Environmental Toxicology and Chemistry, Vol. 26, No. 4, pp. 764–772, 2007
䉷 2007 SETAC
Printed in the USA
0730-7268/07 $12.00 ⫹ .00
EFFECTS OF BROMINATED FLAME RETARDANTS AND BROMINATED DIOXINS ON
STEROIDOGENESIS IN H295R HUMAN ADRENOCORTICAL CARCINOMA CELL LINE
LING DING,† MARGARET B. MURPHY,‡ YUHE HE,‡ YAN XU,‡ LEO W.Y. YEUNG,‡ JINGXIAN WANG,†
BINGSHENG ZHOU,*† PAUL K.S. LAM,‡ RUDOLF S.S. WU,‡ and JOHN P. GIESY§
†State Key Laboratory of Freshwater Ecology and Biotechnology, Institute of Hydrobiology, Chinese Academy of Sciences, Wuhan 430072,
People’s Republic of China
‡Center for Coastal Pollution and Conservation, Department of Biology and Chemistry, City University of Hong Kong,
Hong Kong Special Administrative Region, People’s Republic of China
§National Food Safety and Toxicology Center, Department of Zoology, and Institute for Environmental Toxicology,
Michigan State University, East Lansing, Michigan 48824, USA
( Received 31 July 2006; Accepted 30 October 2006)
Abstract—Brominated flame retardants (BFRs) and brominated dioxins are emerging persistent organic pollutants that are ubiquitous
in the environment and can be accumulated by wildlife and humans. These chemicals can disturb endocrine function. Recent studies
have demonstrated that one of the mechanisms of endocrine disruption by chemicals is modulation of steroidogenic gene expression
or enzyme activities. In this study, an in vitro assay based on the H295R human adrenocortical carcinoma cell line, which possesses
most key genes or enzymes involved in steroidogenesis, was used to examine the effects of five bromophenols, two polybrominated
biphenyls (PBBs 77 and 169), 2,3,7,8-tetrabromodibenzo-p-dioxin, and 2,3,7,8-tetrabromodibenzofuran on the expression of 10 key
steroidogenic genes. The H295R cells were exposed to various BFR concentrations for 48 h, and the expression of specific genes—
cytochrome P450 (CYP11A, CYP11B2, CYP17, CYP19, and CYP21), 3␤-hydroxysteroid dehydrogenase (3␤HSD2), 17␤-hydroxysteroid dehydrogenase (17␤HSD1 and 17␤HSD4), steroidogenic acute regulatory protein (StAR), and 3-hydroxy-3-methylglutaryl
coenzyme A reductase (HMGR)—was quantitatively measured using real-time polymerase chain reaction. Cell viability was not
affected at the doses tested. Most of the genes were either up- or down-regulated, to some extent, by BFR exposure. Among the
genes tested, 3␤HSD2 was the most markedly up-regulated, with a range of magnitude from 1.6- to 20-fold. The results demonstrate
that bromophenol, bromobiphenyls, and bromodibenzo-p-dioxin/furan are able to modulate steroidogenic gene expression, which
may lead to endocrine disruption.
Keywords—Brominated flame retardants
Bromodioxins
Steroidogenesis
H295R
Gene expression
centrations in the environment, including in birds [7], marine
mammals [8], and humans [9].
The principal sources of polychlorinated dibenzo-p-dioxins
and dibenzofurans (PCDD/Fs) are known to be combustion
and thermal processes, but polybrominated dibenzo-p-dioxins
and dibenzofurans (PBDD/Fs) are emitted into the atmosphere
primarily from electronic waste recycling facilities and other
thermal treatment of BFRs [10]. For example, significant
amounts of these compounds are produced by combustion of
television and circuit boards in electronic waste incinerators,
accidental fires, municipal and industrial waste incineration,
and burning of automotive fuels containing BFRs, such as
PBDEs [11]. In addition, PBDD/Fs can be formed in the process of manufacturing bromophenols and TBBPA. Recent
studies have shown that bromophenols can form great yields
of PBDD/Fs from high-temperature oxidation [12]. Because
bromine inhibits combustion, the incomplete combustion likely would produce a significant amount of PBDD/Fs. Indeed,
PBDD/Fs were detected in flue gas, fly ash, and bottom ash
of municipal solid waste incineration plants [13], and considerable levels of PBDD/Fs have been detected in the atmosphere
[14], in blue mussels (Mytilus edulis) [15], and in the liver
and eggs of common cormorants (Phalacrocorax carbo) [7].
Although the detected levels of PBDD/Fs were relatively less
those of than PCDD/Fs, the rapid increase in the combustion
of electrical waste may raise environmental concerns. The
World Health Organization assessment of PBDD/Fs and mixed
INTRODUCTION
Brominated flame retardants (BFRs), such as polybrominated biphenyls (PBBs), bromophenols, tetrabromobisphenol
A (TBBPA), polybrominated diphenyl ethers (PBDEs) and
hexabromocyclododecane have been used in large quantities
to reduce fire risk in electrical equipment, cars, construction
materials, plastics, foams, and textiles. These compounds are
ubiquitous in sediments and biota, including marine mammals
and humans [1]. Recently, BFRs have caused greater concern,
because concentrations of some BFRs measured in biota have
increased significantly [1,2].
Among these BFRs, bromophenols, such as 2,4-dibromophenol (DBP), 2,4,6-tribromophenol (TBP), and pentabromophenol (PBP), represent high-volume flame retardants [3].
