Inhibition and Induction of Aromatase (CYP19) Activity by Brominated F
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《毒物学科学杂志》
Institute for Risk Assessment Sciences, IRAS Utrecht University, Utrecht, The Netherlands
Institut National de la Recherche Scientifique, Institut Armand-Frappier (INRS-IAF), Universite du Quebec, Montreal, Quebec, H9R 1G6, Canada
National Wildlife Research Centre, Canadian Wildlife Service, Environment Canada, Ottawa, Ontario K1A 0H3 Canada
Department of Environmental Chemistry and Analytical Chemistry, Stockholm University, SE-106 91 Stockholm, Sweden
ABSTRACT
Brominated flame retardants (BFRs) are persistent and ubiquitous chemicals in the environment, and they are found at increasing levels in tissues of wildlife and humans. Previous in vitro studies with the BFR class of polybrominated diphenyl ethers (BDEs) have shown endocrine-disrupting properties. Our study assessed the potential effects of nineteen BDEs, five hydroxylated BDEs (OH-BDEs), one methoxylated BDE (CH3O-BDE), tetrabromobisphenol-A (TBBPA), its dibromopropane ether derivative (TBBPA-DBPE), and the brominated phenols/anisols 2,4,6-tribromophenol (TBP), 4-bromophenol (4BP) and 2,4,6-tribromoanisole (TBA) on the catalytic activity of the steroidogenic enzyme aromatase (CYP19) in H295R human adrenocortical carcinoma cells. Effects were studied in the concentration range from 0.5 to 7.5 μM; exposures were for 24 h. Both 6-OH-BDE47 and 6-OH-BDE99 showed an inhibitory effect on aromatase activity at concentrations >2.5 μM and >5 μM, respectively. However, 6-OH-BDE47 also caused a statistically significant increase in cytotoxicity (based on mitochondrial MTT reduction and lactate dehydrogenase-leakage [LDH]) at concentrations >2.5 μM that could explain in part the apparent inhibitory effect on aromatase activity. Compared to 6-OH-BDE47, the methoxy analog (6-CH3O-BDE47) did not elicit a cytotoxic effect, whereas significant inhibition of aromatase remained. TBP caused a concentration-dependent induction of aromatase activity between 0.5 and 7.5 μM (with a maximum of 3.8-fold induction at 7.5 μM). This induction was not observed when a OH– group replaced the CH3O– group or when bromine atoms adjacent to this OH– group were absent. These in vitro results provide a basis for studies of more detailed structure–activity relationships between these brominated compounds and the modulation of aromatase activity.
Key Words: brominated flame retardants (BFRs); BDEs; OH-PBDEs; CH3O-PBDEs; aromatase (CYP19); human adrenocortical (H295R) cell line.
INTRODUCTION
Flame retardants are chemicals that are added to materials to inhibit or suppress ignition and are incorporated during the manufacture of e.g., electronic equipment, furniture, construction materials, and textiles. Brominated organic compounds comprise a significant portion of brominated classification otherwise known as brominated flame retardants (BFRs). From an environmental point of view, BFRs have become an increasingly important group of organohalogen compounds, which includes, among others, polybrominated diphenyl ethers (PBDEs), tetrabromobisphenol A (TBBPA), and hexabromocyclododecane (HBCD) isomers that are all high production volume chemicals (de Wit, 2002). At the beginning of this century the global demand for PBDEs was estimated to be close to 70,000 metric tons per year (Sjodin et al., 2003).
Several decades ago, however, PBDEs were detected in the environment (Alaee et al., 2003), and at that time even the presence of the now controversial 2,2',3,3',4,4',5,5',6,6'-decabromoDE (BDE209) was found in soil and sludge samples collected from areas surrounding PBDE manufacturing facilities (DeCarlo, 1979). Temporal assessment of PBDE levels in wildlife species (e.g., the Lake trout species, Salvelinus namaycush) from the Great Lakes of North America estimate doubling times of 2–5 years (Norstrom et al., 2002), although recent studies have shown that PBDE levels seem to decrease more slowly in some parts of the world (Sellstrom et al., 2003). In addition, PBDE concentrations have been rapidly increasing in human breast milk from Swedish and North American women (Meironyte et al., 1999; Betts, 2002).
Hydroxylated PBDEs (OH-BDEs) are purported metabolites of PBDEs (Hakk and Letcher, 2003) and have been reported in the blood of wildlife and humans. OH-BDEs were present in blood of rats dosed with PBDE mixtures (Malmberg, 2004). Blood plasma of Swedish Atlantic salmon (Salmo salar), herring (Clupea harengus), and commercial fish oils were also found to contain OH-BDEs and/or CH3O-PBDEs (Haglund et al., 1997; Asplund et al., 1999). OH-BDEs that are retained in biological systems consist mainly of 6-OH-2,2',4,4',-tetrabromoDE (6-OH-BDE47). This major metabolite was recently reported in the plasma of various fish species from the Detroit River in the Great Lakes basin (Valters, 2004). It has been suggested that observed OH-BDEs in the blood plasma of fish may be due to cytochrome P450–mediated metabolism of PBDEs and/or accumulation of naturally occurring OH-BDEs, depending on factors such as marine versus freshwater ecosystems and species (Hakk and Letcher, 2003). In this respect, it should also be noted that OH-BDE congeners such as 6-OH-BDE47 and 2'-OH-BDE68 have been shown to be of natural origin in marine fish (Marsh, 2003).
