United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
NCEA-J-0836
September 2000
SAB Review Draft
Chapter 9. Toxic
Equivalency 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(Do Not
Cite or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.
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NCEA-I-0836
September 2000
SAB Review Draft
www.epa.gov/ncea
Chapter 9. Toxic Equivalency Factors (TEF) for Dioxin
and Related Compounds
Exposure and Human Health Reassessment
of 2,3,7,8-Tetrachlorodibenzo-/i-Dioxin (TCDD)
and Related Compounds
Part II: Health Assessment for 2,3,7,8-Tetrachlorodibenzo-/7-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
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DISCLAIMER
This document is a 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.
Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.
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TABLE OF CONTENTS - OVERVIEW
Exposure and Human Health Reassessment
of l^S-Tetrachlorodibenzo-p-Dioxin (TCDD)
and Related Compounds
Part I: Estimating Exposure to Dioxin-Like Compounds (Draft Final)
(EPA/600/P-00/001 Bb, Be, Bd) September 2000
Volume 1 : Executive Summary (EPA/600/P-00/001Ba) ^klsis not iMuded^ri'tMs: draft)
Volume 2: Sources of Dioxin-Like Compounds in the United States (EPA/600/P-00/001Bb)
Chapters 1 through 13
Also included On^his
Database of Sources of Environmental Releases of Dioxin-
Like Compounds in the United States (Draft Final)
(EPA/600/P-98/GQ2B) September 2000
Volume 3: Properties, Environmental Levels, and Background Exposures
(EPA/600/P-OO/OOlBc)
Chapters 1 through 6
Volume 4: Site-Specific Assessment Procedures (EPA/600/P-00/001Bd)
Chapters 1 through 8
Part II: Health Assessment for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related
Compounds (Draft Final)
(EPA/600/P-00/001Be) September 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. Carcinogenicity of TCDD in Animals
Chapter?. Epidemiology/Human Data
Chapters. Dose-Response Modeling for 2,3 ,7, 8 -TCDD
(SAB Review Draft, September 2000)
Chapter 9. Toxic Equivalency Factors (TEF) for Dioxin and Related Compounds
(SAB Review Draft, September 2000)
Part III: Integrated Summary and Risk Characterization for
2,3,7,8-TetrachIorodibenzo-/j-Dioxin (TCDD) and Related Compounds
(SAB Review Draft, September 2000) (EPA/600/P-00/001Bg)
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CONTENTS
9. TOXIC EQUIVALENCY FACTORS (TEFs) FOR DIOXIN AND RELATED
COMPOUNDS 9-1
9.1. INTRODUCTION 9_1
9.2. HISTORICAL CONTEXT OF TEFs 9-1
9.2.1. TEFs for PCDDs and PCDFs 9-2
9.2.2. TEFs for PCBs 9-3
9.2.3. The Most Recent Evaluation of TEFs for PCDDs, PCDFs, and PCBs 9-5
9.2.4. Illustrative Examples of the Data Used for Deriving the TEF Values 9-8
9.3. SPECIFIC ISSUES 9.9
9.3.1. Ah Receptor and Toxicity Factors 9-9
9.3.2. The Role of the AhR in the Toxicity of Dioxin-Like Compounds 9-9
9.3.3. Species Comparison of the AhR 9-10
9.3.4. Ah Receptor Ligands 9-13
9.3.4.1. Industrial/Synthetic AhR Ligands 9-13
9.3.4.2. Naturally Occurring AhR Ligands 9-15
9.3.4.3. Industrial/Synthetic vs. Natural AhR Ligands 9-16
9.4. TOTAL TEQ AND THE ADDITIVITY CONCEPT 9-19
9.4.1. Examination of Laboratory Mixtures of PCDDs and PCDFs 9-20
9.4.2. Examination of Commercial or Laboratory-Derived Mixtures
of PCDDs, PCDFs, and PCBs 9-23
9.4.3. Examination of Environmental Samples Containing PCDDs,
PCDFs, and/or PCBs 9-26
9.4.4. Nonadditive Interactions With Non-Dioxin-Like Compounds 9-27
9.4.5. Examination of the TEF Methodology in Wildlife 9-30
9.4.6. Toxic Equivalency Functions 9-31
9.4.7. Endpoint and Dose-Specific TEFs 9-32
9.5. UNCERTAINTY 9-32
9.6. IMPLICATIONS FOR RISK ASSESSMENT 9-33
9.7. SUMMARY 9-34
REFERENCES FORCHAPTER9 9-39
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LIST OF TABLES
9-1. Estimated relative toxicity of PCDD and PCDF isomers to 2,3,7,8-T4CDD 9-35
9-2. Toxic equivalency factors (TEFs) 9-36
9-3. The range of the in vivo REP values for the major TEQ contributors 9-37
LIST OF FIGURES
9-1. Structures of poly chlorinated dibenzo-^-dioxin, dibenzofurans,
and biphenyls
9-38
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9. TOXIC EQUIVALENCY FACTORS (TEFs) FOR
DIOXIN AND RELATED COMPOUNDS
9.1. INTRODUCTION
Previous risk assessments of dioxin and dioxin-like compounds 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 toxic equivalent (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 World
Health Organization (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 Part II, Chapters 3-6), to wildlife (Giesy
and Kannan, 1998; Ross, 2000), and to humans (IARC, 1997) (see also Part II, Chapter 7 ),
hazard characterization and risk assessment activities have tended to focus on a subset of
polychlorinated dibenzo-^-dioxin (PCDDs), polychlorinated dibenzofurans (PCDFs), and
polychlorinated biphenyls (PCBs)(Figure 9-1). The subset of compounds, shown in Figure 9-1,
are known as "dioxin-like" and have been assigned TEF values by WHO. In this chapter, the
development of TEFs for these and other PHAHs is discussed.
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1 9.2.1. TEFs for PCDDs and PCDFs
2 The first use of a TEF-like method was described by Eadon et al. (1 986) as a means to
3 estimate potential health risks associated with a PCB transformer fire in Binghamton, NY. In
4 1983, the Ontario Ministry of the Environment produced a Scientific Criteria Document for
5 PCDDs and PCDFs which concluded, based on a review of available scientific information, that
6 dioxin and dibenzofurans were structurally similar compounds that shared a common cellular
7 mechanism of action (activation of the Ah receptor [AhR]) and induced comparable biological
8 and toxic responses, and that the development of environmental standards for human health
9 concerns should be based on a "toxic equivalency" approach with 2,3,7,8-tetrachlorodibenzo-j?-
1 0 dioxin (TCDD) as the prototype (OME, 1 984). The final recommendation divided all
1 1 PCDD/PCDF congeners into their respective homologue groups and assigned to each group a
1 2 toxicity factor relative to TCDD (Table 9-1). These numerical factors could then be applied to
1 3 transform various concentrations of PCDDs and PCDFs into equivalent concentrations of
14 2,3,7,8-TCDD.
1 5 Following up on an initial risk assessment methodology designed to address the emission
1 6 of dioxins and furans from waste incinerators, EPA also concluded that TEFs were the best
1 7 available interim scientific policy for dealing with complex mixtures of these contaminants.
1 8 With the mandate to develop active research programs that would address the limitations
1 9 inherent to this risk management technique, the Agency recommended TEFs for specific
20 congeners, rather than isomeric groups (Table 9-2; U.S. EPA, 1987). In an analogous fashion to
21 OME's approach, concentrations of PCDDs and PCDFs would be analytically determined, the
22 concentration of each congener would be multiplied by its respective TEF value, and all the
23 products would be summed to give a single 2,3,7,8-TCDD equivalent. This approach has been
24 described mathematically as:
25
Total Toxic Equivalency (TEQ) = Cn * TEFn
26 Cn equals the concentration of the individual congener in the complex mixture under analysis.
27 TEFs were determined by inspection of the available congener-specific data and an assignment of
28 an "order of magnitude" estimate of relative toxicity when compared to 2,3,7,8-TCDD. In vitro
29 Ah receptor binding and in vitro and in vivo toxicity studies were considered in setting individual
30 TEFs. Scientific judgment and expert opinion formed the basis for these TEF values. External
31 review of the toxicity and pharmacokinetic data utilized by EPA in setting these TEFs supported
32 the basic approach as a "reasonable estimate" of the relative toxicity of PCDDs and PCDFs
33 (Olson etaL, 1989).
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1 A 3-year study conducted by the North Atlantic Treaty Organization Committee on the
2 Challenges of Modern Society (NATO/CCMS) also concluded that the TEF approach was the
3 best available interim measure for PCDD/PCDF risk assessment. On the basis of examination of
4 the available data dealing with exposure, hazard assessment, and analytical methodologies
5 related to dioxin and furans, an International Toxiciry Equivalency Factor (I-TEF) scheme was
6 presented (Table 9-2; NATO/CCMS, 1988). This review also concluded that "data strongly
7 support the role of the AhR in mediating the biologic and toxic responses elicited by 2,3,7,8-
8 TCDD and related PCDDs and PCDFs and provide the scientific basis for the development of
9 TEFs for this class of compounds." Various refinements to previous efforts included selection of
|10 TEF values based more on in vivo toxicities, assigning TEF values to octachlorodibenzo-/>-
11 dioxin and octachlorodibenzofuran, and removing any TEF values for all non-2,3,7,8-substituted
12 congeners. Although it was indicated that, theoretically, it may be possible to detect nearly all of
13 the 210 PCDD/DF isomers in the environment, seventeen 2,3,7,8-substituted congeners were
14 known to be preferentially retained and bioaccumulated. For example, when fish or a variety of
15 rodent species were exposed to a complex mixture of PCDDs/PCDFs from incinerator fly ash,
[ 16 the 2,3,7,8-substituted congeners, which were minor components of the original mixture,
17 predominated in the analysis of their tissues (Kuehl et al., 1986; van den Berg et al., 1994). In
18 addition, when humans were exposed to a complex mixture of more than 40 different PCDF
19 congeners during the Oriental rice oil poisoning episodes, only the 2,3,7,8-substituted congeners
20 were detected in subsequent blood and adipose tissue analysis (Ryan et al., 1990). EPA, which
21 had participated in the NATO/CCMS exercise, officially adopted the revised I-TEFs in 1989,
'22 with the caveat that this risk assessment approach remains interim and continued revisions
23 should be made (U.S. EPA, 1989; Kutz et al., 1990). The use of the TEF model for risk
24 assessment and risk management purposes has been formally adopted by a number of countries
25 (Canada, Germany, Italy, the Netherlands, Sweden, the United Kingdom, U.S.A.) (Yrjanheiki,
26 1992), and as guidance by international organizations such as the International Programme on
27 Chemical Safety, WHO.
28
29 9.2.2. TEFs for PCBs
30 During the period of TEF development for PCDDs/PCDFs, a considerable body of
31' experimental evidence was also being generated regarding the structure-activity relationships
32 between the different polychlorinated biphenyl homologue classes (Safe, 1990, 1994). Following
33 the synthesis of analytical standards for all 209 theoretical PCS congeners by 1984, subsequent
34 analysis of a variety of commercial samples was able to identify all but 26 (Jones, 1988).
