ELSEVIER
Effects of fifteen PBDE metabolites, DE71, DE79 and TBBPA on steroidogenesis in the H295R cell line
Renfang Song a,d,e, Yuhe He ª, Margaret B. Murphy ª, Leo W.Y. Yeung ª, Richard M.K. Yua, Michael H.W. Lamª, Paul K.S. Lama,*, Markus Hecker b,f, John P. Giesy a,b,c, Rudolf S.S. Wua, Wenbing Zhangd, Guoying Shengª, Jiamo Fud
a Department of Biology and Chemistry, City University of Hong Kong, 83 Tat Chee Avenue, Kowloon, Hong Kong
b Department of Veterinary Biomedical Sciences and Toxicology Centre, University of Saskatchewan, Saskatoon, Saskatchewan, Canada S7K 3J8 ” Department of Zoology, National Food Safety and Toxicology Center, Center for Integrative Toxicology, Michigan State University, East Lansing, MI 48824, USA
d State Key Laboratory of Organic Geochemistry, Guangzhou Research Center of Mass Spectrometry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China
e Graduate School of the Chinese Academy of Sciences, Beijing, China ENTRIX, Inc., RR5 Site 515 Box 15, Hidden Ridge Estates, Saskatoon, Canada SK S7K 3J8
Received 16 October 2007; received in revised form 5 January 2008; accepted 17 January 2008 Available online 3 March 2008
Abstract
Polybrominated diphenyl ethers (PBDEs) and tetrabromobisphenol A (TBBPA) are brominated flame retardants that are produced in large quantities and are commonly used in construction materials, textiles, and as polymers in electronic equipment. Environmental and human levels of PBDEs have been increasing in the past 30 years, but the toxicity of PBDEs is not fully understood. Studies on their effects are relatively limited, and show that PBDEs are neurotoxins and potential endocrine disrupters. Hydroxylated (OH-) and meth- oxylated (MeO-) PBDEs have also been reported in the adipose tissue, blood and milk of wild animals and humans. In the present study, 15 PBDE metabolites, two BDE mixtures (DE71 and DE79), and TBBPA were studied individually to determine their effects on ten steroidogenic genes, aromatase activity, and concentrations of two steroid hormones (testosterone and 17ß-estradiol) in the H295R human adrenocortical carcinoma cell line. Exposure to 0.05 uM 2’-OH-BDE-68 significantly induced the expression of CYP11A, CYP11B2, CYP17, CYP21, 3ßHSD2,17BHSD1, and 17ßHSD4, and the expression of StAR was induced by 6-OH-BDE-90 at the three exposure concentrations. Exposure to DE71 and DE79 resulted in dose-dependent trend towards induction, but these effects were not significant. Exposure to 0.5 uM 2-OH-BDE-123 and 2-MeO-BDE-123 resulted in significantly greater aromatase activity. However, none of the compounds affected sex hormone production at the concentrations tested. Generally, OH-BDEs had a much stronger ability to affect steroidogenic gene expression than MeO-BDEs. @ 2008 Elsevier Ltd. All rights reserved.
Keywords: PBDEs; TBBPA; Steroidogenesis; H295R cells
1. Introduction
Polybrominated diphenyl ethers (PBDEs) and tetrabro- mobisphenol A (TBBPA) are commonly used as bromi- nated flame retardants (BFRs) in construction materials,
textiles, and as polymers in electronic equipment (Brown et al., 2004). These compounds are hydrophobic, stable and persist in the environment and have the potential to bioaccumulate (Hites, 2004; Birchmeier et al., 2005). PBDEs have been found in the environment, wildlife, as well as in human blood, tissues, hair and breast-milk (de Wit, 2002), and TBBPA has been found in surface waters, wastewater and sediments (Blanco et al., 2005; Verslycke
*Corresponding author. Tel .: +852 2788 7681; fax: +852 2788 7406. E-mail address: bhpksl@cityu.edu.hk (P.K.S. Lam).
et al., 2005). The most predominant PBDE congeners mea- sured in humans are BDE 47, followed by BDEs 28, 99, and 100. DE71, a pentabrominated mixture, and DE79, an octabrominated mixture, are also commonly found in biological samples. Less brominated PBDEs, such as tetra-, penta- and hexa-BDEs, demonstrate high affinity for lipids and can bioaccumulate in the tissues of wildlife and humans (Zhou et al., 2001). The use of DE71 and DE79 has been banned in Europe and in several states in the US, but there are no restrictions on the use of the deca-BDE mixture and TBBPA (EU, 2003; BSEF, 2006).
The presence of methoxylated (MeO-) and hydroxyl- ated (OH-) PBDEs in blood, adipose tissues, and liver of fish, birds, and mammals has been reported in a few studies (Marsh et al., 2004; Sinkkonen et al., 2004; Teuten et al., 2005). Current understanding indicates that MeO- BDEs found in wildlife are the consequence of accumula- tion via natural sources in the marine environment (Marsh et al., 2004; Teuten et al., 2005), whereas OH-BDEs can have a natural origin and/or result from metabolism of PBDEs (Örn and Klasson-Wehler, 1998; Letcher et al., 2000; Hakk and Letcher, 2003).
