Environmental Toxicology and Chemistry, Vol. 31, No. 11, pp. 2545–2556, 2012
# 2012 SETAC
Printed in the USA
DOI: 10.1002/etc.1975
INCIDENCE OF JAW LESIONS AND ACTIVITY AND GENE EXPRESSION OF
HEPATIC P4501A ENZYMES IN MINK (MUSTELA VISON) EXPOSED TO DIETARY
2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN, 2,3,7,8-TETRACHLORODIBENZOFURAN,
AND 2,3,4,7,8-PENTACHLORODIBENZOFURAN
STEVEN J. BURSIAN,*yz JEREMY MOORE,y JOHN L. NEWSTED,§ JANE E. LINK,y SCOTT D. FITZGERALD,k#
NORA BELLO,yy VIRUNYA S. BHAT,zz DENISE KAY,§ XIAOWEI ZHANG,§§ STEVE WISEMAN,kk
ROBERT A. BUDINSKY,## JOHN P. GIESY,kk and MATTHEW J. ZWIERNIKy
yDepartment of Animal Science, Michigan State University, East Lansing, Michigan, USA
zCenter for Integrative Toxicology, Michigan State University, East Lansing Michigan, USA
§Cardno Entrix, Okemos, Michigan, USA
kDepartment of Pathobiology and Diagnostic Investigation, Michigan State University, East Lansing, Michigan, USA
#Diagnostic Center for Population and Animal Health, Michigan State University, Lansing, Michigan, USA
yyDepartment of Statistics, Kansas State University, Manhattan, Kansas, USA
zzToxicology Services Department, NSF International, Ann Arbor, Michigan, USA
§§State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, China
kkDepartment of Biomedical Veterinary Sciences and Toxicology Centre, University of Saskatchewan, Saskatoon, Canada
##Dow Chemical Company, Midland, Michigan, USA
(Submitted 9 March 2012; Returned for Revision 29 April 2012; Accepted 2 July 2012)
Abstract— This study assessed the effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), 2,3,4,7,8-pentachlorodibenzofuran
(PeCDF), and 2,3,7,8 tetrachlorodibenzofuran (TCDF) on the incidence of jaw lesions and on hepatic cytochrome P4501A (CYP1A)
endpoints in mink (Mustela vison). Adult female mink were assigned randomly to one of 13 dietary treatments (control and four
increasing doses of TCDD, PeCDF, or TCDF) and provided spiked feed for approximately 150 d (60 d prior to breeding through weaning
of offspring at 42 d post-parturition). Offspring were maintained on their respective diets for an additional 150 d. Activity of hepatic
CYP1A enzymes in adult and juvenile mink exposed to TCDD, PeCDF, or TCDD was generally greater compared with controls, but
changes in other CYP1A endpoints were less consistent. Histopathology of the mandible and maxilla of juvenile mink suggested a doserelated increase in the incidence of jaw lesions. The dietary effective doses (ED) for jaw lesions in 50% of the population (ED50) were
estimated to be 6.6, 14, and 149 ng/kg body weight (bw)/d for TCDD, PeCDF, and TCDF, respectively. The relative potencies of PeCDF
and TCDF compared with TCDD based on ED10, ED20, and ED50 values ranged from 0.5 to 1.9 and 0.04 to 0.09, respectively. These
values are within an order of magnitude of the World Health Organization toxic equivalency factor (TEFWHO) values of 0.3 and 0.1 for
PeCDF and TCDF, respectively. Environ. Toxicol. Chem. 2012;31:2545–2556. # 2012 SETAC
Keywords—Mink
Polychlorinated dibenzofuran
Biomarker
CYP1A
Jaw lesion
Several structurally analogous TCDD-like congeners exist
that can bind to the aryl hydrocarbon receptor (AhR) with
different affinities. The relative proportions of these congeners
change as a function of time due to weathering in the environment and selective accumulation and metabolism. Therefore, an
equivalency system has been developed to simplify the complexities of exposure to these compounds. In this equivalency
approach developed by the World Health Organization (WHO)
[14,15], the potency of a mixture of AhR-active congeners can
be expressed as an equivalent amount of TCDD, which was
considered at the time to be the most potent congener. 2,3,7,8Tetrachlorodibenzo-p-dioxin equivalents (TEQWHO) are calculated as the sum of the product of the concentration of each
congener multiplied by its respective consensus TCDD equivalency factor (TEFWHO), which is based on multiple endpoints.
Sampling of wild mink residing in the Tittabawassee River
basin demonstrated the presence of AhR-active PCDD and
PCDF congeners in their tissues. Using the TEFWHO approach,
the mean concentration of TEQWHO in livers of mink harvested
downstream of Midland Michigan, USA, which is the location
of the source of PCDDs and PCDFs entering the Tittabawassee
River, was 400 ng TEQWHO/kg, wet weight. In the upstream
INTRODUCTION
Studies indicating the presence of polychlorinated dibenzop-dioxins (PCDDs) and polychlorinated dibenzofurans
(PCDFs) in the vicinity of the Tittabawassee River, Michigan,
USA [1] have led to concerns regarding the potential health
risks they may pose to terrestrial and aquatic organisms. The
mink (Mustela vison) is a sentinel species of special interest
because it inhabits and consumes prey in both terrestrial and
aquatic habitats and thus has a great potential for exposure [2].
In addition, laboratory studies have shown that mink are
among the most sensitive species to the effects of 2,3,7,8tetrachlorodibenzo-p-dioxin (TCDD) and TCDD-like compounds, not only in terms of impaired reproduction and reduced
survival of offspring [3–6], but also in the development of a
unique jaw lesion characterized as mandibular and maxillary
squamous epithelial proliferation. This jaw lesion has been
reported in both laboratory [7–11] and field studies [12,13].
* To whom correspondence may be addressed
(bursian@msu.edu).
Published online 3 August 2012 in Wiley Online Library
(wileyonlinelibrary.com).
2545
2546
Environ. Toxicol. Chem. 31, 2012
locations, the mean TEQWHO concentration in liver was 20 ng
TEQWHO/kg, wet weight [16]. Of the total TEQWHO measured
in downstream mink, approximately 73% was contributed by
PCDFs and 5% was contributed by PCDDs. The individual
congeners contributing the greatest proportion of TEQWHO
were 2,3,4,7,8-pentachlorodibenzofuran (PeCDF) and 2,3,7,8tetrachlorodibenzofuran (TCDF).
Estimates of exposure based on measurements of PCDDs/
PCDFs in both the diet and tissues of mink were such that, based
on the current understanding of the toxicological potency of
these mixtures derived from laboratory studies, mink should be
experiencing adverse effects with hazard quotients (ratio of
estimated chemical intake [dose] and a reference dose below
which adverse health effects are unlikely) ranging from less
than 1.0 to 10 [10,11,17,18]. No pathology was reported,
however, in any of the 48 wild mink collected from within
the basin, and measures of the population, including abundance
and demographics, indicated that mink populations are stable
and at or close to carrying capacity for the Tittabawassee River
[19].
Reasons for the observed inconsistency between the apparently healthy population and hazard quotients greater than 1.0
are unclear. Possible explanations include (1) TEFWHO that are
too conservative, resulting in an overestimation of risk; (2) the
toxicity reference values used to estimate hazard quotients for
mink are inaccurate due to a lack of toxicological information
for the two dominant PCDF congeners, necessitating use of
toxicity reference values derived for other classes of chemicals
such as TCDD-like PCBs; and (3) pharmacokinetics and rates of
metabolism of TCDF and PeCDF were different compared with
those of TCDD or TCDD-like PCBs that have been studied in
mink [3,4,6,20].
Through a series of studies, several of the uncertainties
outlined above were addressed. Specifically, the pharmacokinetics of TCDF and PeCDF in mink were examined [20] and the
toxicological potency of PCDD and PCDF at the molecular and
organismal level in mink was evaluated [21,22]. Uncertainties
regarding the potencies of TCDF and PeCDF were also
addressed by conducting a dietary reproduction study, the
overall objective of which was to assess the effects of TCDD,
PeCDF, or TCDF on reproductive performance of adult female
mink and the survival and growth of their offspring. The effects
of TCDD, PeCDF, and TCDF on reproduction, survival, and
growth are reported in Moore et al. [23]. Additional objectives
of the reproduction study, which are the subject of the present
study, were to (1) determine the relationships between dietary
exposure to TCDD, PeCDF, or TCDF on mRNA transcription
and protein synthesis of CYP1A1 and CYP1A2, as well as
ethoxyresorufin-O-deethylase (EROD) and methoxyresorufin–
O-demethylase (MROD) activities in the liver, which are
sensitive, functional biomarkers of exposure, in adult mink
and their 27-week-old offspring; (2) determine the incidence
and severity of the jaw lesion in 27-week-old juvenile mink; and
(3) determine the relative potencies of PeCDF and TCDF to
TCDD based on CYP1A endpoints and jaw lesion responses.
MATERIALS AND METHODS
Chemicals and reagents
2,3,7,8-Tetrachlorodibenzo-p-dioxin, TCDF, and PeCDF
were obtained from AccuStandard and dissolved in hexane
(OmniSolv, EMD Chemicals) to produce a stock solution for
each congener. For enzyme analyses, 7–ethoxyresorufin was
obtained from Molecular Probes, and 7-methoxyresorufin and
S.J. Bursian et al.
resorufin were obtained from Sigma-Aldrich. All other biochemical reagents, including reduced nicotinamide adenine
dinucleotide phosphate, were obtained from Sigma-Aldrich
and were reagent grade or better, unless stated otherwise.
