Pilot Study On
International Information Exchange
On Dioxins and Related
Compounds
Scientific Basis for the Development of the
International Toxicity Equivalency Factor
(I-TEF) Method of Risk Assessment for
Complex Mixtures of Dioxins and
Related Compounds
Report Number 178
December 1988
North Atlantic Treaty Organization
Committee on the Challenges of Modern Society
EPA
600
6
90
015
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fell-1X
qc-IO-
0003
ACKNOWLEDGMENTS
This report of the NATO/CCMS Pilot Study on International Information
Exchange on Dioxins and Related Compounds was prepared by the U.S. Envi-
ronmental Protection Agency (EPA), Office of Research and Development, and
Versar Inc., under Contract No. 68-02-4254. Mr. Erich W. Bretthauer of the
U.S. EPA was the Study Director for the Exposure and Hazard Assessment
Working Group, and Dr. Donald G. Barnes of the U.S. EPA was the Chairman
of the TEF Subgroup. Dr. Stephen H. Safe of Texas A&M University was the
principal author of this report. Contributing authors also included
Dr. Frederick W. Kutz of the U.S. EPA and Mr. David P. Bottimore of Versar
Inc. Peer reviewers of this report are listed below as are the lead dele-
gates and TEF Subgroup members that reviewed and concurred with the publi-
cation of this report.
Dr. Linda S. Birnbaum
Dr. Brendan Birmingham
Ms. Sigrid Louise Bjornstad
Dr. Martin J. Boddington
Mr. E.A. Cox
Dr. Alessandro di Domenico
Dr. Donald L. Grant
Prof. Dr. med. Helmut Greim
Dr. Arne Grove
Dr. G.K. Matthew
Ms. Christa Morawa
Dr. James R. Olson
Ms. Frances Pollitt
Dr. Ellen Silbergeld
Dr. C.A. van der Hiejden
Dr. Job A van Zorge
Dr. James Wilson
National Institute of Environmental Health
Sciences (United States)
Ontario Ministry of the Environment (Canada)
State Pollution Control Authority (Norway)
Environment Canada (Canada)
Inspectorate of Pollution, Department of
the Environment (United Kingdom)
Istituto Superiore di Sanita (Italy)
Health and Welfare Canada (Canada)
GSF Muenchen Institute fur Toxikologie
(Federal Republic of Germany)
Kemiteknik, Teknologisk Institut (Denmark)
Department of Health and Social Security
(United Kingdom)
Umweltbundesamt (Federal Republic of
Germany)
State University of New York - Buffalo
(United States)
Department of Health and Social Security
(United Kingdom)
Environmental Defense Fund (United States)
National Institute of Public Health and
Environmental Hygiene (The Netherlands)
Ministry of Housing, Physical Planning and
Environment (The Netherlands)
Monsanto Chemical Company (United States)
HEADQUARTERS U««Y
ENVIRONMENTAL PR0TBCT10H
WASHINGTON, D.C. 20460
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SCIENTIFIC BASIS FOR THE DEVELOPMENT OF THE INTERNATIONAL TOXICITY
EQUIVALENCY FACTOR (I-TEF) METHOD OF RISK ASSESSMENT FOR
COMPLEX MIXTURES OF DIOXINS AND RELATED COMPOUNDS
Table of Contents
Page No.
List of Tables . ii
List of Figures i ii
1. INTRODUCTION 1
2. PCDDs AND PCDFs - ENVIRONMENTAL IMPACT 3
3. PCDDs AND PCDFs - PROBLEMS IN RISK ASSESSMENT 6
4. 2,3,7,8-TCDD AND RELATED COMPOUNDS - MECHANISM OF ACTION .... 9
5. DEVELOPMENT OF TEFs FOR PCDDs AND PCDFs 17
5.1 Lethal i ty 18
5.2 In Vivo Biologic and Toxic Responses 18
5.3 In Vitro Potencies 24
5.4 Toxicity Equivalency Factors 32
6. BIOASSAYS FOR HAZARD ASSESSMENT OF PCDD AND PCDF MIXTURES ... 41
REFERENCES 46
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LIST OF TABLES
Multiplicity of PCDD and PCDF Isomers and Congeners
Page No.
7
Table 1.
Table 2. Toxicity of 2,3,7,8-TCDD and Related Compounds:
Species Differences 10
Table 3. Cooperative LDc0 Values for Several PCDD and PCDF
Congeners in the Guinea Pig, Mouse, and Rat 19
Table 4. A Summary of the Dose-Response In Vivo Biologic and
Toxic Effects of Several Halogenated Aromatics in the
Immature Male Rat and Guinea Pig 21
Table 5. £055 Values for PCDF Congeners in the Mouse:
Immunotoxicity and Teratogenicity 23
Table 6. PCDDs and PCDFs as Competitive Ligands for the Rat
Hepatic Cytosol Receptor SARs 26
Table 7. PCDDs and PCDFs as Inducers of AHH and EROD Activities
in Rat Hepatoma H-4-II E Cells in Culture: SARs 28
Table 8. International Toxicity Equivalency Factors (I-TEFs):
Comparison of Relative Potency Data for the 2,3,7,8-
Substituted PCDDs and PCDFs 35
Table 9. Brominated Aromatic Flame Retardant Pyrolysis and
Municipal Fly Ash Extracts: In Vitro and In Vivo
Determination of 2,3,7,8-Equivalents 42
Table 10. Calculated 2,3,7,8-TCDD Equivalent Concentrations of
the Binghamton Soot for Various Dose-Related
Endpoints Following Subchronic Exposure 45
11
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LIST OF FIGURES
Page No.
Figure 1.
Figure 2.
Figure 3.
Figure 4,
Figure 5.
Figure 6.
Figure 7,
Chemical Sturucture of PCDDs and PCDFs
Proposed Mechanism of Action of 2,3,7,8-TCDD and
Related Compounds
Relative Potencies of Three Isomeric TetraCDDs (Rat) .
Correlation Between the -log ECgQ Values (m) for AHH
Induction Activity in Ret Hepatoma H-4-II E Cell vs.
the -log ED5Q Values (mol/kg) for Hepatic Microsomal
AHH Induction Activity (top panel), Inhibition of Body
Weight Gain (middle panel), and Thymic Atrophy (bottom
pane') in Male Wistar Rats for Several PCDD and PCDF
Congeners
Correlation Between the -log EC5Q Values (m) for AHH
Induction Activity in Ret Hepatoma H-4-II E Cell vs.
the -log £050 Values (mol/kg) for Hepatic Microsomal
AHH Induction Activity (top panel), Inhibition of Body
Weight Gain (middle panel), and Thymic Atrophy (bottom
panel) in Male Wistar Rats for Several PCDD, PCDF, PCB,
and PBDD Congeners
Correlation Between -Log EC^n (m)AHH Induction
(in Rat Hepatoma H-4-II-E cells) vs. -Log £050
(mol/kg) Inhibition of Body Weight Gain (Guinea Pig) ..
Correlation Between -Log ECcQ (m)AHH Induction (In
Rat Hepatoma H-4-II-E Cells) vs. -Log ED50 (mol/kg)
Immunotoxicity (C57BL/6 Mice)
12
14
30
Figure 8. 2378-Substituted PeCDFs
31
33
34
40
in
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1.
INTRODUCTION
The Pilot Study on International Information Exchange on Dioxins and
Related Compounds was initiated to apply the cooperative efforts of sever-
al nations to address issues associated with polychlorinated dibenzo-
p-dioxins (PCDDs), dibenzofurans (PCDFs), and related compounds. The
three-year study was conducted under the aegis of the Committee on the
Challenges of Modern Society (CCMS) of the North Atlantic Treaty Organiza-
tion (NATO). Participating nations included Canada, Denmark, the Federal
Republic of Germany, Italy, the Netherlands, Norway, the United Kingdom,
and the United States. In addition, the governments of Sweden and
Austria, although not members of NATO, requested to be informed of the
progress of the study. International organizations that participated as
observers included the World Health Organization, the United Nations
Environmental Programme, the Organization for Economic Cooperation and
Development, and the Commission of European Communities.
Numerous information exchange activities were undertaken to promote
cooperative research and to identify duplicative efforts and knowledge
voids so that better informed decisions can be made concerning future
activities. When the project was initiated in 1985, it was divided into
three areas of study: exposure and hazard assessment, technology assess-
ment, and management of accidents. The Exposure and Hazard Assessment
Working Group, chaired by the United States, was charged with several
tasks concerning research and risk assessment. The working group devel-
oped the International Toxicity Equivalency Factor (I-TEF) method to
assess the risks posed by exposures to complex mixtures of PCDDs and
PCDFs. The objective was to develop an updated TEF scheme based on exist-
ing methods and the most current data available (Bellin and Barnes, 1987).
The I-TEFs were reported in NATO/CCMS Report Number 176 and were presented
in the Special Session on Prospective Dioxin Research and Regulatory
Issues sponsored by the United States Environmental Protection Agency and
the CCMS at the Eighth International Symposium on Chlorinated Dioxins and
Related Compounds (Dioxin '88) in Umea, Sweden.
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The I-TEFs were presented as an interim procedure that should be of
practical assistance in addressing a number of questions associated with
PCDOs and PCDFs. The thrust of Report Number 176 was to endorse the
toxicity equivalency factor (TEF) concept as an interim measure of pru-
dent science policy and to recommend specific I-TEF values to facilitate
consistency among participating nations in communicating analytical re-
sults and performing risk assessments. That report also included examples
of how the procedure is applied to analytical data and how the various
TEF schemes differ. This report presents a detailed discussion of the
scientific basis for the development of toxicity equivalency factors and
provides the toxicity data and methodology used to determine the I-TEFs.