35 However, once released into the environment, PCBs are subject to a variety of photolysis and
36 biodegradation processes, to the extent that only 50-75 congeners are routinely detected in higher
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1 trophic level species (van den Berg et al., 1995). Initial structure-activity relationship studies
2 revealed that those congeners substituted in only the meta and para positions were approximate
3 isostereomers of TCDD. Subsequent toxicological studies confirmed that these non-ortho-
4 substituted, "coplanar" PCBs (e.g., PCS 77, 81, 126, 169) did induce a variety of in vitro and in
5 vivo effects similar to TCDD (Leece et al., 1985). Maximum TCDD-like activity is obtained for
6 PCBs when there are no ortho, two or more meta, and both para positions occupied (Figure 9-1).
7 Introduction of a single ortho substituent to the biphenyl (mono-ortho "coplanars") results in a
8 diminishing, but not elimination, of TCDD-like activity and toxicological responses resembling
9 commercial mixtures of PCBs. The addition of a single ortho substituent also increases the non-
10 dioxin-like activity of the compound. Several congeners from this group are prevalent in both
11 commercial PCBs and a wide variety of environmental samples. Some of the more persistent
12 mono-ortho substituted PCBs (PCBs 105,118,156) can be found in human serum and adipose -
13 samples at levels up to three orders of magnitude higher than the "coplanar" PCBs, PCDDs and
14 PCDFs (Patterson et al., 1994). In limited studies a third group of PCB congeners, the di-ortho
15 "non-coplanars," has exhibited only minor amounts of dioxin-like activity (if any), usually 4-6
1Q orders of magnitude less potent than TCDD (Safe, 1990). Recent studies have demonstrated that
17 some of the earlier methods of preparation of these di-ortho non-coplanar PCBs had trace
18 contaminants of PCDFs, which may account for the weak dioxin-like activity of these
19 compounds (van der Kolk et al., 1992). In 1991, EPA convened a workshop to consider TEFs
20 for PCBs (Barnes et al., 1991). The consensus was that a small subset of the PCBs displayed
21 dioxin-like activity and met the criteria for inclusion in the TEF methodology. Such proposals
22 for the TEF methodology also seem to have utility in assessing risks to wildlife (van den Berg et
23 al., 1998; Giesey and Kannan, 1998; Ross, 2000).
24 PCBs are often classified into two categories: "dioxin-like" and "non-dioxin-like." The
25 dioxin-like PCBs bind to the AhR and produce dioxin-like effects in experimental animals. All
26 other PCBs then fall into the non-dioxin-like classification. Although the dioxin-like PCBs are
27 generally more potent at inducing biological effects, they constitute only a minor portion of the
28 mass of PCBs found in environmental and biological samples. The non-dioxin-like PCBs
29 account for a majority of the mass of the PCBs found in environmental and biological samples.
30 The use of the term non-dioxin-like PCBs is not necessarily useful. The PCBs not included in
31 the TEF scheme (i.e., the non-dioxin-like PCBs) are not a single class of compounds and have
32 multiple toxicities with separate structure-activity relationships (Barnes et al., 1991). Not enough
33 congener-specific research has been performed to adequately characterize or classify these
34 compounds. For example, the "neurotoxic" PCBs have been typically defined by stracture-
35 activity relationships for decreasing dopamine concentrations or alterations in intracellular
36 calcium in cell culture (Shain et al.3 1991; Kodavanti et al., 1996). However, few of these
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1 congeners have been examined in vivo to determine the predictive ability of these in vitro
2 screens.
3 As part of the joint WHO European Centre for Environmental Health (WHO-ECEH) and
4 the International Programme on Chemical Safety (IPCS) project to harmonize TEF schemes for
5 dioxin-like compounds, a database was generated consisting of all available relevant
6 toxicological data for PCBs up to the end of 1993. Of almost 1,200 peer-reviewed publications,
7 146 were selected and analyzed on the basis of the following criteria: at least one PCB congener
8 was investigated; TCDD or a reference coplanar PCB (77,126, 169) was used during the
9 experiment or results were available from previous experiments (same author, laboratory,
110 experimental design); and the endpoint in question was affected by both the reference compound
111 and the PCB congener in question (i.e., dioxin-specific). TEFs were then determined from a total
112 of 60 articles/manuscripts on the basis of the reported results for 14 different
113 biological/toxicological parameters. Following scientific consultation by 12 experts from 8
14 different countries, interim TEF values were recommended for 13 dioxin-like PCBs (Table 9-2),
15 based on four inclusion criteria: (1) the compound should show structural similarity to PCDDs
16 and PCDFs; (2) it should bind to the Ah receptor; (3) it should induce dioxin-specific
17 biochemical and toxic responses; and (4) it should be persistent and accumulate in the food chain
18 (Ahlborg et al., 1994). Increased consideration was given to selection of a TEF value based on
19 repeat-dosing in vivo experiments, when available.
20 There is experimental evidence to suggest that a limited number of PCB congeners
21 classified as weak or non-AhR agonists could effect concentration-dependent nonadditive
22 interactions with dioxin-like compounds (Safe, 1990; 1994). Both antagonistic (Safe, 1990;
23 Morrissey et al., 1992; Smialowicz et al., 1997b) and synergistic (Safe, 1990; van Birgelen et al.,
24 1996a,b; van Birgelen et al., 1997) interactions between TCDD and PCBs have been observed in
25 experimental systems. The non-additive interactions of the PCBs are thought to be mediated
26 through non-AhR pathways (Smialowicz et al., 1997). These interactions usually occur at
27 extremely high doses of the PCBs that are not environmentally relevant, and thus the nonadditive
28 interactions are thought not to significantly detract from the TEF methodology (van den Berg et
29 al., 1998; Birnbaum, 1999).
30
31 9.2.3. The Most Recent Evaluation of TEFs for PCDDs, PCDFs, and PCBs
32 An additional recommendation from the first WHO PCB TEF consultation was that the
33 current database should be expanded to include all relevant information on PCDDs, PCDFs, and
34 other dioxin-like compounds that satisfied the four inclusion criteria. Prior to the second WHO-
35 ECEH consultation in 1997, various terminologies or definitions applicable to TEFs were
36 reviewed and standardized. Whereas previously the term TEF had been used to describe all
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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 alkoxyresorufm 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-magnitude consensus estimate of the toxicity of a compound relative to the
toxicity of TCDD that is derived using careful scientific judgment of all 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/EC]02550).
• Calculated from ratios of tumor promotion indexes or maximal enzyme induction
levels.
• Calculated from ratios of Ah 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 compound, 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 compound
. following subchronic exposures; a TEF based on administered dose is reduced to
0.0001.
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1 • New in vivo enzyme induction potency and structural similarity with OCDD
2 support the TEF change to 0.0001 for OCDF.
3 • REPs from an in vivo subchronic toxicity study (enzyme induction, hepatic retinol
4 decreases) support reducing the TEF to 0.0001 for PCS 77.
5 • A TEF value of 0.0001 was assigned for PCB 81. Even though PCB 81 was not
6 assigned a TEF value at the 1993 WHO consultation because of lack of human
7 residue and experimental data, more recent data demonstrate similar qualitative
8 structural activity results compared to PCB 77.
9 • Because of the lack of in vivo enzyme induction (CYP 1A1/A2) and reproductive
110 toxicity with structurally similar congeners (PCB 47 and PCB 153), the previous
111 interim TEF values for the di-ortho-substituted PCBs 170 and 180 were
12 withdrawn.
13
14 Although a number of uncertainties associated with the TEF concept have been identified
15 (nonadditive interactions with non-dioxin-like PCBs, natural ligands for the Ah receptor,
16 questionable low-dose linearity of REP responses), the 1997 WHO expert meeting decided that
17 an additive TEF model remained the most feasible risk assessment method for complex mixtures
18 ofdioxin-likePHAHs.
19 The WHO working group acknowledged that there are a number of other classes of
20 chemicals that bind and activate the Ah receptor. The chemicals include, but are not limited to,
21 polyhalogenated naphthalenes, diphenyl ethers, fluorenes, biphenyl methanes, quaterphenyls, and
22 others. In addition, a number of brominated and chloro/bromo-substituted dioxin analogues of
23 the PCDDs and PCDFs have been demonstrated to cause dioxin-like effects. The WHO working
24 group concluded that "at present, insufficient environmental and toxicological data are available
25 to establish a TEF value for any of the above compounds" (van den Berg etal., 1998).
26 In January 1998, EPA and the U.S. Fish and Wildlife Service sponsored a meeting
27 entitled "Workshop on the Application of 2,3,7,8-TCDD Toxicity Equivalency Factors to Fish
28 and Wildlife." The major objective of the workshop was to address uncertainties associated with
29 the use of the TEF methodology in ecological risk assessment. Twenty-one experts from
30 academia, government, industry, and environmental groups participated in the workshop. The
31 consensus of the workgroup was that while there are uncertainties in the TEF methodology, the
32 use of this method decreases the overall uncertainty in the risk assessment process. However,
33 quantifying the decrease in the uncertainty of a risk assessment using the TEF methodology
34 remains ambiguous, as does the exact uncertainty in the TEF methodology itself (U.S. EPA,
35 2000).
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1 This first section has outlined the process of assessing the relative potency of chemicals
2 and the assignment of a consensus TEF value. There are still many questions on the use of the
3 TEF method and the validity of some of the underlying assumptions. A detailed discussion and
4 review of the data supporting the development and use of the TEF method, as well as the data
5 relating to the issue of additivity, is included within the specific issues section that follows.
6
7 9.2.4. Illustrative Examples of the Data Used for Deriving the TEF Values
8 The TEF scheme includes 17 PCDDs and PCDFs and 13 PCBs. However, in human
9 tissue samples and food products, only five of these congeners, TCDD, 1,2,3,7,8-PCDD,
10 1,2,3,6,7,8-HxCDD, 2,3,4,7,8-PeCDF, and PCB 126, account for over 70% of the TEQ. There is
11 considerable data on the relative potency of these compounds in both in vitro and in vivo studies.
12 Table 9-3 provides a summary of the REPs from in vivo data available for the compounds that
13 account for approximately 80% of the TEQs in humans (see Part I, Volume III, Section 4.2.).
14 This uiformation was obtained from the WHO database used to derive TEF values for PCDDs,
15 PCDFs, and PCBs (Van den Berg et al, 1998). The WHO database contains duplicate
16 recordings of studies for several of the compounds. The data in Table 9-3 does not include the
17 duplicates. In addition, the WHO database also contains studies that used a single dose level of
18 the test chemical, and REP values were not estimated for these studies. For example, in the
19 WHO database for PCB 126, there are 144 in vivo endpoints. Of these 144, 50 do not have REP
20 values associated with the entry because the study used only a single dose level. In other cases,
21 for a given endpoint from a particular study, the REP value is presented as estimated by the
22 authors as well as by alternative analyses by members of the WHO workgroup. In total, there are
23 62 data sets that have dose-response relationships sufficient enough to estimate the relative
24 potency of PCB 126. These data sets examine enzyme induction, changes in organ and body
25 weights, immunotoxicity, developmental toxicity, thyroid hormones, renal and hepatic retinoids,
26 and tumor promotion. The WHO database for 1,2,3,7,8-PCDD contained studies examining
27 enzyme induction, changes in organ and body weights, hepatic porphyria, hepatic retinoids, and
28 tumor promotion. The WHO database for 2,3,4,7,8-PCDF contained studies examining enzyme
29 induction, changes in organ and body weights, immunotoxicity, developmental toxicity, thyroid
30 hormones, hepatic retinoids, hepatic porphyria, and tumor promotion. The data presented in
31 Table 9-3 for 1,2,3,6,7,8-HxCDD is from U.S. EPA (1989) because the WHO database contained
32 no new in vivo data for this compound. There are only three in vivo studies on the effects of
33 1,2,3,6,7,8-HxCDD, one of which is the NTP carcinogenicity study on a mixture of 31 %
34 1,2,3,6,7,8-HxCDD and 67% l,2,3,7,8,9HxCDD (NTP, 1980).
35 The REPs for 1,2,3,7,8-PCDD in the in vivo studies vary by approximately a factor of
36 five. A TEF value was assigned to 1,2,3,7,8-PCDD based on the REP for tumor promotion
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which ranged from 0.8-1.0. The REPs for 2,3,4,7,8-PCDF and PCB 126 have a greater
variability, but the assigned TEF values are similar to the means of the REP values. The
meanistandard deviation for all in vivo REP values for 2,3,4,7,8-PCDF is 0.4±0.7. If only
subchronic studies are examined, the mean±standard deviation of the REP values is 0.2±0.13.