For the molecular analyses, the key reagents used included
the Agilent Total RNA Isolation Mini Kit (Agilent Technologies), Superscript III first-strand synthesis SuperMix (Invitrogen), and SYBR Green master mix (Applied Biosystems). For
immunoblots of the CYP1A1 and CYP1A2 proteins, protease
inhibitor cocktail and bicinchoninic acid reagent were purchased from Sigma-Aldrich. All electrophoresis reagents and
molecular weight markers were from Bio-Rad. b-Actin antibody (mouse anti b-actin monoclonal antibody) and
CYP1A1 þ 1A2 antibody (goat-anti CYP1A1 þ 1A2 polyclonal antibody) were from Abcam. The secondary antibodies to
b-actin and CYP1A1 þ 1A2 were alkaline phosphatase conjugated to either goat anti-mouse IgG for b-actin (Abcam) or
rabbit anti-goat for CYP1A1 þ 1A2 (Abcam). Nitroblue tetrazolium and 5-bromo-4-chloro-3-indolyl phosphate salt were
obtained from Fisher Scientific.
Dietary treatments
The study, which was approved by the Michigan State
University (MSU) Institutional Animal Care and Use Committee (Animal Use Form 12/06-140-00), was conducted at the
MSU Experimental Fur Farm. Housing of mink complied with
guidelines specified in the Standard Guidelines for the Operation of Mink Farms in the United States [24]. A standard
dietary mix was used throughout the study with the specific
ingredients, mix ratios, and nutritional data presented in Moore
et al. [23]. The base diet was used as the control diet, with
treatment diets differing only in the supplemental TCDD,
TCDF, or PeCDF added. The dosing regime was set to bracket
a wide range of concentrations designed to elicit responses in
one or more measurement endpoints. Lower doses bracketed
environmentally relevant concentrations and were expected to
result in no observable effect levels for all but the most sensitive
biological responses. The highest dose for each congener was
greater than the median predicted environmental exposure for
the Tittabawassee River of 3.85 ng TEQWHO/kg body weight
(bw)/d. Based on TEFWHO-normalized PCB and PCB-mixture
feeding studies [6,10,11,17,25–27], the highest dose of each
compound was expected to result in complete reproductive
failure. 2,3,7,8-Tetrachlorodibenzo-p-dioxin, TCDF, or PeCDF
were dissolved in hexane to produce stock solutions, and
aliquots of the stock were then diluted with approximately
100 ml corn oil. The corn oil solutions were then added to
the mink diet and mixed well prior to adding other feed
ingredients. The feed was mixed for approximately 30 min
and then stored in 3.8-L aluminum containers. Grab samples
were taken from each batch of feed and stored in I-Chem jars for
congener concentration determination. All samples were stored
in a freezer (208C) until analysis.
Animal husbandry
Adult female mink were housed individually in wire breeder
cages (76 cm L 46 cm W 38 cm H) suspended above the
ground in an open-sided mink shed. A wooden nest box
(38 cm L 25 cm W 29 cm H) bedded with aspen shavings
and excelsior (wood wool) was attached to the outside of each
cage. When the resulting offspring were 18 weeks of age, they
were moved to wire grower cages (61 cm L 25 cm W 38 cm
H) suspended above the ground in an open-sided mink
shed until the end of the study. A wooden nest box
Biomarkers in mink exposed to TCDD-like chemicals
Environ. Toxicol. Chem. 31, 2012
(25 cm L 20 cm W 29 cm H) bedded with aspen shavings
was suspended within each grower cage. Feed (125 g) was
placed on a grid on the top of the cage on a daily basis. Water
was available ad libitum.
2547
EROD and MROD quantification
Liver microsomes were prepared by homogenizing 0.5 g of
liver in Tris buffer (0.05 M Tris and 1.15% KCl, pH 7.5) and
centrifuging to obtain the microsomal fraction. The microsomal
pellet was resuspended in microsomal stabilization buffer (20%
glycerol, 0.1 M KH2PO4, 1 mM EDTA, and 1 mM dithiothreitol,
pH 6.25), and aliquots were stored at 808C. EthoxyresorufinO-deethylase and MROD activities were measured using a
modification of methods described by Kennedy and Jones
[28] and as detailed in Moore et al. [22]. The assays were
optimized and conducted in 96-well plates (Corning Costar),
where both microsomal cytochrome P450 activity and protein
concentration were measured simultaneously using a Fluoroscan Ascent microplate fluorometer (Thermo Fisher Scientific).
Ethoxyresorufin O-deethylase and MROD activities were determined kinetically from the linear range of the time-curves for
each well, and the results were expressed as pmol substrate
converted per minute, per millegram protein (pmol/min/mg).
Experimental design
A total of 117 first-year (virgin) and second-year (proven
breeder), natural dark, female mink were assigned randomly to
one of 13 treatment groups (Table 1). Nine animals per treatment group were assigned randomly to a bank of nine cages
separated from the next bank of nine cages by an empty cage.
Assigning treatments to banks of cages was done to prevent
cross-contamination between doses and compounds. Untreated,
natural dark, male mink were used for breeding purposes only.
Mink were fed the control feed for 10 d prior to being fed treated
diets. All surviving adults were euthanized (CO2) when kits
were 6 weeks old. The adult females were necropsied, and
samples of selected tissues were collected for analytical and
histological assessment. Nine to 29 kits from three to six litters
for each of the 13 treatment groups were maintained on their
respective diets until they were 27 weeks old, at which time they
were euthanized and processed as above.
Total RNA isolation and reverse transcription polymerase
chain reaction
Total RNA was extracted from the livers of individual mink
using the Agilent Total RNA Isolation Mini Kit according to the
manufacturer’s protocol. Purified RNA was stored at 808C
until analysis. First-strand cDNA synthesis and reverse transcription was performed as described in Zhang et al. [21]. The
cDNA synthesis reactions were stored at 208C until further
analysis.
Necropsy
Immediately following euthanasia, body weight and length
were recorded. A board-certified veterinary pathologist
assessed the overall physical condition of the animal as well
as gross appearance of organs. The thyroid gland, thymus, heart,
adrenal glands, kidneys, spleen, mesenteric lymph node, reproductive organs (uterus with ovaries/testes), liver, and brain were
removed, weighed, and placed in buffered formalin for subsequent histological assessment. A 2-g subsample of liver was
distributed between two cryovials (Corning Costar) that were
placed in liquid nitrogen for subsequent determination of microsomal EROD and MROD activities and two vials containing
RNAlater (Qiagen) for subsequent CYP1A1 and CYP1A2 gene
expression analysis. The flesh of each skull was removed, and
the skull was placed in formalin for subsequent histological
examination of the maxilla and mandible.
Real-time polymerase chain reaction
Quantitative real-time polymerase chain reactions (Q-RTPCRs) were performed using an ABI 7900 high-throughput
real-time PCR System in 384-well PCR plates (Applied Biosystems). Polymerase chain reaction primers used in these
reactions were: CYP1A1 sense CCT AAG CAC CTG GAA
AGC; CYP1A1 anti-sense CTA AGT GTC AGA GGG ATT
GG; CYP1A2 sense ACA GCA GRG AGC ACA GAT GG;
CYP1A2 anti-sense CCA GAG TAC CAG GCA GAA GAG; bactin sense GAT GTG GAT CAG CAA GCA GGA G; b-actin
anti-sense GCC AGC AGT CCG TTT AGA AGC [21]. Poly-
Table 1. Concentrations of TCDD, PeCDF, and TCDF in the diet and corresponding doses
Consumed dosea
Dietary concentration
Treatment
TCDD
PeCDF
TCDF
a
nb
ng/kg feed
ng TEQWHO/kg feedc
ng/kg bw/d
ng TEQWHO/kg bw/d
6
6
12
6
9
6
9
6
9
9
9
9
23 1.5
53 3.3
77 6.7
101 9.6
166 9.5
288 11.7
363 69.5
619 29.8
679 65.7
1464 106
2402 452
2866 490
23 1.5
53 3.3
77 6.7
101 9.6
50 2.9
86 3.5
109 20.8
186 8.9
68 6.6
146 10.6
240 45.2
287 49.0
2.1
4.6
6.0
8.4
13
25
30
49
52
116
214
246
2.1
4.6
6.0
8.4
4.0
7.6
9.0
15
5.2
12
21
25
d
Consumed dose based on estimated feed consumed from a daily allotment of 125 g feed and mean body weights (bw) of adult female mink through week 15 of
the study.
b
n refers to the number of dietary samples analyzed.
c
TEQWHO were determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
d
Data are presented as mean standard deviation.
TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin; PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran; TEQWHO ¼ TCDDtoxic equivalents based on World Health Organization (WHO) toxic equivalency factors.
