
Pharmacol 44:336-345.
Grassman, JA; Masten, SA; Walker, NJ; et al. (1998) Animal models of human response to dioxms. Environ Health
Perspect 106 (Suppl 2):761-75.
Hahn, ME. (1998) The aryl hydrocarbon receptor: a comparative perspective. Comp Biochem Physiol C Pharmacol
Toxicol Endocrinol 121(l-3):23-53.
Hahn, ME; Karchner, SI. (1995) Evolutionary conservation of the vertebrate Ah (dioxin) receptor: amplification and
sequencing of the PAS domain of a teleost Ah receptor cDNA. Biochem J 310:383-387.
Hankinson, O. (1995) The aryl hydrocarbon receptor complex. Ann Rev Pharmacol Toxicol 35:307-340.
Harper, PA; Giannone, JV; Okey, AB; et al. (1992) In vitro transformation of the human Ah receptor and its binding
to a dioxin response element. Mol Pharmacol 42:603-612.
Harper, N; Connor, K; Steinberg, M; et al. (1995) Immunosuppressive activity of polychlorinated biphenyl mixtures
and congeners: nonadditive (antagonistic) interactions. Fundam Appl Toxicol 27(1):131-139.
5/22/00
9-37 DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
55
56
57
Harris M; Zacharewski, T; Safe, S; (1993) Comparative potencies of Aroclors 1232, 1242, 1248, 1254, and 1260 in
SSSSS toxic equivalency factor (TEF) approach for ^chl°
Hornung, MW; Zabel EW; Peterson, RE. (1996) Toxic equivalency factors of polybrominated dibenzo-p-dioxin,
dibenzpfuran, biphenyl, and polyhalogenated diphenyl ether congeners based on rainbow trout early life stage
mortality. Toxicol Appl Pharmacol 140(2):227-234.
^C monographs on the evaluation of carcinogenic risks to humans: polychlorinated dibenzo-para-dioxins and
P?ess VoTel dlbenZOforanS- (1"7) McGre§OT> DB; Partensky, C; Wilbourn, J; et al., eds. Lyon, France: IARC
Jones KG. (1998) Determination of polychlorinated biphenyls in human foodstuffs and tissues: suggestions for a
selective congener analytical approach. Sci Total Environ 68:141-159.
Kleman, M; Gustafsson, JA. (1996) Interactions of procarcinogenic heterocyclic amines and indolocarbazoles with
the dioxm receptor. Biol Chem 377(1 1):741-762.
Kleman, MI; Poellinger, L; Gustafsson, JA. (1994) Regulation of human dioxin receptor function by
indolocarbazoles, receptor ligands of dietary origin. J Biol Chem 269(7):5 137-5 144.
Kodavanti PR; Ward, TR; McKinney, JD; et al. (1996) Inhibition of microsomal and mitochondrial
Ca2+-sequestration in rat cerebellum by polychlorinated biphenyl mixtures and congeners. Structure-activity
relationships. Arch Toxicol 70(3-4): 150-157. <",uvuy
Koopman-Esseboom C; Morse, DC; Weisglas-Kuperus, N; et al. (1994) Effects of dioxins.and polychlorinated
Dipnenyis on thyroid hormone status of pregnant women and their infants. Pediatr Res 36:468-473.
Koopman-Esseboom, C; Huisman, M; Touwen, BC; et al. (1997) Newborn infants diagnosed as neurolo-ncally
n° °n t0 PCB 3nd dioxin exP°sure a*1*1 me«- thyroid-hormone status [letter]. Dev Med Child Neurol
ies °f PCDDs ^ PCDFs * fresh water
Kutz, FW; Barnes, DG; Bretthauer, EW; et al. (1990) The International Toxicity Equivalency Factor (I-TEF) method
for estimating risks associated with exposures to complex mixtures of dioxins and related compounds. Toxicol
bnviron Chem 26:99-109.
Lahvis, GP; Bradfield CA. (1998) Ahr null alleles: distinctive or different? Biochein Pharmacol 56(7):78 1-787.