These REP values for 2,3,4,7,8-PCDF are similar to the TEF value of 0.5. The REPs for PCB
126 range over two orders of magnitude with a mean for all in vivo responses of 0.2±0.2. The
mean REP for subchronic studies examining PCB 126 is 0.13±0.13. The TEF for PCB 126 is
0.1, which is slightly lower than the mean of the REP values. With the exception of 1,2,3,6,7,8-
HxCDD, the REPs are based on several studies from different laboratories examining different
endpoints.
The proceeding discussion compares the TEF value with the mean of the REP values.
This information should be used cautiously. The consensus TEF values are derived based on the
weighted criteria cited above. Comparisons of unweighted means of the REP values ignores the
criteria and the scientific judgment used by the WHO workgroup.
16 9.3. SPECIFIC ISSUES
17 9.3.1. Ah Receptor and Toxicity Factors
18 Issues relating to the role of the Ah receptor as the common mediator of toxicity of
19 dioxin-like compounds and the cross-species comparability of AhR structure and function
20 frequently arise when the TEF approach is discussed. Recent data relating to each of these issues
21 are discussed below.
22
23 9.3.2. The Role of the AhR in the Toxicity of Dioxin-Like Compounds
24 The general basis for the TEF scheme is the observation that the AhR mediates most if
25 not all of the dioxin-like biological and toxic effects induced by compounds included in the TEF
I 26 scheme (Safe, 1990; Okey et al, 1994; Birnbaum, 1994; Hankinson, 1995). Binding to the
27 receptor is necessary, but not sufficient, to generate the wide variety of toxic effects caused by
28 dioxin-like HAHs (Sewall and Lucier, 1995; De Vito and Birnbaum, 1995) (for additional review
29 references, see Part II, Chapter 2). There are several lines of evidence that the Ah receptor is
30 important in the toxicity of the dioxin-like compounds. A brief discussion of this evidence shall
31 be presented in the following section. Those wishing a more detailed discussion of this issue are
32 referred to Part II, Chapter 2.
33 Initial studies on the toxicity of PAHs demonstrated that the sensitivity to these
34 compounds varied by strain of mice and segregated with the Ah locus. The Ah locus was then
35 found to encode a receptor designated as the aryl hydrocarbon receptor or AhR. Sensitive strains
36 of mice expressed receptors with high binding affinity for these compounds, while the resistant
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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 AhR in the toxicity of dioxin-like compounds
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 tunes 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 compounds.
9.3.3. Species Comparison of the AhR
Although binding to the AhR initiates a cascade of molecular and cellular events leading
to toxicity, the exact mechanism of action of dioxin-like compounds 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.
There are several lines of evidence suggesting that the AhR is an important factor in
developmental and homoeostatic 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
bHLH-PAS superfamily consists of a growing 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 in sensing oxygen tension (Hif-1, EPAS-1/HLF) (Hahn, 1998). 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 et al., 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).
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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 at critical periods 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 (Fernandez-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,
|18 the females have difficulty maintaining conceptus during pregnancy, surviving pregnancy and
19 lactation, and rearing pups to weaning (Abbott et al., 1999). It should be noted that the AhR
|20 knockout mice are resistant to the toxic effects of TCDD.
121 Comparisons between the AhR of experimental animals (primarily rodents) and the
22 human AhR have revealed a number of similarities in terms of ligand and DNA binding
23 characteristics and biochemical functions. Tissue-specific patterns of expression of AhR mRNA
24 are similar in rats, mice, and humans, with highest levels generally detected in lung, liver,
25 placenta, and thymus (Dolwick et al., 1993; Dohr et al,, 1996). Nuclear AhR complexes isolated
26 from human and mouse hepatoma cells (Hep G2 and Hepa Iclc7, respectively) have similar
27 molecular weights. Although the human AhR appears more resistant to proteolytic digestion by
28 trypsin or chymotrypsin, the major breakdown products were similar between the two species,
29 and photolabeling analysis with TCDD suggested common features in the ligand binding portion
30 of the receptors (Wang etal., 1992).
31 Limited analysis has suggested the average human AhR exhibits a lower binding affinity
32 for various HAHs than "responsive" rodent strains. However, similar to a variety of
33 experimental animals, human populations demonstrate a wide variability in AhR binding affinity
34 (Micka et al., 1997). Recent determination of AhR binding affinity (Kd) toward TCDD in 86
35 human placenta samples showed a greater than twenty-fold range in the binding affinity, and this
36 range encompasses binding affinities similar to those observed in sensitive and resistant mice
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1 (Okey et al., 1997). Whereas the concentration of various ligands required to activate a human
2 AhR reporter gene construct was higher than required with rodent cell cultures, the actual rank
3 order of binding affinities was in agreement (Rowlands and Gustafsson, 1995). Although
4 comparisons have been made of the TCDD binding affinity to the AhR of different species,
5 caution should be used when attempting to predict species sensitivity to TCDD and related
6 compounds. For mice, the sensitivity to the biochemical and toxicological effects of TCDD and
7 related compounds is associated with the relative binding affinity of TCDD to the AhR in the
8 different strains (Birnbaum et al., 1990; Poland and Glover, 1990). However, the relative
9 binding affinity of TCDD to the AhR across species does not aid in the understanding of
10 interspecies differences in the response or sensitivity to TCDD (DeVito and Birnbaum, 1995).
11 The human AhR also demonstrates other slight differences when compared to the AhR
12 from experimental animal species. The molecular mass of the human AhR ligand-binding
13 subunit appears to be greater than the AhR subunit from certain TCDD "responsive" mouse
14 strains but similar to the receptor molecular mass for rats (Poland and Glover, 1987). Currently
15 there has been no association established between differences in the molecular mass of the AhR
16 and sensitivity to a particular biochemical or toxicological response across species (Okey et al.,
17 1994). The non-liganded human AhR appears thermally more stable compared to AhR from
18 various rodent species, whereas the reverse situation exists with the liganded human AhR (Nakai
19 and Bunce, 1995). Transformation of the ligand-bound human AhR receptor (isolated from
20 colon adenocarcinoma cells) to the DNA-binding state, unlike rodent hepatic AhR, is
21 temperature dependent (Harper et al., 1992). The importance of these species differences in
22 transformation and stability of the AhR in the species sensitivity to TCDD remain uncertain.
23 However, in critical areas of receptor function such as ligand recognition, transformation, and
24 interaction with genomic response elements, the human AhR is comparable to the AhR isolated
25 from experimental animals.
26 Ligand-bound or transformed AhR from a variety of mammalian species, including
27 humans, bind to a specific DNA sequence or "dioxin response element" with similar affinities
28 (Bank et al., 1992; Swanson and Bradfield, 1993). The bHLH structure of receptor proteins such
29 as AhR ensures appropriate contact and binding with DNA recognition sites. Amino acid
30 sequence analysis between mouse and human AhR shows an overall sequence homology of
31 72.5%, whereas the bHLH domain shows 100% amino acid concordance (Fujii-Kuriyama et al.,
32 1995). In comparison, the deduced amino acid composition of the AhR from killifish was 78%-
33 80%, similar to the amino acid sequence of rodent and human AhR (Hahn and Karchner, 1995).
34 These studies demonstrate a concordance between the structure of the receptor and its function
35 across species.
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1 The majority of scientific evidence to date supports the theory that binding to the AhR is
12 a necessary first step prior to dioxin-like compounds eliciting a response, as discussed in Part II,
3 Chapter 2. Current research has identified the AhR in a variety of human tissues and cells that
14 appear to function in a similar manner to the AhR from experimental animals, including fish,
5 birds, and mammals. When multiple endpoints are compared across several species, there exists
6 a high degree of homogeneity in response and sensitivity to TCDD and related compounds
7 (DeVito et al., 1995). Therefore, these data provide adequate support for the development of the
8 TEF methodology. However, these data also reflect the true complexity of intra- and interspecies
9 comparisons of biochemical and toxicological properties. Continued research into the variety of
10 additional cytoplasmic and nuclear proteins capable of interacting with the AhR signaling
1 pathway will ultimately lead to a better understanding of the observed species and strain
12 variability in the response to dioxin-like chemicals and may be useful in further refining TEFs.
I13
|14 9.3.4. Ah Receptor Ligands
|15 A wide variety of structurally diverse anthropogenic and natural chemicals are capable of
116 interacting with the AhR. These chemicals also have a broad range of potencies at inducing
M 7 dioxin-like effects in experimental systems. One of the major differences between the
18 anthropogenic chemicals included in the TEF methodology and the natural AhR ligands is their
19 pharmacokinetics. The anthropogenic chemicals included in the TEF methodology are persistent
120 and bioaccumulate in wildlife and humans. In contrast, most if not all of the natural AhR ligands
21 are rapidly metabolized and eliminated from biological systems. The following section will
22 examine the differences between the chemicals included in the TEF methodology and remaining
23 AhR ligands not included in this approach.
24
25 9.3.4.1. Industrial/Synthetic AhR Ligands
26 The synthetic compounds that bind to AhR include a number of different classes of
27 chemicals, most notably the PCDDs, PCDFs, and PCBs. Other synthetic AhR ligands include
28 industrial chemicals (polybrominated biphenyls, polychlorinated napthalenes, chlorinated
29 paraffins, etc.), pesticides (hexachlorobenzene), and contaminants (polybrominated dioxins,
30 dibenzofurans, and napthalenes) associated with various manufacturing, production, combustion,
31 and waste disposal processes. In addition, pyrolysis of organic material can produce a number of
32 non-halogenated polycyclic aromatic hydrocarbons (PAHs) with moderate to high affinity for
33 AhR (Poland and Knudson, 1982; Nebert, 1989; Chaloupka et al., 1993).
34 Not all of the anthropogenic sources of dioxin-like compounds are included in the TEF
35 methodology. Many of these chemicals, such as hexachlorobenzene and the brominated diphenyl
36 ethers, are only weakly dioxin-like and have significant toxicological effects that are not
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1 mediated by the AhR. For these chemicals, it is not clear that adding them to the TEF
2 methodology would decrease the uncertainty in the risk assessment process. For other classes of
3 chemicals, such as the chlorinated napthalenes, environmental concentrations and human
4 exposures are largely uncertain.
5 The PAHs are one class of anthropogenic chemicals not included in the TEF scheme
6 despite evidence for AhR binding. The PAHs are not included in the TEF methodology because
7 of their short half-lives and relatively weak AhR activity. In addition, the role of the Ah receptor
8 in the toxicity of the PAHs is uncertain. For example, both benzo[a]pyrene and chrysene induce
9 CYP1A1 activity through an AhR-mediated mechanism (Silkworth et al, 1995). However,
10 while the Ah receptor also plays a role in the immune suppressive effects of benzo[a]pyrene it
11 does not appear to be involved in the immune suppression induced by chrysene (Silkworth et al.,
12 1995). Furthermore, PAHs are DNA reactive and mutagenic and these mechanisms play a large
13 role in both the carcinogenicity and immunotoxicity of the PAHs (Ross and Nesnow, 1999). In
14 contrast, TCDD and other dioxin-like compounds are not DNA reactive. While there are PAHs
15 that bind to the AhR, the role of AhR or other competing pathways in the toxicity of these
16 compounds has not been clearly defined.