2548
Environ. Toxicol. Chem. 31, 2012
merase chain reaction mixtures for 100 reactions contained
500 ml of SYBR Green master mix (Applied Biosystems), 2 ml
of 10 mM sense/anti-sense gene-specific primers, and 380 ml of
nuclease-free distilled water. A final reaction volume of 10 ml
was made up with 2 ml of diluted cDNA and 8 ml of PCR
reaction mixtures using a Biomek automation system (Beckman
Coulter). The PCR reaction mix was denatured at 958C for
10 min before the first PCR cycle. The thermal cycle profile was
as follows: denaturation for 15 s at 958C; annealing for 30 s at
608C; and extension for 30 s at 728C. A total of 45 PCR cycles
was used. Polymerase chain reaction efficiency, uniformity, and
linear dynamic range of each Q-RT-PCR assay were assessed
by the construction of standard curves using DNA standards.
The CYP1A1 and CYP1A2 mRNA expression levels in the
liver of mink were measured by semiquantitative RT-PCR using
b-actin as the reference gene. Polymerase chain reaction efficiency and linear dynamic range of each Q-RT-PCR assay was
assessed by construction curves using DNA standard [21].
Western blot analysis
Samples of microsomes isolated from mink livers were
prepared in the following manner for use in Western analyses
of CYP1A proteins: The concentrations of protein in the microsomes were determined by the bicinchoninic acid protein assay
method using bovine serum albumin as the standard according
to the manufacturer’s instructions (Sigma-Aldrich). Samples
were adjusted to a final concentration of 2 mg/ml, and 40 mg
of protein was separated on 8% polyacrylamide gels using the
discontinuous buffer system of Laemmli [21]. Proteins were
transferred (20 V for 20 min) onto a 0.45-mM nitrocellulose
membrane (Bio-Rad) using Trans-blot SD semidry electrophoretic transfer cell (Bio-Rad). A 5% solution of nonfat dry milk in
1 TTBS (2 mM Tris, 30 mM NaCl, 0.01% Tween 20, pH 7.5)
was used as a blocking agent (1 h at room temperature) and for
diluting antibodies. The blots were incubated with primary
antibodies for 1 h at room temperature, followed by 1-h incubation with an alkaline phosphatase–conjugated secondary
antibody. The membranes were washed after incubation in
either primary (2 15 min washes in TTBS [2 mM Tris,
30 mM NaCl, pH 7.5, 0.1% Tween-50]) or secondary antibodies
(2 15 min in TTBS followed by 1 10 min in TBS). The
antibody used in the present study was raised against CYP1A1
and CYP1A2 in rats and was not raised specifically against
CYP1A1 and CYP1A2 in mink. A preliminary analysis using
microsomes isolated from livers of rats and mink showed that
the banding patterns in the two samples were identical. In the
western blots, there were two clearly distinguishable bands
representing CYP1A1 and CYP1A2. Proteins of interest were
detected using 5-bromo-4-chloro-3-indolyl-phosphate/nitro blue
tetrazolium color substrate (Bio-Rad). Images were captured on
a VersaDoc Imaging system (Bio-Rad). The protein bands were
quantified using Quantity One Software (Bio-Rad).
Quantification of chemicals
To ensure that co-contaminants were not a factor in the
study, the concentrations of 17 2,3,7,8-substituted PCDF and
PCDD congeners and 12 PCB congeners were measured in the
dietary items, feed samples, and liver tissue as described in
Zwiernik et al. [20]. Concentrations of TEQs were calculated as
the sum of the products of the concentrations of congeners
multiplied by their respective TEFWHO [17]. A surrogate value
of one-half the method detection limit (MDL) was used for
concentrations less than the MDL. Liver tissues were extracted
S.J. Bursian et al.
following a modification of U.S. Environmental Protection
Agency (U.S. EPA) Method 1613B.
Histological analysis
Histological examination of tissues was performed at MSU’s
Diagnostic Center for Population and Animal Health. A boardcertified veterinary pathologist examined slides of the maxilla
and mandible of each mink sampled at necropsy for squamous
epithelial proliferation, which was scored using a modification
of the scoring system described in [12] where 0 ¼ normal
morphology, no squamous epithelial cell proliferation; 1 ¼
mild lesion, focal (one or two) squamous epithelial cysts
invading jaw bone, restricted to a single region of the dental
arcade (molar, premolar, incisor); 2 ¼ moderate lesion, multiple
squamous epithelial cysts (three or more), adjacent to several
teeth or in two or more regions of the dental arcade; and
3 ¼ severe lesion, multiple sites of squamous epithelial proliferation that are larger in size or coalescing (Fig. 1). Histological
results for other tissues sampled in the study are presented in
Moore et al. [23].
Statistical analysis
Dietary concentrations of compounds are presented in
Table 1 as means standard deviations for the number of times
the particular diet was mixed and analyzed. Due to the unbalanced experimental design (unequal sample sizes), least square
mean estimates and their corresponding estimated standard
errors are reported for the various endpoint analyses. Statistical
models for adult female liver and adipose TEQWHO concentrations, and for hepatic cytochrome P4501A endpoints,
included the fixed effects of treatment and dose within treatment; for juveniles, the statistical models included the fixed
effects of treatment, dose within treatment, and sex, as well as
the random effect of litter. When treatment effects were statistically significant, differences among treatment groups were
tested with the Tukey–Kramer test to prevent inflation of type
I error rate while accounting for differences in sample size
among the treatments. Differences among treatments were
considered significant at p < 0.05. The MIXED procedure of
SAS (ver 9.2, SAS Institute) was used for these analyses.
The presence or absence of jaw lesions in the 27-week-old
kits was analyzed using a generalized linear mixed model fitted
to the binary outcome using a logit link function. The model
included the fixed effects of chemical, dose (in TEQWHO), and
their two-way interaction in order to evaluate chemical-specific
susceptibility in mink. World Health Organization TEFs of 1.0,
0.3, and 0.1 were used to scale the doses of TCDD, PeCDF, and
TCDF, respectively, to TEQWHO. Also included in the statistical
models was the random effect of litter nested within treatment to
account for technical replication in the design and to recognize
litter as the experimental unit for treatment. Differences
between treatment groups were considered significant at
p < 0.05. The GLIMMIX procedure of SAS was used for this
analysis.
United States Environmental Protection Agency benchmark
dose (BMD) software (ver 2.2; http://www.epa.gov/ncea/bmds/
index.html) was used to determine effect doses or concentrations (as different percentages of the population; ED10/EC10,
ED20/EC20, and ED50/EC50) for jaw lesion incidence in
27-week-old juvenile mink, with exposure metrics being average daily dose or liver concentrations (when appropriate) that
were expressed as analytical concentration or TEQWHO. The
best-fit model for each compound was chosen collectively considering the following criteria: the highest p value, chi-squared
Biomarkers in mink exposed to TCDD-like chemicals
Environ. Toxicol. Chem. 31, 2012
2549
Fig. 1. View of a normal mink jaw (A). Cysts (cy) of squamous epithelial proliferation within the jaw adjacent to teeth (T), scored as mild (B); moderate (C); or
severe (D). (50 magnification).
residual values less than 2, the lowest Akaike’s Information
Criterion value, the smallest margin between the benchmark
dose (BMD) and benchmark dose lower 95% confidence limit
(BMDL) curves, and a good visual fit of the BMD and BMDL
curves. Prior to calculation of EC values, the hepatic concentrations of TCDD, PeCDF, or TCDF as a function of dietary
dose were modeled using the U.S. EPA BMD software. For
congeners having an adequate dose–response relationship based
on measures of statistical tests of fit for continuous models (chisquared parameters and likelihood ratios), EC values were then
determined.
RESULTS
Exposure and tissue concentrations
Dietary concentrations and consumed doses of TCDD,
TCDF, or PeCDF expressed as ng chemical and ng TEQWHO
are presented in Table 1. Calculations of dose were based on the
average estimated feed consumption and body mass of adult
females in each treatment group through the first 15 weeks of
the trial [23]. Concentrations of TEQWHO contributed by
TCDD, PeCDF, and TCDF in livers and adipose tissue of adult
and juvenile mink are presented in Tables 2 and 3, respectively.
In general, TEQ concentrations increased in a dose-related
manner in liver and adipose tissue of adult and juvenile mink
exposed to TCDD or PeCDF. Hepatic TEQ concentrations in
adult females exposed to TCDF were significantly greater
compared with controls, but there was no evidence that concentrations increased with increasing dose. In contrast, hepatic
TEQ concentrations of juveniles exposed to TCDF increased in
a dose-related manner. Adipose TEQ concentrations in adults
and juveniles exposed to TCDF were significantly greater
compared with controls only at the greatest dose (246 ng
TCDF/kg bw/d or 25 TEQWHO-TCDF/kg bw/d.