Greater concentrations of 2,4,6-TBP have been detected in
marine mussels and mammals [4] as well as in human milk
and serum [5]. A commercial flame retardant containing PBBs
accidentally added to animal feed in Michigan, USA, in 1975
resulted in loss of livestock and long-term impacts on the
health of people who consumed the PBB-contaminated meat
[6]. Although PBBs were phased out from the market during
the 1970s in the United States and in the year 2000 in Europe,
they are more resistant to degradation than polychlorinated
biphenyls (PCBs) and have been measured at significant con* To whom correspondence may be addressed
(bszhou@ihb.ac.cn).
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Environ. Toxicol. Chem. 26, 2007
Gene expression induced by BRFs in H295R cells
Fig. 1. Schematic representation of the steps involved in steroid hormone synthesis and genes measured in the present study. StAR ⫽
steroidogenic acute regulatory protein; HMGR ⫽ 3-hydroxy-3-methylglutaryl–coenzyme A reductase; CYP ⫽ cytochrome P450; 3␤HSD
⫽ 3␤-hydroxysteroid dehydrogenase isomerase; 17␤HSD ⫽ 17␤-hydroxysteroid dehydrogenase; DHEA ⫽ dehydroepiandrosterone.
brominated-chlorinated dioxins and dibenzofurans concluded
that PBDD/Fs are persistent and toxic and that humans and
the environment should be protected from these compounds.
However, few studies on the toxicity of brominated dioxins
have been undertaken.
The BFRs can cause toxicity through thyroidogenic, dioxinlike, and estrogenic mechanisms [16]. Classic 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD)–like responses, such as antiestrogenic activity, have been reported for some BFRs. Evidence exists that BFRs, particularly PBDEs, disrupt thyroid
hormone function and interact with thyroid hormone receptor
[16]. Some bromophenol congeners are suspected of disrupting
endocrine function through thyroid hormone–like activity. For
example, 2,4,6-TBP is suspected to have thyroid hormone–
like activity by binding to transthyretin with greater affinity
than that of thyroxine itself [16], whereas 2,4-DBP binds to
the human estrogen receptor [17]. The PBBs also interact with
the endocrine system, and exposure of rats and pigs has resulted in decreases in serum triiodothyronine and thyroxine
[18]. In addition, several studies have demonstrated that some
BFRs exert their effects on steroidogenic enzymes [19,20]. A
recent study showed that exposure to PBDE 99 resulted in
decreases in sex hormones (e.g., testosterone and 17␤-estradiol) in male rates in vivo, leading to interference with sexual
development and dimorphic behavior [21].
A novel bioassay based on the human adrenocortical carcinoma cell line (H295R) has been used for screening endocrine-disrupting chemicals. The H295R cells have physiological characteristics of zonally undifferentiated human fetal adrenal cells and has been shown to have the ability to produce
most of the steroid hormones of the three phenotypically distinct zones found in the adult adrenal cortex [22] (Fig. 1).
Heneweer et al. [23] showed that exposure of H295R cells to
atrazine resulted in a threefold induction of aromatase activity,
but no induction of this enzyme activity occurred in rat Leydig
cell carcinoma cell line (R2C), suggesting that H295R cells
are more sensitive than R2C cells. Although the mechanisms
of disruption of corticosteroid production are currently still
765
unknown, Sanderson et al. [24] studied the induction of aromatase (CYP19) activity by various classes of pesticides, suggesting that the mechanism of aromatase induction is mediated
through inhibition of phosphodiesterase activity [24]. The
H295R human adrenocortical carcinoma cell line has been
suggested as a potential model for screening adrenocortical
toxicity and steroidogenesis [25,26], and these cells have been
used to examine the effects of PCBs on adrenal aldosterone
biosynthesis [27], various pesticides and synthetic flavonoid
compounds on CYP19 activity [28,29], and phthalate esters
and alkylphenols [30] on steroid enzyme activities. Recently,
a gene expression–based assay has been used for assessment
of chemicals as well as environmental samples on steroidogenic gene expression using this cell line [25,31–33].
In the present study, we examined the effects of BFRs and
related chemicals (e.g., PBDD/Fs) on the expression of key
steroidogenic genes to determine whether these compounds
could disrupt endocrine function through modulation of steroidogenesis. These genes include CYP11A (side-chain cleavage enzyme), CYP11B2 (aldosterone synthetase), CYP17 (steroid 17␣ -hydroxylase/17,20-lyase), CYP19 (aromatase),
CYP21 (steroid 1-hydroxylase), 3␤HSD2 (3␤-hydroxysteroid
dehydrogenase isomerase), 17␤HSD (17␤-hydroxysteroid dehydrogenase), StAR (steroidogenic acute regulatory protein),
and HMGR (3-hydroxy-3-methylglutaryl–coenzyme A reductase).
MATERIALS AND METHODS
Chemicals
Bromophenols, 2-bromophenol (2-BP; purity, ⬎99.0%),
2,4-DBP (purity, ⬎99.5%), 2,6-DBP (purity, ⬎ 98.0%),
2,4,6-TBP (purity, ⬎98.6%), PBP (purity, ⬎99.0%), and
3,3,4,4⬘,5,5⬘-hexabromobiphenyl (PBB 169; purity, ⬎95.0%)
were obtained from Sigma-Aldrich (St. Louis, MO, USA). The
3,3 ⬘,4,4 ⬘ -tetrabromobiphenyl (PBB 77; purity, ⬎ 99.7%),
2,3,7,8-tetrabromodibenzo-p-dioxin (TBDD; purity, ⬎99.0%),
and 2,3,7,8-tetrabromodibenzofuran (TBDF; purity, ⬎99.0%)
were purchased from AccuStandard (New Haven, CT, USA).