Leece, B; Denomme, MA; Towner, R; et al. (1985) Polychlorinated biphenyls: correlation between in vivo and in
vitro quantitative structural-activity relationships (QSARs). J Toxicol Environ Health 16:379-388.
Lipp HP; Schrenk, D; Wiesmuller, T; et al. (1992) Assessment of biological activities of mixtures of
polychlorinated dibenzo-p-dioxins (PCDDs) and their constituents in human HepG2 cells. Arch Toxicol
66(3):220-223.
Manson, MM; Hudson, EA; Ball, HW; et al. (1998) Chemoprevention of aflatoxin Bl-induced carcinogenesis by
mdole-3-carbinol in rat liver-predicting the outcome using early biomarkers. Carcinogenesis 19(10):1829-1836.
IndUCti°n of estradio1 metabolism by dietary indole-3-carbinol in humans. J
Michnovicz, JJ; Bradlow, HL. (1991) Altered estrogen metabolism and excretion in humans following consumption
of indole-3-carbinol. Nutr Cancer 16:59-66.
5/22/00
9-38 DRAFT—DO NOT CITE OR QUOTE
-------
Micka, J; Milatovich, A; Menon, A; et al. (1997) Human Ah receptor (AHR) gene: localization to 7pl5 and
suggestive correlation of polymorphism with CYP1A1 inducibility. Pharmacdgenetics 7:95-101.
Miller, CP; Birnbaum, LS. (1986) Teratologic evaluation of hexabrominated naphthalenes in C57BL/6N mice.
Fundam Appl Toxicol 7(3):398-405.
Mimura, J; Yamashita, K; Nakamura, K; et al. (1997) Loss of teratogenic response to
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in mice lacking the Ah (dioxin) receptor. Genes Cells 2(10):645-654.
Morrissey, RE; Harris, MW; Diliberto, JJ; et al. (1992) Limited PCB antagonism of TCDD-induced malformations
in mice. Toxicol Lett (l):19-25.
Nagao, T; Golor, G; Hagenmaier, H; et al. (1993) Teratogenic potency of 2,3,4,7,8-pentachloro-
dibenzofuran and of three mixtures of polychlorinated dibenzo-p-dioxins and dibenzofurans in mice. Problems with
risk assessment using TCDD toxic-equivalency factors. Arch Toxicol 67(9):591-597.
Nakai, JS; Bumce, NJ. (1995) Characterization of the Ah receptor form human placental tissue. J Biochem Toxicol
10(3):151-159. :
Nagayama, J; Okamura, K; lida, T; et al. (1998) Postnatal exposure to chlorinated dioxins and related chemicals on
thyroid hormone status in Japanese breast-fed infants. Chemosphere 37(9-12):1789-1793.
NATO/CCMS. (1988) Scientific basis for the development of the International Toxicity Equivalency Factor (I-TEF)
method of risk assessment for complex mixtures of dioxins and related compounds. Report No. 178, Dec. 1988.
Nebert, DW. (1989) The Ah locus: genetic differences in toxicity, cancer, mutation and birth defects. CRC Crit Rev
Toxicol 20:153-174.
Okey, AB; Giannone, JV; Smart, W; et al. (1997) Binding of 2,3,7-8-tetrachlqrodibenzo-p-dioxin to AH receptor in
placentas from normal versus abnormal pregnancy outcomes. Chemosphere 34(5-7):1535-1547.
Okey, AB; Riddick, DS; Harper, PA. (1994) The Ah receptor: mediator of the toxicity of 2,3,7,8-tetrachlorodibenzo-
p-dioxin (TCDD) and related compounds. Toxicol Lett 70:1-22.
Olson, JR; McGarrgle, BP. (1992) Comparative developmental toxicity of 2,3,7,8-tetrachloro-dibenzo-p-dioxin
(TCDD). Chemosphere 25:71-74.
Olson, JR; Bellin, JS; Barnes, DG; et al. (1989) Reexamination of data used for establishing toxicity equivalency
factors (TEFs) for chlorinated dibenzo-p-dioxins and dibenzofurans (CDDs and CDFs). Chemosphere 18(1-6):371-
381.