17 Brominated dioxins, dibenzofurans, biphenyls, and napthalenes also induce dioxin-like
18 effects in experimental animals (Miller and Birnbaum, 1986; Zacherewski et al., 1988;
19 Birnbaum et al., 1991; Hornung et al., 1996; DeVito et al., 1997; Weber and Greim, 1997). The
20 brominated dioxins and dibenzofurans may be more or less potent than their chlorinated
21 orthologues, depending on the congener (Birnbaum et al., 1991; DeVito et al., 1997). The
22 sources of the brominated dioxin-like compounds are not well characterized. Some of the
23 chemicals, such as the brominated biphenyls and their contaminants the brominated
24 naphthalenes, were synthesized and sold as commercial flame retardants. The manufacture and
25 use of polybrominated biphenyls has been prohibited. Brominated dibenzofurans are produced
26 as byproducts of synthesis and pyrolysis of some brominated flame retardants. There is some
27 evidence of human exposure to brominated dioxins and dibenzofurans from extruder operators
28 (Ott and Zober, 1996). Polybrominated, polychlorinated, and mixed bromo- and chloro- dioxins
29 and dibenzofurans have been found in soot from textile processing plants (Sedlak et al., 1998).
30 Although these chemicals have been found in humans, these studies are limited to a small
31 population and exposure to the general population remains undetermined. Future examinations
32 of the TEF methodology should include a more detailed discussion of the of the brominated
33 dioxins and dibenzofurans.
34
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|1 9.3.4.2. Naturally Occurring AhR Ligands
12 The evolutionary conservation of AhR and its biological function following activation by
3 dioxin-like compounds have led to the hypothesis that there must be an endogenous or
14 physiological ligand(s) for this receptor. Presently, the endogenous ligand remains
5 undetermined. However, efforts to discover the natural ligand have led to the discovery of a
6 number of naturally occurring AhR ligands. A number of naturally occurring chemicals present
7 in the diet are capable of binding to AhR and inducing some dioxin-like effects in experimental
8 animals (Bradfield and Bjeldanes, 1984; 1987) and humans (Michnovicz and Bradlow, 1991;
9 Sinha et al., 1994). The question of how the interaction of these chemicals relates to the toxicity
10 of those chemicals designated as dioxin-like has become the subject of much debate.
1 One class of naturally occurring chemicals that activate the AhR is the indole derivatives.
12 Indole derivatives, naturally present in a variety of cruciferous vegetables, are capable of
13 modulating the carcinogenicity of PAHs (Wattenberg and Loub, 1978). Indole-3-carbinol (I-3-C)
114 and 3,3'-diindolylmethane (DIM) are major secondary metabolites found in cruciferous
ll 5 vegetables and induce both phase I and II metabolic enzymes (CYP1 A-dependent glutathione and
116 glucuronyl transferases, oxidoreductases) in experimental animals (Bradfield and Bjeldanes,
117 1984, 1987), human cell lines (Bjeldanes et al., 1991; Kleman et al., 1994), and humans
|18 (Michnovich and Bradlow, 1990,1991). Although both-compounds induce CYP450 enzymes
119 under AhR transcriptional control, they exhibit relatively low binding affinity for the Ah receptor
|20 (Gillner et al., 1985). Further investigation revealed that I-3-C is relatively unstable in the acidic
21 environment of the digestive tract and readily forms DIM. In turn, DIM can participate in acid
122 condensation reactions to form indolocarbazoles (ICZs) (Chen et al., 1995). ICZs are also
23 produced by bacterial metabolism of the common dietary amino acid tryptophan. ICZs, in
24 particular indolo[3,2b]carbazole, exhibit high binding affinity for the rodent AhR, approximately
25 equipotent to 2,3,7,8-tetrachlorodibenzofuran, and can induce CYP1 Al activity in cultured cells
126 (Bjeldanes et al., 1991; Gillner et al., 1993; Chen et al., 1995). ICZ and a methylated derivative,
27 5,ll-dimethylindolo[3,2b]carbazole (MICZ), are also capable of binding to and activating the
28 AhR in human hepatoma cells (HepG2) (Kleman et al., 1994). With considerably lower efficacy,
29 I-3-C and DIM can partially displace TCDD from the AhR from human breast cancer cells
30 (T47D) (Chen et al., 1996). These results would suggest that this group of compounds may
31 represent a class of physiologically active AhR ligands derived from natural sources, which could
32 either mimic dioxin-like compounds in their action or act as competitors for AhR binding.
33 In addition to the plant-derived indoles, experimental animals consuming thermally
34 treated meat protein as well as humans fed cooked meat can exhibit induced CYP1A2 activity
35 (Degawa et al., 1989). High-temperature cooking (250°C, 22 minutes) of ground beef resulted in
36 the formation of a number of heterocyclic aromatic amines (HAAs) in part-per-billion levels,
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which were thought to be responsible for the observed CYP1A2 induction in human volunteers
(Sinha et al.5 1994). Mechanistic analysis of one particular HAA, 2-amino-3,8-
dimethylimidazo[4,5-f]quinoxaline (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 compounds.
Other diet-derived chemicals that can interact with the AhR include oxidized essential
amino acids. UV-oxidized tryptophan is capable of inducing CYP1 Al activity in mouse
hepatoma cells through an AhR-dependent mechanism (Sindhu et al., 1996). Rats exposed to
UV-oxidized tryptophan in vivo also exhibited induction of hepatic and pulmonary CYP1A1
activity. Both in vitro and in vivo enzyme induction were transient, with the oxidized tryptophan
possibly being metabolized by CYP1A1 (Sindhu et al., 1996). Tryptanthrins, biosynthetic
compounds produced from the metabolism of tryptophan and anthranilic acid by yeast commonly
found in food, are agonists for the rat AhR (Schrenk et al.,' 1997). Various tryptanthrins also
induce CYP1 Al-related enzyme activity in mouse hepatoma cells with the approximate efficacy
of indolo[3,2b]carbazole.
Recent studies have demonstrated that physiological chemicals can bind to the AhR.
Bilirubin was recently found to transform the AhR from mouse hepatoma cells into its DNA-
binding state, resulting in CYP1 Al induction. Hemin and biliverdin can also be metabolically
converted to bilirubin, resulting in AhR-dependent gene activation (Sinai and Bend, 1997).
Despite these results, there is no clear evidence that these are the physiological ligands for the
AhR, nor is there evidence that these compounds can modulate the activity of dioxin-like
compounds or lead to dioxin-like toxic effects in humans or animals.
9.3.4.3. Industrial/Synthetic vs. Natural AhR Ligands
A number of "natural" or dietary compounds have been identified, which in certain in
vitro cases can function as AhR agonists with similar potency when compared to various
halogenated aromatics. It has been postulated that the endogenous ligands could be the major
Contibutors to the daily dose of TEQs, because of their higher estimated intakes (Safe, 1995).
The "natural" ligands tend to have short half-lives and do not accumulate. The PCDDs/PCDFs
and PCBs included in the TEF methodology clearly bioaccumulate. If contributions to the total
TEQ are estimated on steady-state body burdens of these chemicals instead of daily intake, then
TCDD and other PCDDs/PCDFs and PCBs contribute more than 90% of the total TEQ
compared to the "natural" ligands (DeVito and Birnbaum, 1996). The difference in the results of
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11 these analyses demonstrates our uncertainty of the relative potencies, exposures and dose metrics
12 used in the comparisons of the synthetic dioxins vs. the natural AhRligands.
13 When a comparison is attempted between the perceived relative risk from natural vs.
14 anthropogenic AhR agonists, a number of factors should be taken into consideration. The
5 potency of AhRligands depends on several factors, including AhR binding affinity and
6 pharmacokinetic properties. When estimating the relative potency of a chemical compared to
7 TCDD, the larger the difference in pharmacokinetic properties, the greater the effect that study
8 design has on the relative potency. Initial studies comparing the potency of
9 indolo[3,2b]carbazole to TCDD demonstrate the importance of the pharmacokinetic differences
10 between these chemicals. In Hepa-1 cells exposed for 4 hours, the relative potency for induction
1 of CYP1A1 mRNA of indolo[3,2b]carbazole compared to TCDD is 0.1 (Chen et al, 1995). If
12 the relative potency is determined after 24 hours of exposure, the potency of
13 indolo[3,2b]carbazole drops 1,000-fold to 0.0001 (Chen et al., 1995). In addition, the dioxin-like
14 effects of low doses of indolo[3,2b]carbazole in Hepa-1 cells are transient. Similar transient
15 effects of other dietary-derived AhR ligands have also been reported (Xu and Bresnick, 1990;
16 Berghard et al., 1992; Riddick et al., 1994). These data demonstrate that the relative potencies of
17 these chemicals compared to TCDD are dependent upon the pharmacokinetic properties of the
18 chemicals and the experimental design used in the comparisons. In addition, these data also
i 9 demonstrate that for rapidly metabolizable AhR ligands, the effects are transitory and not
10 persistent like TCDD. It appears that the transient nature of the effect is due to the transient
>1 concentrations of these chemicals in these experimental systems. These data also demonstrate
122 our uncertainty of the relative potency of the dietary-derived AhR ligands.
J23 The chemicals included in the TEF scheme are those that not only bind to AhR but also
J24 bioaccumulate and have long biological half-lives in humans, typically on the order of years. In
125 contrast, the pharmacokinetics of the endogenous or natural group are not well studied, but these
126 chemicals tend to be short-lived, with half-lives on the order of minutes to hours. Although both
27 PAHs and the halogenated aromatics bind to AhR and induce cytochrome P450-related enzyme
28 activities, only the latter group produces the additional dioxin-like spectrum of toxicological
29 responses. These toxicities are thought to be due to the persistent exposures attributable to the
30 long half-lives of these chemicals (Riddick et al., 1994).
31 One caution when comparing the relative exposures to dietary AhR ligands and the
32 anthropogenic AhR ligands is that few in vivo studies have examined the relative potency of the
33 dietary or natural AhR ligands for toxic responses. Using the criteria of the WHO workgroup for
34 PCDDs, PCDFs, and PCBs results in only two in vivo studies of I-3-C which compared the REP
35 to TCDD (Wilker et al., 1996; Bjeldanes et al., 1991). In an in vivo study of the developmental
3 6 effects of 1-3 -C, in utero exposure of rats to 1-3 -C resulted in a number of reproduction-related
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1 abnormalities in male offspring, only some of which resemble those induced by TCDD (Wilker
2 et al., 1996). Because of the different spectrum of effects of 1-3 -C compared to TCDD in these
3 developmental studies, it is likely that mechanisms other than AhR activation are involved in
4 these effects. 1-3-C and some of its acid catalyzed oligomerization products alter androgen and
5 estrogen metabolism (Wilson et al., 1999; Telang et al., 1997), which may contribute to the
6 developmental effects of 1-3-C. While a number of in vitro studies have demonstrated dioxin-
7 like enzyme induction of the indole derivatives, in order to have REP values that adequately
8 describe the in vivo potency of these chemicals, future in vivo studies examining toxic responses
9 are required.
10 Although there are limited data on the in vivo biochemical and toxicological effects of
1T these ligands, the effects of mixtures of anthropogenic and natural AhR ligands is altogether
12 lacking. There is one study examining the interactions of I-3-C and DIM on the effects of
13 TCDD in cell culture systems. However, it is uncertain how to extrapolate these in vitro
14 concentrations to present human in vivo exposures. The limited data available do not adequately
15 address the interactions between these chemicals. Future in vivo studies are required in order to
16 better understand the potential interactions between these classes of AhR ligands.
17 Another limitation in comparing the natural AhR ligands to the dioxins is the multiple
18 effects induced by the natural AhR ligands. In vivo and in vitro studies of I-3-C indicate that it
19 induces a number of biochemical alterations that are not mediated through the AhR (Broadbent
20 and Broadbent, 1998). The activation of these additional pathways creates difficulties in making
21 direct comparisons with TCDD and related chemicals. Similarly, the PAHs also have non-AhR-
22 mediated biochemical and toxicological effects that also complicate direct comparisons with
23 TCDD and related dioxins. For example, co-exposure to TCDD and PAHs have demonstrated
24 both synergistic and antagonistic interactions in mice depending upon the endpoint examined
25 (Silkworth et al., 1993).