There is increasing concern about the role of BFRs and their metabolites as potential endocrine-disrupting compounds (EDCs) in humans and wildlife (Darnerud et al., 2001). So far, in vitro interactions between PBDEs and their hydroxylated metabolites have been observed with transthyretin, the human thyroxine transport protein, due to competition with the thyroid hormone thyroxine or competition with natural estrogens on the estrogen receptor (Meerts et al., 2000; Olsen et al., 2002). Furthermore, hexabromocyclodecane (HBCD), 2,4,6-tribromophenol (TBP), and TBBPA have found to be quite toxic to aquatic organisms (Gribble, 1996; de Wit, 2002). Based on various animal studies with EDCs, including some BFRs, possible adverse effects on development, reproduction, and immune function in wildlife, and possibly humans, have been suggested or can no longer be excluded (Darnerud et al., 2001; Darnerud, 2003). Furthermore, mice studies have shown that neonatal exposure to tetra- and penta-brominated diphenyl ethers can cause dose-related neurobehavioral effects (Eriksson et al., 2001).
At present, many important data gaps for BFRs exist with respect to, e.g., reproductive or developmental in vivo effects, which cause many uncertainties in the risk characterization and assessment for these compounds. In addition, it should also be recognized that many of the endocrine-disrupting effects of these BFRs have been observed either in vitro or in vivo at relatively high concentrations (Darnerud, 2003). The significance of these findings are unclear with reference to concentrations found under normal environmental conditions.
One of the key processes in the effects mentioned above is regulated by aromatase (CYP19), the enzyme that mediates the conversion of androgens to estrogens through a bioconversion process known as aromatization. In humans, this enzyme activity is expressed in various tissues, ovaries, breast, testes, placenta, and fetal adrenals and brain (Pezzi et al., 2003), and it is directly involved in reproduction, development, behavior, and estrogen-dependent carcinogenesis (Simpson et al., 2002). Recently, our laboratory has identified the interaction of several pesticides and suspected in vivo EDCs with aromatase in the H295R human adrenocortical carcinoma cell line (Sanderson et al., 2000). This cell line expresses a significant number of steroidogenic enzymes that are involved in the synthesis of androgens, estrogens, and corticoid steroids (Gazdar et al., 1990).
In this study, using the H295R cell line, we assessed the possible effects and structure–activity relationships of several environmentally relevant BFRs and some of their possible metabolites on the expression or catalytic activity of aromatase.
MATERIALS AND METHODS
Selected brominated flame retardants or their metabolites.
H295R cells were exposed to a range of BFRs, i.e., TBBPA, tetrabromobisphenol A-bis (2,3) dibromopropylether (TBBPA-DBPE), TBP, and a number of PBDEs and their hydroxy or methoxy derivatives (Table 1). With the exception of (TBBPA-DBPE), which was a gift from Broomchemie, Terneuzen, (The Netherlands), chemicals were synthesized at the Wallenberg laboratory of the Stockholm University (Sweden). All BFRs were highly purified (>99% purity), and the presence of brominated dibenzo-p-dioxins or dibenzofurans was eliminated by applying a charcoal column as described elsewhere (rn, 1996; Marsh, 2003). Stock solutions of 2.5 mM were used for further dilution to experimental concentrations that ranged from 0.5 μM to 7.5 μM. This concentration range was initially selected based on earlier experiments in our laboratory with this cell line and a variety of different phytochemicals and pesticides (Sanderson et al., 2001; Heneweer et al., 2004).
Cell culture and treatment.
H295R cells were obtained from the American Type Culture Collection (ATCC #CRL-2128; Manassas, VA) and grown under culture conditions published previously by Sanderson et al. (2000). Briefly, the cells were grown in 1:1 Dulbecco's modified Eagle's medium/Ham's F-12 nutrient mix (DMEM/F12) (GibcoBRL 31300–038). The medium was supplemented with 6.7 μg/l sodium selenite, 10 mg/l insulin, and 5.5 ml/l transfer agent (ITS-G, GibcoBRL 41400–045), 100 U/l penicillin/streptomycin (GibcoBRL 15140–114), and 1% serum Ultroser SF (Sopachem, France). Cells were cultured in 24-well plates (Greiner, The Netherlands) and seeded with 1 ml of cell suspension per well at 37°C and 5% CO2. The culture medium was changed 24 h after plating, during which time the cells attached to the plate and reached near confluence. After this time period the cells were exposed to test chemicals (see Table 1) that were added to the wells at various concentrations (0.5–7.5 μM) using 2.5 mM of stock solutions in DMSO (0.1% v/v).
As positive control for aromatase induction, cells were exposed to 100–300 μM of 8-bromo-cyclic-adenosine-monophosphate (8-Br-cAMP, Sigma B5386). 4-Hydroxyandrostenedione (4-HA, 10 μM) was used as a positive control for aromatase catalytic inhibition, as described elsewhere (Heneweer et al., 2004).
Aromatase assay.