Dimethyl sulfoxide (DMSO) was first added into all standards,
and the original solvent (toluene, methanol, and iso-octane)
was then blown out under a gentle stream of nitrogen before
preparing their dilution series and stored at ⫺20⬚C. The final
solvent concentration in the culture medium did not exceed
0.5% (v/v).
Cell culture
The H295R human adrenocortical carcinoma cells (American Type Culture Collection, Beltsville, VA, USA) were cultured in 1:1 (v/v) Dulbecco’s modified Eagle’s medium/F-12
Ham nutrient mixture. The culture medium contained 15 mM
4-(2-hydroxyethyl)-1-piperazineethanesulfonic acid (Hepes)
buffer (Sigma-Aldrich), 1% insulin-transferring sodium selenite plus Premix (BD Bioscience, San Jose, CA, USA), 2.5%
Nu-Serum (BD Bioscience), 100 U/ml of penicillin, and 100
␮g/ml of streptomycin. The H295R cells were cultured at 37⬚C
in 5% CO2. The culture medium was changed every 2 to 3 d.
Cell viability assay and exposure conditions
To test for cytotoxicity, the H295R cells were seeded in
96-well plates (Falcon, Franklin Lakes, NJ, USA) at a concentration of 1 ⫻ 105 cells/well. When the cells reached approximately 70% confluence, DMSO dissolved standards were
diluted with culture medium to the desired concentrations and
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Environ. Toxicol. Chem. 26, 2007
L. Ding et al.
First-strand cDNA synthesis
Fig. 2. Actual threshold cycle (Ct ) values for ␤-actin in response to
the test compounds and the solvent control. Values are the means of
three determinations on each of three replicate exposures (mean ⫾
standard deviation). DMSO ⫽ dimethyl sulfoxide; BP ⫽ bromophenol; DBP ⫽ dibromophenol; TBP ⫽ tribromophenol; PBP ⫽ pentabromophenol; PBB ⫽ polybrominated biphenyl; TBDD ⫽ 2,3,7,8tetrabromodibenzo-p-dioxin; TBDF ⫽ 2,3,7,8-tetrabromodibenzofuran.
First-strand cDNA synthesis was performed by use of the
Superscript娂 first-strand synthesis system (Invitrogen, Carlsbad, CA, USA). Total RNA (2 ␮g) was combined with 0.5 ␮g
of biotinylated oligo (dT)12–18 and 0.5 mM deoxynucleotide
triphosphate nucleotides, then RNase-free water was added to
a final volume of 10 ␮l. Samples were denatured at 65⬚C for
5 min and quickly chilled on ice for another 5 min. Reverse
transcription was performed in a 9-␮l reaction mixture containing the following buffer: 2 ␮l of 10⫻ RT buffer, 4 ␮l of
25 mM MgCl2, 1␮l of RNaseOUT (40 U/␮l; Invitrogen), and
2 ␮l of RNase-free H2O. The mixtures were incubated at 42⬚C
for 2 min, then 50 U of SuperScript II RT (Invitrogen) were
added. The reaction was incubated at 42⬚C for 50 min and
then inactivated by heating at 70⬚C for 15 min. To digest RNA,
1 ␮l of RNase H (2 U/␮l) was added to each tube and incubated
at 37⬚C for 20 min.
Real-time polymerase chain reactions
added to each well. Three wells were used for each treatment
and control as triplicates. After 48 h of exposure, the cell
viability was assayed by using the Sulforhodamine B assay
(Sigma-Aldrich). All plates included a negative control with
no treatment (blank) and a solvent control treated with 0.5%
DMSO. For gene expression examination, 1 ⫻ 106 cells were
seeded in a six-well plate (Nalgene Nunc, Rochester, NY, USA)
and cultured for 48 h. The culture medium was removed, fresh
medium added, and the cells exposed to three concentrations
of the test compounds for another 48 h. Each treatment was
tested in triplicate. The cell condition was checked regularly
under a microscope.
RNA isolation
Total RNA was isolated with the SV Total RNA Isolation
system (Promega, Madison, WI, USA) following the manufacturer’s introductions. Purity of the prepared RNA was verified by agarose gel electrophoresis, and the quantity of the
RNA sample was determined using ultraviolet photometry
(Eppendorf, Hamburg, Germany). The purified RNA was used
immediately for reverse transcription (RT) or stored at ⫺80⬚C
until analysis. The housekeeping gene, ␤-actin, which is ubiquitously expressed and the expression level of which does not
vary under experimental conditions in the present study (Fig.
2), was used as an internal control to standardize the amount
of sample RNA.
Real-time polymerase chain reaction (PCR; quantitative
PCR) was performed by using the ABI 7500 Fast Real-Time
PCR System (Applied Biosystems, Foster City, CA, USA) in
sterile, 96-well PCR plates (Applied Biosystems). The cDNA
samples were then diluted five times with RNase-free H2O.