Section 2 of this report discusses the sources and environmental im-
pact of PCDDs and PCDFs and provides a background for the utility of risk
assessment techniques for these chemicals with regard to their presence
in humans and the environment. Section 3 presents some of the problems
encountered in performing risk assessments for complex mixtures of PCDDs
and PCDFs and the need for the development of toxicity equivalency fac-
tors. Section 4 presents the mechanisms of toxicity for these compounds
and the structure-activity relationships of the various isomers/congeners
of PCDDs and PCDFs. Section 5 provides a discussion of the development
of TEFs from experimental toxicity data. In addition, this section pre-
sents the data and methodology used to develop the I-TEFs by the NATO/CCMS
Pilot Study on International Information Exchange on Dioxins and Related
Compounds. Section 6 presents bioassay techniques that can be used to
estimate and/or verify the toxic equivalents calculated from analytical
data for a variety of mixtures including PCDDs and PCDFs as well as poly-
brominated dibenzo-p-dioxins (PBDDs) and polybrominated dibenzofurans
(PBDFs).
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2.
PCDDs AND PCDFs - ENVIRONMENTAL IMPACT
Polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs)
(Figure 1) are members of a class of organic pollutants that include the
polychlorinated biphenyls (PCBs), naphthalenes, azobenzenes, and ter-
phenyls; the polybrominated biphenyls (PBBs); and other mixed polychloro/
bromo aromatics. PCBs were widely used as industrial compounds, whereas
the PCDDs and PCDFs are not industrial or commercial chemicals but are
formed as by-products of diverse processes. For example, PCDDs and PCDFs
are formed during the synthesis of chlorinated phenols and their derived
products (Rappe et al. 1978, 1979; Nilsson et al. 1978). PCDDs and PCDFs
are also formed during the incomplete combustion or incineration of
organohalogen-cortaining wastes (Buser et al. 1978; Rappe et al. 1979,
1983; Buser, 1979; Lustenhouwer et al. 1980; Ballschmiter et al. 1983;
01ie et al. 1977) and in automobile combustion processes (Marklund et al.
1987). Polybrominated dibenzo-p-dioxins (PBDDs), dibenzofurans (PBDFs),
and halogenated (mixed bromo/chloro) dibenzo-p-dioxins and dibenzo-
furans have recertly been identified as by-products of waste incineration
in which there are sources of chlorine and bromine (Thoma et al. 1986,
Schafer and Ballschmiter 1986). In addition, PCDFs have also been
reported as contaminants in commercial PCBs (Vos et al. 1970, Bowes
et al. 1975, Kunita et al. 1984).
PCDDs, PCDFs, PCBs, and related halogenated aryl hydrocarbons are
highly stable, lipophilic chemicals and this is paralleled by their envi-
ronmental persistence and preferential bioaccumulation in higher trophic
levels of the food chain. Recent studies have identified PCDDs and PCDFs
in river and lake sediments (Czuczwa et al. 1984a, 1984b) and at subparts-
per-billion levels in fish, wildlife, human adipose tissue, and milk (Van
den Berg et al. 1985; Rappe et al. 1981, 1987; Ryan et al. 1985; Ryan
1986, Ono et al. 1986; Stalling et al. 1985; Kahn et al. 1988). High
resolution gas chromatographic-mass spectrometric (GC-MS) analysis has
demonstrated that the PCDDs and PCDFs present in environmental matrices
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Cl,
Polychlorinated dibenzo-p-dioxins (PCDDs)
Cl,
Polychlorinated dibenzofurans (PCDFs)
FIGURE 1. Chemical Structure of PCDDs and PCDFs
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are highly complex mixtures of isomers and congeners, and this correlates
with the composition of PCDDs and PCDFs identified in industrial or
combustion by-products.
Several studies have reported the high resolution GC-MS analysis of
PCDDs and PCDFs in human samples (Ryan et al. 1985, Van den Berg et al.
1985) and it was evident that humans are also primarily exposed to complex
mixtures of these compounds. In contrast, human exposure to 2,4,5-tri-
chlorophenol and its derived products (e.g., 2,4,5-T, Agent Orange) can
result primarily in exposure to a single PCDD congener, namely the highly
toxic 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD). In situa-
tions where there is excess exposure to 2,3,7,8-TCDD, the mixture of
PCDDs and PCDFs will contain higher levels of this compound. For
example, "Vietnam veterans who were heavily exposed to Agent Orange
exceeded matched control subjects in both blood and adipose tissue levels
of 2,3,7,8-TCDD" (Kahn et al. 1988). Nevertheless, even in this study,
2,3,7,8-TCDD was only one of 13 PCDD and PCDF congeners identified in
these human samples.
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3.
PCODs AND PCDFs - PROBLEMS IN RISK ASSESSMENT
It is apparent that risk assessment of PCDD or PCDF mixtures require
prior information on the identities of the specific congeners and their
concentration in any analyte. An analytical approach to this problem
requires the availability of all the appropriate PCDF and PCDD congeners
(see Table 1} and methods that can both separate and quantitate each indi-
vidual compound in a mixture. Several laboratories have made significant
progress in this field, using sophisticated cleanup procedures followed
by high resolution capillary GC and/or GC-MS analysis, which can separate,
identify, and quantitate individual PCDD and PCDF congeners {Buser and
Rappe 1980, Crummett 1983). In addition to the general lack of availabil-
ity of all the PCDD and PCDF standards, these analytical techniques
require highly sophisticated equipment and training and are therefore very
costly. Moreover, the use of analytical data for risk assessment requires
prior information on the relative potencies of the individual components
in the mixture, and it must also be assumed that the interactive effects
of the components in the mixture are not significant. This assumption may
not always be correct since several studies have reported antagonistic
interactions between 2,3,7,8-TCDD and the commercial PCB, Aroclor 1254,
and several 1,3,6,8-substituted dibenzofurans (Keyes et al. 1986, Haake
et al. 1987, Bannister et al. 1987, Davis and Safe, 1988). Antagonistic
interactions were observed only when the relative concentrations of the
antagonists were high.
Regulatory risk assessments for individual compounds routinely use
the results of long-term cancer bioassays for generating various "effect
levels" that can be used in standard setting. Rodent carcinogenicity
studies for 2,3,7,8-TCDD (Kociba et al. 1978) and a hexachlorodibenzo-
p-dioxin mixture (NTP 1980) have been reported, and the data obtained
for 2,3,7,8-TCDD have been used extensively by numerous regulatory
agencies (USEPA 1985, Kimbrough et al. 1984, Van der Heijden et al. 1982,
Bell in and Barnes 1987, Ontario, 1982, OME, 1985) to calculate virtually
safe doses (VSDs) or allowable daily intakes (ADIs). Carcinogenicity
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8366H
Table 1. Multiplicity of PCOO and PCOF Isomers and Congeners
PCDDs
PCDFs
Toxic PCOOsa
Toxic PCDFs8
Total 2.3.7.8-CDDs/CDFs
Total non -2.3.7.8-CDDs/CDFs
TOTAL
Number of i softie rs
C1L C12 C13 C14 C15 C16 Clj C18 Total
2 10 14 22 14 10 2 1 75
4 16 28 38 28 16 4 1 135
11311 7
1 2 4 2 1 10
17
193
210
2,3,7,8 substituted congeners (Mason et al. 1985. Safe 1986).
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studies on the other environmentally relevant individual PCODs and PCDFs
or mixtures of these compounds have not been determined. Thus, it is not
possible to calculate carcinogenicity-derived VSD or ADI values for PCDD
or PCDF isomers and congeners (other than 2,3,7,8-TCDD), and this has
necessitated the development of toxicity equivalency factors (TEFs) for
this class of compounds (Bellin and Barnes, 1987; Safe 1987). This
approach utilizes the available in vivo and in vitro data on the toxic
potencies of individual PCDDs and PCDFs and assigns a fractional number
comparing their toxic potency to that of the most toxic member of this
class of compounds, namely 2,3,7,8-TCDD (fractional number = 1.0). The
rationale for the use of TEFs is based on extensive research on the mech-
anism of action of 2,3,7,8-TCDD and related halogenated aryl hydrocarbons,
The following section of this report briefly highlights some of the
results that provide strong support for the use of TEFs in the risk
assessment process for PCDDs and PCDFs.
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4.
2,3,7,8-TCDO AND RELATED COMPOUNDS - MECHANISM OF ACTION
PCDDs, PCDFs.. and related aryl hydrocarbons elicit a broad spectrum of
biologic and toxic responses in diverse animal species and mammalian cells
in culture (reviewed in Poland et al. 1979; Poland and Knutson 1982; Safe
1986; Whitlock 1986, 1987). Some of these effects include a wasting syn-
drome, chloracne and related dermal lesions, hepatotoxicity and porphyria,
reproductive toxicity, immunologic alterations, and tumor promotion. In
addition, treatment of laboratory animals and mammalian cells with halo-
genated aromatics causes a number of responses including (a) the induction
of diverse Phase I and Phase II drug-metabolizing enzymes, ornithine
decarboxylase, and enzymes associated with steroid metabolism; (b) modula-
tion of diverse cellular hormone receptor levels; and (c) the depletion of
hepatic vitamin A levels.
2,3,7,8-TCDD, the most potent halogenated aryl hydrocarbon, has served
as a prototype for investigating the effects and mechanism of action of
this important class of chemicals. Utilizing 2,3,7,8-TCDD as a probe, it
has been shown that the effects and relative potency of this compound and
related halogenated aromatics are dependent on several factors, including
(a) the age and sex of the animal, (b) the species and strain used, and
(c) the structure of the chemical used. Although the wasting syndrome
and lymphoid involution are routinely observed in most laboratory animals
exposed to 2,3,7,8-TCDD and related compounds, many other responses are
highly species-specific (Table 2). For example, chloracne and related
dermal lesions are observed in humans, monkeys, some genetically inbred
strains of mice e.g., hairless (HRS/J) mice, cattle and rabbits, but not
in other animal species.