The catalytic activity of aromatase was determined based on the tritiated water-release method of Lephart and Simpson (1991). This method measures the production of 3H2O, which is formed as a result of the aromatization of the substrate [1-3H]-androstenedione. Cells were exposed to 54 nM [1-3H] androstenedione, New England Nuclear Research Products, Boston, MA) and dissolved in serum-free (Ultroser SF-free) culture medium (Sanderson et al., 2001). Following the procedure described elsewhere (Lephart and Simpson, 1991), 200 μl of culture medium was extracted and used for measuring the level of radioactivity after a substrate incubation period of 90 min. Corrections were made for background radioactivity, dilution factor, and specific activity of the substrate.
Protein determination.
The cells were stored at 4°C for measurement of protein content within 2–4 days, according to established methods (Lowry et al., 1951). Protein levels were extrapolated from a standard curve that was generated using bovine serum albumin (Sigma A7030).
Cytotoxicity measurements.
Cell viability, as an indicator of cytotoxicity, was determined by measuring the capacity of H295R cells to reduce MTT (3-[4,5-dimethylthiazol-2-yl]-2,5-diphenyltetrazolium bromide) to formazan (Denizot and Lang, 1986). MTT is reduced to blue-colored formazan by the mitochondrial enzyme succinate dehydrogenase, which is considered a sensitive measure of the mitochondrial function. Briefly, the cells in each well on the 24-well plate were incubated for 30 min, at 37°C with 0.5 ml of MTT (1 mg/ml) dissolved in culture medium without serum. Then, the MTT solution was removed, after which the cells were washed twice with PBS. The formazan formed in the cells was extracted by adding 1 ml of isopropanol and incubation for 10 min at room temperature. The alcohol fraction was measured spectrophotometrically at 595 nm (FLUOstar Galaxy V4.30–0/ Stacker Control V1.02–0, BMG Labtechnologies).
Lactate dehydrogenase (LDH) leakage was also used as an indicator of cytotoxicity for measuring the membrane integrity of the cells. The amount of LDH was measured according to the method of (Bergmeyer et al., 1965). The percentage of LDH leakage was determined by measuring the amount present in the well medium, relative to the total sum of the LDH in the medium and the cellular lysate.
Data analysis.
All experiments were done in triplicate, and within an individual experiment each concentration was tested in quadruplicate. Experiments were found to be reproducible. Data shown are those of one representative experiment. All results are presented as means with their standard deviations. Statistically significant differences from control groups were calculated using a two-tailed t-test analysis.
RESULTS
Induction of Aromatase Activity
Thirty BFRs and derivatives were studied individually to determine the effects on aromatase (CYP19) in the H295R cell line. Nineteen different BDEs were tested, and none of them showed a significant inductive effect on aromatase activity (Table 1). Only BDE28 and BDE38 (Table 1) showed minor induction of aromatase, which was observed only at the highest concentration of 7.5 μM, and exceeded no more than two times the control activity.
However, significant dose-dependent induction of aromatase was observed for TBP in the concentration range 0.5–7.5 μM. A maximum of 3.8-fold induction of aromatase (Fig. 1) was measured at 7.5 μM. We also studied the structure–activity relationship for bromophenol-mediated aromatase induction with respect to the presence of bromine atoms adjacent to the OH-group, and the replacement of the OH-group with a CH3O-group.
Structurally related 2,4,6-tribromoanisole (TBA) and 4-bromophenol (4-BP) were used for this purpose (Table 1). At 2.5, 5, and 7.5 μM, TBA and 4-BP did not show any induction of aromatase (data not shown). This indicates that the hydroxy group, combined with adjacent bromine(s), is an essential structural requirement for the induction of aromatase in the H295R cells. No change in protein content or cytotoxicity was observed in experiments with TBP, 2,4,6-TBA, or 4-BP treatments. As a positive control for these induction experiments, 8-Br-cAMP was used at 100 μM, which caused a fourfold induction of aromatase activity as reported elsewhere (Heneweer et al., 2004).
Inhibition of Aromatase Activity
Of the 19 (Table 1) different BDEs tested at concentrations ranging from 2.5 to 7.5 μM, none of the congeners showed a significant inhibitory effect on aromatase activity except BDE206 and BDE209, but only at 7.5 μM (Table 1). For BDE206, the aromatase activity was 61% of the control activity at a concentration of 7.5 μM. For some of these congeners, e.g., BDE99, significant cytotoxicity was found at the highest concentrations (>7.5 μM). Initial range-finding experiments were done with five different OH-BDEs. Some of these compounds, e.g., 6-OH-BDE99 and 6-OH-BDE47, caused a significant reduction of aromatase activity, 46.4% and 1.8%, respectively, at 7.5 μM (Table 1).