The PCR reaction mixtures (20 ␮l) contained 1⫻ SYBR
Green娂 PCR Master Mix (Applied Biosystems), 0.2 to 0.4
␮M sense and antisense gene-specific primers (Table 1), and
a 1-␮g cDNA sample. Samples were denatured at 95⬚C for 10
min, followed by 40 cycles of denaturation at 95⬚C for 15 s,
annealing with extension for 1 min at 60⬚C, and a final cycle
of 95⬚C for 15 s, 60⬚C for 1 min, and 95⬚C for 15 s. Melting
curve analyses were performed after the 60⬚C stage of the final
cycle to differentiate between desired PCR products and primer–dimers or DNA contaminants.
The quantity of product amplicon was derived from the
threshold cycle (Ct) number, which is determined as the lowest
cycle at which significant signal over background noise (setup
threshold) is observed. The Ct number is proportional to the
starting amount of the nucleic acid. The fold-change can be
calculated using the following formulae:
⌬Ct ⫽ Ct target ⫺ Ct ␤-actin
⌬(⌬Ct ) ⫽ ⌬Ct exp ⫺ ⌬Ct con
Fold change ⫽ x exp /xcon ⫽ 2⫺⌬(⌬Ct)
Table 1. Primers for quantitative reverse transcription–polymerase chain reactiona
Gene
␤-Actin
CYP11A
CYP11B2
CYP17
CYP19
CYP21
3␤HSD2
17␤HSD1
17␤HSD4
StAR
HMGR
a
Product
Concn.
length (bp) (␮M)
100
137
146
134
128
108
95
136
121
168
152
0.2
0.4
0.4
0.4
0.2
0.4
0.4
0.4
0.4
0.2
0.4
Sense primer
Antisense primer
5⬘-CACCTTCCAGCCTTCCTTCC-3⬘
5⬘-GAGATGGCACGCAACCTGAAG-3⬘
5⬘-TCCAGGTGTGTTCAGRAGTTCC-3⬘
5⬘-AGCCGCACACCAACTATCAG-3⬘
5⬘-AGGTGCTATTGGTRCATCTTGCTC-3⬘
5⬘-CGTGGTGCTGACCCGACTG-3⬘
5⬘-TGCCAGTCTTCATCTACACCAG-3⬘
5⬘-CTCCCTCTGACCAGCAACC-3⬘
5⬘-TGCGGGATCACGGATGACTC-3⬘
5⬘-GTCCCACCCTGCCTCTGAAG-3⬘
5⬘-TGCTTGCCGAGCCTAATGAAAG-3⬘
5⬘-AGGTCTTTGCGGATGTCCAC-3⬘
5⬘-CTTAGTGTCTCCTTGATGCTGGC-3⬘
5⬘-GAAGCCATCTCTGAGGTCTGTG-3⬘
5⬘-TCACCGATGCTGGAGTCAAC-3⬘
5⬘-TGGTGGAATCGGGTCTTTATGG-3⬘
5⬘-GGCTGCATCTTGAGGATGACAC-3⬘
5⬘-TTCCCAGAGGCTCTTCTTCGTG-3⬘
5⬘-TGTGTCTCCCACGCAATCTC-3⬘
5⬘-GCCACCATTCTCCTCACAACTC-3⬘
5⬘-CATACTCTAAACACGAACCCCACC-3⬘
5⬘-AGAGCGTTCGTGGGTCCATC-3⬘
bp ⫽ base pairs; CYP⫽ cytochrome P450; 3␤HSD2 ⫽ 3␤-hydroxysteroid dehydrogenase isomerase; 17␤HSD ⫽ 17␤-hydroxysteroid dehydrogenase; StAR ⫽ steroidogenic acute regulatory protein; HMGR ⫽ 3-hydroxy-3-methylglutaryl-coenzyme A reductase.
Environ. Toxicol. Chem. 26, 2007
Gene expression induced by BRFs in H295R cells
767
where ⌬Ct represents the value of each tested gene normalized
by the Ct for ␤-actin; ⌬(⌬Ct) represents the difference between
⌬Ct values for the solvent control and chemically exposed
cells; xexp and xcon represent the degree of expression in the
exposed and control samples, respectively; and xexp/xcon represents the fold-induction [31]. All data were reported and
statistically analyzed as fold-induction between exposed and
solvent control cells. Gene expression levels were measured
in triplicate for each treatment and controls, and each exposure
was repeated three times in the same experiment.
Statistical analysis
Statistical analysis of gene expression profiles was performed using SPSS (Ver 12.0; SPSS, Chicago, IL, USA). The
gene expressions were checked for homogeneity of variances
by Levene’s test. Because the data failed to pass the tests, a
logarithmic transformation was performed on the data, and the
data were checked again for homogeneity of variances. When
the assumptions for homogeneity of variances were met, differences in gene expression were evaluated by one-way analysis of variance followed by Tukey’s test. The criterion for
statistical significance was p ⬍ 0.05.
RESULTS
The H295R cells were exposed to three concentrations of
five brominated phenols, two PBBs, and TBDD/TBDF for 48
h, and relative responses of 10 genes involved in the steroidogenic pathway were evaluated. Cell viability was not affected
by exposure to the tested concentrations. Expression profiles
of 10 steroidogenic genes varied in response to these chemicals.
Effects of brominated phenols on adrenocortical
steroidogenesis
Exposure to 0.17 ␮M 2-BP significantly up-regulated
CYP17, CYP19, 3␤HSD2, and 17␤HSD1 by 1.8-, 1.9-, 6.6-,
and 1.7-fold, respectively, compared to the control (Fig. 3A).