EROD and MROD activity
Hepatic EROD and MROD activities in adult mink (Table 4)
exposed to dietary TCDD or PeCDF via the diet were greater
than activities in nonexposed controls, but, in general, there was
no evidence for a dose-dependent increase in enzyme activities
in any of the treatment groups with the exception of MROD
activity in females exposed to TCDD. Activity of MROD was
significantly greater in animals exposed to 6.0 and 8.4 ng
TCDD/kg bw/d (6.0 and 8.4 ng TEQWHO-TCDD/kg bw/d) compared with those exposed to 2.1 and 4.6 ng TCDD/kg bw/d (2.1
and 4.6 ng TEQWHO-TCDD/kg bw/d). In adult females exposed
to TCDF, EROD activity was significantly greater compared
with control activity only at 214 ng TCDF/kg bw/d (21 ng
TEQWHO-TCDF/kg bw/d), and MROD activity was significantly
greater compared with control activity at 52 and 214 ng TCDF/
kg bw/d (5.2 and 21 ng TEQWHO-TCDF/kg bw/d) but not at
116 and 246 ng TCDF/kg bw/d (12 and 25 ng TEQWHO-TCDF/kg
bw/d). In 27-week-old juvenile mink (Table 5), EROD activity
was significantly greater compared with control activity at all
doses of TCDD, PeCDF, and TCDF, but there was no evidence
of a dose–response relationship. Activity of MROD was significantly greater compared with control activity at all doses
of TCDD (no evidence of a dose–response relationship) and
at all doses of PeCDF except 30 ng PeCDF/kg bw/d (9.0 ng
TEQWHO-PeCDF)/kg bw/d. The activity of MROD was also
significantly greater compared with control activity for all
doses of TCDF, with activity at 214 ng TCDF/kg bw/d
(21 ng TEQWHO-TCDF/kg bw/d) being greater than activity at
52 ng TCDF/kg bw/d (5.2 ng TEQWHO-TCDF/kg bw/d).
Hepatic CYP1A mRNA expression
Exposure of adult female mink to PeCDF resulted in a
significant increase in hepatic CYP1A1 and CYP1A2 mRNA
2550
Environ. Toxicol. Chem. 31, 2012
S.J. Bursian et al.
Table 2. Concentrations of TEQWHO in liver of adult and 27-week-old juvenile mink exposed to TCDD, PeCDF, or TCDF via the dieta
Liver (ng TEQWHO/kg, wet wt)b
Treatment
Control
TCDD
PeCDF
TCDF
Consumed dose
(ng TEQWHO/kg bw/d)
n
Adults
SE
n
Juveniles
SE
0.0
2.1
4.6
6.0
8.4
4.0
7.6
9.0
15
5.2
12
21
25
4
4
3
3
3
4
3
3
2
3
4
3
4
1.08 A
59.2 B
225 BC
224 BC
331 C
546 B
901 BC
1207 C
2755 D
4.55 B
4.62 B
8.29 B
12.7 B
69.9
34.4
39.7
39.7
39.7
70.5
81.4
81.4
100
1.85
1.61
1.85
1.61
10
7
8
7
6
6
6
3
4
6
6
6
6
1.35 A
62.0 B
162 C
255 D
340 E
501 B
934 BC
1388 C
2464 D
4.30 B
16.3 C
22.9 D
23.7 D
79.5
16.8
15.1
14.4
14.0
97.3
97.3
113
113
1.18
1.18
1.18
1.18
a
TEQWHO determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
Data are presented as least squares mean and standard error (SE). Adult female mink and juvenile mink were sampled approximately 9 and 27 weeks
postpartum, respectively. Means within the same column with different uppercase letters are significantly different at p < 0.05.
TEQWHO ¼ TCDD-toxic equivalents based on World Health Organization (WHO) toxic equivalency factors; TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin;
PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran.
b
Table 3. Concentrations of TEQWHO in adipose tissue of adult and 27-week-old juvenile mink exposed to TCDD, PeCDF, or TCDF via the dieta
Adipose (ng TEQWHO/kg, wet wt)b
Treatment
Control
TCDD
PeCDF
TCDF
Consumed dose
(ng TEQWHO/kg bw/d)
n
Adults
SE
n
Juveniles
SE
0.0
2.1
4.6
6.0
8.4
4.0
7.6
9.0
15
5.2
12
21
25
6
7
7
7
5
6
5
7
7
3
3
4
3
2.03 A
344 B
540 B
969 C
1,418 D
394 B
511 BC
719 CD
902 D
28.9 A
31.8 A
56.8 AB
79.1 B
52.5
53.2
53.2
53.2
62.9
63.3
69.4
58.6
58.6
3.8
7.8
7.8
6.7
7
3
4
3
3
6
5
3
3
3
3
3
3
1.16 A
376 B
836 C
1,105 CD
1,560 D
290 B
461 C
575 C
871 D
32.3 A
54.4 A
63.3 AB
75.6 B
57.8
96.0
95.6
113
113
45.6
49.8
64.1
64.1
2.99
2.99
2.99
2.99
a
TEQWHO were determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
Data are presented as least squares mean and standard error (SE). Adult female mink and juvenile mink were sampled approximately 9 and 27 weeks
postpartum, respectively. Means within the same column with different uppercase letters are significantly different at p < 0.05.
TEQWHO ¼ TCDD-toxic equivalents based on World Health Organization (WHO) toxic equivalency factors; TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin;
PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran.
b
Table 4. Hepatic EROD and MROD activities in adult female mink exposed to TCDD, PeCDF, or TCDF via the diet
Treatment
Control
TCDD
PeCDF
TCDF
a
Consumed dosea
(ng TEQWHO/kg bw/d)
n
2.1
4.6
6.0
8.4
4.0
7.6
9.0
15
5.2
12
21
25
4
4
3
3
3
4
3
3
2
3
4
3
4
ERODb
(pmol/min/mg)
36.6 A
138 AB
209 B
195 B
236 B
298 B
317 B
252 B
320 B
134 A
106 A
298 B
142 A
SE
MRODb
(pmol/min/mg)
SE
40.7
40.7
47.0
47.0
47.0
40.7
47.0
47.0
57.6
47.0
40.7
47.0
40.7
6.53 A
21.4 B
18.4 AB
34.4 C
38.2 C
32.2 B
30.2 B
30.7 B
29.4 B
22.9 B
18.2 AB
30.3 B
17.6 AB
4.28
4.28
4.94
4.94
4.94
4.28
4.94
4.94
6.05
4.94
4.28
4.94
4.28
TEQWHO were determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
Data are presented as least square mean and standard error (SE). Means within the same column with different letters are significantly different from one
another ( p < 0.05). Adult female mink were sampled at approximately 9 weeks postpartum.
EROD ¼ ethoxyresorufin-O-deethylase; MROD ¼ methoxyresorufin-O-demethylase; TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin; PeCDF ¼ 2,3,4,7,8pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran; TEQWHO ¼ TCDD-toxic equivalents based on World Health Organization (WHO) toxic
equivalency factors.
b
Biomarkers in mink exposed to TCDD-like chemicals
Environ. Toxicol. Chem. 31, 2012
2551
Table 5. Hepatic EROD and MROD activities in 27-week-old juvenile mink exposed to TCDD, PeCDF, or TCDF via the diet
Treatment
Consumed dosea
(ng TEQWHO/kg bw/d)
n
ERODb
(pmol/min/mg)
SE
2.1
4.6
6.0
8.4
4.0
7.6
9.0
15
5.2
12
21
25
10
7
8
7
6
6
6
3
3
6
6
6
6
75.3 A
177 B
248 B
220 B
166 B
213 B
214 B
199 B
226 B
179 B
226 B
254 B
211 B
30.5
28.2
30.5
28.2
30.5
30.5
30.5
43.1
37.3
30.5
30.5
30.5
33.4
Control
TCDD
PeCDF
TCDF
MRODb
(pmol/min/mg)
10.3
30.4
31.5
28.3
29.7
39.8
27.2
23.8
28.0
24.9
27.3
39.6
26.6
A
B
B
B
B
B
B
AB
B
B
BC
C
BC
SE
4.76
4.40
4.76
4.40
4.76
4.76
4.76
6.73
5.82
4.76
4.76
4.76
5.21
a
TEQWHO were determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
Data presented as least squares mean and standard error (SE). Means within the same column with different letters are significantly different from one another
( p < 0.05). Juvenile mink sampled at approximately 27 weeks post partum.
EROD ¼ ethoxyresorufin-O-deethylase; MROD ¼ methoxyresorufin-O-demethylase; TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin; PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran; TEQWHO ¼ TCDD-toxic equivalents based on World Health Organization (WHO) toxic
equivalency factors.
b
expression compared with controls at 13 and 30 ng PeCDF/kg
bw/d (4.0 and 9.0 ng TEQWHO-PeCDF/kg bw/d), but not at 25 and
49 ng PeCDF/kg bw/d (7.6 and 15 ng TEQWHO-PeCDF /kg bw/d).
CYP1A1 and CYP1A2 mRNA levels in TCDD- and TCDFtreated mink were not significantly different from those in
control mink (Table 6). For juvenile mink, expression of
CYP1A1 and CYP1A2 mRNA was significantly greater compared with controls in animals exposed to TCDD or PeCDF, but
not TCDF (Table 7). In animals exposed to TCDD, expression
of both CYP1A1 and CYP1A2 was greater compared with
controls at doses greater than 2.1 ng TCDD/kg bw/d (2.1 ng
TEQWHO-TCDD/kg bw/d). In animals exposed to PeCDF,
expression of CYP1A1 and CYP1A2 mRNA was significantly
greater compared with controls at doses of 13 and 30 ng PeCDF/
kg bw/d (4.0 and 9.0 ng TEQWHO-PeCDF/kg bw/d), but not at 25
and 49 ng PeCDF/kg bw/d (7.6 and 15 ng TEQWHO-PeCDF/kg
bw/d).