The effects of PBDE and TBBPA exposure have been assessed in several systems. Some studies have suggested that tetra- and penta-BDEs were more toxic and bioaccu- mulative compounds compared with octa- and deca-cong- eners (Siddiqi et al., 2003). PBDEs have been shown to bind to the AhR as measured using ethoxyresorufin-O- deethylase (EROD) activity (Meerts et al., 2001; Mundy et al., 2004), and some PBDEs were found to activate the AhR signal transduction pathway at moderate to high con- centrations (Brown et al., 2004). Exposure to PBDEs induced thyroid hyperplasia and altered thyroid hormone production and transport in vitro (Marsh et al., 1998; Meerts et al., 2000, 2001; Siddiqi et al., 2003). As the struc- ture of PBDEs is similar to those of PCBs, PBDEs and their metabolites may act as endocrine disrupters by inter- fering with thyroid hormone homeostasis (Hooper and McDonald, 2000; Zhou et al., 2001). Previous studies sug- gested that liver, thyroid gland, pancreas, and kidney could be the endpoints for identifying endocrine disrupters of cytochrome P450 isozymes (Hallgren et al., 2001; Darne- rud et al., 2005). Some PBDEs and their derivatives (including OH-BDEs and MeO-BDEs) were reported to be able to induce or inhibit aromatase (CYP19) and CYP17 activity in H295R cells (Cantón et al., 2005, 2006). TBBPA is also considered a potential endocrine dis- ruptor (Kitamura et al., 2002, 2005). An in vitro study of BFRs showed that TBBPA binding to TTR was ten times more effective than that of T4 to TTR (Meerts et al., 2000). However, information about the endocrine activity of PBDE metabolites or derivatives is still limited.
Despite the widespread occurrence of PBDEs and TBBPA in the environment, limited information is avail- able on their effects on steroidogenesis (Zhou et al., 2002; Darnerud et al., 2005). The H295R human adrenocortical carcinoma cell line has been characterized in detail and
shown to express all key enzymes necessary for steroido- genesis (Gazdar et al., 1990). This system has been used in mechanistic studies and to provide relevant data for risk assessment based on measuring effects on specific enzymes and hormone production (Hecker et al., 2006, 2007). Therefore, this cell line can be used as a model for in vitro screening for both adrenocortical toxicity and ste- roidogenesis. In the present study, the expression of 10 key steroidogenic genes was measured (CYP11A, CYP11B, CYP17, CYP19, CYP21,3ßHSD2, 17BHSD1, 17ßHSD4, HMGR and StAR) using quantitative real-time PCR (Q- RT-PCR). This technique offers a more specific, sensitive, comprehensive, and reliable method for the investigation of multiple functionally related genes (Sanderson et al., 2002; Hilscherova et al., 2004; Zhang et al., 2005).
To date, there is little information on the effects of PBDEs and their derivatives relevant to human exposure levels (1-462 ng PBDEs/g in adipose tissue) (Gill et al., 2004). The purpose of this study is to determine the effects of 15 PBDE metabolites, two mixtures, and TBBPA on ste- roidogenesis using the H295R cell line at relevant human exposure levels.
2. Materials and methods
2.1. Test chemicals
Of the chemicals tested, the two commercial mixtures (DE71 and DE79) and TBBPA were purchased from Sigma Chemical Co. (St. Louis, MO), while the PBDE metabolites were synthesized in the Department of Biology and Chemistry of City University of Hong Kong according to the methods described in Marsh et al. (2003) with puri- ties >98%. Each compound was dissolved in dimethyl sulf- oxide (DMSO), and the final DMSO concentration in the exposure medium was 0.5% (v/v).
2.2. Cell culture
H295R cells were obtained from the American type cul- ture collection (ATCC CRL-2128; ATCC, Manassas, VA) and cultured in 75 cm2 flasks at 37 ℃ in a 5% CO2 atmo- sphere. The culture medium was a 1:1 mixture of Dul- becco’s modified Eagle’s medium and Ham’s F-12 Nutrient mixture (Sigma Chemical Co., St Louis, MO), supplemented with 1% ITS + Premix (BD Biosciences, San Jose, CA), 2.5% Nu-Serum (BD Biosciences), 0.5% antibiotics (5000 µg/ml penicillin and 5000 µg/ml strepto- mycin), and 1.2 g/l Na2CO3. The medium was changed twice a week, and cells were detached from flasks for sub- culturing using trypsin/EDTA (Life Technologies Inc., Grand Island, NY).
2.3. Cell exposures
In order to measure the effects of BFR exposure on ste- roidogenic gene expression, H295R cells were plated onto
6-well culture plates at an initial cell density of 1 × 106 cells/ml in 2 ml cell suspension per well. After 24 h, cells were exposed to chemicals at concentrations of 0.025, 0.05 and 0.5 uM for another 48 h and then total RNA was isolated from the cells. For the aromatase and hormone measurements, cells were plated in 24-well culture plates at a density of 3 × 105 cells/ml in 1 ml per well. After 24 h, the medium was replaced and cells were exposed to chemicals at the concentrations listed above for another 48 h, after which the culture medium was collected and fro- zen at -80 ℃ for hormone measurement. The cells in the plate were gently washed twice with PBS before being used in the aromatase assay.
2.4. Cell viability assay
H295R cells were seeded into 96-well view plates at a concentration of 1 × 10° cells/ml in 200 ul of medium per well. After 24 h, cells were dosed with the PBDE metabo- lites. Cell viability was measured using the sulforhodamine B (SRB) assay after 48 h of exposure (Körner et al., 1998). Absorbance was subsequently measured using a plate-read- ing measurement system (Molecular Devices Spectra Max 340 PC).