Treatment with greater concentrations (1.74 ␮M) further resulted in a significant increase of 17␤HSD4 induction by 2.3fold (Fig. 3A). Exposure to the greatest tested concentration
of 17.4 ␮M resulted in increased expression of CYP11A,
3␤HSD2, and StAR by 1.9-, 6.7-, and 1.6-fold, respectively
(Fig. 3A). Among the genes examined, 3␤HSD2 was the most
markedly up-regulated (more than sixfold), whereas CYP11B2,
CYP21, and HMGR mRNA levels remained unchanged after
exposure to all concentrations.
Treatment with 2,4-DBP resulted in up-regulation of most
of the steroidogenic genes measured and showed a good concentration-dependent response (Fig. 3B). The 3 ␤ HSD2 ,
17␤HSD1, and 17␤HSD4 genes were induced after exposure
←
Fig. 3. Effects of bromophenols on steroidogenic gene expression in
the H295R cells. (A) 2-Bromophenol (2-BP). (B) 2,4-dibromophenyl
(2,4-BDP). (C) 2,6-dibromophenyl (2,6-BDP). (D) 2,4,6-Tribromophenol (2,4,6-TBP). (E) Pentabromophenol (PBP). Relative mRNA
expression represents the fold-change with exposure to 0.5% dimethyl
sulfoxide (DMSO). Expression of steroidogenic genes was normalized
to the expression of ␤-actin. Fold-induction represents the increase
in expression compared to the relevant solvent control. Values are the
means of three determinations on each of three replicate exposures
(mean ⫾ standard deviation). An asterisk indicates a statistically significant (p ⬍ 0.05) difference from the DMSO control. See Figures
1 and 2 for definitions of the other abbreviations.
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Environ. Toxicol. Chem. 26, 2007
to all concentrations (0.12, 1.2, and 12 ␮M), and CYP11A,
CYP11B2, and CYP17 expression increased moderately with
exposure to 1.2 and 12 ␮M 2,4-DBP. Significantly up-regulated
expression of CYP19, StAR, and HMGR of 3.2, 1.8-, and 1.5fold, respectively, was observed only at the greatest concentration (12 ␮M). As with 2-BP, 3␤HSD2 was the most significantly up-regulated gene (more than fivefold). Expression
of CYP21 was not significantly altered by any concentration
of 2,4-DBP.
In the 2,6-DBP exposure, suppressed expression of most
of the tested genes generally was observed after treatment with
the lowest concentration (0.2 ␮M) (Fig. 3C), but no significant
difference in expression was seen after exposure to greater
concentrations (2 and 20 ␮M), with the exception of 3␤HSD2,
which was the only gene that was significantly up-regulated
at all tested concentrations (1.9- to 8.4-fold). Expression of
the CYP19 gene was not significantly altered.
In the exposure to 2,4,6-TBP, 3␤HSD2 was the most markedly induced gene, with greater than 15-fold induction at all
test concentrations (0.09, 0.9, and 9 ␮M) (Fig. 3D). The
CYP11A, CYP19, and StAR genes also were significantly upregulated by approximately 1.5-fold at the greatest concentration (9 ␮M). In contrast, suppression of CYP11B2 and CYP17
(1.6-fold) was observed at all three concentrations, and CYP21
was suppressed (1.6-fold) after treatment at concentrations of
0.9 and 9 ␮M. None of the other genes was significantly affected.
Treatment with PBP resulted in a complicated gene expression profile (Fig. 3E). Up-regulation of gene expression
generally was observed. The 3␤HSD2 gene showed the greatest PBP-induced activation (10.3-fold) among all the genes
examined. Expression of the CYP11A, CYP11B2, CYP17,
CYP19, CYP21, and 17␤HSD1 genes was less induced depending on the exposure concentrations. The CYP21 and
17␤HSD1 genes were less induced (1.5- to 1.7-fold), whereas
the CYP11B2 gene was moderately induced by 5.7- to 7.3fold. For HMGR, a significant induction of gene expression
(1.5-fold) was observed at 0.14 ␮M, and a slight but significant
suppression of the gene (1.4-fold) was observed at 20 ␮M. No
effect of PBP on 17␤HSD4 and StAR was observed.
Effects of PBBs on adrenocortical steroidogenesis
In the exposure to PBB 77, up-regulation of gene expression
generally was observed in most of the genes examined (Fig.
4A). Expression of the CYP11A, CYP11B2, CYP19, CYP21,
3␤HSD2, and 17␤HSD1 genes was elevated by more than 1.6fold compared to the solvent control. The StAR mRNA levels
increased by 2.4-fold, and CYP11B2 was the most induced,
by more than 20-fold at the greatest concentration (3.75 ␮M).
No significant induction of 17␤HSD4 and HMGR gene expression was observed on exposure to the lowest concentration
(0.06 ␮M), but significantly greater gene expression was observed at the median concentration (0.5 ␮M).
After PBB 169 exposure, 3␤HSD2 was the most stimulated
among the steroidogenic genes examined, with a 16-fold induction at the greatest concentration (4.0 ␮M) (Fig. 4B). The
CYP11A, CYP11B2, CYP17, CYP19, and 17␤HSD1 genes
also were up-regulated depending on the concentration (Fig.
4B). However, complex gene expression profiles were observed for 17␤HSD1, StAR, and HMGR, in which both upregulation and suppression of gene expression were observed
depending on the exposure concentrations.