Some of the early research on 2,3,7,8-TCDD illustrated the differen-
tial species sensitivity to the lethality of 2,3,7,8-TCDD as evidenced by
the following LD^g values: guinea pig (0.6-2.0 ^g/kg), rat (22-45
M9/kg), chicken (25-50 ^g/kg), monkey (70 /^g/ kg), rabbit (115 /^g/kg),
dog (100-200 M9/kg), mouse (114-284 //g/kg), bull frog (> 1000 M9/kg), and
hamster (1157-5000 /ig/kg) (Kociba and Schwetz 1982). More recent studies
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8366H
Table 2. Toxicity of 2.3.7,8-TCDD and Related Compounds: Species Differences
Animal species
Effect
Chloracne and related
dermal lesions
Body weight loss
Developmental toxic ity
Liver damage
Edema
Thymic atrophy
Guinea
Human Monkey Pig
J J 0
J 4
J J
J 0
J 0
J J
Mouse Chicken
Jb o
4 J
J J
J 4
J J
J 4
Rat
0
4
4
4
0
4
4 Effect observed in animal species.
0 Effect not observed.
a Poland and Knutson (1982).
Observed only in hairless HRS/J mice.
10
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have also demonstrated that the mink are also susceptible to 2,3,7,8-TCDD
induced lethality (Hochstein et al, 1988) (LD5Q 4.2 /*g/kg) and that
there are remarkable intraspecies strain differences in the toxicity of
this compound (Chapman and Schiller 1985; Walden and Schiller 1985;
Pohjanvirta et a". 1988). These results illustrate the greater than 5000-
fold difference ~n the sensitivity of the guinea pig and hamster to the
lethal effects of 2,3,7,8-TCDD; however, it is not clear whether this
species sensitivity is observed for other effects caused by 2,3,7,8-TCDD
and related compounds.
The mechanisn of action of 2,3,7,8-TCDD and related compounds has been
extensively investigated using a number of different in vivo and in vitro
models. Unlike nany other toxic and genotoxic hydrocarbons, 2,3,7,8-TCDD
does not act via metabolic activation processes or by covalent modifica-
tion of cellular protein, RNA, and DNA {reviewed in Shu et al. 1987).
However, a multitude of studies strongly support a receptor-mediated mech-
anism of action for 2,3,7,8-TCDD and related halogenated aryl hydrocarbons
(Figure 2). The following lines of evidence support this mechanism.
Studies with genetically inbred "responsive" and "nonresponsive"
strains of mice (e.g., C57BL/6 and DBA/2 mice, respectively) and their
backcrosses showed that many of the biochemical (e.g., induction of aryl
hydrocarbon hydroxylase (AHH) (Poland and Glover 1973, 1974, 1975; Poland
et al. 1974) and toxic (porphyria, immunotoxicity, thymic atrophy, body
weight loss, epidermal hyperplasia and hyperkeratosis, and lethality)
responses caused by 2,3,7,8-TCDD segregate with a specific genetic locus,
namely, the aryl hydrocarbon (Ah) locus (Poland and Glover 1980; Vecchi
et al. 1980, 1983; Nagarkatti et al. 1984; Jones and Sweeney 1980; Luster
et al. 1986). The initial hypothesis for this locus, which codes for the
Ah receptor protein, was proposed to explain the differences in AHH
inducibility observed in genetically inbred strains of mice by aryl
hydrocarbons and 2,3,7,8-TCDD.
Poland and coworkers (1976) first identified a saturable, high-affini-
ty, low-capacity Ah receptor protein in murine hepatic cytosol. Subse-
quent studies in several laboratories have extensively characterized the
11
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Nuclear Binding Sites
Cytoplasm
Cytochrome
P-4501A1 mRNA
Cytochrome P-4501A1
Induction (AHH ind
other monooxygenases)
. "Induced Proteins" .
Pleiotroplc Response*
' -2,3,7,8-TCDD and
related isostereomert
-PAH
FIGURE 2. Proposed Mechanism of Action of 2,3,7,8-TCDD amd Related Compounds.
12
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molecular properties of this intracellular protein in several animal
species including humans (reviewed in Safe 1988). Furthermore, the direct
binding of the Ah receptor with other aryl hydrocarbons (Okey et al.
1984), including radiolabeled PCDF congeners has also been reported
(Parrel 1 et al. 1987).
Endogenous receptor ligands such as steroid hormones and neurotrans-
mitters interact with receptors that are located within specific tissues
or cells. Tissue specificity has also been demonstrated for the 2,3,7,8-
TCDD receptor ir rats and mice (Okey et al. 1983, 1984); Carlstedt-Duke
1979; Denison et al. 1986; Gasiewicz et al. 1984). C57BL/6J mice and
Sprague-Dawley rats, which are highly responsive to 2,3,7,8-TCOD, exhibit
tissue-dependent concentrations of the receptor that vary from 0 to 54
fmol/mg cytosolic protein. Relatively high levels of receptor have been
identified in tissues (e.g., liver, thymus) that are responsive to the
effects of 2,3,7,8-TCDD and related compounds. In contrast, nondetectable
levels of the receptor protein are observed in cytosol from DBA/2J mice,
which are relatively nonresponsive to the effects of 2,3,7,8-TCDD and
related toxic halogenated aryl hydrocarbons. However, the correlation
between tissue susceptibility and the receptor levels requires further
study.
Stereoselective interactions between receptors and their respective
hormone agonists are one of the important characteristics of the receptor-
mediated processes. The effects of structure on the biologic and toxic
potencies of PCDDs and PCDFs have been extensively investigated (reviewed
in Poland et al. 1975, 1979; Poland and Knutson 1982; Safe 1986).
Figures 3 and 4 depict the effects of variable lateral 2,3,7, and 8 chlo-
rine substituents on the toxicity of a series of PCDD and PCDF congeners
(Safe 1986, 1987; Mason et al. 1985, 1986). Figure 3 summarizes the
relative potencies of three tetrachlorodibenzo-p-dioxin (tetraCDD)
isomers, namely 2,3,7,8-, 1,3,7,8-, and 2,3,6,8-tetraCDD, on body weight
loss and thymic atrophy in the rat and their relative activities as induc-
ers of AHH in rat hepatoma H-4-II E cells and rat hepatic microsomes. The
13
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2,3,7,8-TCDD-
1,3,7,8-TCDD_~
2,3,6,8-TCDD
<5.0
<4.0
I I
in vitro
in vivo (rat)
8
-log ED 50
10
12
2,3,7,8-TCDD _
1,3,7,8-TCDD-
2,3,6,8-TCDD_
<3.0
<3.0
body wt. loss
thymic atrophy
I
5
I
7
-log ED50
FIGURE 3. Relative Potencies of 3 Isomeric TetraCDDs (Rat) [Upper
Panel; Induction of AHH in Rat Hepatoma H-4-NE Cells (In
Vitro) and Rat Hepatic Microsomes fin Vivo): Lower Panel;
Toxicities Observed in Male Rats (Safe 1986, 1987)]. The
In Vivo ED50 Values Are mol/kg and the jn Vitro Values for
Induction Are Concentrations (M).
14
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major structural diversity in these tetraCDD isomers is the number of
lateral chlorine substituents (4, 3, and 2, respectively). The EDgQ
values for 2,3,7,8-TCDO are approximately 100-fold lower than the values
for 1,3,7,8-tetraCDD, and the only structural difference in the two
isomers involves the shift of the C-2 lateral Cl group to a nonlateral
(C-l) position. The third isomer, 1,3,6,8-tetraCDD, is not toxic at
doses up to 10 mol/kg. Comparable SARs have been observed for PCDF
congeners (Bandiera et al. 1984; Mason et al. 1985). Structure-activity
relationships (SARs) for PCDDs and PCDFs have been observed for a number
of Ah receptor-Tiediated responses (e.g., teratogenicity, immunotoxicity,
lethality, enzyne induction, body weight loss, thymic atrophy, Ah receptor
binding, and various dermal lesions) in several animal species (guinea
pig, rat, and mouse) and mammalian cells in culture (see Section 5).
These data strongly support the role of the Ah receptor in mediating
the biologic and toxic responses elicited by 2,3,7,8-TCDD and related
PCDDs and PCDFs and provide the scientific basis for the development of
TEFs for this class of compounds. It should also be noted that the devel-
opment of toxic responses after exposure to PCDDs and PCDFs is a complex
multicomponent process. For example, cellular models have been developed
for investigating 2,3,7,8-TCDD-mediated epidermal hyperplasia, hyperkera-
tosis, sebaceous gland metaplasia, and tumor promotion in HRS/J hairless
(hr/hr) and haired (hr/+) mice (Knutson and Poland 1980, 1982; Poland
et al. 1982, 1933, 1984). Their results indicate that a second genetic
locus (the hr lacus) is involved in the development of these epidermal
responses and indicate that the Ah locus is necessary but not sufficient
for mediating the effects caused by 2,3,7,8-TCDD and related compounds.
However, despite these complicating factors, the structure-toxicity rela-
tionships were also observed for these responses. It is likely that the
mechanisms involved in other effects elicited by PCDDs and PCDFs also re-
quire other genetic loci and other factors; however, this should not alter
the structure-dapendent potencies of these compounds.