2.5. Real-time PCR assay
The expression levels of 10 steroidogenic genes plus one housekeeping gene (ß-actin) were measured following Hil- scherova et al. (2004). Total RNA was isolated from exposed cells using the SV total RNA isolation system (Promega, San Luis, CA). Two microgram of cellular total RNA for each sample was used for reverse transcription using the SuperScript™M First-Strand Synthesis System for RT-PCR (Invitrogen, Carlsbad, CA). The ABI 7500 fast real-time PCR System (Applied Biosystems, Foster City, CA) was used to perform quantitative real-time PCR. PCR reaction mixtures (20 ul) contained 1 ul (0.2- 0.4 uM) of forward and reverse primers, 5 ul of cDNA sample, and 10 ul of 2x SYBR Green™M PCR Master Mix (Applied Biosystems). The thermal cycle profile was denatured at 95 ℃ for 10 min, followed by 40 cycles of denaturation for 15 s at 95℃, annealing together with extension for 1 min at 60 ℃; and a final cycle of 95 ℃ for 15 s, 60 ℃ for 1 min, and 95 ℃ for 15 s. Melting curve analyses were performed during the 60 ℃ stage of the final cycle to differentiate between desired PCR products and primer-dimers or DNA contaminants.
To quantify the RT-PCR results, the cycle at which the fluorescence signal was first significantly different from background (Ct) was determined for each reaction. The expression level of a target gene was normalized with refer- ence to the ß-actin endogenous control gene to derive the mean normalized expression (MNE) value (Eq. (1))
MNE = (Ereference) CT (Etarget ) CT reference, mean
target, mean ;
(1)
where Ereference and Etarget represent the PCR efficiencies (=10-1/slope) determined from the slopes of the standard curves constructed using gene-specific RNA standards of known copy number (Simon, 2003). Gene expression levels were measured in triplicate for control and exposed cells. Levels of expression relative to solvent control were calcu- lated (Eq. (2))
N-fold change = MNEexp/MNEcon- (2)
2.6. Aromatase activity assay
The activity of CYP19 (aromatase) was determined based on the method described by Lephart and Simpson (1991) with modifications. After medium removal and washing, the cells were exposed to 54 nM [1]-3H] andro- stenedione (PerkinElmer, Boston, MA) in serum-free med- ium for 90 min at 37 ℃ and 5% CO2 (Sanderson et al., 2002). Two hundred microlitre of the medium was extracted and used for measuring the level of radioactivity. Corrections were made for background radioactivity, dilu- tion factor, and specific activity of the substrate. 8-Bromo- cyclic-adenosine-monophosphate (8-Br-cAMP, 100 µM) was used as a positive control/aromatase inducer, while 4-hydroxyandrostenedione (4-HA, 10 µM) was used as a negative control/aromatase inhibitor, as described previ- ously (Heneweer et al., 2004).
2.7. Hormone measurement
Testosterone (T) and 17ß-estradiol (E2) concentrations were measured using methods described by Hecker et al. (2006). Frozen medium was thawed on ice, and 500 ul medium was extracted twice with 2.5 ml diethyl ether. All samples were spiked with 10 ul 1,2,6,7-3H-labeled T (0.0002 µCi/ul) (PerkinElmer) prior to extraction to deter- mine extraction recoveries. The ether phase containing the target hormones was evaporated under nitrogen, and the dried extract was reconstituted in 250 ul ELISA assay buf- fer and frozen at -80 ℃ for subsequent analysis. The extracts were diluted 1:50 and 1:5 for T and E2 analysis, respectively, and the hormones levels were measured by competitive ELISA following manufacturer protocols (Cayman Chemical Company, Ann Arbor, MI).
2.8. Statistical analyses
All experiments were conducted in duplicate, and tripli- cate measurements were used for each individual exposure. Statistical analyses were conducted using SPSS 13 (SPSS Inc., Chicago, IL, USA). As the results were found to be reproducible and consistent between exposures, data from one representative exposure are presented as means ± stan- dard deviations. Differences in gene expression, aromatase activity, and hormone production between control and exposed cells were evaluated by ANOVA with post-hoc
Dunnett’s tests. Differences with p < 0.05 were considered significant.
3. Results
3.1. Cytotoxicity
None of the tested compounds were found to be cyto- toxic at the exposure concentrations tested.
3.2. Effects of PBDEs and TBBPA on the expression of CYP family genes
Fifteen PBDE metabolites, two BDE mixtures (DE71 and DE79), and TBBPA were examined individually at three concentrations to study their effects on ten steroido- genic genes in H295R cell line. Five chemicals had no signif- icant effect on any of these genes at all three concentrations (Table 1). CYP11A expression was only up-regulated by exposure to 0.05 µM 2’-OH-BDE-68 (1.7-fold). Exposure to 6-MeO-BDE-47 (0.025 uM), 5-C1-6-OH-BDE-47 (0.5 µM), 5-C1-6-MeO-BDE-47 (0.025 µM), and 2’-OH- BDE-68 (0.05 µM) significantly up-regulated CYP11B2 expression by 1.4-, 1.4-, 1.4-, and 2.1-fold, respectively. In contrast, exposure to 6-OH-BDE-47 (0.025, 0.05, and 0.5 µM) and 2-OH-BDE-123 (0.05 and 0.5 µM) signifi- cantly down-regulated CYP11B2 expression. The expres- sion of CYP17 was up-regulated 1.7-fold by exposure to 0.05 µM 2’-OH-BDE-68, whereas its expression was down-regulated by exposure to 5-Cl-6-OH-BDE-47 (0.025 µM), 5-C1-6-MeO-BDE-47 (0.5 µM), and 4’-OH- BDE-49 (0.05 and 0.5 µM). Exposure to 0.05 µM 6-OH- BDE-47 resulted in up-regulation of CYP19 expression by 1.4-fold. Moreover, CYP21 expression was also up-regu- lated by exposure to 6-OH-BDE-47, 2’-OH-BDE-68, 6’- C1-2’-OH-BDE-68, 2-MeO-BDE-123, DE71, DE79, and TBBPA at various concentrations.