26 Presently, there are several limitations in our understanding of the importance of naturally
27 occurring dioxin-like compounds vs. the dioxin-like compounds included in the TEF
28 methodology. First is the limited data available on the dioxin-like toxicities of the natural
29 ligands. In addition, there is a lack of data on the interactions between these classes of
30 chemicals. Few if any mixtures of natural AhR ligands and PCDDs or PCDFs examining a toxic
31 response have been published. Many of the natural AhR ligands have multiple mechanisms of
32 action that presently cannot be accounted for in the TEF methodology. For example, I-3-C has
33 anticarcinogenic properties in tumor promotion studies, and these effects may or may not be
34 mediated through AhR mechanisms (Manson et al., 1998). The lack of data and the role of non-
35 AhR mechanisms in the biological effects of these chemicals prohibit a definitive conclusion on
36 the role of natural vs. anthropogenic dioxins in human health risk assessment. Though it is
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1 important to address these issues, the available data do not lend themselves to an appropriate
2 quantitative assessment of these issues.
3 Although Safe has suggested that exposure to natural AhR ligands is 100 times that of
4 TCDD and other dioxin-like compounds (Safe, 1995), the impact of the natural AhR ligands
5 remains uncertain. Epidemiological studies suggest that human exposures to TCDD and related
6 chemicals are associated with adverse effects, such as developmental impacts and cancer. In
7 many of these studies, the exposed populations have approximatley 100 times more TCDD
8 exposure than background populations (see Part II, Chapter 7). If the exposure to natural AhR
9 ligands is included in these comparisons, then the exposed populations should have
110 approximatley double the total TEQ exposures than the background population. It seems
11 unlikely that epidemiological studies could discriminate between such exposures. These data
12 suggest that the estimates of the contribution of the natural AhR ligands to the total TEQ
13 exposure are overestimated. In addition, regardless of the background human exposure to
14 "natural" AhR ligands, the margin of exposure to TCDD and related chemicals between the
15 background population and populations where effects are observed remains a concern.
16
17 9.4. TOTAL TEQ AND THE ADDITIVITY CONCEPT
18 The issue of the scientific defensibility of additivity in determining total TEQ has been
19 raised since the onset of the use of TEFs. Arguments regarding this approach include the
20 presence of competing agonists or antagonists in various complex mixtures from environmental
21 sources, interactions based on non-dioxin-like activities (inhibition or synergy), and the fact that
22 dose-response curves for various effects may not be parallel for all congeners assigned TEFs.
23 Although comparative pharmacokinetics have also been raised as an issue, this has generally
24 been accounted for by the heavier weight accorded to in vivo studies in the assignment of TEFs.
25 Despite these concerns, empirical data support the use of the additivity concept, recognizing the
26 imprecise nature of the TEFs per se. A substantial effort has been made to test the assumptions
27 of additivity and the ability of the TEF methodology to predict the effects of mixtures of dioxin-
28 like compounds. These efforts have focused on environmental, commercial, and laboratory-
29 derived mixtures. In addition, endpoints examined ranged from biochemical alterations, such as
30 enzyme induction, to toxic responses such as tumor promotion, teratogenicity, and
31 immunotoxicity. A brief summary of some of the more important work is given and discussed in
32 the following section.
33 The TEF methodology has been examined by testing mixtures of chemicals containing
34 dioxins and sometimes other chemicals. These mixtures have either been combined and
35 produced in the laboratory or were actual environmental samples. Researchers have also used
36 different approaches in estimating the TCDD equivalents of the mixtures. Some researchers
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1 have determined the REP of the components of the mixture in the same system in which the
2 mixture was tested and have used these REPs to estimate TCDD equivalents. These studies can
3 provide insight into the validity of the assumption of additivity of the TEF methodology. Other
4 researchers have used consensus TEF values to estimate the TCDD equivalents of the mixture. It
5 is not clear if these studies can be considered true tests of the additivity assumption. The
6 consensus TEF values have been described as conservative estimates of the relative potency of a
7 chemical in order to protect humans and wildlife. If the consensus TEF values are conservative
8 and protective, then they should overestimate the potency of mixtures tested in an experimental
9 system. In essence, using the consensus TEF values should generally over predict the potency of
10 a mixture (and therefore under predict the response) when compared to the equivalent
11 concentrations of TCDD in an experimental system. In the following discussion of the studies
12 examining the assumption of additivity, these differences in study design and their implications
13 for interpretation of the data must be considered.
14
15 9.4.1. Examination of Laboratory Mixtures of PCDDs and PCDFs
16 Bock and colleagues evaluated the TEF methodology in several systems using both
17 individual congeners as well as laboratory-derived mixtures (Lipp et al., 1992; Schrenk et al.,
18 1991,1994). REPs or toxic equivalents or "TEs" (as designated by the authors) were determined
19 for 2,3,7,8-substituted PCDDs based on enzyme induction in human HepG2 cells, rat H4IIE
20 cells, and primary rat hepatocytes. Three laboratory-defined mixtures (Ml, M2, and M3) were
21 prepared and then examined in these same cell culture systems. TCDD contributed between
22 6%-8% of the TEQs for Ml and M2, but was not present in M3. In Ml, 1,2,3,4,6,7,8-HpCDD
23 contributes approximately 60% of the TEQ, and 1,2,3,7,8-PCDD and 1,2,3,4,7,8-HxCDD
24 contribute 10% each. In M2, 1,2,3,4,6,7,8-HpCDD contributes 45%, 1,2,3,7,8-PCDD and
25 1,2,3,4,7,8-HxCDD contribute 15% each; and TCDD, 1,2,3,6,7,8-HxCDD, and 1,2,3,7,8,9-
26 HxCDD contribute less than 10% to the total TEQ. The TEQs in M3 are derived predominately
27 from 1,2,3,4,7,8-HxCDD (50%); 1,2,3,4,7,8-HxCDD (20%); and 1,2,3,6,7,8-HxCDD (18%).
28 These mixtures also contain up to 49 chlorinated dibenzo-j>-dioxins. The TCDD equivalents of
29 the mixtures were determined on the basis of the assumption of additivity using the TEF
30 methodology and the laboratory derived REPs or TEs as well as experimentally by comparing the
31 ECJ0s of the mixtures with that of TCDD. According to the authors, in all three systems the data
32 demonstrated that the components of the mixture act in an additive manner (Lipp, 1991; Schrenk
33 et al., 1991). For example, in the human HepG2 cells the EC50 of a mixture of 49 different
34 PCDDs was determined experimentally at 0.034 pg TEQ/plate, compared to the calculated or
35 predicted EC50 of 0.028 pg TEQ/plate. Interestingly, the TEF methodology accurately predicted
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1 the effects of M3, a mixture containing predominately OCDD, some heptaCDDs and hexaCDDs,
2 and no pentaCDDs or TCDD (Schrenck et al, 1991).
3 Bock and colleagues also tested a mixture of 49 PCDDs in a rat liver tumor promotion
4 study. The mixture, designated as M2, was the same mixture used in the cell culture studies
5 described above and TCDD contributed approximatley 8% of the TEQs of this mixture. In
6 theses studies, rats received an estimated 2-200 ng TCDD/kg/d or 200-20,000 ng mixture/kg/d.
7 The doses of the mixture were equivalent to the TCDD doses using a TE of the mixture of 0.01
8 based on enzyme induction in rat hepatocytes (Schrenk et al., 1991). A comparison of the
9 relative potency of the mixture was based on liver concentrations of the chemicals followed by
110 TEQ calculations using the I-TEFs (NATO/CCMS, 1988). According to the authors, in the low-
11 dose region (2-20 ng TCDD/kg/d) the I-TEFs accurately predict the enzyme-inducing activity of
12 the mixture but tend to overestimate the potency of the mixture at the higher doses (20-200
113 ng/kg/d). Also, according to the authors, the I-TEFs provide a rough estimate of the tumor-
14 promoting potency of the mixture but overestimate the mixture's potency . However, the authors
15 did not quantify or qualify the magnitude of the overestimation.
16 In the studies by Schrenk and colleagues, the TEQs were based on tissue dose, not
17 administered dose. Recent studies by DeVitq et al. (1997b, 2000) indicate that the REP for
18 dioxin-like compounds can differ when determined based on administered or tissue dose. The
19 higher chlorinated dioxins tend to accumulate in hepatic tissue to a greater extent than does
20 TCDD, and their REPs tend to decrease when estimated based on tissue dose (DeVito et al.,
21 1997b, 2000). Because the I-TEFs are based on an administered dose, they may not predict the
22 response when the TEQ dose is expressed as liver concentration. If the TEQ dose in the data by
23 Schrenk et al. (1994) is compared on an administered dose, then the dose-response relationship
24 for increases in relative volume of preneoplastic ATPase-deficient hepatic foci (% of liver) are
25 comparable between TCDD and the mixture, indicating that additive TEFs provided an
26 approximation of the tumor-promoting ability of a complex mixture of PCDDs (Schrenck et al.,
27 1994). In addition, because TCDD contributed less than 10% of the total TEQ in these mixtures,
28 these data indicate that the assumption of additivity reasonably predicts the response of complex
29 mixtures of dioxins.
30 In responsive mouse strains, induction of cleft palate and hydronephrosis by TCDD
31 occurs at doses between 3 and 90 //g TCDD/kg (Nagao et al., 1993; Weber et al., 1985;
32 Birnbaum et al., 1985, 1987, 1991). Several groups have examined the assumption of additivity
33 using teratogenic effects of dioxins as an endpoint. Birnbaum and colleagues examined TEF
34 methodology using mouse teratogenicity as an endpoint (Weber et al., 1985; Birnbaum et al.,
35 1985, 1987, 1991). REPs were derived for 2,3,7,8-TCDF, 2,3,4,7,8-PeCDF, 1,2,3,7,8-PeCDF,
36 and 1,2,3,4,7,8-HxCDF (Weber et al., 1984, 1985; Birnbaum et al., 1987). Analysis of the dose-
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1 response for these chemicals, based on administered dose, demonstrated parallel slopes.
2 According to the authors, dose-response analysis of two mixtures containing either TCDD and
3 2,3,7,8-TCDF or 2,3,4,7,8-PeCDF and 1,2,3,4,7,8-HxCDF demonstrated strict additivity
4 (Birnbaum et al., 1987; Weber et al, 1985).
5 Nagao et al. (1993) also examined the TEF methodology using teratogenicity in mice as
6 an endpoint Mice were exposed to a single dose of TCDD (5-90 //g/kg) or a mixture of PCDDs,
7 or one of two different mixtures of PCDFs. The mixtures contained no detectable TCDD. Thel-
8 TEFs were used to determine the TEQ of the mixtures. According to the authors, the I-TEFs
9 predicted the potency of the PCDD mixture, and the dose-response relationship was consistent
10 with the assumption of additivity. The I-TEFs overestimated the potency of the PCDF mixtures
11 by two- or fourfold. All three mixtures contained significant concentrations of non 2,3,7,8-
12 chloro-substituted PCDDs and PCDFs in addition to the dioxin-like compounds present. In the
13 studies by Bimbaum and colleagues (Weber et al., 1985; Birnbaum et al., 1985,1987,1991) and
14 Nagao et al. (1993) examining the assumption of additivity using teratogenicity as an endpoint,
15 the TEF methodology proves useful in estimating the effects of these mixtures.
16 Rozman and colleagues have examined the assumption of additivity of PCDDs in both
17 acute and subchronic studies. In acute studies, TCDD (20-60 //g/kg), 1,2,3,7,8-PCDD (100-300
18 y"g/kg), 1,2,3,4,7,8-HxCDD (700-1,400 /ug/kg), and 1,2,3,4,6,7,8-HpCDD (3,000-8,000 ^g/kg)
19 were administered to male rats, and REP values were determined for lethality. A mixture of all
20 four chemicals at equally potent concentrations was then prepared and dose-response studies
21 were performed with the mixture at doses that would produce 20%, 50%, and 80% mortality.