It should be noted, however, that information questions the unique
validity of extending the Ah receptor mechanism for TCOD toxicity to
15
-------
species other than the two strains of mice (C57BL/6J and DBA/2J) in which
the correlations are well established. For example, acute lethality
(Pohjanvirta et al. 1988), microsomal enzyme induction (Rozman et al.
1985a, 1985b, Henry and Gasiewicz 1986), thymic involution {Gorski et al.
1988), induction of cleft palate (Lamb et al. 1986), induction of hyper-
keratosis (Puhvel and Sakamotot 1987), suppression of antibody response
to sheep red blood cells (Pazdernik and Rozman 1985), as a consequence of
2,3,7,8-TCDD administration have been shown not to correlate with binding
affinity to the Ah-high affinity protein in other strains/species. Also,
{Gasiewicz and Rucci 1984) indicate that the correlation between sensitiv-
ity to 2,3,7,8-TCDD toxicity and Ah receptor binding affinity seen in
mice is not necessarily applicable to other species. Therefore, whenever
possible, TEFs are also to be based upon whole animal data, rather than
solely "Ah-receptor" binding affinities or enzyme induction. Such data
may be developed from other methods such as the early life stage bioassay
being developed in the Netherlands (Helder and Seinen 1985).
16
-------
5.
DEVELOPMENT OF TEFs FOR PCDDs AND PCDFs
The basic premise of this approach is that the TEF for a given com-
pound is assigned based upon its toxic potency relative to 2,3,7,8-TCDD
{which is assigned a value of 1). The development and validation of TEFs
for individual PCDDs and PCDFs require a detailed examination of all the
quantitative SARs (QSARs) that have been reported for this class of
compounds. These data can then be used to derive TEFs for all those
compounds that have been tested. Moreover, if an individual PCDD or PCDF
can be compared to 2,3,7,8-TCDD for more than one Ah receptor-mediated
response or in more than one animal species, the TEF values will undoubt-
edly encompass a range of values. Regulatory agencies can either develop
TEF ranges for individual PCDDs and PCDFs or use a graded system that
selects specific TEF values derived from criteria-based toxic or biologic
endpoints. For example, Health and Welfare Canada and Ontario Ministry
of the Environment (communication from D.L. Grant) have developed an
evaluation procedure in which the following responses are weighted in the
following descending order of rating priority:
1. Evidence for carcinogencity based on long-term animal studies.
2. Where evidence from long-term studies is absent, data from studies
of reproductive effects. Toxicological evidence indicates that
doses or exposure levels resulting in reproductive effects overlap
the range of doses or exposure levels that result in carcinogenic
effects,
3. Other subchronic toxic effects, e.g., thymic atrophy, body weight
loss, general toxicity.
4, Acute toxicity studies, e.g., LD50.
5. In vivo or in vitro biological effects, e.g., receptor binding,
enzyme induction.
It is clear that because of the general lack of carcinogencity data
for PCDDs and PCDFs, the effects noted in 2 through 5 will have to be
utilized to generate most of the TEFs.
This section of the report will briefly summarize the major studies
that have investigated in vivo or in vitro QSARs for PCDDs and PCDFs.
17
-------
5.1 Lethality
Table 3 summarizes the acute LD5Q values for 2,3,7,8-TCDD and
related PCDD isomers and congeners. The SARs are comparable to those
observed for other Ah receptor-mediated biologic and toxic responses (Safe
1986), and these studies provided some of the first experimental evidence
demonstrating SARs for PCDDs. Although there are considerable interspe-
cies differences with respect to the absolute values for the acute LDcn
values, the SARs for these compounds are comparable in all three species.
For example, in the guinea pig, the relative potency ratios for 2,3,7,8-
tetraCDD/l,2,3,7,8-pentaCDD/ 1,2,3,4,7,8-hexaCDD were 1, 0.2, and 0.008,
respectively (using the LDcQ of 0.6 pg/kg). The absolute potency ratios
derived from the mouse LDrQ values were different from those observed in
the guinea pig, however, the order of relative potencies of the three com-
pounds were the same in both animal species.
5.2 In Vivo Biologic and Toxic Responses
Several studies have reported the toxic effects of PCDDs, PCDFs and
related toxic halogenated aryl hydrocarbons for several receptor-mediated
responses (Mason et al. 1985, 1986, 1987; Bandiera et al. 1984; Nagayama
et al. 1985; Poland and Knutson 1982; Birnbaum et al. 1987a, 1987b; Davis
and Safe 1988; Pluess et al. 1988a, 1988b). Most of these studies are
limited with respect to the number of congeners used; however, in all
cases the qualitative SARs are comparable to those observed for other
receptor-mediated responses. Safe (1987) and coworkers have carried out
QSAR studies for a series of PCDD, PCDF, PCB, polybrominated dibenzo-
p-dioxins (PBDDs), and halogenated (Br-Cl) dibenzo-p-dioxins in
male Wistar rats (inhibition of body weight gain, thymic atrophy), male
Hartley strain guinea pigs (inhibition of body weight gain) and male
C57BL/6 mice (immunotoxicity) (Safe 1986, 1987; Bandiera et al. 1984,
Mason et al. 1985, 1986, 1987; Leece et al. 1985; Holcomb et al. 1988).
In addition, the structure-dependent effects of PCDDs, PCDFs, and related
compounds on the induction of rat hepatic microsomal AHH and related
18
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8366H
Table 3. Comparative LD50 Values for Several PCDD and PCDF
Congeners in the Guinea Pig, Mouse, and Rata
Congener
2.3.7.8 TetraCOD
2.3-DiCDD
2.7-OiCOO
2.8-OiCDD
1.3.7-TriCDO
2,3,7-TriCDD
1.2.3.4-letraCDD
1.3,6.8-TetraCDD
1.2,3,7,8 PentaCOD
1,2,4,7.8-PentaCOO
1,2,3,4,7,8-HexaCDD
l,2,3,6,7,8-Hexa:DD
1,2.3.7.8.9 HexaCOO
1,2.3,4.6,7.8-Hepta COD
OCDD
2.8-DiCDF
2,4,8-TriCDF
2,3.7.8-TetraCDF
2,3.4,7,8-PtmtaCOF
2,3.4,6.7.8 HexaCDF
Guinea pig
Ug/kg)
0.6 2
-
-
>300.000
29.444
-
>15.000.00Q
3.1
1.125
72.5
70-100
60 100
>600
-
-
5-10
3-10
120
LD values
Mouse
Ug/kg)
114-284; 182; 2,570
-
>2. 000. 000
8.470,000
>15.000,000
>3.000
-
>2. 987. 000
337.5
>5,000
825
1,250
>1,440
>4. 000. 000
>1 5. 000. 000
>15,000.000
>6,000
-
-
Rat
Ug/kg)
22-45; 164-409; >1400
>1. 000, 000
>1. 000, 000
>5, 000, 000
>5, 000, 000
>1. 000, 000
>1, 000, 000
>10,000,000
-
-
_
-
_
> 1,000, 000
>15.000.000
>1 5, 000, 000
>1.000
916
~
a Kociba and Schwetz 1982. Schwetz et al. 1973. McConnell et al. 1978, Kociba and Cabey
1985, Moore et al. 1979, McKinney et al. 1981; Brewster et al 1988; Chapman and Schiller
1985; Walden and Schiller 1985; Pohjanvirta et al. 1987.
19
-------
monooxygenases have also been determined. The results of all these
studies are summarized in Tables 4 and 5.
The SARs were comparable for the responses in the three animal spe-
cies. In all cases the most toxic compounds were the 2,3,7,8-substituted
PCDDs and PCDFs; the removal of lateral Cl groups or the addition of non-
lateral Cl substituents resulted in decreased toxic potencies. However,
a comparison of the relative toxic potencies of some of the more toxic
2,3,7,8-substituted compounds illustrates some important species and
response-dependent differences in relative toxicities. For example, in
the rat, 2,3,4,7,8-pentaCDF is consistently more toxic than 1,2,3,7,8-
pentaCDF, and the 2,3,4,7,8-pentaCDF/l,2,3,7,8-pentaCDF relative potency
ratios were 2.5, 8.4, and 39.7 for inhibition of body weight gain, thymic
atrophy, and the induction of hepatic microsomal AHH. Thus, based on
studies on the rat, 2,3,4,7,8-pentaCDF exhibited 2.5 to 39.7 times greater
potency than 1,2,3,7,8-pentaCDF. These values are well reflected in the
relative potency ratio of 10:1 (0.5/0.05), which was assigned for these
two isomers in the International-TEF (I-TEF) method. Recent in vivo
studies on the subchronic toxicities of 2,3,7,8-TCDD, 1,2,3,7,8-pentaCDD,
1,2,3,6,7,8-hexaCOF, 2,3,4,7,8-pentaCDF and 1,2,3,7,8-pentaCDF in rats
gave TEFs of 1.0, 0.4, 0.1, 0.4, and 0.01, respectively (Pluess et al.
1988a, 1988b). The 2,3,4,7,8-pentaCDF/l,2,3,7,8-pentaCDF ratio (i.e.,
40:1} was closer to the I-TEF value. A closer examination of the relative
activity of 1,2,3,7,8- and 2,3,4,7,8-pentaCDFs in the guinea pig (Table 4)
shows that the former isomer is more potent in this species (i.e., for
inhibition of body weight gain). The reasons for these species-dependent
differences in relative potencies are unknown.