CYP1A protein in liver
No evidence was found for any consistent effect of exposure
to TCDD, PeCDF, or TCDF on concentrations of CYP1A
protein (Tables 6 and 7). In adult female mink treated with
TCDD, only the 4.6- and 8.4-ng TCDD/kg bw/d (4.6 and 8.4 ng
TEQWHO-TCDD /kg bw/d) groups had CYP1A protein concentrations that were greater than controls. For PeCDF, mink dosed
with the greatest concentration (49 ng PeCDF/kg bw/d or 15 ng
TEQWHO-PeCDF/kg bw/d) had significantly greater protein concentrations compared with controls and with animals in the
13-ng PeCDF/kg bw/d (4.0 ng TEQWHO-PeCDF/kg bw/d) group.
Exposure to TCDF had no significant effect on CYP1A1 protein
concentrations in the liver. In juvenile mink, there was no
evidence that exposure to TCDD or TCDF had an effect on
CYP1A protein concentrations. Exposure to the greatest dose
of PeCDF (49 ng PeCDF/kg bw/d or 15 ng TEQWHO-PeCDF/kg
Table 6. Hepatic CYP1A endpoints in adult female mink exposed to TCDD, PeCDF, or TCDF via the diet
Treatment
Control
TCDD
PeCDF
TCDF
a
Consumed dosea
(ng TEQWHO/kg bw/d)
2.1
4.6
6.0
8.4
4.0
7.6
9.0
15
5.2
12
21
25
n
CYP1A1 mRNA expressionb
(fold-change)
4
4
3
3
3
4
3
3
2
3
4
3
4
2.02 A
8.09 A
3.92 A
11.7 A
15.0 A
16.8 B
11.6 AB
16.1 B
6.89 AB
5.62 A
5.10 A
14.0 A
4.16 A
SE
CYP1A2 mRNA expressionb
(fold-change)
3.75
3.25
4.59
3.75
3.75
3.75
3.75
3.75
4.59
4.59
3.75
3.75
3.75
3.78 A
8.32 A
4.83 A
14.3 A
25.8 A
20.1 B
16.4 AB
23.0 B
11.7 AB
8.07 A
9.06 A
22.3 A
7.56 A
SE
CYP1A proteinb
(fold-change)
SE
5.38
4.66
6.59
5.38
5.38
5.38
5.38
5.38
6.59
9.31
5.38
5.38
5.38
0.22 A
0.60 A
4.42 B
0.75 A
4.46 B
0.91 A
2.22 AB
2.09 AB
3.60 B
0.61 A
0.87 A
1.02 A
0.45 A
0.66
0.66
0.93
0.76
0.76
0.66
0.76
0.76
0.93
0.93
0.76
0.76
0.66
TEQWHO were determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
Fold change in gene expression was calculated by the delta cycle threshold (Ct) method where CYP1A1 and CYP1A2 levels were first normalized to the b-actin
Ct values and then all samples were normalized to the mean of the biological control samples. Data are presented as least square mean and standard error (SE).
Means within the same column with different letters are significantly different from one another ( p < 0.05). Adult female mink were sampled at approximately
9 weeks postpartum.
CYP1A ¼ cytochrome P4501A; TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin; PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran; TEQWHO ¼ TCDD-toxic equivalents based on World Health Organization (WHO) toxic equivalency factors.
b
2552
Environ. Toxicol. Chem. 31, 2012
S.J. Bursian et al.
Table 7. Hepatic CYP1A endpoints in 27-week-old juvenile mink exposed to TCDD, PeCDF, or TCDF via the diet
Treatment
Control
TCDD
PeCDF
TCDF
Consumed dosea
(ng TEQWHO/kg bw/d)
n
2.1
4.6
6.0
8.4
4.0
7.6
9.0
15
5.2
12
21
25
10
7
8
7
6
6
6
3
3
6
6
6
6
CYP1A1 mRNA expressionb
(fold-change)
1.30
3.05
5.27
6.51
6.45
4.01
4.71
4.25
6.44
1.74
2.56
3.06
2.65
A
A
B
B
B
B
BC
BC
C
A
A
A
A
SE
0.60
0.75
0.66
0.68
0.58
0.68
0.68
0.82
0.80
0.62
0.69
0.62
0.58
CYP1A2 mRNA expressionb
(fold-change)
1.17
2.41
4.28
5.15
4.70
2.54
4.00
3.23
5.61
1.69
1.73
1.99
2.24
A
A
B
B
B
AB
BC
B
C
A
A
A
A
SE
CYP1A proteinb
(fold-change)
SE
0.45
0.55
0.48
0.53
0.49
0.53
0.53
0.70
0.64
0.51
0.54
0.51
0.49
0.040 A
0.78 A
1.18 A
1.07 A
1.15 A
1.23 AB
1.49 AB
1.46 AB
1.75 B
1.00 A
1.55 A
1.02 A
0.51 A
0.58
0.33
0.47
0.37
0.47
0.47
0.33
0.82
0.41
0.37
0.47
0.47
0.41
a
TEQWHO were determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
Fold change in gene expression was calculated by the delta cycle threshold (Ct) method where CYP1A1 and CYP1A2 levels were first normalized to the b-actin
Ct values and then all samples were normalized to the mean of the biological control samples. Data presented as least squares mean and standard error (SE).
Means within the same column with different letters are significantly different from one another ( p < 0.05). Juvenile mink sampled at approximately 27 weeks
post partum.
CYP1A ¼ cytochrome P4501A; TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin; PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran; TEQWHO ¼ TCDD-toxic equivalents based on World Health Organization (WHO) toxic equivalency factors.
b
bw/d) resulted in a significant increase in hepatic CYP1A
protein concentrations in juvenile mink compared with
controls.
Jaw lesions
Histopathological assessment of the mandible and maxilla of
27-week-old juvenile mink indicated the presence of jaw
lesions in mink exposed to TCDD, TCDF, or PeCDF. Lesion
severity was predominantly mild to moderate (lesion scores of 1
and 2, respectively). One animal in the 49-ng PeCDF/kg bw/d
(15 ng TEQWHO-PeCDF/kg bw/d) group had a lesion score of 3
(severe) (Fig. 2). A significant p value ( p ¼ 0.0047) for the
interaction between chemical treatments and dose on the prob-
ability of jaw lesion indicates that the toxicity response in terms
of TEQ doses is specific to each chemical (Fig. 3). The effective
consumed doses resulting in 10, 20, and 50% of the animals
having an increased incidence of the lesion over background
expressed as estimated ED10, ED20, and ED50 values are
presented in Table 8. Assessment of dose–response relationships between consumed dose and hepatic concentrations of
TCDD, PeCDF, or TCDF indicated that concentrations of
TCDD and PeCDF in the liver could be used as a reliable
internal dose metric for jaw lesions (as a surrogate for external
consumed dose), but hepatic TCDF concentration could not
because there was no evidence of a dose–response relationship.
Thus, EC10, EC20, and EC50 values for TCDD and PeCDF
only are presented in Table 8.
Fig. 2. Observed incidence and severity of maxillary and mandibular squamous epithelial proliferation in 27-week-old juvenile mink treated with TCDD,
PeCDF, or TCDF (ng TEQWHO/kg bw/d). Total number of animals per treatment: n ¼ 16 for control; n ¼ 10 for PeCDF 4.0, TCDF 12 and 25; n ¼ 9 for TCDD
2.1, 4.6, 6.0, 8.4, PeCDF 9.0, 15, TCDF 5.2 and 21; and n ¼ 7 for PeCDF 12 ng TEQWHO/kg bw/d. TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin;
PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran; TEQWHO ¼ TCDD-toxic equivalents based on World Health
Organization (WHO) toxic equivalency factors.
Biomarkers in mink exposed to TCDD-like chemicals
Environ. Toxicol. Chem. 31, 2012
Fig. 3. Incidence of jaw lesion by dose. Because the dose is expressed in
TEQWHO, a significant p value for the interaction between dose and chemical
treatments ( p ¼ 0.0047) indicates evidence for heterogeneous slopes in
the toxicity response to dose for each chemical. TCDD ¼ 2,3,7,8tetrachlorodibenzo-p-dioxin; PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran;
TCDF ¼ 2,3,7,8-tetrachlorodibenzofuran; TEQWHO ¼ TCDD-toxic equivalents
based on World Health Organization (WHO) toxic equivalency factors.
DISCUSSION
Cytochrome P450
Given the ability of AhR agonists such as TCDD, TCDF, or
PeCDF to upregulate CYP1A genes [29,30], three different
measures of CYP1A expression were evaluated in the current
study. These included mRNA transcription levels, concentrations of catalyst proteins, and enzymatic activities. The purpose
2553
of evaluating these endpoints was to further characterize their
utility as functional indicators of exposure to TCDD-like compounds in mink [30]. In the current study, only EROD and
MROD activities were increased by exposure to each of the
three chemicals in both adult and juvenile mink, and only
exposure to PeCDF resulted in an increase in all three endpoints
in adults and juveniles, although there was no evidence that the
increases were related to dose. The absence of a consistent
effect of exposure to TCDD, PeCDF, or TCDF on CYP1A
metrics may have been due to several factors including metabolism or sequestration, or both, of these compounds in the liver,
thus altering their biological activity at the site of action. The
experimental design also could have influenced effects on
CYP1A metrics. For example, in the present study, the time
from the start of exposure until collection of tissues was
relatively long in that both adult and juvenile mink were
exposed for longer than 150 d. This might have allowed
concentrations of the test chemicals in the liver to reach a
steady state at which CYP1A expression was near or at maxima.