The most potent OH-BDE, 6-OH-BDE47, was then further tested in more detail at various concentrations ranging from 0.5 μM to 7.5 μM, which confirmed a strong concentration-dependent inhibitive effect on aromatase (Fig. 2A). However, cell viability measurements using MTT and LDH clearly indicated that the decreased aromatase activities occurred at concentrations of 6-OH-BDE47 that were also cytotoxic (Fig. 2B). In contrast, no cytotoxic effects were observed for 6-OH-BDE99, not even at the highest concentrations (Table 1). These results suggest that position of the OH-group and possible adjacent bromine atoms can have a differential influence on the inhibition of aromatase as well as cytotoxicity in this in vitro system. The CH3O-analog of 6-OH-BDE47, 6-CH3O-BDE47, showed no cytotoxicity, which demonstrates that cytotoxicity of 6-OH-BDE47 is a function of the OH-group. Nevertheless, an aromatase inhibition of approximately 50% still remained with 6-CH3O-BDE47 at concentrations of 5 μM and 7.5 μM (Fig. 3). This experiment indicates that the presence of a 6-CH3O-group reduces, but does not eliminate completely, the inhibitive potency of 6-OH-BDE47 on aromatase activity.
Two brominated bisphenol A compounds were also tested for inhibition of aromatase. TBBPA did not have a significant effect on aromatase activity and was not cytotoxic at concentrations of 2.5 μM and 7.5 μM. However, TBBPA-DBPE decreased the catalytic activity of aromatase with 50% at 7.5 μM (Fig. 4), and this was not caused by cytotoxicity, which was absent at the concentrations tested (data not shown).
DISCUSSION
Brominated flame retardants are widespread environmental contaminants that can commonly be found in the abiotic as well as biotic environment. A number of these compounds, such as certain PBDE congeners and HBCD isomers (e.g., -HBCD), bioaccumulate in the (human) food chain and are found in human milk and blood at concentrations that approach those of PCBs (Bocio et al., 2003).
Detailed information about possible mechanisms of toxicological or biological action is mostly lacking for these BFRs. However, both in vitro and in vivo experiments with, for example, PBDEs and TBPPA, have shown that these compounds or their hydroxylated metabolites or analogs can have an effect on thyroid hormones, the estrogen receptor, and neurobehavioral development (Meerts et al., 2000; Legler and Brouwer, 2003; Viberg et al., 2003). Based on these earlier results, some BFRs are considered to be potential EDCs (Darnerud et al., 2001), for which more information is necessary with respect to mechanism of action and concentration–effect relationships. In view of the accumulation of some PBDEs in humans and wildlife, there is concern about possible effects on reproduction and development by these compounds. One of the obvious end points to assess possible effects on the steroidogenesis by BFRs is aromatase, the key enzyme for the production of estrogens that is involved in many developmental and reproductive processes.
In our study we observed interactions of some BFRs and related compounds with aromatase activity in the H295R human adrenocortical carcinoma cell line, which also possesses a wide range of steroidogenic enzymes (CYP17, 3-HSD, among others). Using the same H295R cell line, our laboratory previously demonstrated that a number of pesticides, herbicides and fungicides are capable of both inhibiting or inducing aromatase activity (Sanderson et al., 2001, 2002; Heneweer et al., 2004).
Only a few of the 30 different BFRs (mainly PBDEs or their structural analogs) had an inhibiting or inducing effect on aromatase activity. However, the presence and position of an OH-group, or substitution with a CH3O-group, had a modulating effect on the potency of the inhibition. Two lower brominated PBDEs (28 and 38) and TBP could induce aromatase in vitro. Whereas both PBDE congeners induced aromatase slightly, TBP was found to be a more potent inducer, closely resembling 8-Br-cAMP in H295R cell line (Sanderson et al., 2002; Heneweer et al., 2004). It was also found that the OH-group in combination with adjacent bromine atoms plays an essential role in this induction mechanism. This was indicated by further experiments with 2,4,6-tribromoanisole (TBA) or 4-bromophenol (4-BP), neither of which caused induction of aromatase.
Another of the brominated triphenols, TBP, is widely found in the environment. It is both a naturally occurring compound in the marine environment (Gribble, 1996) and a synthetic flame retardant. Other studies with TBP showed that it can bind to the estrogen receptor and transthyretin, a thyroid hormone transport protein (Olsen et al., 2002). Our experiments now show that TBP can, at least in vitro, interfere in steroidogenesis by inducing aromatase. Thus, this compound clearly shows a number of in vitro effects that would classify it as a potential in vivo endocrine disruptor that needs to be investigated more. Furthermore, the exact mechanism of induction of aromatase in these H295R cells has to be elucidated to determine which different promoter genes can be responsible for the endocrine disruption (Heneweer et al., 2004). The human aromatase gene is under the control of different tissue-specific promoters with an apparently different physiological and biological relevance (Simpson et al., 2001). In H295R cells several promoters are involved in the induction of aromatase activity through different activation pathways, depending on the inducer added to the cells (e.g., cAMP or prostaglandins). Previous studies in our laboratory showed that amplification responses of aromatase pII and I.3 promoters could be measured in H295R cells after exposure to different inducers (Heneweer et al., 2004). Based on those results, it could be suggested that induction of aromatase by TBP might take place via one of these promoter regions (p.II or pI.3), but further experiments are needed to elucidate the possible induction pathway.
a physiological point of view, induction of aromatase activity will increase estrogen production. Estrogens are involved in numerous processes that play a crucial role in reproduction, development, and hormone-dependent carcinogenesis, among other processes (Clemons and Goss, 2001). Our results provided a basis for more research on BFRs and their potential modulating role in in vivo estrogen synthesis. For most BFRs studied in our experiments, an antiestrogenic effect caused by inhibition appears to be more likely, with the exception of TBP, which was found to be a clear inducer. However, further in vivo experiments should confirm if such an effect is relevant.