22 The mixture studies demonstrated strict additivity of these four chemicals for biochemical and
23 toxicological effects (Stahl et al., 1992; Weber et al, 1992a,b). Following the acute studies,
24 Viluksela et al. (1998a,b) prepared a mixture of these chemicals and estimated the TEQ based on
25 the REPs from the acute studies. A loading/maintenance dose regimen was used for 90 days and
26 the animals were followed for an additional 90 days. According to the authors, the assumption of
27 additivity predicted the response of the mixture for lethality, wasting, hemorrhage, and anemia,
28 as well as numerous biochemical alterations such as induction of hepatic EROD activity and
29 decreases in hepatic phosphenolpyruvate carboxykinase and hepatic tryptophan 2,3-dioxygenase
30 (Viluksela et al., 1997; 1998). Increases in serum tryptophan concentrations and decreases in
31 serum thyroxine concentrations were also predicted by the TEF methodology (Viluksila et al.,
32 1998a).
33 Rozman and colleagues followed up these initial studies by examining the assumption of
34 additivity of the effects of PCDDs as endocrine disrupters (Gao et al., 1999). Ovulation is a
35 complex physiological phenomenon that requires the coordinated interaction of numerous
36 endocrine hormones. In a rat model, ovulation can be inhibited by TCDD at doses between 2 to
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|1 32 //g/kg (Gao et al., 1999). Dose-response analysis of TCCD, 1,2,3,7,8-PeCDD, and
|2 1,2,3,4,7,8-HxCDD demonstrate that the slopes are parallel and the REPs are 0.2 and 0.04,
p respectively. According to the authors, the dose response for a mixture of these chemicals, in
J4 which the components were at equally potent concentrations, further demonstrated the response
5 additivity of mixtures of PCDDs and the predictive ability of the TEF methodology (Gao et al.,
6 1999).
7 The research on the interactions between mixtures of PCDDs and PCDFs has taken two
8 approaches. The first is to derive REP values in the same system in which the mixtures shall be
9 tested. These studies confirm that the assumption of additivity can predict the response of
1 0 mixtures of PCDDs and PCDFs. A second approach is to use the I-TEFs to assess the potency of
1 a mixture. These studies tend to indicate that the I-TEFs overestimate the potency of a mixture
2 by factors of two to four. Recently, the WHO TEFs have been described as "order of magnitude"
1 3 estimates of the potency of dioxin-like compounds. However, the studies using consensus TEFs
1 4 demonstrate that for mixtures of PCDDs and PCDFs, the TEF methodology will predict within a
5 half-order of magnitude or less (Schrenck et al., 1994; Nagao et al., 1993). In either case, the
1 6 TEF methodology accurately predicts the responses of experimentally defined mixtures of
1 7 PCDDs and PCDFs. Furthermore, several of these studies described the effects of mixtures
1 8 containing either no TCDD or with TCDD contributing less than 1 0% of the TEQ in the presence
1 9 of significant concentrations of non-2378- CDDs and CDFs. These studies strongly support the
10 use of the TEF methodology.
9.4.2. Examination of Commercial or Laboratory-Derived Mixtures of PCDDs, PCDFs,
andPCBs
Commercial mixtures of PCBs elicit a broad spectrum of biological and toxicological
responses in both experimental animals and humans. Some of the observed effects resemble
those induced by dioxin and furans (enzyme induction, immunotoxiciry, teratogenicity, endocrine
alterations, etc.). Attempts to expand the TEF approach to risk assessment of PCBs have
investigated the ability of both commercial PCBs and individual congeners, selected on the basis
of structure-activity relationships, to induce dioxin-like effects and to interact with TCDD. One
of the first studies to examine the interactions of individual PCB congeners with TCDD used
mouse teratogencity as an endpoint (Birnbaum et al., 1985, 1987). A mono-ortho PCB
(2,3,4,5,3',4'-HxPCB or PCB 156) at doses of 20 mg/kg or higher (Birnbaum, 1991) induced
hydronephrosis and cleft palate in mice. When mice were co-exposed to PCB 156 and 3.0 //g
TCDD/kg the interactions resulted in strict additivity.
The interaction of TCDD with dioxin-like PCBs has been examined by van Birgelen et al.
(1 994a,b) in subchronic rat feeding studies. Concentrations of PCB 1 26 in the diet between 7
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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
and 180 ppb induced several dioxin-like effects, including CYP1A1 induction, thymic atrophy,
liver enlargement, and decreases in hepatic retinol concentrations, body weight gains, and plasma
thyroxine concentrations. The REP for PCS 126 was estimated by the authors at between 0.01
and 0.1 (van Birgelen et al., 1994a). Co-exposure to PCB 126 and TCDD (0.4 or 5.0 ppb) in the
diet demonstrated additivity for all responses except induction of C YP1A2 and decreases in
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 also induced 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 coplanar 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
CYP1A1, hi liver, lung, and skin; ACOH activity, a marker for CYP1A2, in liver; and hepatic
porphyrins (DeVito et al., 1997a; 2000; van Birgelen et 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 at high
doses, 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 palrnitate 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 palrnitate, they
also can be used to predict responses of mixtures in another species.
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1 The commercial PCB mixtures induce a variety of dioxin-like effects. Rats exposed to
2 commercial Aroclors and observed for 2 weeks exhibited dose-dependent induction of hepatic
3 CYP1A activity (EROD) but no thymic atrophy (Harris et al., 1993). Using REP values derived
4 for EROD induction in rats, the TEF methodology provided good agreement with experimental
5 estimates of the ED50 for enzyme induction. However, use of the conservative TEF values of
6 Safe (1990) overestimated the potency of the Aroclor mixutres (Harris et al., 1993). In contrast,
7 similar studies examining immunotoxicity as an endpoint demonstrate that both experimentally
8 derived REP values and the conservative TEF values of Safe (1990) overestimate the potency of
9 the Aroclor mixtures by a factor of 1.2 - 22 (Harper et al., 1995). These data demonstrate that
110 there are nonadditive interactions between dioxin-like compounds and the non-dioxin-like PCBs
11 and that these interactions are response specific and most likely are not due to AhR antagonism.
12 In in vitro systems, using H4IIe cells and rat hepatocytes, Schmitz et al. (1995, 1996)
13 examined the assumption of additiviry for individual congeners as well as commercial mixtures.
14 After deriving REP values for enzyme induction, the authors concluded that a laboratory mixture
15 of PCBs 77, 105, 118, 126, 156, and 169 demonstrated perfect additive behavior in these cell line
16 systems (Schmitz et al., 1995). However, when the mixture was combined with a tenfold surplus
17 of a mixture containing non-dioxin-like PCBs (PCB 28, 52, 101, 138,153 and 180), the mixture
18 , demonstrated an approximate threefold higher TEQ than predicted. The authors concluded that a
19 moderate synergistic interaction is responsible for the increased enzyme-inducing potency of the
20 mixture containing dioxins and non-dioxin-like PCBs. Further studies by Schmitz et al. (1996)
21 also demonstrated a slight synergistic deviation (less than threefold) from strict additivity when
22 the calculated TEQ based on chemical analysis of Aroclor 1254 and Clophen A50 was compared
23 to the CYPlA-induction TEQ derived in an established rat hepatoma cell line (H4IIE) (Schmitz
24 etal., 1996).
25 Researchers have evaluated the applicability of the TEF methodology to mixtures
26 containing dioxin-like PCBs by examining the interactions of binary mixtures, laboratory-derived
27 mixtures, or commercial mixtures of PCBs. The studies examining the binary mixtures or
28 laboratory-derived mixtures have demonstrated that the assumption of additivity provides good
29 estimates of the potency of a mixture of PCBs and other dioxin-like compounds. In contrast,
30 studies using commercial mixtures of PCBs suggest that the assumption of additivity may be
31 endpoint specific, and that both synergistic and antagonistic interactions may occur for some
32 mixtures of dioxins and PCBs for certain endpoints. A more detailed examination of these issues
33 follows in the section on nonadditive interactions with non-dioxin-like compounds.
34
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1 9.4.3. Examination of Environmental Samples Containing PCDDs, PCDFs, and/or PCBs
2 One of the first tests of the TEF methodology examined soot from a transformer fire in
3 Binghamton, NY (Eadon et al, 1986). Benzene extracts of soot from a PCB transformer fire
4 which contained a complex mixture of PCDDs, PCDFs, PCBs, and polychlorinated
5 biphenylenes were administered to guinea pigs as single oral doses, and LD50 values were
6 compared to TCDD. Relative potency values for the PCDDs and PCDFs based on guinea pig
7 LDSO values were used to estimate the TCDD equivalents of the mixture. Eadon and co-workers
8 exposed guinea pigs to either TCDD alone or the soot and determined their LD50s. With these
9 relative potency values, the soot extract had a TCDD equivalent concentration of 22 ppm.
10 Comparison of the LD50s for TCDD and the soot led to a TCDD equivalent of 58 ppm for the
11 mixture. Other endpoints examined included alterations in thymus weight, body weight, serum
12 enzymes, and hepatotoxiciry. Experimentally the TCDD equivalents of the soot varied from 2 to
13 58 ppm. The authors concluded that because the benzene extract of the soot contained hundreds
14 of chemicals including PCDDs, PCDFs, and PCBs, the difference between the calculated TEQ of
15 22 ppm and the experimentally derived TEQs between 2 and 58 seems minimal. (Note: the
16 initial analytical TEQ value of soot [22 ppm] was calculated on the basis of guinea pig LD50
17 values of the respective components; using the current recommended TEF scheme [van den Berg
18 et al., 1998], the "calculated" TCDD TEQ would be approximately 17 ppm.)
19 Shortly after the studies on the Binghamton transformer fire soot, investigators applied the
20 TEF methodology to the leachate from Love Canal, NY. The organic phase of the leachate
21 consisted of more than 100 different organic compounds including PCDDs and PCDFs. The
22 leachate did not contain PCBs or PAHs. The authors estimated the TEQ of the mixture on the
23 basis of REP values for teratogenicity (cleft palate and hydronephrosis in mice) for the PCDDs
24 and PCDFs present in the leachate. The authors state that the leachate contained the equivalent
25 of 3 fj% TCDD/g and that more than 95% of the TEQ was contributed by TCDD. There were two
26 other PCDFs present in the leachate, and their contribution to the total TEQ was approximately
27 5% (Silkworth et al., 1989). When the TEQ of the mixture was based on dose-response analysis
28 of the mixture compared to TCDD, the leachate was estimated to contain between 6.6 and 10.5
29 ng TCDD/g (Silkworth et al., 1989). The authors concluded there was a good agreement
30 between the experimental TCDD equivalents (6.6-10.5 //g TCDD/g) and the analytical TEQs
31 (3 //g TCDD/g). In addition, these studies illustrate that the non-AhR components of the leachate
32 did not interfere with receptor-mediated teratogenicity (Silkworth etal., 1989). Additional
33 investigations have shown that the same complex mixture of non-AhR agonists slightly
34 potentiated TCDD-induced thymic atrophy and immunosuppression (plaque-forming cells/spleen
35 response) while decreasing the hepatic CYPlA-inducing ability of the TCDD component
36 (Silkworth etal., 1993).
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|1 The assumption of additivity was also examined using a PCDD/PCDF mixture extracted
\2 from fly ash from a municipal waste incinerator (Suter-Hofmann and Schlatter, 1989). As a
purification step, rabbits were 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, lyophilized and
pulverized liver was added to the standard rat diet. This diet was fed to rats for 13 weeks and
17 body weights and terminal thymus weights were recorded. The authors concluded that the
8 mixture of PCDDs and PCDFs produced equivalent toxicities as TCDD, and the assumption of
9 additivity was confirmed.