Table 5 summarizes the relative toxicities of 2,3,7,8-TCDD and several
PCDF congeners in the mouse (Birnbaum et al. 1987a, 1987b; Davis and Safe,
1988). These results further illustrate some species and response-
dependent variations in TEFs. For example, 2,3,4,7,8-pentaCDF and
2,3,7,8-TCDD are equipotent as immunotoxins (i.e., inhibition of the
20
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8366H
Table 4. A Sunnary of the Dose-Response _!_£ Vivo Biologic and Toxic Effects of
Several Halogenated Anomalies in the Immature Male Rat and Guinea Piga
Compounds
A. PCDDs. PBDDs. and Br-PCDDs
2,3,7,8 tetraCDD
1,2,3,7,8-pentaCDO
1,2,3,4,7,8-hexaCDD
1,3, 7, 8- tetraCDD
1,2,4,7.8-pentaCDD
2,3,7,8-tetraBDD
2,3-dibroitio-7,8-diCDD
2,bromo-3,7,8-triCDD
1.2.3.7.8-pentaBOD
1.2.4.7.8-pentaBDO
1.3,7,8-tetraBDO
B. PCDFs
2,3,4,7.8-pentaCDF
1,2,3,4,7,8-hexaCOF
1,2,3,7,8-pentaCOF
2.3,4,6,7,8-hexaCDF
1.2,3,6,7.9-hexaCDF
2,3,7,8-tetraCOF
1.3.4.7.8-pentaCDF
2.3.4.7.9-pentaCDF
2.3.4.7-tetraCDF
1,2,3,7,9-pentaCDF
1.2.4.7,8-pentaCOF
1.2.3,7-tetraCDF
2,3.4,8-tetraCOF
1,2,4,6.7-pentaCDF
1.2.3.6-tetraCOF
PCOF mixtureb
Inhibition of
body weight gain
Gu i nea
Rat pig
0.05 0.0056
0.62
1.63
132 0.66
34.0 0.061
0.068
0.012
0.12
0.87
12.9
252
1.04 0.012
1.30
2.64 0.0059
2.80
3.20
3.20
26.1
22.0 0.040
34.0
49.3 0.160
49.3
86.9
137
>150
>250
0.52
In Vivo ED (jimol/kq)
50
Thvmiu atroohv
Guinea
Rat pig
0.09 ND
0.17
1.07
100 ND
11.0 ND
0.034
0.0073
0.035
0.39
6.17
35.5
0.21 ND
0.5
1.76 ND
0.93
0.93
3.60
0.70
5.5 ND
7.84
23.0 ND
46.4
110
>150
>150
>250
0.65
Hepatic AHH
Rat
0.004
0.031
0.03
31.2
2.82
0.00076
0.00049
0.0025
0.025
0.195
6.50
0.037
0.293
1.47
0.265
0.347
0.652
3.49
6.96
46.1
14.7
7.80
110
>150
>250
0.016
Induction
Guinea
pig
0.00028
1.6
0.049
0.0012
0.0059
0.026
0.14
21
-------
r
8366H
Table 4. (continued)
In Vivo ED
5()
Compounds
Inhibition of
body weiqht qain
Gu i nea
Rat pig
Thymic atrophy
Guinea
Rat pig
Hepatic AHH Induction
Guinea
Rat pig
C. PCBs
2.3.4,4',5-pentaCBP
2.3,3',4.4'.S'-hexaCBP
2,3,3't4,4'.5-nexaCBP
2,3,3'.4.4'-pentaCBP
2,3',4,4',5-pentaCBP
2.3,4,4'.5.-pentaCBP
S.S'^^'.S.S'-hexaCBP
3.3'.4.4'.5-pentaCBP
180
220
180
750
1.120
370
15
3.3
200
225
180
1.030
1,550
2.790
8.9
0.95
30
6
25
65
165
130
0.50
1.10
a Mason et al. 1985, 1986; Leece et al. 1985; Hoicomb et al. 1988).
This mixture contained the following congeners: 2,3,7,8-tetra-, 1,2,4,7,8-penta-
2,3.4,7.8-penta-, and 1,2,3,4.7.8-hexachlorodiberuofuran.
KD = non detect.
1.2.3.7.8-penta-,
22
-------
B3S6H
Table 5. ED50 Values for PCDF Congeners in the House:
Iirmunotoxicily and Teratogenicily3
ED5Q values (pmol/kg)
Compound
2,3,7,8-TCOD
2.3,4,7.8-pentaCDF
1,2,3,7,8-pentaCOF
2,3,7,8-tetraCDF
1.2,3.7,9-pentaCDF
1.2.3.4.7.8-hexaCDF
Teratogenicity
0.011
0.09
0.35
0.22
--
0.84
Imnunotoxicity
0.0024
0.0030
-
0.014
0.710
_
1,3,6,8-tetraCDF
35.7
"Davis and Safe 1988. Birnbaum et al. 1987a. 1987b.
23
-------
splenic plaque-forming cell response to sheep red blood cells) but the
teratogenic activity of 2,3,7,8-TCDD is eight-fold greater than 2,3,4,7,8-
pentaCDF. For the teratogenic response (i.e., cleft palate) (Birnbaum et
al. 1987b), the 2,3,4,7,8/1,2,3,7,8-pentaCDF potency ratio was less than
4, which is significantly lower than the 10:1 ratio derived from the I-TEF
values (i.e., 0.5/0.05). Examination of the data in Tables 4 and 5 illu-
strates that several other relative potency ratios are different from the
I-TEF values. The species- and response dependent differences in toxic
potencies are due to several factors including pharmacokinetic and meta-
bolic differences (Brewster and Birnbaum 1987, 1988), route, and duration
of administration (Couture et al. 1988). This latter factor is particu-
larly important for higher chlorinated PCDDs and PCDFs (Couture et al.
1988) and is discussed later in this document.
5.3
In Vitro Potencies
Several groups have developed the use of bioassays for detecting and
quantitating toxic halogenated arotnatics (Bradlaw and Casterline 1979;
Bradlaw et al. 1980; Eadon et al. 1986; Gierthy and Crane 1985; Knutson
and Poland 1980, 1982; Gierthy et al. 1984; Jansing and Shain 1985; Safe
1987). These assay systems primarily utilize mammalian cells in culture,
which measure a specific Ah receptor-mediated response (e.g., keratiniza-
tion, changes in cell morphology, receptor binding, or enzyme induction)
associated primarily with the toxic PCB, PCDF, and PCDD congeners.
Bradlaw and coworkers (1979, 1980) first reported that PCBs, PCDDs, and
PCDFs readily induce AHH activity in rat hepatoma H-4-II E cells in cul-
ture, and they demonstrated the utility of this assay system for detecting
toxic halogenated aromatics in diverse matrices including fish extracts,
PCB/PCDF-contaminated rice oil, and diverse food extracts including
gelatin samples containing pentachlorophenol and trace levels of higher
chlorinated PCDDs. In addition, relative Ah receptor binding assays for
halogenated aryl hydrocarbons also constitute a potential in vitro assay
system (Safe 1987).
24
-------
Table 6 summarizes the competitive Ah receptor binding affinities of
several PCODs and PCDFs using rat hepatic cytosol and [3H]-2,3,7,8-TCDD
as a radioligand. The qualitative SARs for these responses were compar-
able to those observed for the structure-toxicity relationships (Table 4);
however, it was apparent that there were major differences in their QSAR
data (Bandiera et al. 1984; Mason et al. 1985, 1986). This assay system
is limited by solubility problems and the narrow window of relative activ-
_p _C
ities {i.e., 10 to 10 M using the sucrose density gradient assay
system), and recent studies show that relative binding affinity EC50
values using the competitive assay system are not appropriate for quanti-
tatively determining structure-dependent differences (Mason et al. 1985,
1986; Bunce et al. 1988).
Table 7 summarizes the in vitro induction of AHH and ethoxyresorutin
0-deethylase (EROD) activities by several PCDD and PCDF congeners in rat
hepatoma H-4-II E cells in culture. Again, the SARs for this highly char-
acteristic Ah receptor-mediated response were comparable to those observed
in other studies (see Tables 4 and 6). The monooxygenases are readily
induced by low levels of the more toxic halogenated aryl hydrocarbons
(10" to 10" M), and differences in congener potencies of > 10
can be determined.