L. Ding et al.
Fig. 4. The effects of polybrominated biphenyls (PBBs) on steroidogenic gene expression in the H295R cells. (A) PBB 77. (B) PBB 169.
Expression of steroidogenic genes was normalized to the expression
of ␤-actin. Fold-induction represents the increase in expression compared to the relevant solvent control. Values are the means of three
determinations on each of three replicate exposures (mean ⫾ standard
deviation.). An asterisk indicates a statistically significant (p ⬍ 0.05)
difference from the dimethyl sulfoxide control. See Figures 1 and 2
for definition of the other abbreviations.
Effects of TBDD/TBDF on adrenocortical steroidogenesis
Only CYP11B2 and 3␤HSD2 were up-regulated after exposure to TBDD (⬎1.5-fold) (Fig. 5A). Expression of the
CYP11A, CYP17, CYP21, and 17␤HSD4 genes was significantly suppressed by 1.6- to 2.2-fold. None of the other genes
was induced compared to the control.
The TBDF exposure induced 3␤HSD2 expression by 15to 20-fold (Fig. 5B). A lesser concentration (0.005 ␮M) induced greater mRNA expression in CYP11A, CYP11B2,
CYP19, 17␤HSD1, 17␤HSD4, StAR, and HMGR. The CYP17
gene was not affected, and a 1.8-fold suppression of CYP21
was observed with exposure to all concentrations (0.005, 0.05,
and 0.5 ␮M) (Fig. 5B). Moreover, a good concentration-dependent response was observed for CYP19, 17␤HSD4, and
HMGR.
DISCUSSION
In the present study, we used the H295R cells to demonstrate the effects of BPs, PBBs, and TBDD/TBDF on mRNA
expression of steroidogenic genes. Previous studies have
shown that these toxicants can affect thyroid hormones, the
aryl hydrocarbon receptor, and the estrogen receptor and,
hence, have endocrine-disruptive activities [16,34]. Therefore,
the present study further supports the mechanism of endocrine
disruption via a non–receptor mediated pathway as indirect or
direct stimulators or inhibitors of the key enzymes involved
in the metabolism of steroid hormones [24]. By using H295R
cells, Sanderson et al. [24] studied the induction of CYP19
activity by various classes of pesticides, suggesting that the
Gene expression induced by BRFs in H295R cells
Fig. 5. The effects of bromodioxin/furan on steroidogenic gene expression in the H295R cells. (A) 2,3,7,8-Tetrabromodibenzo-p-dioxin
(TBDD). (B) 2,3,7,8-Tetrabromodibenzofuran (TBDF). Expression of
steroidogenic genes was normalized to the expression of ␤-actin as
indicated. Fold-induction represents the increase in expression compared to the relevant solvent control. Values are the means of three
determinations on each of three replicate exposures (mean ⫾ standard
deviation). An asterisk indicates a statistically significant (p ⬍ 0.05)
difference from the dimethyl sulfoxide control. See Figures 1 and 2
for definitions of the other abbreviations.
mechanism of aromatase induction is mediated through inhibition of phosphodiesterase activity.
Among the steroidogenic genes examined, the 3␤HSD2
gene was up-regulated the most by almost all the tested chemicals. The gene is responsible for oxidation and isomerization
of 5-ene-3␤-hydroxy steroids to 4-ene-3-ketosteroids, an obligate step in the biosynthesis of not only androgens and estrogens but also mineralocorticoids and glucocorticoids [35].
In humans, two types of 3␤HSD (3␤HSD1 and 3␤HSD2) have
been identified, of which 3␤HSD2 is exclusively expressed in
the adrenal cortex and gonads [36]. The present results demonstrated that bromophenols, in addition to showing thyroid
hormone–like activity [16] and weakly binding to human estrogen receptor (e.g., 2,4-DBP) [17], exert endocrine-disruptive potential by affecting steroidogenesis in vitro. It should
be noted that the gene expression profiles varied depending
on the number and position of substituted bromines as well as
on the dosing concentrations.
Bromophenols also may affect neuroendocrine function
through other mechanisms. For example, a recent study using
a neuroendocrine cell line (PC12) showed that bromophenols
(e.g., 2,4-DP and 2,4,6-TBP) disrupt the endocrine system by
affecting cell calcium homeostasis [37]. In the present study,
PBB 169 and TBDF induced greater gene expression (⬎15fold and ⬎25-fold, respectively), whereas PBB 77 and TBDD
induced moderate gene expression (⬎4-fold and ⬎2.5-fold,
respectively), suggesting greater inductive potency of the two
compounds. In contrast, exposure of the H295R cells to PCB
149 resulted in a moderate induction of 3␤HSD1 (greater than
Environ. Toxicol. Chem. 26, 2007
769
twofold), whereas 3␤HSD2 was comparably up-regulated by
PCBs (PCBs 149, 110, and 101) [32]. The discrepancy of the
magnitude in gene expression may suggest that the inductive
potencies of PBBs are stronger than those of PCBs. Clearly,
exposure respective structure of chlorinated and brominated
biphenyls would compare the inductive potency of these compounds. In addition to up-regulation of 3␤HSD2, suppression
of 3␤HSD2 expression also was observed in the H295R cells
after exposure to sediment extracts containing mainly PAHs
and persistent chlorinated chemicals [33]. It should be noted
that differential expression of 3␤HSD2 and 3␤HSD1 also was
observed in the present study, suggesting distinct functional
roles of the two isogenes in steroidogenesis. Nevertheless, the
up-regulation of 3␤HSD2 in the present study suggested that
the tested compounds could alter sex steroid hormone synthesis.