Ontario Ministry of the Environment (OME). (1984) Scientific criteria document for standard development, No. 4-
84. Polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs).
Ott, MG; Zober, A. (1996) Morbidity study of extruder personnel with potential exposure to brominated dioxins and
furans. II. Results of clinical laboratory studies. Occup Environ Med 53(12):844-846.
Patterson, DG, Jr; Todd, GD; Turner, WE; et al. (1994) Levels of non-o:rtho-substituted (co-planar), mono- and di-
ortho-substituted polychlorinated biphenyls, dibenzo-p-dioxins, and dibenzofurans in human serum and adipose
tissue. Environ Health Perspect 102(Suppl 1): 195-204.
Peters, JM; Narotsky, MG; Elizondo, G; et al. (1999) Amelioration of TCDD-induced teratogenesis in aryl
hydrocarbon receptor (AhR)-null mice. Toxicol Sci 47(l):86-92.
Poland, A; Glover, E. (1987) Variation in the molecular mass of the Ah receptor among vertebrate species and
strains of rats. Biochem Biophys Res Commun 146(3): 1439-1449.
5/22/00
9-39
DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
55
56
57
Poland, A; Knutson, JC. (1987) 2,3,7,8-Tetrachlorodibenzo-p-dioxin and related halogenated aromatic
hydrocarbons: examination of the mechanism of toxicity. Ann Rev Pharmacol Toxicol 22:517-554.
Poland, A; Glover, E. (1990) Characterization and strain distribution pattern of the murine Ah receptor specified by
the Ahd and Ahb-3 alleles. Mol Pharmacol 38(3):306-312.
Powell-Coffinan, JA; Bradfield, CA; Wood, WB. (1998) Caenorhabditis elegans oithologs of the aryl hydrocarbon
receptor and its heterodimerization partner the aryl hydrocarbon receptor nuclear translocator Proc Natl Acad Sci
USA 95(6):2844-2849.
Putzrath, RM; (1997) Estimating relative potency for receptor-mediated toxicity: Devaluating the toxicity
equivalence factor (TEF) model. Regul Toxicol Pharmacol 25:68-78.
Restum, JC; Bursian, SJ; Giesy, JP; et al. (1998) Multigenerational study of the effects of consumption of PCB-
contaminated carp from Saginaw Bay, Lake Huron, on mink. 1. Effects on mink reproduction, kit growth and
survival, and selected biological parameters. J Toxicol Environ Health 54(5):343-375.
Riddick, DS; Huang, Y; Harper, PA; et al. (1994) 2,3,7,8-tetrachlorodibenzo-p-dioxin versus 3-methylcholanthrene-
comparative studies of Ah receptor binding, transformation, and induction of CYP1A1 J Biol Chem
269(16):12118-12128.
Ross, P; De Swart R; Addison, R; et al. (1996) A contaminant-induced immunotoxicity in harbour seals- wildlife at
risk? Toxicology 112(2):157-169.
Ross, DO; van Birgelen, A; DeVito, MJ; et al. (1997) Relative potency factors derived from CYP1A induction in
mice are predictive for alterations in retinoid concentrations after subchronic exposure to mixtures of PCDDs,
PCDFs, and PCBs in female Sprague Dawley rats. Organohalogen Compounds 34:281-287.
Ross, PS. (2000) Marine mammals as sentinels in ecological risk assessment. Human Ecol Risk Assess.
Rowlands, JC; Gustafsson, JA. (1995) Human dioxin receptor chimera transactivation in a yeast model system and
studies on receptor agonists and antagonists. Pharmacol Toxicol 76:328-333.
Ryan, JJ; Gasiewicz, TA; Brown, JF, Jr. (1990) Human body burden of polychlorinated dibenzofurans associated
with toxicity based on the Yusho and Yucheng incidents. Fundam Appl Toxicol 15:722-731.
Ryu, DY; Levi, PE; Femandez-Salguero, P; et al. (1996) Piperonyl butoxide and acenaphthylene induce cytochrome
P450 1A2 and 1B1 mRNA in aromatic hydrocarbon-responsive receptor knock-out mouse liver. Mol Pharmacol
50(3):443-446.