3.3. Effects on hydroxysteroid dehydrogenase gene expression
Among the 18 chemicals tested, only 2’-OH-BDE-68 affected hydroxysteroid dehydrogenase gene expression. Exposure to 0.05 µM 2’-OH-BDE-68 resulted in significant up-regulation of 3ßHSD2, 17BHSD1, and 17BHSD4 expression by 2.1-, 2.0-, and 1.8-fold, respectively (Table 1).
3.4. Effects on StAR and HMGR gene expression
Exposure to 2’-OH-BDE-68 up-regulated the expression of StAR by 1.7-fold, while 6-OH-BDE-47 down-regulated its gene expression by about 0.3-fold. Moreover, the expression of StAR was up-regulated (2.2-fold) by expo- sure to 6-OH-BDE-90 at all three concentrations tested. In contrast, HMGR expression did not change in any of the treatments (Table 1).
| No. | Chemicals | Genes affected | Magnitude of effect (1/4 fold-change) | Effective concentration (HM) |
|---|---|---|---|---|
| OH-BDEs | ||||
| 1 | 2'-OH-BDE-25 | None | N.e. | N/A |
| 2 | 2'-OH-BDE-28 | None | N.e. | N/A |
| 3 | 6-OH-BDE-47 | CYP11B2 | 40.3 | 0.5, 0.05, 0.025 |
| CYP21 | 11.5 | 0.5 | ||
| CYP19 | 11.4 | 0.05 | ||
| StAR | 10.3 | 0.025 | ||
| 4 | 5-Cl-6-OH-BDE-47 | CYP17 | Į0.4 | 0.025 |
| CYP11B2 | 11.4 | 0.5 | ||
| 5 | 4'-OH-BDE-49 | CYP17 | 10.4 | 0.5, 0.05 |
| 6 | 2'-OH-BDE-68 | CYP11A | 11.7 | 0.05 |
| CYP11B2 | 12.1 | 0.05 | ||
| CYP17 | 11.7 | 0.05 | ||
| CYP21 | 11.9 | 0.05, 0.025 | ||
| 3฿HSD2 | 12.1 | 0.05 | ||
| 17ßHSD1 | 12.0 | 0.05 | ||
| 17ßHSD4 | 11.8 | 0.05 | ||
| StAR | 11.7 | 0.05 | ||
| 7 | 6'-C1-2'-OH-BDE-68 | CYP21 | 11.6 | 0.05 |
| 8 | 6-OH-BDE-85 | None | N.e. | N/A |
| 9 | 6-OH-BDE-90 | StAR | 12.2 | 0.5, 0.05, 0.025 |
| 10 | 2-OH-BDE-123 | CYP11B2 | 40.3 | 0.5, 0.05 |
| MeO-BDEs | ||||
| 11 | 6-MeO-BDE-47 | CYP21 | 11.4 | 0.5 |
| CYP11B2 | 11.4 | 0.025 | ||
| 12 | 5-Cl-6-MeO-BDE-47 | CYP11B2 | 11.4 | 0.5 |
| CYP17 | 10.4 | 0.5 | ||
| 13 | 4'-MeO-BDE-49 | None | N.e. | N/A |
| 14 | 2'-MeO-BDE-68 | None | N.e. | N/A |
| 15 | 2-MeO-BDE-123 | CYP21 | 11.5 | 0.5 |
| BDE mixtures | ||||
| 16 | DE71 | CYP21 | 11.6 | 0.5 |
| 17 | DE79 | CYP21 | 11.6 | 0.5, 0.05 |
| Other BFR | ||||
| 18 | TBBPA | CYP21 | 11.4 | 0.5 |
Symbols indicate fold-differences relative to control. 1 = increased; Į = decreased; N/A = not applicable; and N.e. = No effect.
3.5. Effects on aromatase activity and sex hormone production
Most of the chemicals tested in this study did not sig- nificantly affect aromatase activity. Exposure to 0.05 µM DE71 resulted in significantly higher aromatase activity (mean activity was 252% of solvent control), and expo- sure to 0.5 M DE71 caused a trend towards induction (393% of solvent control), although the response to this chemical was variable (Fig. 1). Exposure to the octa- BDE mixture DE79 also caused a trend towards induc- tion of aromatase activity. Moreover, exposure to 0.5 µM 2-OH-BDE-123 and 2-MeO-BDE-123 resulted in slightly but significantly greater aromatase activity (Fig. 1). However, none of the tested compounds signifi- cantly affected sex hormone production at the exposure concentrations tested.