In addition to the 3␤HSD2 enzyme, many others catalyze
the biosynthetic pathway from cholesterol to cortisol, including CYP11A, CYP11B, CYP17, and CYP21 [36]. The
CYP11A, also known as CYP450scc, is responsible for catalyzing the side-chain cleavage of cholesterol to pregnenolone
and also is the first step controlling the conversion of cholesterol to steroid hormones. In the present study, most of the
tested chemicals significantly induced the expression of this
gene, whereas the significant inhibition of this gene was observed after TBDD exposure.
The CYP11B2 enzyme is responsible for the rate-limiting
final conversion in the aldosterone biosynthetic process. In the
present study, PBB 169 caused significant dose-dependent induction of the CYP11B2 gene. Most of the tested chemicals
altered the expression of the CYP11B2 gene at different concentrations. Because CYP11B enzymes are important in adrenal steroid synthesis, their modulation could significantly
alter various physiological processes controlled by cortisol and
aldosterone. Using H295R cells, Li et al. [27] showed that
CYP11B2 is sensitive to coplanar PCB 126 and that significantly increased expression was accompanied by elevated production of aldosterone. Likewise, a recent study by Xu et al.
[32] showed that the greatest transcriptional activation of this
gene (29.9-fold) was caused by PCB 110 in H295R cells. These
results suggested that PCB congeners may exert toxic effects
by increasing aldosterone production. In the present study,
PBBs 77 and 169 also greatly induced CYP11B2 gene expression (⬎20-fold and 7-fold, respectively), which suggests
that some PBB congeners may have similar toxic mechanism
to PCBs.
The CYP17 enzyme catalyzes the conversion of aldosterone
to corticosteroid substrates, and for cortisol synthesis in adrenocortical cells. In the present study, gene expression patterns varied depending on the toxicants and concentrations.
For example, 2,4,6-TBP and TBDD suppressed CYP17 gene
expression, whereas 2-BP, 2,4-DBP, PBP, and PBB resulted in
up-regulation. Nakajin et al. [30] showed that some alkylphenols, such as 4-t-octylphenol and 4-nonylphenol inhibited
CYP17 enzyme activity in H295R cells. Although fewer studies have addressed the effects of brominated compounds on
CYP17, Moran et al. [38] showed that 2,3,7,8-TCDD significantly inhibited CYP17 activity, which in turn led to a proportional decrease in estradiol secretion in human luteinizing
granulosa cells. A more recent study [20] showed that CYP17
activity was significantly inhibited by almost all tested hydroxylated PBDEs (e.g., 4-OH-BDE 49, 6-OH-BDE 47, and
6-OH-BDE 99) in H295R cells, and those authors suggested
770
Environ. Toxicol. Chem. 26, 2007
that the inhibition of CYP17 activity would result in a decrease
of the synthesis of weak androgens and, consequently, affect
testosterone and estradiol production in the testis and ovary,
respectively.
The CYP21 gene product is required for the synthesis of
both aldosterone and corticosteroids. The suppression of this
gene would likely result in deficiencies of both cortisol and
aldosterone and be accompanied by overproduction of androgens. In the present study, 2,4,6-TBP moderately suppressed,
and PBP induced, CYP21 gene expression, whereas other
bromophenols showed no significant responses, suggesting
that CYP21 generally was not sensitive to bromophenols.
However, TBDF and TBDD suppressed CYP21 gene expression, but PBB 77, PBB 169, and PBP resulted in significantly
increased expression. An increase in CYP21 gene expression
may lead to an increase in the synthesis of cortisol and aldosterone, which could result in an increase in the substrate
available for androgen and estrogen production. Nakajin et al.
[30] showed that 4-t-octylphenol and 4-nonylphenol inhibited
CYP21 enzyme activity in H295R cells, indicated that some
alkylphenols behaved as inhibitors of cortisol synthesis, and
showed good correlation between cortisol secretion and inhibition of the enzyme activity. Another study showed that
CYP21 gene expression also was significantly suppressed on
exposure to sediment extracts in H295R cells [33]. In contrast,
CYP21 gene expression was induced by PCBs congeners (e.g.,
PCBs 101, 110, and 149) and by some of the metabolites
(methylsulfonyl PCBs) in H295R cells [32].
In addition to expression of the 3␤HSD2 gene, expression
of the 17␤HSD gene also was significantly affected by bromophenols. The 17␤HSD enzyme plays a key role in the final
step of sex steroid biosynthesis, which controls estrogen and
androgen concentrations. Up-regulation of this gene may lead
to elevated estrogen production. Likewise, depending on the
number and position of bromines, the levels of up-regulation
significantly varied for 17␤HSD1 and 17␤HSD4. In the present
study, we observed that 2,4-DBP and PBP induced greater
gene expression, whereas 2,4,6-TBP showed no induction for
both 17␤HSD1 and 17␤HSD4 and PBP had no effect on
17␤HSD4. In addition, PBB 77 significantly induced expression of the two genes. Hence, increased gene expression would
likely affect estrogen biosynthesis and could lead to elevated
estrogen production. In addition, 2,3,7,8-TBDD resulted in a
slight but significant suppression of 17␤HSD4 gene expression.
Among the examined genes, CYP19 (aromatase) has received great attention. The CYP19 enzyme catalyzes the conversion of androgens to estrogens, is expressed in various tissues, and is directly involved in reproduction, behavior, development, and estrogen-dependent carcinogenesis [19].