Figure 4 illustrates the correlation between the -log EC50 value for
AHH induction (in vitro) vs. the -log ED^Q values for in vivo body
weight loss, thymic atrophy, and AHH induction (in the rat) for the PCDDs
and PCDFs combined (see Table 4). The linear correlation coefficients
were r = 0.89, 0.83, and 0.82, respectively. The correlations observed
(Figure 5) between the in vitro induction bioassay results and the
complete list of halogenated aryl hydrocarbons shown in Table 4 gave even
higher r values (0.91, 0.92, and 0.84). It was also reported that the
only individual compounds that did not fit in these correlations were
compounds that contained two adjacent unsubstituted positions. These
compounds tend to be more rapidly metabolized in vivo (Safe 1980) and are
less active in the in vivo assays (note: the results for these congeners
25
-------
r
6366H
Table 6. PCDDs and PCDFs as Competitive Ligands for
the Rat Hepatic Cytosol Receptor SARs
Congener
Dibenzofuran
2-monoCDF
3-monoCDF
4-raonoCOF
2.3-diCDF
2,6-dlCDF
2.8-diCDF
1,3.6-triCDF
1,3,8-triCDF
2.3,4-trlCDF
2,3,8-triCOF
2,6.7-triCDF
2.3.4,6-tetraCDF
2,3,4.8-tetraCDF
2,3,6.8-tetraCDF
2,3,7.8-tetraCDF
1,2.4.8-tetraCOF
1.2.3.6-tetraCOF
1.2.3,7-tetraCOF
1.3,4,7,8-pentaCDF
2.3,4,7,9-pentaCDF
1,2,3.7.9-pentaCDF
1.2.4.6.7-pentaCDF
1,2.4.7.9-pentaCOF
1.2,3.4.8-pentaCDF
1.2,3,7.8-pentaCDF
1,2,4.7.8-pentaCDF
2,3,4.7.8-pentaCDF
1,2.3 4 7 8-hexaCOF
1.2,3.6.7,8-hexaCDF
1.2.4,6,7.8-hexaCDF
2.3,4,6.7.8-hexaCOF
2.3,7,8-TCDO
2,3.7-triCDD
2,8-diCDD
1,2,3,7.8-pentaCDD
2.3.6.7-tetraCDD
2.3.6-triCDD
1,2.3.4,7,8-hexaCOO
1.3.7.8-tetraCDD
1,2.4,7.8-pentaCOO
Receptor binding
affinities (EDSQ)
(M)
<10'3
2.8 x 10~4
4.2 ± 0.6 x 10"5
< 10~3
4.72 x 10"6
2.46 x 10~4
2.57 x 10~4
4.40 x l(f6
8.50 x 10~5
1.9 x 10"5
1.0 ± 0.1 x 10"G
4.5 x 10~7
3.5 x 10~7
2.0 x 10~7
2.2 x 10"7
4.1 ± 0.6 x 10"8
>io-5
3.5 x 10"7
1.1 x 10~7
2.0 x 10"7
2.0 x 10"7
4.0 x 1(T7
6.77 x 10~8
2.0 x 10"5
1.2 x 10"7
7.45 t 2.04 x 10"8b
1.3 x 10"6
flK
1.5 ± 0.1 x 10 HD
2.3 x 10~7
2.7 ± 1.0 x 10~7b
8.3 x 10~6
4.7 ± 0.4 x 10"8b
1.0 x 10"8
7.1 x 10"8
3.2 x 10"6
7.9 x 10"8
1.6 x 10"7
2.2 x 10"7
2.8 x 10~7
7.9 x 10~7
1.1 x 10"6
26
-------
8366H
Table 6. (continued)
Congener
Receptor binding
affinities (E05Q)
(M)
1.2,3,4-tetraCDD
1.2.3.4./-pentaCDD
1.2.4-t.riCDD
OCDD
1 monoCDD
1.3 x 10
6.4 x 10
1.3 x 10
-6
-5
>1.0 x 10
>1.0 x 10~
-5
Competitive EDrn values were determined using the sucrose density
gradient assay system.
Dose-response experiments carried out in triplicate and expressed
as means + SD.
27
-------
8366H
Table 7. PCDDs and PCDFs as Inducers of AHH and EROD Activities
in Rat Hepatoma H-4-II E Cells in Culture: SARs
Congeners
A. PCDFs
Oibenzofuran
2-monoCDF
3-monoCDF
4-monoCDF
2.3-diCOF
2,6-diCDF
2.8-diCDF
1.3.6-triCDF
1.3.8-triCDF
2,3,4-triCDF
2,3,8-triCDF
2.6.7-triCDF
2,3,4,6-tetraCDF
2.3,4,8-tetraCDF
2.3,6,8-tetraCDF
2,3,7,8-tetraCDF
1,2,4,8-tetraCDF
1,2,3,6-tetraCDF
1.2.3.7-tetraCOF
1.3,4,7,8-pentaCDF
2.3,4,7,9-pentaCDF
1,2.3.7,9-pentaCDF
1.2.4,6.7-pentaCDF
1,2,4.7,9-pentaCDF
1.2.3.4.8-pentaCDF
1,2.3,7.8-pentaCDF
1,2,4,7.8-pentaCDF
2.3.4,7.8-pentaCDF
1,2,3.4.7.8-hexaCOF
1,2.3.6,7.8-hexaCDF
1,2,4.6,7,8-hexaCDF
2,3,4.6.7.8-hexaCDF
B. PCOOs
2,3,7,8-tetraCDO
1,2,3,7,8-pentaCDD
2,3,6,7-tetraCDO
2,3,6-triCOD
1,2,3,4,7,8-hexaCDD
AHH (H)
ND
ND
ND
1.0 x 10"5
2.19 x 10"B
6.17 x 10"5
3.95 x 10~5
2.53 x 10~6
1.94 x 10"5
1.51 x 10"7
2.49 x 10"6
2.80 x 10"6
1.32 x 10~6
4.14 x 10"8
1.04 x 10~6
3.91 x 10"9
1.20 x 10"5
>10~*
2.70 x 10"5
1.60 x 10"9
7.90 x 10~9
8.60 x 10~8
3.25 x 10~7
3.77 x 10"8
2.09 x 10~7
2.54 x 10~9
1.06 x 10~7
2.56 x 10~10
3.56 x 10~10
1.47 x 10"9
4.24 x 10~8
6.87 x 10'10
7.2 x 10"11
1.1 x 10~8
6.1 x 10'8
-_
2.1 x 10"9
EROD (M)
ND
ND
ND
1.71 x 10"5
4.84 x 10"6
6.31 x 10"5
4.0 x 10"5
3.37 x 10"6
3.02 x 1C"5
2.48 x 10"7
1.56 x 10"6
3.13 x 10"6
1.13 x 10~6
3.76 x 10~8
7.79 x 10"7
2.02 x 10"9
9.26 x 10"5
>10^4
6.30 x 10"5
5.80 x 10~9
5.80 x 10~9
8.60 x 10"8
3.48 x 10~7
3.84 x 10~8
1.63 x 10~7
3.06 x 10"9
1.48 x 10"7
1.34 x 10"10
3.79 x 10~10
1.24 x 10"9
2.93 x 10"8
5.75 x 10"10
1.9 x 10"10
1.7 x 10~8
1.1 x 10"8
__
4.1 x 10"9
28
-------
8366H
Table 7. (continued)
Congeners
AHH (M)
EROD (M)
B. PCOOs (continued)
1.3.7,8-tetraCOO
1,2,4,7,8-pentaCDD
1.2.3.4-tetraCDD
2.3.7-triCOO
2,8-diCDO
1.2,3,4.7-pentaCDD
1.2.4-triCDO
OCDD
1-monoCDD
5.9 x 10
2.1 x 10
3.7 x 10
3.6 x 10
-8
-6
-7
>1.0 x 10
-4
-7
6.6 x 10
4.8 x 10"
>1.0 x 10
>1.0 x 10"
-4
3.2 x 10
1.1 x 10
2.4 x 10"
1.4 x 10
r7
-8
-7
>1.0 x 10
8.2 x 10
2.2 x 10
-7
-6
>1.0 x 10
>1.0 x 10
-4
-4
29
-------
AHH INDUCTION: in vitro vs. In vivo
PCDDs and PCDFs
11
10 -
9 -
e -
7-
R a 0.82
5678
•toy HMO AHH INDUCTION (BAT)
*«
!* -
- «
AHH INDUCTION vs. BODY WEIGHT LOSS
PCDDs and PCDFs
e
*>
e
3
11
10-
9 -
e
7H
6
Re 0.89
3 4 5 6 7 C
-log ED-50 BODY WEJGHT LOSS (RAT)
AHH INDUCTION vs. THYMIC ATROPHY
PCDDs and PCDFs
e
•S
I
s
ti
4567
•Jos EO-50 THYUIC ATROPHY (RAT)
30
c E o^T
^ ^* « Q> LL.
z «2 io
.
0> C- O
- ag§
Sf=> Sa
— LU Q. Sli
E
o
I
X — » «g
i c S
•r 5* 2.
*i°i
o&Iil
SxSic
£«2^§
*; OC u -5 «
<3<=!i<5E
UJ
DC
O
-------
AHH INDUCTION: In vitro vs. In vivo
PCBs, PCDDs, PCDFs, PBDDs
o
I
s
11
10 -
9-
8-
7-
6-
5-
4
R 00.84
56789
•log ED-50 AHH INDUCTION (RAT)
10
AHH
I
'>
c
5
INDUCTION vs. BODY WEIGHT LOSS
PCBs, PCDDs, PCDFs, PBDDs
34567
-log ED-50 BODY WBQHT LOSS (RAT)
AHH INDUCTION vs. THYMIC ATROPHY
PCBs, PCDDs, PCDFs, PBDDs
|
X
3:
S?
ti
= cc
in
UJ
DC
O
345678
-log ED-50 THYUIC ATROPHY (RAT)
31
-------
are not included in the Figures 4-7). It should also be noted that there
are good correlations between the in vitro AHH induction potencies for
PCDDs and PCDFs and their toxic potencies (body weight loss and immunotox-
icity) in other animals (guinea pigs and mice), as illustrated in
Figures 6 and 7. The excellent correlation between the in vivo and
in vitro potencies for PCDDs and PCDFs is consistent with the proposed
receptor-mediated mechanism of action of PCDDs, PCDFs, and related
compounds. Moreover, the correlation observed between the in vitro AHH
induction bioassay results and the in vivo toxicity data {Figures 4
through 7) confirms the utility of both in vivo and in vitro assay
systems for estimating the TEF for mixtures of PCDDs, PCDFs, and related
compounds. It should also be noted that other in vitro bioassays that
measure an Ah receptor-mediated response could also be used to estimate
TEFs for toxic halogenated aryl hydrocarbons.
5.4
Toxicity Equivalency Factors
Several groups have reported TEF values for the 2,3,7,8-substituted
PCDDs and PCDFs, and these have been summarized (Bellin and Barnes 1987,
NATO/CCMS 1988). Prior to the development of the I-TEF values by the
Pilot Study on International Information Exchange on Dioxins and Related
Compounds, numerous slightly-different TEF schemes were used throughout
the world. As a result of the existence of so many methods, the communi-
cation of the toxicological significance of a set of analytical data was
often hindered by the absence of a standardized approach. One of the
goals of the Pilot Study was to achieve consensus on a specific TEF meth-
odology. In addition to the addition of the most recent toxicological
data, the I-TEFs were developed with simplicity as one of the underlying
principles. Therefore, in developing the I-TEFs, order of magnitude
values were generally used rather than the precise numeric values.