700
DE71 O DE79 @ 2-OH-BDE-123 2-MeO-BDE-123
Aromatase activity % (DMSO control = 100%)
600
500
400
*
300
200
*
T
T
*
100
L
L
£
0
0.005 µM
0.05 µM
0.5 µM
Dosed concentrations
4. Discussion
PBDEs are widely used, and both the parent compounds and their metabolites have been shown to have endocrine disrupting effects in several test systems (e.g. Meerts et al., 2001; Cantón et al., 2005; Harju et al., 2007). How- ever, their toxicity, and especially their effects on steroido- genesis, are not fully understood to this point. In the present study, the expression of ten key steroidogenic genes was measured by quantitative RT-PCR to observe the effects of the selected PBDEs, PBDE mixtures, and TBBPA on the H295R cell line. Thirteen of the 18 chemicals tested, especially the OH-BDE compounds, affected the expression of some genes; this observation was similar to those of pre- vious studies in which CYP19 enzyme activity was mea- sured (Cantón et al., 2005, 2008), and a study of the effects of other PBDE metabolites at relatively high expo- sure concentrations (He et al., 2008).
CYP11A catalyzes the side-chain cleavage of choles- terol, which is the starting point of steroid synthesis and it is also the rate-limiting step. Therefore, only a small change in the expression of this gene may have large effects on steroidogenesis. However, in the present study only 2’- OH-BDE-68 caused a significant up-regulation of CYP11A expression at a concentration of 0.05 µM.
CYP11B2 catalyzes the biosynthesis of the glucocorticoid cortisol from 11-deoxycortisol, which has numerous meta- bolic, developmental, immunosuppressive, anti-inflamma- tory and other functions in the body (Hilscherova et al., 2004). In the present study, four OH-BDEs (6-OH-BDE- 47; 5-C1-6-OH-BDE-47; 2’-OH-BDE-68; 2-OH-BDE-123) and two MeO-BDEs (6-MeO-BDE-47; 5-Cl-6-MeO-BDE- 47) significantly affected CYP11B2, demonstrating their potential to interfere with the glucocorticoid synthesis path- way in vitro (Table 1). Exposure to 2’-OH-BDE-68 had the strongest effect on CYP11B2 expression, possibly because its -OH group is adjacent to a bromine atom (Cantón
et al., 2005). This metabolite has been found in several spe- cies, including fish, seals, algae, and seabirds (Olsson et al., 2000; Marsh et al., 2004), and thus its potential effects on endocrine function should be investigated further in vivo.
CYP17 catalyzes the conversion of aldosterone to corti- costeroid substrates and ultimately to sex steroid sub- strates, which is the initial step of cortisol biosynthesis. Therefore, it is possible that this enzyme can redirect ste- roid output from mineralocorticoids to glucocorticoids or weak androgens, whereas inhibition of CYP17 would have the opposite effect. A recent study reported that some OH- or MeO- PBDEs significantly inhibited CYP17 activity in H295R cells (Cantón et al., 2006). In the current study, the fact that exposure to 5-Cl-6-OH-BDE-47 resulted in down-regulation of CYP17 and up-regulation of CYP11B2 expression by 1.4-fold might be evidence of this shift in the steroidogenic pathway.
CYP19 is responsible for the final conversion of andro- gens to estrogens. Only 0.05 µM 6-OH-BDE-47 had an effect on CYP19 expression, and most of the chemicals tested in this study did not affect aromatase activity at any of the three dosed concentrations. Four chemicals, DE71, DE79, 2-OH-BDE-123 and 2-MeO-BDE-123, showed some ability to induce aromatase activity, although the induction levels observed were limited and variable. This result is not unexpected in the context of previous studies, as the dosed concentrations used in the current study were relatively low. A previous study reported that exposure of H295R cells to 6-OH-BDE-47 reduced aroma- tase (CYP19) activity at concentrations greater than 2.5 µM (Cantón et al., 2005), and another recent study reported that exposure of the same cell line to 10 uM 5- Cl-6-OH-BDE-47 significantly decreased aromatase activ- ity and E2 production (He et al., 2008). Therefore, the envi- ronmentally relevant concentrations tested in the current study are likely too low to cause effects on aromatase activ- ity in the H295R cell line.
Exposure to three OH-BDEs, two MeO-BDEs, two BDE mixtures and TBBPA were observed to induce CYP21 expression at different concentrations in the present study (Table 1). The CYP21 gene is required for the synthe- sis of both aldosterone and corticosteroids. Induction of CYP21 may lead to an increase in the synthesis of cortisol and aldosterone, and may result in a decrease in the avail- ability of substrates for androgen and estrogen production. In the present study, CYP21 was the only gene affected by TBBPA. The lack of effects caused by TBBPA is most likely due to its short biological half life (WHO/ICPS, 1995). This substrate effect may be reflected in the aroma- tase induction that was observed in cells exposed to DE71, DE79 and 2’-MeO-BDE-123; more of the aromatase enzyme may be produced by the cells as a way to compen- sate for a reduction in the amount of androgens available for aromatization, although no change in CYP19 expres- sion was observed after exposure to these compounds.