Apparent hydroxymetabolites of BDE47 and BDE99 (OH-BDE47/OH-BDE99) and a derivative of TBBPA, TBBPA-DBPE, showed potent inhibitory effects, which caused as much as a 95% reduction of the aromatase activity at a concentration of 7.5 μM for 6-OH-BDE47. However, simultaneous cytotoxicity tests in the same concentration range demonstrated a significant reduction of MTT and an increase in LDH leakage by 6-OH-BDE47, indicating that the inhibition of aromatase by 6-OH-BDE47 may be explained, at least in part, by the cytotoxicity of this compound.
To differentiate possible cytotoxic effects from aromatase inhibition by the hydroxy group in a PBDE molecule, we also studied the methoxyderivative of 6-OH-BDE47. Our experiments with 6-CH3O-BDE47 did not show cytotoxicity, but a significant inhibition of aromatase activity still remained. The two OH-BDE metabolites with an OH-group in the 6 position, 6-OH-BDE47 and 6-OH-BDE99, which only differ by one additional adjacent bromine atom, showed differences in inhibition and cytotoxicity. This structural difference is apparently sufficient to convert the highly cytotoxic 6-OH-BDE47 into the noncytotoxic compound 6-OH-BDE99, while conserving its aromatase-inhibiting properties.
Combined, our results indicate that, at least in vitro, some environmentally relevant OH-BDEs can influence steroidogenesis through aromatase inhibition and concurrent cytotoxicity. In our studies, however, this influence depended strongly on the position of the OH-group and bromine atoms. Several hydroxylated and methoxylated BDE derivatives have been found in Atlantic salmon, herring, and ringed and gray seals from the Baltic Sea, and their concentrations were of the same order of magnitude as PBDEs present in the sample (Haglund et al., 1997; Marsh et al., 2004; Valters, 2005). Furthermore, 6-CH3O-BDE47 has been found in human blood (Hovander et al., 2002; Marsh et al., 2004; Valters, 2005). However, the occurrence of OH-BDEs in the environment is complicated by the fact that many now appear to be natural products, at least in marine environments. Particularly OH-BDEs with the OH-group in an ortho-position are known endogenous compounds in organisms from the marine environment (Marsh et al., 2004). In addition, both OH-BDEs and TBBPA have been found to have the necessary structural properties to interact with TTR and thyroid receptor, which could lead to a cascade of effects on the thyroid hormones (Legler and Brouwer, 2003).
The question arises: can the observed in vitro inhibition potency of OH-PBDEs in our study possibly be relevant for the in vivo situation To have a first indication of the relevance of this effect for the human situation, a comparison with pharmacological aromatase inhibitors can be made. Earlier, in vitro experiments in our laboratory with the H295R human adrenocortical carcinoma cell line and the aromatase inhibitor Fadrozole showed an IC50 value of 3 nM (Heneweer et al., 2004). Thus, the inhibition potency of Fadrozole is many order of magnitudes higher than those observed for some OH- or CH3O-derivatives of PBDEs and TBBPA-BDPE in the present study.
No information about the occurrence of BDE-47 derivatives in human blood is yet available. From literature it is known that certain PBDEs, like BDE47, can be measured in human blood and serum samples at levels ranging between 4 and 16 ng/g lip.w. (Sjodin et al., 1999; Thomsen et al., 2001).
As a worst-case scenario, the hydroxy and methoxylated BDEs would be present in the human body at concentrations similar as those of BDE47. In this case, a maximum concentration of metabolites that might be found in whole blood would be about 0.35 nM (parameters used for calculation: plasma weight = 1.02 kg/l, plasma fat = 2%, total blood volume = 5.1 l, of which 2.8 l is plasma). This estimated blood concentration is still several orders of magnitude lower than the medium concentrations of hydroxylated and methoxylated BDEs (2.5–7.5 μM) that had a significant inhibitory effect in our in vitro experiments. Thus, assuming that median concentrations in our in vitro experiments can be compared with our maximal estimated blood concentrations in humans, our calculations could indicate that in vivo aromatase interaction by these PBDE metabolites would not easily be achieved. However, a more precise assessment can only be made when more in vivo information is available—e.g., about human tissue and plasma concentrations of these PBDE metabolites and TBBPA-BDPE.
In conclusion, we have identified some BFRs that are able either to inhibit or to induce in vitro aromatase activity in the H295R human adrenocortical carcinoma cell line. The importance of a phenolic structure, as well as the position of the hydroxyl group and the bromine atoms was established, but based on these results a complete structure–activity relationship cannot yet be determined. At present insufficient information is available about these BFRs and their metabolites, specifically in humans, to actually determine the possible in vivo relevance of this effect.
ACKNOWLEDGMENTS
This work described in this paper was fully supported by FIRE European project under contract number QLRT-2001–00596.
REFERENCES
Alaee, M., Arias, P., Sjodin, A., and Bergman, A. (2003). An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 29, 683–689.