Table 8 summarizes the I-TEF values for the 2,3,7,8-substituted PCDDs and
PCDFs and the ranges of TEF values that have been derived from the data
presented in Tables 3 through 7. The choice of a single TEF for an indi-
vidual PCDD or PCDF was based on several factors, including those noted by
Health and Welfare Canada and Ontario Ministry of the Environment. Two
32
-------
o
'
O
O
S
a
CB
O
11
10-
9-
4
8-
7-
2,3,7.8-TCDD
•
2.3,4,7,8-PeCDF ^
2.3,4.7,9-PeCDF
\
1.2.4,7,8-PeCDD
1,2,3,7,8-PeCDF
1,2,3,7,9-PeCDF
1,3.7,8-TCDD
R = 0.93
789
•log ED-50 BODY WEIGHT LOSS - GUINEA PIG
FIGURE 6. Correlation Between -log ECM (M) for AHH Induction in
Rat Hepatoma H-4-H E Cells vs. -log ED^ (moI/kg) Inhibition
of Bodyweight Gain (Guinea Pig)
33
-------
LU
s*
2,3.7.8-TCDD
2,3,4.7.8-TCDD -
56789
-tog ED-50 {mol/kg) IMMUNOTOXICJTY - C57BL/6 MICE
FIGURE 7. Correlation Between -log ECM (M) for AHH Induction in
Rat Hepatoma H-4-II E Cells vs. -log EDW (mol/kg) Immuno-
toxicity (C57BL/6 Mice)
34
-------
8366H
Table 8. International Toxicity Equivalency Factors (I-TEFs):* Comparison
of Relative Potency Data for the 2,3,7,8-Substituted PCDDs and PCOFs
Congener
2.3.7.8-tetraCDD
1,2,3,7,8-pentaCDD
1.2.3.4.7.8-hexaCOO
1.2.3.7.8.9-hexaCDDNTP
l,2,3,6,7,8-hexaCDDNTP
1,2,3,4,6,7,8-heptaCOO
OCDO
2.3.7,8-tetraCOf
2.3.4.7,8-pentaCDF
1.2.3,7.8-pentaCDF
1,2,3,4,7,8-hexaCOF
1,2,3,6,7.8-hexaCOF
Observed
I-TEF in vivo toxic it ies
1 1
0.5 0.59 - 0.053
(0.59m1. 0.42gl.
0.081r, 0.053r)
0.1 0.24 - 0.018
(0.24m1. 0.084r.
0.018gl. 0.13r)
0.1 0.14 - 0.016
(0.016g1. 0.14m1)
0.1 0.16 - 0.015
(0.16m1. 0.015g1)
O.Cll
0.001
0.1 0.17 - 0.016
(0.017gl. 0.17m1. 0.05mt.
0.025r. 0.016r)
O.S 0.8 - 0.048
(0.8m1, 0.479, 0.43r, 0.13g1
(0.12"*. 0.0481")
0.05 0.95 - 0.019
(0.95s. 0.05r.
0.031mt. 0.019r)
0.1 0.18 - 0.013
(0.18r. 0.0381",
0.013mt)
0.1 0.097 - 0.016
(0.097r. 0.016r)
TEF ranges
AHH
in vivo
1
0.13
(0.13r)
0.13
(0.13r)
-
0.0002C
0.006
(0.006r)
0.23 - 0.11
(0.239 - O.ll1")
0.047 - 0.003
(0.0479, 0.003r)
0.014
(0.014r)
0.012
(0.0121")
induction
in vitro
1
0.011 - 0.0065
(0. 011/0. 0065h)
0.046 - 0.034
(0. 034/0. 046h)
0.008
(0.008hB)
0.012
(0.012hB)
0.003
(0.003hB)
0.0006 ,
till'
(0.0006nB
0.09 - 0.018
(0. 018/0. 09h)
1.41 - 0.28
(0.28/1.41h)
0.06 - 0.028
(0.06/0,028h)
0.50 - 0.20
(O^O/O.SO*1)
0.153 - 0.04
(0. 049/0. 153h)
35
-------
8366H
Table 8. (continued)
Observed TEF ranges
Congener
I-TEF In vivq toxic Hies
AHH induction
in vivo
in vitro
1.2.3.7.8,9-hexaCDF 0.1
2,3.4.6.7,8-hexaCDF 0.1
1.2,3,4,6,7.8-heptaCDF 0.01
1,2.3,4,7.8.9-heptaCOF 0.01
OCDF 0.001
0.097 - 0.011
(0.097r. 0.018r. 0.0119)
0.015
(0.015r)
0.33 - 0.11
(0.11/0.33h)
^Guinea pig and rrat data (Table 4).
91
Guinea pig and mouse lethalities {Table 3).
Mouse teratogenicity and ""mouse imnunotoxicity (Table 5).
W hepatoma data (AHH/EROD) (Table 7).
hBRat hepatoma data (AHH) (Bradlaw and Casterline 1979).
°Couture et al. 1988.
MTP,
NTP (1980) carcinogenic potency values - 0.04.
Taken from NATO/CCNS 1988.
36
-------
conservative approaches would be (a) to utilize the highest observed TEF
for each compound or (b) to use the TEF values for in vitro AHH (or EROO)
induction since these values would minimize the species-dependent pharma-
cokinetic differences in congener potencies. Another option would utilize
a TEF range for each compound. These are some of the factors that were
considered in finalizing the I-TEF values given in Table 8.
One of the most noteworthy deviations from previous TEF methods is
the assignment of values to OCDD and OCDF. In previous schemes, these
congeners received values of zero on the basis of limited short-term in
vivo and in vitro data. In a recently published study (Couture et al.,
1988), however, male rats were exposed to low levels of OCDD for 13 weeks.
At the end of the experiment, the animals were beginning to show signs of
toxicity that were reminiscent of "dioxin toxicity." Detectable levels
of OCDD had accumulated in the organism. These data suggest that OCDD
exhibits minimal toxicity in short-term studies simply because so little
of the compound is absorbed in a short time. However, after multiple
exposures for longer periods, the animals appear to absorb and bio-
accumulate sufficient amounts of OCDD to manifest "dioxin-like" effects.
Based on these new data a value of 0.001 has been assigned to both OCDD
and OCDF in the I-TEF scheme.
Another modification made in the I-TEF scheme is the elimination of a
factor for non-2378 substituted dioxins and furans. In several past
methods, including the EPA method (Bellin and Barnes 1987), the non-2378-
substituted congeners were assigned a value of 1% of the value of the re-
spective 2378-substituted compound. During the past two years, scientists
have gathered additional data indicating that nearly all of the 210 PCDDs/
PCDFs can be found at very low levels in many parts of the environment.
However, it appears that the 2378-substituted congeners are selectively
absorbed and/or retained in higher animals; e.g., fish, humans, and other
mammals. That is, of the PCDDs/PCDFs detected in a variety of samples of
biological tissues (e.g., human, mammalian animals, and fish) the 2378-
PCDDs/PCDFs congeners clearly predominate over the non-2378-substituted
37
-------
congeners. This is true even when the source of the PCDDs/PCDFs is rela-
tively low in the concentration of 2378-substituted congeners.
For example, flyash from municipal waste combustors (MWCs) generally
contains detectable amounts of PCDDs/PCDFs. In most instances, the amount
of non-2378-substituted congeners vastly outweighs the amount of 2378-
substituted PCDDs and PCDFs in such samples. However, when mice or fish
are exposed to MWC flyash and their tissues are subsequently analyzed for
the presence of PCDDs/PCDFs, essentially only the 2378-substituted are
detected (Kuehl et al., 1986; Van den Berg et al., 1985). Similarly, the
"background levels" of CDDs/CDFs routinely found in human tissues (fat,
blood, and milk) contain almost exclusively the 2378-substituted compounds
(Rappe et al., 1987). The environmental concern rests primarily with
long-term exposures. Since the non-2378-substituted congeners appear to
be either not absorbed or quickly eliminated by biological systems, it is
the 2378-substituted congeners that seem to pose the greatest long-term
potential. .Therefore, in the interest of keeping the TEF system as simple
as possible, attention is focused exclusively on 2378-congeners in the
I-TEF scheme.
The I-TEF scheme introduces an additional complexity that was not a
part of the previous schemes. In the I-TEF scheme, the 2,3,4,7,8-PeCDF
is assigned a value of 0.5, while the 1,2,3,7,8-PeCDF is assigned a value
of 0.05. This is the only instance in which the I-TEFs depart from the
guiding principle of "simplicity" in which TEFs are expressed as rounded
orders of magnitude. This departure is prompted by a growing body of data
that indicate that 2,3,4,7,8-PeCDF is notably more active than originally
thought.
Based upon the data in Table 8, it can be seen that the 0.5 value for
2,3,4,7,8-PeCDF gains support from the in vivo thymic atrophy data
(0.43) and the mouse immunotoxicity data (0.8). The 0.05 value for
1,2,3,7,8-PeCDF gains support from the in vivo investigations of
thymic atrophy (0.05) and the in vivo and in vitro investigations
of enzyme induction (0.003-0.06). Note that there is one outlier in the
38
-------
eight data points reported for 1,2,3,7,8-PCDF in Table 8. Specifically,
there is a 0.95 value recorded for reduction in body weight gain seen in
guinea pigs. This one experiment in one laboratory should be investigated
further to determine its possible significance. At the present time, how-
ever, the weight of the evidence argues for the lower TEF.