In a dietary pharmacokinetic study [20], a time-dependent
accumulation of TCDF and PeCDF in livers of female mink
was reported with concentrations approaching an apparent
steady state by 90 d for both compounds. Furthermore, although
the time to steady state was similar for both TCDF and PeCDF,
tissue concentration data and pharmacokinetic model parameters showed that there were significant differences in accumulation of TCDF and PeCDF in the liver of exposed mink that
presumably involved induction of CYP1A gene expression. For
instance, the rapid clearance of TCDF from the liver (half-life
less than 15 h) is related to the induction of CYP1A1 enzymatic
activity, whereas for PeCDF, the selective uptake and sequestration into the liver is related to induction and binding to
CYP1A2 protein [31,32]. As a result, exposure to these compounds not only results in the induction of CYP1A genes, but
also influences the accumulation of these compounds in the
liver, which alters the dose–response relationships in a concen-
Table 8. ED10/EC10, ED20/EC20, and ED50/EC50 values for incidence of jaw lesions in juvenile mink based on consumed dose or liver concentrations (as an
internal dose metric) of TCDD, PeCDF, or TCDF and corresponding TEQWHOa
Consumed dose expressed as
TEQWHOb (ng TEQWHO/kg bw/d)
Consumed dose (ng/kg bw/d)
Treatment
TCDD
PeCDF
TCDF
ED10
ED20
ED50
ED10
ED20
ED50
Best-fit model
p value
4.0
2.1 (1.9)c
45 (0.09)
4.7
4.4 (1.1)
73 (0.06)
6.6
14 (0.5)
149 (0.04)
4.0
0.63
4.5
4.7
1.3
7.3
6.6
4.1
15
Log probit
Gamma
Weibull
0.7196
0.5950
0.9881
Liver concentration (ng/kg, wet wt)
Treatment
TCDD
PeCDF
TCDFd
a
Liver concentration expressed as
TEQWHO (ng TEQWHO/kg, wet wt)
EC10
EC20
EC50
EC10
EC20
EC50
Best-fit model
p value
149
299 (0.50)
–
181
634 (0.29)
–
263
1969 (0.13)
–
81
62.4
–
181
190
–
263
591
–
Log probit
Gamma
–
0.9344
0.5812
–
ED10/EC10, ED20/EC20, and ED50/EC50 were calculated using the U.S. Environmental Protection Agency Benchmark Dose (BMD) software (http://
www.epa.gov/ncea/bmds/). Values in the table correspond to the BMD. The best-fit model for each compound was determined by collectively considering the
following criteria: the highest p value, chi-squared residual values of less than 2, the lowest Akaike’s Information Criterion (AIC) value, the smallest margin
between the BMD and the BMDL (lower limit of a one-sided 95% confidence limit of the BMD) curves, and a good visual fit of the BMD and BMDL curves.
b
TEQWHO were determined using mammalian toxic equivalency factors of 1.0, 0.3, and 0.1 for TCDD, PeCDF, and TCDF, respectively [17].
c
Numbers in parentheses refer to the relative potency of PeCDF or TCDF compared with TCDD based on the corresponding ED or EC values.
d
EC values were not calculated for TCDF because liver concentrations did not increase with consumed dose as indicated by inadequate statistical fit using the
EPA BMD software. The lack of a dose–response relationship suggests that hepatic TCDF concentration is not a reliable internal dose metric for jaw lesions.
EDx ¼ effective dose resulting in x% increase in the incidence of the jaw lesion over background; ECx ¼ effective concentration resulting in x% increase in the
incidence of the jaw lesion over background; TCDD ¼ 2,3,7,8-tetrachlorodibenzo-p-dioxin; PeCDF ¼ 2,3,4,7,8-pentachlorodibenzofuran; TCDF ¼ 2,3,7,8tetrachlorodibenzofuran; TEQWHO ¼ TCDD-toxic equivalents based on World Health Organization (WHO) toxic equivalency factors.
2554
Environ. Toxicol. Chem. 31, 2012
tration-dependent manner. This was observed in EROD activities that increased up to maximal values until concentrations
of PeCDF and TCDF in liver reached concentrations of approximately 350 to 400 ng TEQWHO-TCDF /kg and 5 to 6 ng
TEQWHO-TCDF/kg wet weight, respectively [20]. In the present
study, concentrations of these two congeners in liver either
approached or exceeded these values in both adult and juvenile
mink.
Jaw lesion
The most consistent histopathological lesion observed in the
present study was the presence of the mandibular and maxillary
squamous epithelial proliferation that occurred in all treatment
groups except the control and 2.1-ng TCDD/kg bw/d (2.1 ng
TEQWHO-TCDD/kg bw/d) groups. Initial studies from our laboratory have shown that young mink fed diets containing PCB
126 and TCDD at concentrations of 2,400 ng TEQWHO/kg feed
(300 ng TEQWHO/kg bw/d) beginning at 6 to 12 weeks of age
exhibited histological signs of a lesion associated with the jaw
[7–9]. Gross lesions consisted of mandibular and maxillary
nodular proliferation of the gingiva and loose teeth. The maxilla
and mandible were porous because of loss of alveolar bone.
Histologically, this osteoporosis was caused by proliferation of
squamous cells that formed infiltrating cords. Subsequent studies demonstrated histological evidence of the lesion in mink
exposed to PCBs derived from contaminated fish both in utero
and during the growth phase [11,17] as well as in mink trapped
in the wild [12,13].
Mandibular and maxillary squamous epithelial proliferation
was induced in 27-week-old juveniles in the present study at
doses as small as 6.0 ng TCDD (6.0 ng TEQWHO-TCDD)/kg bw/d,
13 ng PeCDF (4.0 ng TEQWHO-PeCDF)/kg bw/d, or 52 ng TCDF
(5.2 ng TEQWHO-TCDF)/kg bw/d. Juvenile mink that were fed
diets containing as little as 6.6 ng TEQWHO-PCBs/TCDDs/PCDFs/kg
feed (0.83 ng TEQWHO-PCBs/TCDDs/PCDFs/kg bw/d) provided by
fish collected from the Housatonic River, Berkshire County,
Massachusetts, USA [11] or 36 ng TEQWHO-PCBs/PCDDs/PCDFs/
kg feed (4.5 ng TEQWHO-PCBs/TCDDs/PCDFs/kg bw/d) provided
by fish collected from the Saginaw River, Michigan, USA [17]
had histological evidence of mandibular and maxillary squamous epithelial proliferation. Comparison of the various feeding studies indicate that the jaw lesion has typically been
observed in juvenile or adult mink at a TEQWHO dose approximating 3.4 ng TEQWHO/kg bw/d, which is the geometric mean
of the five lowest observed adverse effect levels from the
present study, the Housatonic River study [11], and the Saginaw
River study [17].
In the present study, the jaw lesion was induced by TCDD,
PeCDF, and TCDF at doses for which there was no evidence of
significant effects on reproduction and offspring growth and
survivability [23]. These results are similar to results from the
Housatonic River study [11] and Saginaw River study [17] in
that the jaw lesion occurred at doses less than the dose that
resulted in reduced kit survivability (Housatonic River) or in the
absence of evidence for effects on reproduction and offspring
growth and survivability (Saginaw River). These results suggest
that proliferation of mandibular and maxillary squamous epithelial cells may be used as a sensitive biomarker of exposure
and effect in relation to TCDD-like chemicals.
In the present study, the dietary ED50s expressed on a
TEQWHO basis were 6.6 ng TEQWHO-TCDD/kg bw/d, 4.1 ng
TEQWHO-PeCDF/kg bw/d, and 15 ng TEQWHO-TCDF/kg bw/d
for TCDD, PeCDF, and TCDF, respectively. The ED50s in
the present study are 2.6- to 9.4-fold greater than the estimated
S.J. Bursian et al.
ED50 (13 ng TEQWHO-PCBs/TCDDs/PCDFs/kg feed or 1.6 ng
TEQWHO-PCBs/TCDDs/PCDFs/kg bw/d) for the Housatonic River
study [11]. In the Housatonic River study, PCB 126 contributed
81% of the dietary TEQWHO.
Relative potency factors
The use of TEFWHO to evaluate potential risks of TCDD-like
compounds to wildlife such as mink must be understood in
terms of the underlying assumptions and uncertainties that are
inherent in this approach. Assumptions include (1) that the most
dominant effects of TCDD-like chemicals are mediated through
the AhR; (2) that all individual congeners are full agonists with
parallel dose–response curves; (3) that TEFs are assumed to be
equivalent for all exposure scenarios and endpoints; and (4) that
the relative ranking in potencies remains the same, even though
the pharmacokinetics of the different congeners may differ
between wildlife species [14,15,33]. However, the toxicity
literature related to TCDD-like chemicals in wildlife species,
including traditional laboratory mammals, indicates that the
uncertainties associated with these assumptions when examined
within the context of individual species’ responses are considerable. These uncertainties are of particular relevance for mink
given their importance in ecological risk assessments of TCDDlike compounds and the fact that few studies have been conducted that specifically examine the appropriateness of these
assumptions. For instance, the AhR has yet to be characterized
in any mustelid, and sequence information for CYP1A1 and
CYP1A2 has only recently become available in the literature
[21,22]. Although some inferences regarding AhR responsiveness can be ascertained from CYP1A enzyme activities that
have been reported in several mink studies, the use of these data
in a quantitative manner to better understand the sensitivity of
the mink to TCDD-like chemicals is difficult due to differences
in experimental design, including exposure scenarios, that can
contribute to age- and sex-based variation in chemical accumulation and organismal responsiveness. In the present study,
the assumption that there is direct proportional bioequivalence
between AhR agonists is brought into question in that the
interaction between chemical and consumed dose expressed
on a TEQ basis on the probability of jaw lesion was significant
( p ¼ 0.0047), indicating that the biological response (presence
of the jaw lesion) in terms of TEQ dose is specific to each of the
chemicals.