Asplund, L. A. M., Sjodin, A., Bergman, A., Borjeson, H. (1999). Organohalogen substances in muscle, egg and blood from healthy Baltic salmon (Salmo salar) and Baltic salmon that produced offspring with the M74 syndrome. Ambio 28, 67–76.
Bergmeyer, H. U., Bent, U., and Hess, B. (1965). Lactic dehydrogenase. In Methods in Enzymatic Analysis (H. U. Bergmeyer, Ed.), Verlag Chemie GmbH, Wienheim, pp. 736–743.
Betts, K. S. (2002). Rapidly rising PBDE levels in North America. Environ. Sci. Technol. 36, 50A–52A.
Bocio, A., Llobet, J. M., Domingo, J. L., Corbella, J., Teixido, A., and Casas, C. (2003). Polybrominated diphenyl ethers (PBDEs) in foodstuffs: Human exposure through the diet. J. Agric. Food Chem. 51, 3191–3195.
Clemons, M., and Goss, P. (2001). Estrogen and the risk of breast cancer. N. Engl. J. Med. 344, 276–285.
Darnerud, P. O. (2003). Toxic effects of brominated flame retardants in man and in wildlife. Environ. Int. 29, 841–853.
Darnerud, P. O., Eriksen, G. S., Johannesson, T., Larsen, P. B., and Viluksela, M. (2001). Polybrominated diphenyl ethers: Occurrence, dietary exposure, and toxicology. Environ. Health Perspect. 109(Suppl. 1), 49–68.
de Wit, C. A. (2002). An overview of brominated flame retardants in the environment. Chemosphere 46, 583–624.
DeCarlo, V. J. (1979). Studies on brominated chemicals in the environment. Ann. N. Y. Acad. Sci. 320, 678–681.
Denizot, F., and Lang, R. (1986). Rapid colorimetric assay for cell growth and survival. Modifications to the tetrazolium dye procedure giving improved sensitivity and reliability. J. Immunol. Methods 89, 271–277.
Eriksson, P., Jakobsson, E., and Fredriksson, A. (2001). Brominated flame retardants: A novel class of developmental neurotoxicants in our environment Environ. Health Perspect. 109, 903–908.
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., and La Rocca, R. V. (1990). Establishment and characterization of a human adrenocortical carcinoma cell line that expresses multiple pathways of steroid biosynthesis. Cancer Res. 50, 5488–5496.
Gribble, G. W. (1996). Naturally occurring organohalogen compounds A comprehensive survey. Fortschr. Chem. Org. Naturst. 68, 1–423.
Haglund, P. S., Zook, D. R., Buser, H. R., and Hu, J. W. (1997). Identification and quantification of polybrominated diphenyl ethers and methoxy-polybrominated diphenyl ethers in Baltic biota. Environ. Sci. Technol. 31, 3281–3287.
Hakk, H., and Letcher, R. J. (2003). Metabolism in the toxicokinetics and fate of brominated flame retardants—A review. Environ. Int. 29, 801–828.
Heneweer, M., van den Berg, M., and 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.
Hovander, L., Malmberg, T., Athanasiadou, M., Athanassiadis, I., Rahm, S., Bergman, A., and Wehler, E. K. (2002). Identification of hydroxylated PCB metabolites and other phenolic halogenated pollutants in human blood plasma. Arch. Environ. Contam. Toxicol. 42, 105–117.
Legler, J., and Brouwer, A. (2003). Are brominated flame retardants endocrine disruptors Environ. Int. 29, 879–885.
Lephart, E. D., and Simpson, E. R. (1991). Assay of aromatase activity. Methods Enzymol. 206, 477–483.
Lowry, O. H., Rosebrough, N. J., Farr, A. L., and Randall, R. J. (1951). Protein measurement with the Folin–phenol reagent. J. Biol. Chem. 193, 265–275.
Malmberg, T. (2004). Identification and characterisation of hydroxylated PCB and PBDE metabolites in blood. Congener specific synthesis and analysis. Department of Environmental Chemistry, Stockholm University.
Marsh, G. (2003). Polybrominated diphenyl ethers and their hydroxy and methoxy derivatives. Department of Environmental Chemistry. Stockholm University.
Marsh, G., Athanasiadou, M., Bergman, A., and Asplund, L. (2004). Identification of hydroxylated and methoxylated polybrominated diphenyl ethers in Baltic Sea salmon (Salmo salar) blood. Environ. Sci. Technol. 38, 10–18.
Meerts, I. A., van Zanden, J. J., Luijks, E. A., van Leeuwen-Bol, I., Marsh, G., Jakobsson, E., Bergman, A., and Brouwer, A. (2000). Potent competitive interactions of some brominated flame retardants and related compounds with human transthyretin in vitro. Toxicol. Sci. 56, 95–104.
Meironyte, D., Noren, K., and Bergman, A. (1999). Analysis of polybrominated diphenyl ethers in Swedish human milk. A time-related trend study, 1972–1997. J. Toxicol. Environ. Health A 58, 329–341.
Norstrom, R. J., Simon, M., Moisey, J., Wakeford, B., and Weseloh, D. V. (2002). Geographical distribution (2000) and temporal trends (1981–2000) of brominated diphenyl ethers in Great Lakes hewing gull eggs. Environ. Sci. Technol. 36, 4783–4789.