Safe, S. (1990) Polychlorinated biphenyls (PCBs), dibenzo-p-dioxms (PCDDs), dibenzofurans
(PCDFs), and related compounds: environmental and mechanistic considerations
which support the development of toxic equivalency factors (TEFs). Crit Rev Toxicol 21(l):51-88.
Safe, S. (1994) Polychlorinated biphenyls (PCBs): environmental impact, biochemical and toxic responses and
implications for risk assessment. Crit Rev Toxicol 24(2):87-149.
Safe, S. (1995) Human dietary intake of aryl hydrocarbon (Ah) receptor agonists: mass balance estimates of
exodioxins and endodioxins and implications for health assessment. Organohalogen Compounds 26:7-13.
Sargent, LM; Sattler, GL; Roloff, B; et al. (1992) Ploidy and specific karyotypic changes during promotion with
phenobarbitai, 2,5,2',5'-tetrachlorobiphenyl, and/or 3,4,3'4'-tetrachlorobiphenyl in rat liver Cancer Res
52(4):95S-962.
Schmidt, JV; Su, GH; Reddy, JK; et al. (1996) Characterization of a murine Ahr null allele: involvement of the Ah
receptor in hepatic growth and development. Proc Natl Acad Sci USA 93(13):6731-6736..
5/22/00
9-40
DRAFT—DO NOT CITE OR QUOTE
-------
Schmitz, HJ; Hagenmaier, A; Hagenmaier, HP; et al. (1995) Potency of mixtures of poly chlorinated biphenyls as
inducers of dioxin receptor-regulated CYP1A activity in rat hepatocytes and H4IIE cells. Toxicology 99(l-2):47-54.
Schmitz, HJ; Behnisch, P; Hagenmaier, A; et al. (1996) CYPlAl-inducing potency in H4IIE cells and chemical
composition of technical mixtures of polychlorinated biphenyls. Environ Toxicol Pharmacol l(l):73-79.
Schrenk, D; Lipp, HP; Wiesmuller, T; et al. (1991) Assessment of biological activities of mixtures of
polychlorinated dibenzo-p-dioxins: comparison between defined mixtures and their constituents. Arch Toxicol
65(2): 114-118.
Schrenk, D; Buchmann, A; Dietz, K; et al. (1994) Promotion of preneoplastic foci in rat liver with
2,3,7,8-tetrachlorodibenzo-p-dioxin, 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxih and a defined mixture of 49
polychlorinated dibenzo-p-dioxins. Carcinogenesis 15(3):509-515.
Schrenk, D; Riebniger, D; Till, M; et al. (1997) Tryptanthrins: a novel class of agonists of the aryl hydrocarbon
receptor. Biochem Pharmacol 54(1): 165-171.
Sedlak, D; Dumler-Gradl, R; Thoma, H; et al. (1998) Polyhalogenated dibenzo-p-dioxins and dibenzofurans in the
exhaust air during textile processings. Chemosphere37(9-12):2071-2076.
Sewall, CH; Lucier, GW. (1995) Receptor-mediated events and the valuation of the Environmental Protection
Agency (EPA) of dioxin risks. Mutat Res 333:111-122.
Shain, W; Bush, B; Seegal, R. (1991) Neurotoxicity of polychlorinated biphenyls: structure-activity relationship of
individual congeners. Toxicol Appl Pharmacol 11 l(l):33-42.
Shipp, EB; Restum, JC; Giesy, JP; et al. (1998a) Muitigenerational study of the effects of consumption of PCB-
contaminated carp from Saginaw Bay, Lake Huron, on mink. 2. Liver PCB concentration and induction of hepatic
cytochrome P-450 activity as a potential biomarker for PCB exposure. J Toxicol Environ Health 54(5):377-401.
Shipp, EB; Restum, JC; Bursian, SJ; et al. (1998b) Muitigenerational study of the effects of consumption of
PCB-contaminated carp from Saginaw Bay, Lake Huron, on mink. 3. Estrogen receptor and progesterone receptor
concentrations, and potential correlation with dietary PCB consumption. J Toxicol Environ Health 54(5):403-420.