The current TEFWHO-TCDF value for TCDF is 0.1 [15], which
is slightly greater than the relative potencies (RePs) of 0.09,
0.06, and 0.04 calculated in the present study based on the
ED10s, ED20s, and ED50s, respectively, for the incidence of
jaw lesions in juvenile mink (Table 8). However, it is important
to remember that TEFWHO values are consensus values that are
half order of magnitude increments on a logarithmic scale of
0.03, 0.1, 0.3, etc. [15]. Furthermore, TEFWHO are given as
point estimates that are presented without any confidence
intervals to better understand the variability associated with
these values. A review of the in vivo and in vitro data used for
establishing TEFs [34] found that the ReP values for TCDF in
mammalian species are variable and range from 0.006 to 0.63.
This span in ReP values encompasses the ReP values determined in the present study. Thus, although the use of the current
TEFWHO-TCDF would tend to slightly overestimate the potential
risks associated with TCDF to mink, the magnitude of this
overestimate would be less than a half-logarithmic unit if one
used the ReP values calculated in the present study.
Relative potency factors for PeCDF have been proposed
based on numerous toxicity endpoints that include acute lethal-
Biomarkers in mink exposed to TCDD-like chemicals
ity, growth metrics, immunotoxicity, and teratogenicity. Based
on in vitro studies, RePs for PeCDF ranged from 0.11 to 0.67,
whereas for in vivo studies, RePs have been shown to range
from 0.12 to 0.8 [33]. For cancer endpoints, a ReP of 0.1 was
estimated from a dietary initiation–promotion study [35],
whereas Harper et al. [36] reported RePs that varied from
0.58 to 4.0 based on plaque-forming immunotoxicity in mice.
Relative potency factors for neoplastic endpoints from a longterm rodent study approximated 0.2 to 0.3, whereas RePs based
on noncarcinogenic endpoints ranged from 0.7 to 1.1 [37]. More
recently, Haws et al. [34] reviewed the WHO database and
reported a minimum ReP of 0.0065, a maximum of 3.7 and a
50th percentile value of 0.2 based on in vivo data. Based on all
in vivo and in vitro data, the overall average ReP was 0.22.
Given that the current TEFWHO for PeCDF is 0.3 [15], the
dietary-based RePs of 1.9, 1.1, and 0.5 for mink based on the
ED10s, ED20s, and ED50s, respectively, for the incidence of
jaw lesions (Table 8) are within the range of values reported for
other mammalian species. Thus, use of the current TEF value of
0.3 would be expected to slightly underestimate the potency
of PeCDF to mink, a value well within the analytical error
associated with the measure of this congener in biological
tissues.
One of the many applications of TEFs in wildlife studies
is calculation of total toxic equivalents of TCDD-like compounds in tissues such as liver [10,11,25–27]. However, it is
important to recognize that TEFs or RePs that have been
calculated using administered dose do not account for pharmacokinetic differences between congeners within a species
and therefore may not accurately portray the relative risks
these mixtures may pose to mammalian species [38]. As a
result, recommendations have been made that TEF values be
developed based on both dietary and tissue concentrations
[29,39,40].
In the present study, dose–response relationships were constructed based on concentrations of two of the three AhR-active
compounds (TCDD and PeCDF) and the incidence of jaw
lesions in juvenile mink to develop tissue-based RePs. The
EC10s, EC20s, and EC50s for TCDD and PeCDF based on
hepatic concentrations were 149, 181, and 263 ng TCDD/kg
liver, wet weight, and 299, 634, and 1,969 ng PeCDF/kg liver,
wet weight. Resulting ReP values for PeCDF are 0.5, 0.29, and
0.13, which are close to the TEF value of 0.3. Because there was
no evidence for a dose–response relationship between consumed dose and hepatic concentration of TCDF, which can
be explained by its bioaccumulation factor of 0.08 [25], EC
values were not calculated for this congener. If EC values had
been calculated based on hepatic concentrations of TCDF, the
values would have been considerably lower than the corresponding values for TCDD, resulting in ReP values estimated to
be 20-fold greater than the TEF value of 0.1. These results
suggest that care must be taken when associating the occurrence
of a deleterious effect with tissue concentrations of the chemical(s) in question because of potential differences in metabolism and sequestration, as has been reported for TCDF and
PeCDF [20].
Results of this study indicated that the relationships among
measures of CYP1A gene expression, protein concentrations,
enzymatic activities, and dietary doses of TCDD, PeCDF, and
TCDF in mink were inconsistent. Exposure to environmentally
relevant concentrations of TCDD, PeCDF, and TCDD induced
mandibular and maxillary squamous epithelial proliferation,
and the relative potencies of PeCDF and TCDF relative to
TCDD based on effective doses and concentrations for the
Environ. Toxicol. Chem. 31, 2012
2555
incidence of jaw lesions were within a half-logarithmic unit
of the TEFWHO values.
Acknowledgement—This work was supported by an unrestricted grant from
The Dow Chemical Company to Michigan State University, a Discovery
Grant from the National Science and Engineering Research Council of
Canada (Project 326415-07), and an instrumentation grant from the Canada
Foundation for Infrastructure to the University of Saskatchewan. The authors
have no conflicts of interest.
REFERENCES
1. Hilscherova K, Kannan K, Nakata H, Hanari N, Yamashita N, Bradley
PW, McCabe JM, Taylor AB, Giesy JP. 2003. Polychlorinated dibenzop-dioxin and dibenzofuran concentration profiles in sediments and floodplain soils of the Tittabawassee River, Michigan. Environ Sci Technol
37:468–474.
2. Basu N, Scheuhammer AM, Bursian SJ, Elliott J, Rouvinen-Watt K,
Chan HM. 2007. Mink as a sentinel species in environmental health.
Environ Res 103:130–144.
3. Hochstein JR, Aulerich RJ, Bursian SJ. 1988. Acute toxicity of 2,3,7,8tetrachlorodibenzo-p-dioxin to mink. Arch Environ Contam Toxicol
17:33–37.
4. Hochstein JR, Bursian SJ, Aulerich RJ. 1998. Effects of dietary exposure
to 2,3,7,8-tetrachlorodibenzo-p-dioxin in adult female mink. Arch
Environ Contam Toxicol 15:348–353.
5. Hochstein JR, Render JA, Bursian SJ, Aulerich RJ. 2001. Chronic
toxicity of dietary 2,3,7,8-tetrachlorodibenzo-p-dioxin to mink. Vet Hum
Toxicol 43:134–139.
6. Beckett KJ, Yamini B, Bursian SJ. 2008. The effects of 3,30 ,4,40 5pentachlorobiphenyl (PCB 126) on mink (Mustela vison) reproduction
and kit survivability and growth. Arch Environ Contam Toxicol 54:123–
129.
7. Render JA, Aulerich RJ, Bursian SJ, Nachreiner RF. 2000. Proliferation
of maxillary and mandibular periodontal squamous cells in mink fed
3,30 ,4,40 ,5-pentachlorobiphenyl (PCB 126). J Vet Diagn Invest 12:477–
479.
8. Render JA, Hochstein JR, Aulerich RJ, Bursian SJ. 2000. Proliferation of
periodontal squamous epithelium in mink fed 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Vet Hum Toxicol 42:85–86.
9. Render JA, Bursian SJ, Rosenstein DS, Aulerich RJ. 2001. Squamous
epithelial proliferation in the jaws of mink fed diets containing
3,30 ,4,40 ,5-pentachlorobiphenyl (PCB 126) or 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Vet Hum Toxicol 43:22–26.
10. Bursian SJ, Sharma C, Aulerich RJ, Yamini B, Mitchell RR, Orazio CE,
Moore DRJ, Svirsky S, Tillitt DE. 2006. Dietary exposure of mink
(Mustela vison) to fish from the Housatonic River, Berkshire County,
Massachusetts, USA: Effects on reproduction, kit growth, and survival.
Environ Toxicol Chem 25:1533–1540.
11. Bursian SJ, Sharma C, Aulerich RJ, Yamini B, Mitchell RR, Beckett KJ,
Orazio CE, Moore D, Svirsky S, Tillitt DE. 2006. Dietary exposure of
mink (Mustela vison) to fish from the Housatonic River, Berkshire
County, Massachusetts, USA: Effects on organ weights and histology
and hepatic concentrations of polychlorinated biphenyls and 2,3,7,8tetrachlorodibenzo-p-dioxin toxic equivalence. Environ Toxicol Chem
25:1541–1550.
12. Beckett KJ, Millsap SD, Blankenship AL, Zwiernik MJ, Giesy JP,
Bursian SJ. 2005. Squamous epithelial lesions of the mandibles and
maxillae of wild mink (Mustela vison) naturally exposed to polychlorinated biphenyls. Environ Toxicol Chem 24:674–677.