The fact that the two 2378-substituted congeners can elicit such dif-
ferent biological responses can be rationalized by examining the stereo-
chemistry of the two chemicals (Bandiera et al., 1984). When superimposed
on the molecular structure of 2,3,7,8-TCDD, the C-4 of the "bent" PeCDF
is more stereochemically a "lateral position" (i.e., closer to C-3 on the
2,3,7,8-TCDD skeleton), while the C-l is even less stereochemically a
"lateral position" (i.e., farther away from C-2) (see Figure 8). There-
fore, the 2,3,4,7,8-PeCDF would theoretically be expected to be more
active than the 1,2,3,7,8-PeCDF since it has more chlorine substituents in
the lateral positions.
39
-------
Cl
Cl
1, 2, 3, 7, 8 - PeCDF on 2, 3, 7, 8 - TCDD
2, 3, 4, 7, 8 - PeCDF on 2, 3, 7, 8 - TCDD
Figure 8. 2378 - Substituted Pentachlorodibenzofurans
(Adapted from Bandlera et al., 1984)
40
-------
6.
BIOASSAVS FOR HAZARD ASSESSMENT OF PCDD AND PCDF MIXTURES
The previous sections of this report have confirmed that the SARs for
PCDDs and PCDFs are similar for most bioassays (in vivo and in vitro).
It is also evident that relative potencies for individual compounds can
be variable and are both assay- and species-dependent. This is not sur-
prising since there are undoubtedly species-dependent differences in the
pharmacokinetics of individual congeners and this affects the delivered
dose to the target organ/cells. Despite these variables, the correlations
between the in vitro and in vivo QSARs are remarkably consistent
(Figures 4 through 7). These findings provide further support for the
proposed Ah receptor mediated mechanism of action for 2,3,7,8-TCDD and
related compounds. Thus, in vitro systems which measure a specific Ah
receptor mediated response such as the in vitro AHH induction assay
(Bradlaw and Casterline, 1979) or the keratinization/flat cell assay
(Gierthy and Crane, 1985) should be useful for determining the TEF values
for complex mixtures.
Two recent studies by Safe and coworkers have utilized in vitro and
in vivo bioassays to estimate the relative toxicities of two extracts con-
taining complex irixtures of halogenated dibenzo-p-dioxins and dibenzo-
furans (Table 9) (Safe et al. 1987, Zacharewski et al. 1988). The dose-
response in vivo and in vitro effects of these extracts were directly
compared to results obtained for 2,3,7,8-TCDD. Using sample dilution
factors, one can determine the relative amount or concentration of analyte
required to elicit one-half maximal response (ECcQ or EDj-Q). This
amount/concentration is equivalent to the known amount/concentration of
2,3,7,8-TCDD required to cause this same magnitude of response. Using
this approach, one can readily calculate the "2,3,7,8-TCDD equivalents"
in an extract.
GC-MS analysis of the PCDDs and PCDFs in a fly ash extract from a
municipal incinerator indicated a total of 3,830 and 5,520 ng/g of these
compounds, respectively. Using the in vitro enzyme induction bioassay
procedure (i.e., rat hepatoma H-4-II E cells), the estimated "2,3,7,8-TCDD
41
-------
8366H
Table 9. Braninated Aromatic Flame Retardant Pyrolysis and Municipal Fly Ash Extracts:
In Vitro and In Vivo Determination of "2,3,7.8-TCDD Equivalents"3
Sample "2.3.7.8-TCDD equivalents" loam)
FireHasterb
Bioassay 300 BA
AHH induction 174
(in vitro)
EROD induction 194
(in vitro)
AHH induction
( in vivo)
EROD induction
(in vivo)
Body weight loss
(in vivo)
Thymic atrophy
(in vivo)
PBDDs plus PBDFs 10.935
(total ppm)c
PCODs plus PCOFs
(total. ppm}c
FireMasterb Bromka1b Bromkalb Bromkalb Fly ash
BP-6 70-5-OE 70-DE GI extract
1,400 2.140 8.780 3.920 0.105
480 4.680 6,740 5,260
540 - 5,200 - 0.075
520 - 3,860
760 - 6.260
1,680 - 8,960
2,070 610.390 268.480 547.700
9.35
a Safe et al. 1987. Zacharewski et al. 1988.
Brominated flame retardants (FireMaster BP-6 is a polybrom mated biphenyl and the remainder of
flame retardants are pnlybrominated diphenylethcrs).
c Determined by GC-MS analysis.
42
-------
equivalents" in this extract were 105 ng/g fly ash. The dose-response in
vivo induction of AHH in the rat was also determined, and the estimated
"2,3,7,8-TCDD equivalents" in this extract were 75 ng/g fly ash. This
confirms the correlation between the in vivo and in vitro bio-assays and
illustrates that the "2,3,7,8-TCDD equivalents" of these PCDD, and PCDF-
containing extracts are significantly lower than the total concentration
of these compounds in the sample extract as determined by GC-MS analysis.
Using a comparable approach, the "2,3,7,8-TCDD equivalents" (in vitro)
were determined for several extracts of pyrolyzed brominated (biphenyl
and diphenylether) flame retardants. The GC-MS analysis of the extracts
revealed a complex mixture of PBDDs and PBDFs for which very few analyti-
cal standards are available. The ranges of "2,3,7,8-TCDD equivalent"
levels (ug/g or ppm) derived from the AHH and EROD bioassays for each of
the pyrolyzed flame retardant samples were 174-194, 480-1,400, 2,140-
4,680, 6,740-8,780, and 3,920-5,260 ppm for FireMaster 300 BA, FireMaster
BP-6, Bromkal 70-5 DE, Bromkal 70-DE, and Bromkal Gl, respectively. The
in vivo dose response effects of two pyrolyzed flame retardant extracts
were determined in immature male Wistar rats and compared to the dose-
response activities of 2,3,7,8-TCDD. The in vivo responses measured
included hepatic microsomal AHH and EROD induction, body weight loss, and
thymic atrophy in the rat (Zacharewski et al. 1988). For the pyrolyzed
FireMaster BP-6 and Bromkal 70 DE samples, the in vivo "2,3,7,8-TCDD
equivalents" (ppm in sample) for the four in vivo bioassays were 520-
1,680 ppm and 3,860-8,960 ppm, respectively, and the in vitro "2,3,7,8-
TCDD equivalents" were 480-1,400 and 6,740-8,780 ppm, respectively. The
excellent overlap between the in vivo and in vitro "2,3,7,8-TCDD equiva-
lents" data for the two flame retardant pyrolysate extracts supports the
utility of the in vitro induction bioassay for quantitatively determining
the relative toxic potencies associated with mixtures containing toxic
halogenated aryl hydrocarbons.
43
-------
Eadon and coworkers (1986) have also summarized studies that compare
the calculated "2,3,7,8-TCDD equivalents" with various dose-related end-
points. Table 10 summarizes the "2,3,7,8-TCDD equivalents" determined in
the guinea pig for several responses using soot from the Binghamton office
building (De Caprio et al. 1983, 1986). The soot was contaminated with
PCDDs, PCDFs, and polychlorinated biphenylenes; the estimated TEF, based
on analytical chemical data, was 22 ppm (i.e., 2,3,7,8-TCDD equivalents).
The experimentally determined "2,3,7,8-TCDD equivalents" from the in vivo
studies in guinea pigs varied from 2 to 21 ppm. This observed range was
response-dependent and correlated well with previous studies in the rat
and with the TEF ranges summarized in Table 8.
These results clearly demonstrate that hazard assessments of toxic
mixtures of PCDDs and PCDFs and related compounds are readily determined
using both in vivo and in vitro bioassays. Invariably, for either indi-
vidual PCDD and PCDF congeners or mixtures of these compounds, multiple
bioassays will give a range of TEFs that exhibit overlap with TEFs esti-
mated by conventional analytical (chemical) approaches. The range of
TEFs obtained from multiple in vivo bioassays will also generally overlap
with TEFs derived from in vitro bioassays. This suggests that the short-
term in vitro bioassays are useful and validated methods for estimating
"2,3,7,8-TCDD equivalents" in complex mixtures of PCDDs, PCDFs, and
related compounds.
44
-------
8366H
Table 10. Calculated 2,3,7,8-TCDD-Equivalent Concentrations
of the Binghampton Soot for Various Dose-Related
Endpoints Following Subchronic Exposure3
Endpoint
Sex
Hethodc
2,3.7,8-TCDD-
equivalent
concentration in
soot (ppm)
Relative thymus
weight (decrease)
Linear 0.951
19"
Percent of initial body
weight (decrease)
Linear 0.994
Serum triglycerides
(increase)
Linear 0.998
Serum ALT
(decrease)
Log
0.960
1BU
Hepatocellular cyto-
plasmic inclusion bodies
Mortality
F
M
ED50
LD50
10 (4-28)9
2 (1-3)9
* Eadon ct al.. 1986.
Response data from the previously reported Binghamton soot (DeCaprio et al.
1983). and B.3,7,8 TCDD subchronic studies, (DeCaprio et al. 1986) were used
to calculate 2,3,7,8-TCDO-equivalent concentrations for the soot.
c Method by which 2,3,7.8-TCDD dose-response standard curve was constructed:
Linear = linear regression analysis of dose level (ppt) vs. response; LOG =
linear regression analysis of log dose level (ppt) vs. response; ED™ =
calculation of ED^g dose level by method of Carmines et al. (1980); LD^Q
for subchronic exposure by method of Carmines et al. (1980).
Correlation coefficient from linear regression analysis.
e Calculated using response data from 1.9-ppm Binghamton soot dose level.
Calculated using response data from 3.9-ppm Binghamton soot dose level.
^ 95 percent confidence limits.
45
-------
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Birnbaum LS, Harris MW, Crawford DO, and Morrissey RE. 1987b.
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