Silkworth, JB; Cutler, DS; Antrim, L; et al. (1989) Teratology of 2,3,7,8-tetrachlorodibenzo-p-dioxin in a complex
environmental mixture from the Love Canal. Fundam Appl Toxicol 13:1-15.
Silkworth, JB; Cutler, DS; Okeefe., PW; et al. (1993) Potentiation and antagonism of 2,3,7,8-tetrachlorodibenzo-p-
dioxin effects hi a complex environmental mixture. Toxicol Appl Pharmacol 119(2):236-247.
Sinai, CJ; Bend, JR. (1997) Aryl hydrocarbon receptor-dependent induction of CYP1A1 by bilirubin in mouse
hepatoma Iclc7 cells. Mol Pharmacol 52:590-599.
Sindhu, RK; Reisz-Porszasz, S; Hankinson, O; et al. (1996) Induction of cytochrome P4501A1 by photooxidized
tryptophan in Hepa Iclc7 cells. Biochem Pharmacol 52(12):1883-1.893.
Sinha, R; Rothman, N; Brown, ED; et al. (1994) Pan-fried meat containing high levels of heterocyclic aromatic
amines but low levels of polycyclic aromatic hydrocarbons induces cytochrome P4501A2 activity in humans. Cancer
Res54(23):6154-6159.
Smialowicz, RJ; DeVito, MJ; Riddle, MM; et al. (1997a) Comparative hnmunotoxic potency of mixtures containing
polychlorinated dibenzo-p-dioxin (PCDDs), dibenzofurans (PCDFs), and biphenyls (PCBs). Toxicologist 31:1350.
Smialowicz, RJ; DeVito, MJ; Riddle, MM; et al. (1997b) Opposite effects of 2,2',4,4',5,5'-hexa-chIorobiphenyl and
2,3,7,8-tetrachlorodibenzo-p-dioxin on the antibody response to sheep erythrocytes hi mice. Fundam Appl Toxicol
37(2):141-149.
5/22/00
9-41
DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
55
56
Stahl, BU; Kettrup, A; Rozman, K. (1992) Comparative toxicity of four chlorinated dibenzo-p-dioxins (CDDs) and
their mixture. Part I: Acutetoxicity and toxic equivalency factors (TEFs). Arch Toxicol 66(7):47 1 -477.
Summer, CL; Giesy JP; Bursian, S3; et al. (1996a) Effects induced by feeding oirganochlorine-contaminated carp
49(^409 SfsT' *° '^ ^^ Legh°m heHS' "' Emblyotoxic ^ toatogenic effects. J Toxicol
Summer, CL; Giesy JP; Bursian, SJ; et al. (1996b) Effects induced by feeding oirganochlorine-contaminated carp
from Saginaw Bay Lake Huron, to laying White Leghorn hens. I. Effects on health of adult hens, egg production
and fertility. J Toxicol Environ Health 49(4):3 89-407. wuufcuon,
Suter-Hofmann, M; Schlatter, CH. (1989) Subchronic relay toxicity with a mixture of polychlorinated dioxins
(PCDDs) and polychlorinated furans (PCDFs). Chemosphere 18:277-282.
Swanson, HI; Bradfield, CA. (1993) The Ah receptor: genetics, structure and function. Pharmacogenetics 3:213-230.
Tillitt, DE; Wright, PJ. (1997) Dioxin-like embryotoxicity of a Lake Michigan lake trout extract to developing lake
trout. Organohalogen Compounds 34:221-225.
U.S. Environmental Protection Agency (U.S. EPA). (1987) Interim procedures for estimating risks associated with
exposures to mixtures of chlorinated dibenzo-p-dioxins and -dibenzofurans (CDDs and CDFs). EPA/625/3-87/012.
U.S. EPA. (1989) Interim procedures for estimating risks associated with exposures to mixtures of chlorinated
dibenzo-p-dioxins and -dibenzofurans (CDDs and CDFs) and 1989 update. EPA/625/3-89/016.