Olsen, C. M., Meussen-Elholm, E. T., Holme, J. A., and Hongslo, J. K. (2002). Brominated phenols: Characterization of estrogen-like activity in the human breast cancer cell-line MCF-7. Toxicol. Lett. 129, 55–63.
rn, U., Erikson, L., Jakobsson, E., and Bergman, . (1996). Synthesis and characterisation of polybrominated diphenyl ethers—Unlabeled and radiolabelled tetra-, penta- and hexabrominated diphenyl ethers. Acta Chem. Scand. 50, 802–807.
Pezzi, V., Mathis, J. M., Rainey, W. E., and Carr, B. R. (2003). Profiling transcript levels for steroidogenic enzymes in fetal tissues. J. Steroid Biochem. Mol. Biol. 87, 181–189.
Sanderson, J. T., Boerma, J., Lansbergen, G. W., and 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.
Sanderson, J. T., Letcher, R. J., Heneweer, M., Giesy, J. P., and van den Berg, M. (2001). Effects of chloro-s-triazine herbicides and metabolites on aromatase activity in various human cell lines and on vitellogenin production in male carp hepatocytes. Environ. Health Perspect. 109, 1027–1031.
Sanderson, J. T., Seinen, W., Giesy, J. P., and van den Berg, M. (2000). 2-Chloro-s-triazine herbicides induce aromatase (CYP19) activity in H295R human adrenocortical carcinoma cells: A novel mechanism for estrogenicity Toxicol. Sci. 54, 121–127.
Sellstrom, U., Bignert, A., Kierkegaard, A., Haggberg, L., de Wit, C. A., Olsson, M., and Jansson, B. (2003). Temporal trend studies on tetra- and pentabrominated diphenyl ethers and hexabromocyclododecane in guillemot egg from the Baltic Sea. Environ. Sci. Technol. 37, 5496–5501.
Simpson, E. R., Clyne, C., Rubin, G., Boon, W. C., Robertson, K., Britt, K., Speed, C., and Jones, M. (2002). Aromatase—A brief overview. Annu. Rev. Physiol. 64, 93–127.
Simpson, E. R., Clyne, C., Speed, C., Rubin, G., and Bulun, S. (2001). Tissue-specific estrogen biosynthesis and metabolism. Ann. N. Y. Acad. Sci. 949, 58–67.
Sjodin, A., Hagmar, L., Klasson-Wehler, E., Kronholm-Diab, K., Jakobsson, E., and Bergman, A. (1999). Flame retardant exposure: Polybrominated diphenyl ethers in blood from Swedish workers. Environ. Health Perspect. 107, 643–648.
Sjodin, A., Patterson, D. G., Jr., and Bergman, A. (2003). A review on human exposure to brominated flame retardants—Particularly polybrominated diphenyl ethers. Environ. Int. 29, 829–839.
Thomsen, C., Leknes, H., Lundanes, E., and Becher, G. (2001). Brominated flame retardants in laboratory air. J. Chromatogr. A 923, 299–304.
Valters, K. M., Alaee, H., Li, G., Marsh, ., Bergman, I. D'Sa, and Letcher, R. J. (2005). Hydroxylated polybrominated and polychlorinated diphenyl ethers in the plasma of Detroit River fish. Environ. Sci. Technol. 39, 5612–5619.
Viberg, H., Fredriksson, A., and Eriksson, P. (2003). Neonatal exposure to polybrominated diphenyl ether (PBDE 153) disrupts spontaneous behaviour, impairs learning and memory, and decreases hippocampal cholinergic receptors in adult mice. Toxicol. Appl. Pharmacol. 192, 95–106.(Rocío F. Canton, J. Thomas Sanderson, Ro)
Institut National de la Recherche Scientifique, Institut Armand-Frappier (INRS-IAF), Universite du Quebec, Montreal, Quebec, H9R 1G6, Canada
National Wildlife Research Centre, Canadian Wildlife Service, Environment Canada, Ottawa, Ontario K1A 0H3 Canada
Department of Environmental Chemistry and Analytical Chemistry, Stockholm University, SE-106 91 Stockholm, Sweden
ABSTRACT
Brominated flame retardants (BFRs) are persistent and ubiquitous chemicals in the environment, and they are found at increasing levels in tissues of wildlife and humans. Previous in vitro studies with the BFR class of polybrominated diphenyl ethers (BDEs) have shown endocrine-disrupting properties. Our study assessed the potential effects of nineteen BDEs, five hydroxylated BDEs (OH-BDEs), one methoxylated BDE (CH3O-BDE), tetrabromobisphenol-A (TBBPA), its dibromopropane ether derivative (TBBPA-DBPE), and the brominated phenols/anisols 2,4,6-tribromophenol (TBP), 4-bromophenol (4BP) and 2,4,6-tribromoanisole (TBA) on the catalytic activity of the steroidogenic enzyme aromatase (CYP19) in H295R human adrenocortical carcinoma cells. Effects were studied in the concentration range from 0.5 to 7.5 μM; exposures were for 24 h. Both 6-OH-BDE47 and 6-OH-BDE99 showed an inhibitory effect on aromatase activity at concentrations >2.5 μM and >5 μM, respectively. However, 6-OH-BDE47 also caused a statistically significant increase in cytotoxicity (based on mitochondrial MTT reduction and lactate dehydrogenase-leakage [LDH]) at concentrations >2.5 μM that could explain in part the apparent inhibitory effect on aromatase activity. Compared to 6-OH-BDE47, the methoxy analog (6-CH3O-BDE47) did not elicit a cytotoxic effect, whereas significant inhibition of aromatase remained. TBP caused a concentration-dependent induction of aromatase activity between 0.5 and 7.5 μM (with a maximum of 3.8-fold induction at 7.5 μM). This induction was not observed when a OH– group replaced the CH3O– group or when bromine atoms adjacent to this OH– group were absent. These in vitro results provide a basis for studies of more detailed structure–activity relationships between these brominated compounds and the modulation of aromatase activity.