13. Haynes JM, Wellman ST, Beckett KJ, Pagano JJ, Fitzgerald SD, Bursian
SJ. 2009. Histological lesions in mink jaws are highly sensitive
biomarker of effect after exposure to TCDD-like chemicals: Field and
literature-based confirmations. Arch Environ Contam Toxicol 57:803–
807.
14. Van den Berg M, Birnbaum L, Bosveld ATC, Brunstrom B, Cook P,
Feeley M, Giesy JP, Hanberg A, Hasegawa R, Kennedy SW, Kubiak T,
Larsen JC, Van Leeuwen FXR, Liem AKD, Nolt C, Peterson RE,
Poellinger L, Safe S, Schrenk D, Tillitt D, Tysklind M, Younes M, Waern
F, Zacharewski T. 1998. Toxicity equivalency factors (TEFs) for PCBs,
PCDDs, PCDFs for humans and wildlife. Environ Health Perspect
106:775–792.
15. Van den Berg M, Birnbaum LS, Denison M, De Vito M, Farland W,
Feeley M, Fiedler H, Hakansson H, Hanberg A, Haws L, Rose M, Safe S,
Schrenk D, Tohyama C, Tritscher A, Tuomisto J, Tysklind M, Walker N,
Peterson RE. 2006. The 2005 World Health Organization reevaluation of
2556
16.
17.
18.
19.
20.
21.
22.
23.
24.
25.
26.
27.
Environ. Toxicol. Chem. 31, 2012
human and mammalian toxic equivalency factors for dioxins and dioxinlike compounds. Toxicol Sci 93:223–241.
Zwiernik MJ, Kay DP, Moore JN, Beckett KJ, Khim JS, Newsted JL,
Roark S, Giesy JP. 2008. Exposure and effects assessment of resident
mink exposed to polychlorinated dibenzofurans and other dioxin-like
compounds in the Tittabawassee River Basin, Midland, MI, USA on wild
mink (Mustela vison). Environ Toxicol Chem 27:2076–2087.
Bursian SJ, Beckett KJ, Yamini B, Martin PA, Kannan KL, Shields KL,
Mohr FC. 2006. Assessment of effects in mink caused by consumption of
carp collected from the Saginaw River, Michigan, USA. Arch Environ
Contam Toxicol 50:614–623.
Blankenship AL, Kay DP, Zwiernik MJ, Holem RR, Newsted JL, Hecker
M, Giesy JP. 2008. Toxicity reference values for mink exposed to
2,3,7,8-tetrachlodibenzo-p-dioxin. Ecotoxicol Environ Saf 69:325–349.
Zwiernik MJ, Beckett KJ, Bursian SJ, Kay DP, Holem RR, Moore JN,
Yamini B, Giesy JP. 2009. Chronic effects of polychlorinated
dibenzofurans on mink in laboratory and field environments. Integr
Environ Assess Manage 5:291–301.
Zwiernik MJ, Bursian SJ, Aylward LL, Kay DP, Moore JN, Rowlands C,
Woodburn K, Shotwell MS, Khim JS, Giesy JP, Budinsky RA. 2008.
Toxicokinetics of 2,3,7,8-TCDF and 2,3,4,7,8-PeCDF in mink (Mustela
vison) at ecologically relevant exposures. Toxicol Sci 105:34–43.
Zhang X, Moore JN, Newsted JL, Hecker M, Zwiernik MJ, Jones PD,
Bursian SJ, Giesy JP. 2008. Sequencing and characterization of mixed
function monooxygenase genes CYP1A1 and CYP1A2 of mink (Mustela
vison) to facilitate study of dioxin-like compounds. Toxicol Appl
Pharmacol 234:306–313.
Moore JN, Newsted JL, Hecker M, Zwiernik MJ, Fitzgerald SD, Kay DP,
Zhang X, Higley EB, Aylward LL, Beckett KJ, Budinsky RA, Bursian
SJ, Giesy JP. 2009. Hepatic P450 enzyme activity, tissue morphology,
and histology of mink (Mustela vison) exposed to polychlorinated
dibenzofurans. Arch Environ Contam Toxicol 57:416–425.
Moore JN, Zwiernik MJ, Newsted JL, Fitzgerald SD, Link JE, Bradley
PW, Kay D, Budinsky R, Giesy JP, Bursian SJ. 2012. Effects of dietary
exposure of mink (Mustela vison) to 2,3,7,8-tetrachlorodibenzo-pdioxin (TCDD), 2,3,4,7,8-pentachlorodibenzofuran (PeCDF) and
2,3,7,8-tetrachlorodibenzofuran (TCDF) on reproduction and offspring
survivability and growth. Environ Toxicol Chem 31:360–369.
Fur Commission USA. 2003. Standard guidelines for the operation of
mink farms in the United States. Coronado, CA.
Heaton SN, Bursian SJ, Giesy JP, Tillitt DE, Render JA, Jones PD,
Verbrugge DA, Kubiak TJ, Aulerich RJ. 1995. Dietary exposure of mink
to carp from Saginaw Bay, Michigan. 1. Effects on reproduction and
survival, and the potential risks to wild mink populations. Arch Environ
Contam Toxicol 28:334–343.
Heaton SN, Bursian SJ, Giesy JP, Tillitt DE, Render JA, Jones PD,
Verbrugge DA, Kubiak TJ, Aulerich RJ. 1995. Dietary exposure of mink
to carp from Saginaw Bay, Michigan. 2. Hematology and liver
pathology. Arch Environ Contam Toxicol 29:411–417.
Tillitt DE, Gale RW, Meadows JC, Zajieck JL, Peterman PH, Heaton SN,
Jones PD, Bursian SJ, Kubiak TJ, Giesy JP, Aulerich RJ. 1996. Dietary
S.J. Bursian et al.
28.
29.
30.
31.
32.
33.
34.
35.
36.
37.
38.
39.
40.
exposure of mink to carp from Saginaw Bay. 3. Characterization of
dietary exposure to planar halogenated hydrocarbons, dioxin equivalents, and biomagnification. Environ Sci Technol 30:283–291.
Kennedy SW, Jones SP. 1994. Simultaneous measurement of
cytochrome P4501A catalytic activity and total protein concentration
with a fluorescence plate reader. Anal Biochem 222:217–223.
Neubert D, Golor G, Neubert R. 1992. TCDD-toxicity equivalencies for
PCDD/PCDF congeners: Prerequisites and limitations. Chemosphere
25:65–70.
Whitlock JP Jr. 1999. Induction of cytochrome P4501A1. Annu Rev
Pharmacol Toxicol 39:103–125.
Brewster DW, Birnbaum LS. 1987. Disposition and excretion of 2,3,4,5pentachlorodibenzofuran in the rat. Toxicol Appl Pharmacol 90:243–
252.
Diliberto JJ, Burgin DE, Birnbaum LS. 1999. Effects of CYP1A2 on
disposition of 2,3,7,8-tetrachlorodibenzo-p-dioxin, 2,3,4,7,8-pentachlorodibenzofuran and 2,20 ,4,40 ,5,50 -hexachlorobiphenyl in CYP1A2
knockout and parental (C57BL/6N and 129/Sv) strains of mice. Toxicol
Appl Pharmacol 159:52–64.
Safe S. 1998. Development validation and problems with the toxic
equivalency factor approach for risk assessment of dioxins and related
compounds. J. Anim Sci 76:134–141.
Haws LC, Su SH, Harris M, De Vito MJ, Walker NJ, Farland WH, Finley
B, Birnbaum LS. 2006. Development of a refined database of
mammalian relative potency estimates for dioxin-like compounds.
Toxicol Sci 89:4–30.
Waern F, Flodstrom S, Busk L, Kornevi T, Nordgren I, Ahlborg UG.
1991. Relative tumour-promoting activity and toxicity of some
polychlorinated dibenzo-p-dioxin-congeners and dibenzofuranscongeners in female Sprague-Dawley rats. Pharmacol Toxicol 69:
450–458.
Harper N, Connor K, Safe S. 1993. Immunotoxic potencies of
polychlorinated biphenyl (PCB), dibenzofurans (PCDF), and dibenzop-dioxin (PCDD) congeners in C57BL/6 and DBA/mice. Toxicology
80:217–227.
Walker NJ, Crockett P, Nyska A, Brix A, Kokinen MP, Sells DM, Hailey
JR, Easterling M, Haseman JK, Yin M, Wyde ME, Bucher JR, Portier CJ.
2005. Dose-additive carcinogenicity of a defined mixture of dioxin-like
compounds. Environ Health Perspect 113:43–48.
Budinsky RA, Paustenbach DJ, Fontaine D, Landenberger BD, Starr TB.
2006. Recommended relative potency factors for 2,3,4,7,8-pentachlorodibenzofuran: The impact of different dose metrics. Toxicol Sci
91:275–285.
DeVito MJ, Diliberto JJ, Ross DG, Menache MG, Birnbaum LS. 1997.
Dose-response relationships for polyhalogenated dioxins and dibenzofurans following subchronic treatment in mice. Toxicol Appl Pharmacol
147:267–280.
U.S. Environmental Protection Agency. 2003. Framework for application of the toxicity equivalence methodology for polychlorinated
dioxins, furans, and biphenyls in ecological risk assessement. EPA/630/
P-03/002A. Washington, DC.