W°rksh0p repOTt On tOXid1y ed,J,/,8-tetrachlorodibenzo-p-dioxin on hepatic porphyrin levels in the rat. Environ Health Perspect 104(5):550-557.
5 N*-Ste^enson> D; Devito> MJ; ^ al. (1996b) Synergistic effects on porphyrin metabolism in
mice after subchromc exposure to a mixture of PCDDs, PCDFs, and PCBs. Organohalogen
Compounds 29:300-305.
van Birgelen, AP; DeVito, MJ; Akins JM; et al. (1996c) Relative potencies of polychlorinated dibenzo-^-dioxins
e blPhenyIs derived from hePatic porphyrin accumulation in mice. Toxicol Appl Pharmacol '
van Birgelen, AP; DeVito, MJ; Birnbaum, LS. (1996d) Toxic equivalency factors derived from cytochrome P-450
mduction in mice are predictive for cytochrome P-450 induction after subchronic exposure to a mixture of PCDDs
PCDFs and PCBs in female b6C3Fl mice and Sprague-Dawley rats. Organohalogen Compounds 29:251-256.
van Birgelen, APJM; Visser, TJ; Kaptein, E; et al. (1997) Synergistic effects on thyroid hormone metabolism in
female Sprague Dawley rats after subchronic exposure to mixtures of PCDDs, PCDFs and PCBs. Organohalogen
Compunds 34:370-375.
5/22/00
9-42 DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
55
van Birgelen, AP; van der Kolk, J; Fase, KM; et al. (1994a) Toxic potency of 3,3',4,4',5-pentachlorobiphenyl relative
to and in combination with 2,3,7,8-tetrachlorodibenzo-p-dioxin in a subchronic feeding study in the rat. Toxicol
Appl Pharmacol 127(2):209-221. !
van Birgelen, AP; van der Kolk, J; Fase, KM; et al. (1994b) Toxic potency of 2,3,3',4,4' 5-hexachlorobiphenyl
relative to and in.combination with 2,3,7,8-tetrachlorodibenzo-^-dioxin in a subchronic feeding study in the rat.
Toxicol Appl Pharmacol 126(2):202-213.
van der Kolk, J; van Birgelen, APJM; Poiger, H. et al. (1992) Interactions of 22',4,4',fi,5'-heXachlOrobiPhenyl and
2,3,7,8-tetrachlorodibenzo-^-dioxin in a subchronic feeding study in the rat. Chemosphere 25(12):2023-202 /.
van den Berg, M; De Jongh, J; Poiger, H; et al. (1994) The toxicokinetics and metabolism of polychlorinated
dibenzo-^-dioxins (PCDDs) and dibenzofurans (PCDFs) and their relevance for toxicity. Cnt Rev Toxicol
24(1)1-74.
van den Berg, M; Sinnige, TL; Tysklind, M; et al. (1995) Individual PCBs as predictors for concentrations of non
and mono-ortho PCBs in human milk. Environ Sci Poll 2(2):73-82.
van den Berg, M; Birnbaum, L; Bosveld, ATC; et al. (1998) Toxic equivalency factors (TEFs) for PCBs, PCDDs,
PCDFs for humans and wildlife. Environ Health Perspect 106(12):775-792.
van der Plas, SA; Haag-Gronlund, M; Scheu, G; etal. (1999) Induction of altered hepatic foci by a mixture of
dioxin-like compounds with and without 2,2',4,4',5,5'-hexachlorobiphenyl in. female Sprague-Dawley rats. Toxicol
Appl Pharmacol 156(l):30-39.
van Leeuwen, FXR. (1997) Derivation of toxic equivalency factors (TEFs) for dioxin-like compounds in humans and
wildlife. Organohalogen Compunds 34:237.
Viluksela, M; Stahl, BU; Birnbaum, LS; et al. (1998a) Subchronic/chronic toxicity of a mixture of four chlorinated
dibenzo-p-dioxins in rats. II. Biochemical effects. Toxicol Appl Pharmacol 151:70-78.