HSD enzymes are members of the short-chain alcohol dehydrogenase (SCAD) enzyme family. HSD enzymes
mainly catalyze two reactions: (1) the oxidation of a sec- ondary alcohol to a ketone; and (2) the reduction of a ketone to a secondary alcohol. 3ßHSD oxidizes a 3ß-OH group to a C3 ketone that is an obligate step in the biosynthesis of androgens and estrogens as well as mineral- ocorticoids and glucocorticoids. 17ß-Hydroxysteroid dehy- drogenases (17HSDs) are a group of enzymes responsible for the interconversion between low-activity 17-ketoste- roids and high-activity 17ß-hydroxysteroids. They act as key enzymes modulating the biosynthesis and metabolism of both estrogens and androgens. In the present study, 2’- OH-BDE-68 significantly increased the expression of 3฿HSD2, 17BHSD1 and 17ßHSD4 by 1.8- to 2-fold, but no clear dose-response relationship was observed.
The protein encoded by the StAR gene plays a key role in the acute regulation of steroid hormone synthesis. It enhances the conversion of cholesterol to pregnenolone by mediating the transport of cholesterol from the outer mitochondrial membrane to the inner mitochondrial mem- brane. Exposure to 2’-OH-BDE-90 caused 2.2-fold induc- tion of StAR expression at all three concentrations, and exposure to 2’-OH-BDE-68 also up-regulated StAR by 1.7-fold. However, exposure to 6-OH-BDE-47 reduced StAR expression by about 0.3-fold. MeO-BDEs showed no effect on StAR expression.
These effects on gene expression support the idea that functional group position,, especially -OH group position, might play an important role in the effects of PBDE metab- olites on steroidogenesis; other studies have also reported that hydroxylated PBDEs have stronger biological effects than either the parent compound or metabolites with other substituents (Meerts et al., 2000, 2001; Cantón et al., 2005, 2008; Harju et al., 2007), possibly due to the role of the -OH group as a hydrogen donor or acceptor (Harju et al., 2007). However, no consistent relationships between -OH group position and steroidogenic effects were found in either the present study or in previous studies (Cantón et al., 2005, 2008). For example, both the tetrabrominated 2’-OH-BDE-68 and 5-OH-BDE-47 have an -OH group adjacent to a bromine atom, but 2’-OH-BDE-68 induced the expression of several genes at 0.025 to 0.5 uM in the current study, while 5-OH-BDE-47 had no effect. Similar inconsistent effects have been reported for methoxylated PBDE metabolites; exposure to 10 µM 2’-MeO-BDE-68 significantly induced aromatase activity in the H295R cells in a previous study, but exposure to 5-MeO-BDE-47 did not (He et al., 2008). It is possible that the reason that structural relationships are difficult to determine for these compounds lies in differences in their metabolism, as some PBDE metabolites may be degraded within minutes while others are more persistent (Harju et al., 2007). Some hydroxylated PBDEs have also been shown to inhibit estra- diol sulfotransferase activity, which could contribute to and prolong their endocrine effects (Harju et al., 2007).
In the current study, none of the tested chemicals signif- icantly affected sex hormone production at any of the expo- sure concentrations tested, and some compounds had
inconsistent or sometimes conflicting effects on different endpoints of the steroidogenic pathway. The lack of effects at the hormone level may be due to the relatively low expo- sure concentrations tested. Sex hormone concentrations are potentially the most functional endpoint measured in the present study, because increased hormone concentrations in vivo are more likely to have effects than elevated enzyme activity, which might change quickly. These results indi- cated that these low levels of PBDEs and their derivatives may pose little risk in vivo, but further studies are clearly needed to address this question.
In conclusion, the present study demonstrated that some OH- and MeO-BDEs are able to induce or inhibit steroido- genic gene expression and aromatase activity in the H295R human adrenocortical carcinoma cell line at environmen- tally relevant concentrations. The position of the OH-group with regard to neighboring bromine atoms may play a key role in the alteration of the expression of these genes and enzyme activities, but clear relationships between structure and toxicity could not be determined. Further research is needed to more fully understand the effects of PBDE meta- bolites and derivatives on the steroidogenic process.
Acknowledgement
Support for this research was provided by grants from the National Science Foundation of China (Nos. 20518002, 40332024 and 40590393), and Strategic Research Grant 7002122 and Hong Kong Research Grants Council Grant CityU 2/06C.
References
Birchmeier, K.L., Smith, K.A., Passino-Reader, D.R., Sweet, L.I., Chernyat, S.M., Adams, J.V., Omann, G.M., 2005. Effects of selected polybrominated diphenyl ether flame retardants on lake trout (Salv- elinus namaycush) thyomcyte viability, apoptosis, and necrosis. Envi- ron. Toxicol. Chem. 24, 1518-1522.
Blanco, E., Casais, M.C., Mejuto, M.C., Cela, R., 2005. Analysis of tetrabromobisphenol A and other phenolic compounds in water samples by non-aqueous capillary electrophoresis coupled to photo- diode array ultraviolet detection. J. Chromatogr. A 15, 205-211.
Brown, D.J., Overmeire, I.V., Goeyens, L., Denison, M.S., Vito, M., Clark, G.C., 2004. Analysis of Ah receptor pathway activation by brominated flame-retardants. Chemosphere 57, 1509-1518.
BSEF (Bromine Science Environmental Forum), 2006. < http:// www.bsef.com>.
Canton, R.F., Sanderson, T., Letcher, R.J., Bergman, Å., van den Berg, M., 2005. Inhibition and induction of aromatase (CYP19) activity by brominated flame retardants in H295R human adrenocortical carci- noma cells. Toxicol. Sci. 88, 447-455.