10
1 9.4.4. Nonadditive Interactions With Non-Dioxin-Like Compounds
12 For a number of toxicological responses, there appears to be evidence for nonadditive
13 interactions in defined dose ranges by both commercial Aroclors and major congeners with little
14 if any AhR agonist activity (i.e., PCB 153). Both commercial Aroclors and a PCB mixture
15 comprised of major congeners found in human breast milk were shown to antagonize the
16 immunotoxic effects of TCDD in mice (Biegel et al., 1989; Davis and Safe, 1989; Harper et al.,
17 1995). When immunotoxicity-derived TEF values for a variety of PCB congeners were used in
18 an additive manner to estimate TCDD TEQs for commercial Aroclors, in comparison to the
19 experimental TEQs, they were approximately predictive for Aroclor 1254 and 1260 (Harper et
IQ al., 1995). However, the TEF approach tended to overestimate the immunotoxiciry of Aroclors
1242 and 1248, suggesting some antagonism.
|22 Typical responses to TCDD exposure in rodents include CYP1 enzyme induction and
thymic atrophy. Rats consuming a diet containing 5 ppb TCDD for 13 weeks exhibited a 33-fold
J24 increase in hepatic CYP1A activity (EROD) and a greater than 50% reduction in relative thymus
J25 weight. Addition of PCB 153 to the diet at concentrations up to 100 ppm had no significant
126 effect on either response (van der Kolk et al., 1992). Mice dosed simultaneously with TCDD and
27 up to a 106-fold molar excess of PCB 153 (1 nmol/kg vs. 1 mmol/kg) exhibited no significant
28 dose-dependent alteration in hepatic CYP1A1/A2 protein compared to the TCDD dose group
29 alone (De Jongh et al., 1995). There was, however, an approximate twofold increase in hepatic
30 EROD activity in the highest combined PCB 153:TCDD dose group. Subsequent tissue analysis
31 revealed that the increase in EROD activity appeared related to PCB 153 increasing hepatic
32 TCDD concentrations. The same PCB congener at high doses (358 mg/kg) is able to almost
33 completely inhibit TCDD-induced suppression of the plaque-forming cell (PFC) response toward
34 sheep red blood cells in male C57BL/6J mice (Biegel et al., 1989; Smialowicz et al., 1997).
35 However, as PCB 153 displays negligible AhR binding affinity, the exact mechanism(s) behind
36 these interactions is unknown. Recently, it has, been shown that PCB 153 at high doses (greater
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1 than 100 mg/kg) actually enhances the PFC response in female B6C3F1 mice, thereby raising the
2 "control" set point. When combined doses of TCDD and PCS 153 are then compared to the
3 elevated PCB 153 response, an apparent block of the immunosuppressive effect of TCDD is
4 observed (Smialowicz et al, 1997). The relevance of this functional antagonism is uncertain, as
5 the doses required to inhibit the TCDD-like effects are at least 100 mg/kg of PCB 153. These
6 doses of PCB 153 seem unlikely to occur in human populations except under extreme conditions.
7 Commercial PCBs and various PCB congeners have been shown to potentiate or
8 antagonize the teratogenicity of TCDD depending upon the dose ranges and response examined
9 (Biegel et al., 1989; Morrissey et al., 1992). Treatment of developing chicken embryos with
10 TCDD and dioxin-like PCBs induces a characteristic series of responses, including embryo
11 lethality and a variety of embryo malformations/deformities. Combined exposure of chicken
12 embryos to both PCB 126 and PCB 153 (2 //g/kg and 25-50 mg/kg, respectively) resulted in
13 protection from PCB 126-induced embryo malformations, edema, and liver lesions, but not
14 mortality (Zhao et al., 1997). In mice, doses of 125 mg PCB 153/kg or higher inhibit the
15 induction of cleft palate by TCDD (Biegel et al., 1989; Morrissey et al., 1992). The induction of
16 hydronephrosis by TCDD was slightly antagonized by PCB 153, but only at doses of 500 mg/kg
17 or higher. Once again, the environmental relevance of exposures of 100 mg/kg of PCB 153 or
18 higher remains quite speculative, and nonadditive interactions are not expected at environmental
19 exposures.
20 Nonadditive interactions have also been observed in rodents exposed to both TCDD and
21 mixtures of various PCB congeners for hepatic porphyrin accumulation and alterations in
22 circulating levels of thyroid hormones. A strong synergistic response was seen with hepatic
23 porphyrin accumulation in female rats following the combined dietary exposure to TCDD and
24 PCB 153 (van Birgelen, 1996a). The mechanism accounting for the interaction was thought to
25 be a combination of both AhR-dependent (CYP1A2 induction) and AhR-independent
26 (8-aminolevulinic acid synthetase [ALAS] induction) events. Additionally, subchronic exposure
27 of mice to a mixture of PCDDs, PCDFs, and dioxin-like PCBs in a ratio derived from common
28 foods also resulted in a highly synergistic response, when compared to an equivalent dose of
29 TCDD alone, for both hepatic porphyrin accumulation and urinary porphyrin excretion (van
30 Birgelen et al., 1996b). PCB 153, although not porphyrinogenic alone, when added to the
31 mixture further enhanced the synergistic response of hepatic porphyrin accumulation. Non-AhR-
32 mediated induction of ALAS activity by both the dioxin-like mono ortho-substituted PCBs in the
33 mixture and by PCB 153 was hypothesized to partially explain the synergism.
34 Decreases in thyroid hormone levels have been observed in both experimental animals and
35 humans following exposure to both dioxin-like and non-dioxin-like compounds (Nagayama et
36 al., 1998; Koopman-Esseboom et al., 1997). It is currently thought that multiple mechanisms,
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11 including induction of specific isozymes of hepatic UDP-glucoronyl transferase (UDPGT) and
2 binding to thyroid hormone transport proteins (thyroid binding globulin, transthryetin) could be
3 involved. Exposure of female rats to a food-related mixture of PCDDs, PCDFs, and dioxin-like
4 PCBs for 90 days resulted in an approximately 85% decrease in decrease in plasma levels of
5 thyroxine. In contrast, the TCDD equivalent dose produced no effect on serum thyroxine (van
16 Birgelen et al., 1997). Increased induction of several isoforms of UDPGT by the HAH mixture
7 as compared to TCDD was thought to only partially explain the observed response with
8 thyroxine levels.
9 Several studies examining the interactions of dioxins and non-dioxins for rat liver tumor
10 promotion and additive and nonadditive interactions have been reported. Synergistic interactions
1 for tumor promotion have been observed for combinations of PCB 77 and PCB 52 (2,2',5,5r-
12 tetrachlorbiphenyl) in rat liver (Sargent et al., 1992). Eager et al. (1995) reported greater than
13 additive interactions of PCBs 126 and 153 in a rat liver tumor promotion model. The assumption
14 of additivity was examined in a laboratory-derived mixture of PCDDs, PCDFs, and PCBs in a rat
|15 liver tumor promotion model (van der Plas et al., 1999). The mixture contained TCDD,
116 1,2,3,7,8-PeCDD, 2,3,4,7,8-PeCDF, and PCBs 126, 118, and 156. In addition, a dose-response
117 study was performed using the mixture with PCB 153 added. Van der Plas and colleagues
[18 concluded that the TEF methodology predicted the tumor-promoting potency of the mixture quite
ll 9 well, within a factor of two (van der Plas et al., 1999).
|20 The interactions of dioxins with non-dioxin-like compounds results in additive and
|21 nonadditive responses. The antagonistic interactions, while endpoint specific, appear to occur at
|22 dose levels that greatly exceed most human exposures and should not affect the overall use of the
23 TEF methodology. One of the difficulties in addressing the nonadditive interactions is
24 understanding the mechanism behind these interactions. For the greater than additive
25 interactions for induction of porphyria and decreases in serum thyroxine, there are hypotheses
26 that may explain these effects. The mechanism of the antagonistic interactions of non-dioxin-
27 like PCBs and TCDD on immunotoxicity and teratogenicity in mice is uncertain. For other
28 responses, such as developmental reproductive toxicity, the interactions of PCDDs, PCDFs, and
29 PCBs have not been examined. In addition, it has also been suggested that antagonism of Ah
30 receptor-mediated events may be species specific. For example, addition of PCB 52, a congener
31 commonly found in biotic samples, inhibited the TCDD-induced expression of a reporter gene
32 under the regulatory control of the Ah receptor in mouse and rat cells, but not in guinea pig or
33 human hepatoma cells (Aarts et al., 1995). Our limited understanding of the interactions
34 between dioxins and non-dioxins for a variety of responses requires further research before their
35 impact on the TEF methodology can be fully understood.
36
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1 9.4.5. Examination of the TEF Methodology in Wildlife
2 Many wildlife species also exhibit toxic effects associated with exposure to halogenated
3 aromatic hydrocarbons. Early life stage (ELS) or sac fry mortality in fish, characterized by
4 edema, structural malformations, and growth reduction prior to fry mortality can be induced in
5 trout species following exposure to dioxin-like PCDDs, PCDFs, and PCBs (Walker and
6 Peterson, 1991). Binary combinations of a variety of PCDDs, PCDFs, and both dioxin and non-
7 dioxin-like PCB congeners injected into fertilized trout eggs were also capable of inducing ELS
8 mortality, with the majority of interactions between the congeners described as strictly additive
9 (Zabel et al., 1995). When a synthetic complex mixture of PCDDs, PCDFs, and PCBs, in
10 congener ratios that approximated Great Lakes fish residues, was tested in the ELS mortality
11 assay, the lethal potency observed for the mixture, compared to TCDD, deviated less than
12 twofold from an additivity approach (Walker et al., 1996). Recently, the TCDD TEQ of an
13 environmental complex mixture of PCDDs, PCDFs, and PCBs extracted from lake trout and
14 applied to the ELS bioassay could also be predicted by an additivity approach (Tillitt and Wright,
15 1997). These results suggest that additional halogenated aromatic compounds, including non-
16 dioxin-like PCBs, present in fish do not significantly detract from an additivity response for this
17 AhR-mediated event.
18 There are also numerous studies that have examined the effects of environmental mixtures
19 in marine mammals and avian species (Ross, 2000; Giesy and Kantian, 1998; Ross et al., 1996;
20 Shipp et al., 1998a,b; Restum et al., 1998; Summer et al., 1996a,b). Ross and colleagues
21 examined captive harbor seals fed herring from either the Atlantic Ocean (low levels of
22 PCDDs/PCDFs/PCBs) or the Baltic Sea (high levels of PCDDs/PCDFs/PCBs). The seals fed
23 herring from the Baltic Sea displayed immunotoxic responses including impaired natural killer
24 cell activity and antibody responses to specific antigens. These effects were correlated with the
\
25 TEQ concentrations in the herring. Using mink as a model, Aulerich, Bursian, and colleagues
26 have also examined the TEF methodology. Minks were fed diets containing carp from Saginaw
27 Bay to provide exposures of 0.25, 0.5, or 1 ppm PCB in the diet. In a series of reports, the
28 authors demonstrated that the diet induced dioxin-like effects ranging from enzyme induction to
29 reproductive and developmental effects, and that these effects were correlated with the dietary
30 intake of TEQs (Giesy and Kannan, 1998). Similar studies in White Leghorn hens also
31 demonstrated that the TEQ approach provided accurate estimates of the potency of the mixtures
32 (Summer etal., 1996).