Viluksela M; Stahl, BU; Birnbaum, LS; et al. (1998b) Subchronic/chronic toxicity of a mixture of four chlorinated
dibenzo-^'-dioxins in rats. I. Design, general observations, hematology.and liver concentrations. Toxicol Appl
Pharmacol 151:57-69.
Wang, X; Santostefano, M; Yu, Y; et al. (1992) A comparison of the mouse;versus human aryl hydrocarbon (Ah)
receptor complex: effects of proteolysis. Chem Biol Interact 85(l):79-93.
Walker MK- Peterson, RE. (1991) Potencies of polychlorinated diberizo-p-dioxin, dibenzoforan and biphenyl
congeners, relative to 2,3,7,8-tetrachlorodibenzo-p-dioxin, for producing early life stage mortality in rainbow trout
(Oncorhynchus mykiss). Aquat Toxicol 21:219-238.
Walker MK; Cook, PM; Butterworth, BC; et al. (1996) Potency of a complex mixture of polychlorinated
dibenzo-p-dioxin, dibenzofuran, and biphenyl congeners compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin m causing
fish early life stage mortality. Fundam Appl Toxicol 30(2): 178-86.
Wattenberg, LW; Loub, WD. (1978) Inhibition of PAH-induced neoplasia by naturally occurring indoles. Cancer
Res 38:1410-1413.
Weber, H; Lamb, JC; Harris, MW. (1984) Teratogenicity of 2.3.7.8-tetrachlorodibenzofuran (TCDF) in mice.
Toxicol Lett 20(2): 183-188.
Weber, H; Harris, MW; Haseman, JK; et al. (1985) Teratogenic potential of TCDD, TCDF and TCDD-TCDF
combinations in C57BL/6N mice. Toxicol Lett 26:159-167.
5/22/00
9-43
DRAFT—DO NOT CITE OR QUOTE
-------
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
1.7
18
19
20
21
22
23
24
25
26
27
28
29
30
>H Ł ^' M;DStahiBU; et al" (1992a> Comparative toxicity of four chlorinated dibenzo-p-dioxins
(CDDs) and their mixture. Part II: structure-activity relationships with inhibition of hepatic phosphoenolpyruvate
carboxykinase, pyruvate carboxylase, and gamma-glutamyl transpeptidase activities. Arch Toxicol 66(7):478-483.
L (1"2b) ComParative touchy of four chlorinated dibenzo-p-dioxins
nh , I:sf cture-activ»y relationship with increased plasma tryptophan levels, but no
relationship to hepatic ethoxyresorufin o-deethylase activity. Arch Toxicol 66(7):484-488
Weber, LWD; Greim, H (1997) The toxicity of brominated and mixed-halogenated dibenzo-p-dioxins and
dibenzofurans: an overview. J Toxicol Environ Health 50: 1 95-2 15.
Wilker, C; Johnson, L; Safe, S. (1996) Effects of developmental exposure to indole-3-carbinol or 2 3 7 8-
tetrachlorodibenzo-^-dioxin on reproductive potential of male rat offspring. Toxicol Appl Pharmacol' 141:68-75.
; (1"2) ReVieW °f ^ m°delS f°r TEFs in assessinS health risks of PCDDs and PCDFs. Toxic Sub J
Zabel, EW; Walker, MK; Hornung, MW; et al. (1995) Interactions of polychlorinated dibenzo-p-dioxin
34'204 2°? C°ngenerS f°r Producing rainbow *«»* e^V U& stage mortality. Toxicol Appl
Zacharewski, T; Harris, M; Safe, S; et al. (1988) Applications of the in vitro aryl hydrocarbon hydroxvlase induction
Ziao, F; Mayura, K; Kocurek, N; et al. (1997) Inhibition of 3,3',4,4',5-pentachlorobIphenyl-induced chicken
embryotoxicity by 2,2',4,4',5,5'-hexachlorobiphenyl. Fundam Appl Toxicol 35(l):l-8.
5/22/00
9-44 DRAFT—DO NOT CITE OR QUOTE
------- |