Cantón, R.F., Sanderson, T., Nijmeijer, S., Bergman, Å., Letcher, R.J., van den Berg, M., 2006. In vitro effects of brominated flame retardants and metabolites on CYP17 catalytic activity: a novel mechanism of action? Toxicol. Appl. Pharmacol. 216, 274-281.
Cantón, R.F., Scholten, D.E.A., Marsh, G., de Jong, P.C., van den Berg, M., 2008. Inhibition of human placental aromatase activity by hydroxylated polybrominated diphenyl ethers (OH-PBDEs). Toxicol. Appl. Pharmacol. 227, 68-75.
Darnerud, P.O., Wong, J., Bergman, A., Ilback, N.G., 2005. Common viral infection affects pentabrominated diphenyl ether (PBDE)
distribution and metabolic and hormonal activities in mice. Toxicology 210, 159-167.
de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemosphere 46, 583-624.
European Union, 2003. Restriction of Hazardous Substances Directive; Directive 2002/95/EC of the European Parliament and of the Council of 27 January 2003 on the Restriction of the Use of Certain Hazardous Substances in Electrical and Electronic Equipment. OJ L37, 13 February 2003, p. 19.
Gazdar, A.F., Oie, H.K., Shackleton, C.H., Chen, T.R., Triche, T.J., Myers, C.E., Chrousos, G.P., Brennan, M.F., Stein, C.A., La Rocca, R.V., 1990. Establishment and characterization of a human adreno- cortical carcinoma cell line that expresses multiple pathways of steroid biosynthesis. Cancer Res. 50, 5488-5496.
Gill, U., Chu, I., Ryan, J.J., Feeley, M., 2004. Polybrominated diphenyl ethers: human tissue levels and toxicology. Rev. Environ. Contam. Toxicol. 183, 55-97.
Hakk, H., Letcher, R.J., 2003. Metabolism in the toxicokinetics and fate of brominated flame retardants-a review. Environ. Int. 29, 801-828.
Hallgren, S., Sinjari, T., Hakansson, H., Darnerud, P.O., 2001. Effects of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphe- nyls (PCBs) on thyroid hormone and vitamin A levels in rats and mice. Arch. Toxicol. 75, 200-208.
Harju, M., Hamers, T., Kamstra, J.H., Sonneveld, E., Boon, J.P., Tysklind, M., Andersson, P.L., 2007. Quantitative structure-activity relationship modeling on in vitro endocrine effects and metabolic stability involving 26 selected brominated flame retardants. Environ. Toxicol. Chem. 26, 816-826.
He, Y., Murphy, M.B., Yu, R.M.K., Lam, M.H.W., Hecker, M., Giesy, J.P., Wu, R.S.S., Lam, P.K.S., 2008. Effects of twenty PBDE metabolites on steroidogenesis in the H295R cell line. Toxicol. Lett. 176, 230-238.
Hecker, M., Newsted, J.L., Murphy, M.B., Higley, E.B., Jones, P.D., Wu, R.S.S., Giesy, J.P., 2006. Human adrenocarcinoma (H295R) cells for rapid in vitro determination of effects on steroidogenesis: Hormone production. Toxicol. Appl. Pharmacol. 217, 114-124.
Hecker, M., Hollert, H., Cooper, R., Vinggaard, A .- M., Akahori, Y., Murphy, M., Nelleman, C., Higley, E., Newsted, J., Wu, R., Lam, P., Laskey, J., Buckalew, A., Grund, S., Nakai, M., Timm, G., Giesy, J., 2007. The OECD validation program of the H295R steroidogenesis assay for the identification of in vitro inhibitors and inducers of testosterone and estradiol production Phase 2: Inter-laboratory pre- validation studies. Environ. Sci. Poll. Res. 14, 23-30.
Heneweer, M., van den Berg, M., Sanderson, J.T., 2004. A comparison of human H295R and rat R2C cell lines as in vitro screening tools for effects on aromatase. Toxicol. Lett. 146, 183-194.
Hilscherova, K., Jones, P.D., Gracia, T., Newsted, J.L., Zhang, X., Sanderson, J.T., Yu, R.M.K., Wu, R.S.S., Giesy, J.P., 2004. Assessment of the effects of chemicals on the expression of ten steroidogenic genes in the H295R cell line using real-time PCR. Toxicol. Sci. 81, 78-89.
Hites, R.A., 2004. Polybrominated diphenyl ethers in the environment and in people: a meta-analysis of concentrations. Environ. Sci. Technol. 38, 945-956.
Hooper, K., McDonald, T.A., 2000. The PBDEs: an emerging environ- mental challenge and another reason for breast-milk monitoring program. Environ. Health Persp. 108, 387-392.
Kitamura, S., Jinno, N., Ohta, S., Kuroki, H., Fujimoto, N., 2002. Thyroid hormonal activity of the flame retardants tetrabromobisphe- nol A and tetrachlorobisphenol A. Biochem. Biophys. Res. Commun. 26, 554-559.
Kitamura, S., Kato, T., Iida, M., Jinno, N., Suzuki, T., Ohta, S., Fujimoto, N., Hanada, H., Kashiwagi, K., Kashiwagi, A., 2005. Anti- thyroid hormonal activity of tetrabromobisphenol A, a flame retardant and related compounds: affinity to the mammalian thyroid hormone receptor, and effect on tadpole metamorphosis. Life Sci. 18, 1589- 1601.