33 In summary, current experimental evidence suggests that for PCDDs, PCDFs, coplanar
34 dioxin-like PCBs, and strictly AhR-mediated events, the concept of TEF additivity adequately
35 estimates the dioxin-like toxicity of either synthetic mixtures or environmental extracts, despite
36 the variations in relative contributions of each congener. Addition of the more prevalent mono-
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11 and di-ortho-substituted PCBs to a mixture, at least in the case of environmental extracts and
2 wildlife, does not seem to significantly detract from this assumption of additivity. Interactions
13 other than additivity (antagonism, synergism) have been observed with a variety of effects
4 (teratogenicity, immunotoxicity, hepatic porphyrin accumulation, thyroid hormone metabolism)
5 in both binary combinations and complex synthetic mixtures of dioxin and partial or non-Ah
6 receptor agonists (commercial PCBs, PCB 153). However, it appears that at these high-dose
7 exposures, multiple mechanisms of action not under the direct control of the Ah receptor are
8 responsible for these nonadditive effects.
[ 9 Additional research efforts should focus on complex mixtures common to both
10 environmental and human samples and the interactions observed with biological and
1 toxicological events known to be under Ah receptor control. In the interim, the additive
12 approach with TEFs derived by scientific consensus of all available data appears to offer a good
13 estimation of the dioxin-like toxicity potential of complex mixtures, keeping in mind that other
14 effects may be elicited by non-dioxin-like components of the mixture.
15 -
16 9.4.6. Toxic Equivalency Functions
17 The TEF methodology has been described as an "interim" methodology. Since this
|18 interim method has been applied, there have been few proposed alternatives. One recent
19 proposal suggests that the TEF value be replaced by a toxic equivalency function (Putzrath,
10 1997). It has been proposed that the REPs for PCDDs/PCDFs are better described by a function
121 as compared to a factor or single-point estimate (Putzrath, 1996). Recent studies have examined
122 this possibility for a series of PCDDs/PCDFs and PCBs (DeVito et al., 1997; DeVito et al.,
J23 2000). For the PCDDs/PCDFs, the data indicate that the REPs estimated from enzyme induction
|24 data in mice are best described by a factor and not a function. For some of the PCBs examined, a
125 function fit better, but the change in the REP was within a factor of two to five for most of the
26 four enzymatic responses examined (DeVito et al., 2000). In addition, the dose dependency was
27 observed only at the high-dose and not in the low-dose region (DeVito et al., 2000).
28 Even though these studies suggest that a TE function may be useful, there are numerous
29 difficulties in applying this method. If the REPs are really functions and not factors, there must
30 be a mechanistic basis for these differences, and these mechanisms would most likely be
31 response specific and perhaps species specific. This would then require that for all critical
32 responses, every chemical considered in the TEF methodology would have to be examined.
33 Once again, it is highly unlikely that 2-year bioassays and multigenerational studies will be
34 performed on all the TEF congeners in the foreseeable future. The use of a TEF function
35 requires extensive data sets that are not available and are unlikely to be collected.
36
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1 9.4.7. Endpoint and Dose-Specific TEFs
2 It is often suggested that species, endpoint, and dose-specific TEFs may be required for the
3 TEF concept to provide accurate estimates of risk. Although these proposals are interesting,
4 specific TEFs would require a much more complete data set than is available at this time. One
5 reason the TEF methodology was developed was because these data were not available, and it
6 was unlikely that all relevant chemicals would be tested for all responses in all species, including
7 humans. For example, it is extremely unlikely that 2-year bioassays for carcinogenesis or multi-
8 generational studies will be performed on all chemicals included in the TEF methodology. Even
9 though there are significant data demonstrating that a number of chemicals produce dioxin-like
10 toxic effects, clearly the data set is not complete. For this reason, WHO recommends revisiting
11 the TEF values every 5 years.
12
13 9.5. UNCERTAINTY
14 TEFs are presented as point estimates, in spite of the fact that variability in supporting
15 experimental data can range several orders of magnitude for a particular congener. It has been
16 proposed that some of this variability can be attributed to differences in exposure regimens, test
17 species, or purity of the test compound; however, the reasons for much of this variability have
18 not been adequately examined experimentally and remain unknown. Because of the multiple
19 methods of deriving the REP values for a particular chemical, it is difficult to estimate the
20 variability or uncertainty of a TEF point estimate. Consequently, the TEQ approach as currently
21 practiced does not provide for a quantitative description of the uncertainty for individual TEF
22 values, nor has any proposed method for incorporating quantitative uncertainty descriptors into
23 TEFs received general support or endorsement from the scientific community. Suggestions have
24 been made to use meta-analytic approaches or Monte Carlo techniques, however (Finley et al.,
25 1999), these approaches are only as good as the data available. Given the incompleteness of the
26 available database, it seems unlikely that these approaches would provide much useful insight at
27 this time.
28 Qualitative statements of confidence are embodied in the discussions associated with the
29 establishment and revision of TEFs. • These qualitative judgments, when examined in the context
30 of a specific risk assessment, can provide valuable insight into the overall uncertainty of some
31 TEQ estimates. For example, using WHO TEFs (van den Berg et al., 1998) to look at
32 background exposure from a typical U.S. diet, it is clear that only a limited number of congeners
33 significantly contributed to the total TEQ. More than 80% of the TEQ-WHO98 associated with
34 background dietary exposure (1 pg/kg/d) comes from only five congeners: 2,3,7,8-TCDD,
35 1,2,3,7,8-PCDD, 2,3,4,7,8-PeCDF, and PCB 126 (see Part I, Volume 3). The variability of the
36 REP values found in the literature for these congeners is much lower than for congeners that are
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minor contributors to background TEQ. Furthermore, the assigned TEF values for the chemicals
contributing 80% to the TEQ intake are similar to the mean of their in vivo REP values. The
confidence in the TEF methodology is also increased by empirical examination. A number of
studies have examined complex mixtures of dioxin and non-dioxin-like compounds and the TEF
methodology consistently results in TEQ estimates within a factor 2-3 for these mixtures. Based
on these mixture studies 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 additiviry. Finally, the uncertainty in TEQ estimates is only one component of the
overall uncertainty in a dioxin risk assessment. The TEQ uncertainty only addresses the
confidences associated in ascribing 2,3,7,8-TCDD equivalents to a mixture. It does not address
the uncertainty associated with quantitatively linking health effects to 2,3,7,8-TCDD exposure, or
the uncertainties associated with exposure estimates themselves.
15 9.6. IMPLICATIONS FOR RISK ASSESSMENT
16 The TEF methodology provides a mechanism to estimate potential health or ecological
17 effects of exposure to a complex mixture of dioxin-like compounds. However, the TEF method
18 must be used with an understanding of its limitations. This methodology estimates the dioxin-
19 like effects of a mixture by assuming dose-additivity and describes the mixture hi terms of an
20 equivalent mass of 2,3,7,8-TCDD. Although the mixture may have the toxicological potential of
21 2,3,7,8-TCDD it should not be assumed for exposure purposes to have the same environmental
22 fate as 2,3,7,8-TCDD. The environmental fate of the mixture is still the product of the
23 environmental fate of each of its constituent congeners. Different congeners have different
24 physical properties such as vapor pressure, practical vapor partition, water octanol coefficient,
25 photolysis rate, binding affinity to organic mater, water solubility, etc. Consequently, both the
26 absolute concentration of a mixture in an environmental medium and the relative concentration
27 of congeners making up an emission will change as the release moves through the environment.
28 For some situations, treating emission as equivalent to exposure, which assumes that modeling
29 fate and exposure can be reasonably accomplished by treating a mixture as if it were all
30 2,3,7,8,-TCDD, is a useful but uncertain assumption. However, for many risk assessments the
31 differences in fate and transport of different congeners must be taken into consideration and TEQ
32 must be calculated at the point of exposure if more accurate assessments are to be achieved.
33 Similarly, many dioxin releases are associated with the release of non-dioxin-like compounds
34 such as pesticides, metals, and non-dioxin-like PHAHs, and their risk potential may also need to
35 be assessed in addition to dioxin-related risk.
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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
There are instances where exposures to PCBs are the major problem. The TEF
methodology provides risk assessors with a useful tool to estimate potential dioxin-related health
risks associated with these exposures. Typically, the congener makeup of environmental
exposures to PCBs does not resemble the congener profile of any of the commercial mixtures
produced. Because the environmental mixtures do not resemble the commercial mixtures, it is
not clear that using total PCB concentrations and comparing them to any of the commercial
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 PCB carcinogenicity (U.S. EPA, 1996; Cogliano, 1998)
9.7. SUMMARY
The AhR mediates the biochemical and toxicological actions of dioxin-like compounds
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 compounds 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.
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Table 9-1. Estimated relative toxicity of PCDD and PCDF isomers to 2,3,7,8-
T4CDDa
Isomer groups
DD
M,CDD
D2CDD
T3CDD
T4CDDb
P5CDD
H6CDD
H7CDD
O8CDD
DF
M,CDF
D2CDF
T3CDF
T4CDF
PSCDF
H6CDF
H7CDF
CXCDF
Toxicity factor relative to 2,3,7,8-T4 CDD
nontoxic
0.0001
0.001
0.01
0.01
0.1
0.1
0.01
0.0001
nontoxic
0.0001
0.0001
0.01
0.5
0.5
0.1
0.01
0.0001
a OME, 1984.
b Excluding 2,3,7,8-T4CDD.
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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
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
PCBs
IUPAC # Structure
77 3,3',4,4'-TCB
81 3,4,4',5-TCB
105 253,3',4,4'-PeCB
114 2,3,4,4',5-PeCB
118 2,3',4,4',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 2,3,3l,4,4',5'-HxCB
167 ^S'A^^S'-HxCB
169 S^'^^'^^'-HxCB
170 2,2',3,3',4,4',5-HpCB
180 2,2',3,4,4',555'-HpCB
189 2,3,3',4,4',5,5T-HpCB
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
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
0.0005
o.oooi
0.0005
0.0001
0.0001
0.1
0.0005
0.0005
0.00001
0.01
0.0001
0.00001
0.0001
1
1
0.1
0.1
0.1
0.01
0.0001
0.1
0.05
0.5
0.1
0.1
0.1
0.1
0.01
0.01
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.
* van Leeuwen, 1997.
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Table 9-3. The range of the in vivo REP values for the major TEQ contributors
Chemical
1,2,3,7,8-
PCDD
2,3,4,7,8-
PCDF
1,2,3,6,7,8-
HxCDD
PCB 126
Number
of
in vivo
endpoints
22
40
3
62
Range of
REPs
(mean±std)
0.16-0.9
(0.5±0.22)
0.018-4.0
(0.4±0.7)
0.015-0.16
0.0024-0.98
(0.20±0.20)
Number of
endpoints
from subchronic
studies
16
20
1
31
Range of
REPs
(mean±std
0.19-0.9
(0.53±0.24)
0.018-0.6
(0.20±0.13)
0.04
0.004-0.18
(0.13±0.13)
TEF
1
0.5
0.1
0.1
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TCDD (2,3,7,8)
TCDF (2,3,7,8)
2,3,3',4>4'-PeCB
Cl Cl
2,2',4,4',5,5'-HCB
Figure 9-1. Structures of polychlorinated dibenzo-p-dioxins, dibenzofurans and biphenyls.
The prototype chemical 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD[2,3,7,8])5 and example of a
dioxin-like dibenzofuran 2,3,7,8-tetrachlorodibenzfaran (TCDF[2,3,7,8]), a mono-ortho dioxin-
like PCS, 2,353',4,4'-pentachlorobiphenyl (233,3',4,4t-PeCB)5 a dioxin-like coplanar PCB,
3,3',4,4',5-pentachlorobiphenyl (3,3',4,4',5-PeCB) and an example of a non-dioxin-like di-ortho
substituted PCB, 2,2',4,4',5,5'-hexachlorobiphenyl (2,2',4,4',5,5'-HCB).
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24
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34
35
36
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39
40
41
42
43
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45
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48
49
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52
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