Hence, it has been hypothesized that the ability of chemicals
to alter CYP19 enzyme activity represents a potential mechanism of endocrine disruption [31]. Indeed, the effects of a
wide range of chemicals on this enzyme have been studied
extensively in H295R cells [19,24,28]. One recent study
showed that 0.5 to 7.5 ␮M 2,4,6-TBP caused a concentrationdependent induction of aromatase activity, whereas 2.5 to 7.5
␮M 4-BP did not show any induction in H295R cells [19]. In
the present study, most bromophenols significantly induced
CYP19 gene expression. The present results therefore showed
that bromophenols potentially interfere in steroidogenesis in
vitro. In the present study, no gene induction was observed
after 2,4-DBP and 2,6-DBP treatment, suggesting different
L. Ding et al.
responses caused by different bromine substitution. Additionally, Cantón et al. [19] measured CYP19 activity in H295R
cells exposed to 30 PBDE congeners (7.5 ␮M) and found that
only BDEs 19 and 28 slightly induced, and BDEs 206 and
209 slightly inhibited, the activity of this enzyme. In the present study, PBB 77, PBB 169, and TBDF all significantly induced CYP19 gene expression, whereas gene expression in the
TBDD treatment remained unchanged, suggesting that TBDF
is a more potent inducer than TBDD. Although measurement
of gene expression patterns offers information at the mRNA
level, measurement of gene expression together with enzyme
activities and steroid hormone concentrations, such as testosterone and estradiol, would provide more detailed information
regarding the effects of chemicals on steroidogenesis. Therefore, based on the present results, further research may be
needed to examine whether these compounds will affect the
same pathway in vivo.
The StAR protein transports cholesterol across the mitochondrial membrane for conversion to pregnenolone by
CYP11A and, thus, controls the rate-limiting step in steroidogenesis. In the present study, most bromophenols induced upregulation of StAR gene expression. However, PBB 169 induced the StAR gene at 0.001 ␮M, whereas a down-regulation
of this gene was clearly observed with exposure to greater
concentrations (0.06 an d 4.0 ␮M). These responses may be
a biphasic response of the gene: The lesser concentration of
the test chemical induced increased gene expression, whereas
at greater concentrations, gene expression was suppressed. Sugawara et al. [39] reported that StAR promoter activity increased with ␤-naphthoflavone concentration, reached a peak,
and then subsequently decreased at increased concentrations.
A recent study showed that StAR gene expression was suppressed by several PCBs [31]. Although StAR gene expression
was up- or down-regulated after exposure to toxicants, Sugawara and Fujimoto [40] demonstrated that the increase in
StAR protein level in monkey kidney cell line (COS-1) is not
a result of an increase in StAR gene expression but, rather, is
a result of both an increase in translation and a longer halflife of the StAR protein.
The HMGR enzyme catalyzes the rate-limiting reaction of
cholesterol biosynthesis, and its expression is regulated by
various hormonal factors, such as concentrations of cholesterol, thyroid hormone, and estradiol [41]. Therefore, HMGR
functions as a cholesterol regulator. Likewise, several of the
tested compounds affected its gene expression depending on
the nature of toxicants, dose concentrations, and substituted
bromines. Among the tested compounds, TBDF induced the
greatest responses, whereas no significant effect of TBDD was
observed. A recent study showed that HMGR gene expression
remained relatively unchanged after exposure to PCBs [32].
In general, transcriptional levels of HMGR and StAR mRNA
were not markedly altered. This is consistent with the results
of a previous study [25], in which the authors hypothesized
that the basal abundance of the mRNA was relatively small.
Because small changes in gene expression may lead to large
changes in the steroidogenic pathway and complex control
points and because multiple enzymes are involved, examination of gene expression together with the measurement of enzyme activity and hormone levels would provide more information regarding the effects of chemicals on steroidogenesis.
In conclusion, the present study demonstrated that some
BFRs are able to alter steroidogenic gene expression in the
H295R human adrenocortical carcinoma cells. Most of the
Environ. Toxicol. Chem. 26, 2007
Gene expression induced by BRFs in H295R cells
genes were either up- or down-regulated, to some extent, by
exposure to the compounds, and 3␤HSD2 induction was the
most pronounced. The present study further supports the mechanism of chemical-induced endocrine disruption via modulation of steroidogenic genes, which could interfere with steroid
hormone production and, in turn, affect hormone balance and
immune function. The H295R cell line provided a good in
vitro model for screening endocrine-disruptive chemicals. Although gene expression was significantly altered by many of
the tested compounds, measurement of the activities of the
relevant enzymes involved in steroidogenesis along with measurement of specific hormone concentrations would provide
more information regarding chemical-induced endocrine disruption. The ubiquity of BFRs, including PBDEs and their
metabolites, new emerging bromodioxins/furans, and mixtures
of chlorodioxins/furans, warrants further study for testing their
endocrine activities. In addition, examining the potential endocrine activities of these compounds in vivo will improve
our understanding of their effects.
Acknowledgement—We would like to thank the Knowledge Innovation Program from the Chinese Academy of Sciences (project KSCX2SW-128), and the City University of Hong Kong matching fund to
Area of Excellence Scheme under the University Grants Committee
of the Hong Kong Special Administrative Region, China (project
9400001), to R. Wu for financial support.
14.
15.
16.
17.
18.
19.
20.
21.
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