Körner, W., Hanf, V., Schuller, W., Bartsch, H., Zwirner, M., Hagenma- ier, H., 1998. Validation and application of a rapid in vitro assay for
assessing the estrogenic potency of halogenated phenolic chemicals. Chemosphere. 37, 2395-2407.
Lephart, E.D., Simpson, E.R., 1991. Assay of aromatase activity. Methods Enzymol. 206, 477-483.
Letcher, R.J., Klasson-Wehler, E., Bergman, Å., 2000. Methyl sulfate and hydroxylated metabolites of polychlorinated biphenyls. The Hand- book Environ. Chem. 3, part K.
Marsh, G., Bergman, Å., Bladh, L.G., Gillner, M., Jakobsson, E., 1998. Synthesis of p-hydroxybromodiphenyl ethers and binding to the thyroid receptor. Organohal. Compds. 37, 305-308.
Marsh, G., Stenutz, R., Bergman, A., 2003. Synthesis of hydroxylated and methoxylated polybrominated diphenyl ethers - Natural products and potential polybrominated diphenyl ether metabolites. Eur. J. Org. Chem. 14, 2566-2576.
Marsh, G., Athanasiadou, M., Bergman, Å., Asplind, L., 2004. Identi- fication of hydroxylated and methoxylated polybrominated diphenyl ethers in Baltic Sea salmon (Salmo salar) blood. Environ. Sci. Technol. 38, 10-18.
Meerts, I.A.T.M., van Zanden, J.J., Luijks, E.A., van Leeuwen-Bol, I., Marsh, G., Jakobsson, E., Bergman, Å., Brouwer, A., 2000. Potent competitive interactions of some brominated flame retardants and related compounds with human transthyretin in vitro. Toxicol. Sci. 56, 95-104.
Meerts, I.A.T.M., Letcher, R.J., Hoving, S., Marsh, G., Bergman, Å., Lemmen, J.G., van der Burg, B., Brouwer, A., 2001. In vitro estrogenicity of polybrominated diphenyl ethers, hydroxylated PBDEs, and polybrominated bisphenol A compounds. Environ. Health Persp. 109, 399-407.
Mundy, W.R., Freudenrich, T.M., Crofton, K.M., Devito, M.J., 2004. Accumulation of PBDE 47 in primary cultures of rat neocortical cells. Toxicol. Sci. 82, 164-169.
Olsson, A., Ceder, K., Bergman, Å., Helander, B., 2000. Nestling blood of the white-tailed sea eagle (Haliaeetus albicilla) as an indicator of territorial exposure to organohalogen compounds - an evaluation. Environ. Sci. Technol. 34, 2733-2740.
Örn, U., Klasson-Wehler, E., 1998. Metabolism of 2,2’,4,4’-tetrabromo- diphenyl ether in rat and mouse. Xenobiotica. 28, 199-211.
Sanderson, J.T., Boerma, J., Lansbergen, G.W., van den Berg, M., 2002. Induction and inhibition of aromatase (CYP19) activity by various classes of pesticides in H295R human adrenocortical carcinoma cells. Toxicol. Appl. Pharmacol. 182, 44-54.
Siddiqi, M.A., Laessig, R.H., Reed, K.D., 2003. Polybrominated diphenyl ethers (PBDEs): new pollutants-old diseases. Clin. Med. Res. 1, 281-290. Simon, P., 2003. Q-Gene: processing quantitative real-time RT-PCR data. Bioinformatics 19, 1439-1440.
Sinkkonen, S., Rantalainen, A .- L., Paasivirta, J., Lahtiperä, M., 2004. Polybrominated methoxy diphenyl ethers (MeO-PBDEs) in fish and guillemots of Baltic Atlantic and Arctic environments. Chemosphere 56, 767-775.
Teuten, E.L., Xu, L., Reddy, C.M., 2005. Two abundant bioaccumulated halogenated compounds are natural products. Science 307, 917-920. Verslycke, T.A., Vethaak, A.D., Arijs, K., Janssen, C.R., 2005. Flame retardants, surfactants and organotins in sediment and mysid shrimp of the Scheldt estuary (The Netherlands). Environ. Pollut. 136, 19-31. WHO/ICPS, 1995. Environmental Health Criteria 172: Tetrabromobi- sphenol A and Derivatives. World Health Organization, Geneva, Switzerland.
Zhang, X.W., Yu, R.M.K., Jones, P.D., Lam, G.K.W., Newsted, J.L., Gracia, T., Hecker, M., Hilscherova, K., Sanderson, J.T., Wu, R.S.S., Giesy, J.P., 2005. Quantitative RT-PCR methods for evaluating toxicant-induced effects on steroidogenesis using the H295R cell line. Environ. Sci. Technol. 39, 2777-2785.
Zhou, T., Ross, D.G., Devito, M.J., Crofton, K.V., 2001. Effects of short- term in vivo exposure to polybrominated diphenyl ethers on thyroid hormones and hepatic enzyme activities in weanling rats. Toxicol. Sci. 61, 76-82.
Zhou, T., Taylor, M.M., DeVito, M.J., Crofton, K.M., 2002. Develop- mental exposure to brominated diphenyl ethers results in thyroid hormone disruption. Toxicol. Sci. 66, 105-116.