Key Words: brominated flame retardants (BFRs); BDEs; OH-PBDEs; CH3O-PBDEs; aromatase (CYP19); human adrenocortical (H295R) cell line.
INTRODUCTION
Flame retardants are chemicals that are added to materials to inhibit or suppress ignition and are incorporated during the manufacture of e.g., electronic equipment, furniture, construction materials, and textiles. Brominated organic compounds comprise a significant portion of brominated classification otherwise known as brominated flame retardants (BFRs). From an environmental point of view, BFRs have become an increasingly important group of organohalogen compounds, which includes, among others, polybrominated diphenyl ethers (PBDEs), tetrabromobisphenol A (TBBPA), and hexabromocyclododecane (HBCD) isomers that are all high production volume chemicals (de Wit, 2002). At the beginning of this century the global demand for PBDEs was estimated to be close to 70,000 metric tons per year (Sjodin et al., 2003).
Several decades ago, however, PBDEs were detected in the environment (Alaee et al., 2003), and at that time even the presence of the now controversial 2,2',3,3',4,4',5,5',6,6'-decabromoDE (BDE209) was found in soil and sludge samples collected from areas surrounding PBDE manufacturing facilities (DeCarlo, 1979). Temporal assessment of PBDE levels in wildlife species (e.g., the Lake trout species, Salvelinus namaycush) from the Great Lakes of North America estimate doubling times of 2–5 years (Norstrom et al., 2002), although recent studies have shown that PBDE levels seem to decrease more slowly in some parts of the world (Sellstrom et al., 2003). In addition, PBDE concentrations have been rapidly increasing in human breast milk from Swedish and North American women (Meironyte et al., 1999; Betts, 2002).
Hydroxylated PBDEs (OH-BDEs) are purported metabolites of PBDEs (Hakk and Letcher, 2003) and have been reported in the blood of wildlife and humans. OH-BDEs were present in blood of rats dosed with PBDE mixtures (Malmberg, 2004). Blood plasma of Swedish Atlantic salmon (Salmo salar), herring (Clupea harengus), and commercial fish oils were also found to contain OH-BDEs and/or CH3O-PBDEs (Haglund et al., 1997; Asplund et al., 1999). OH-BDEs that are retained in biological systems consist mainly of 6-OH-2,2',4,4',-tetrabromoDE (6-OH-BDE47). This major metabolite was recently reported in the plasma of various fish species from the Detroit River in the Great Lakes basin (Valters, 2004). It has been suggested that observed OH-BDEs in the blood plasma of fish may be due to cytochrome P450–mediated metabolism of PBDEs and/or accumulation of naturally occurring OH-BDEs, depending on factors such as marine versus freshwater ecosystems and species (Hakk and Letcher, 2003). In this respect, it should also be noted that OH-BDE congeners such as 6-OH-BDE47 and 2'-OH-BDE68 have been shown to be of natural origin in marine fish (Marsh, 2003).
There is increasing concern about the role of BFRs and their metabolites as potential endocrine-disrupting compounds (EDCs) in humans and wildlife (Darnerud et al., 2001). So far, in vitro interactions between PBDEs and their hydroxylated metabolites have been observed with transthyretin, the human thyroxine transport protein, due to competition with the thyroid hormone thyroxine or competition with natural estrogens on the estrogen receptor (Meerts et al., 2000; Olsen et al., 2002). Furthermore, hexabromocyclodecane (HBCD), 2,4,6-tribromophenol (TBP), and TBBPA have found to be quite toxic to aquatic organisms (Gribble, 1996; de Wit, 2002). Based on various animal studies with EDCs, including some BFRs, possible adverse effects on development, reproduction, and immune function in wildlife, and possibly humans, have been suggested or can no longer be excluded (Darnerud et al., 2001; Darnerud, 2003). Furthermore, mice studies have shown that neonatal exposure to tetra- and penta-brominated diphenyl ethers can cause dose-related neurobehavioral effects (Eriksson et al., 2001).
At present, many important data gaps for BFRs exist with respect to, e.g., reproductive or developmental in vivo effects, which cause many uncertainties in the risk characterization and assessment for these compounds. In addition, it should also be recognized that many of the endocrine-disrupting effects of these BFRs have been observed either in vitro or in vivo at relatively high concentrations (Darnerud, 2003). The significance of these findings are unclear with reference to concentrations found under normal environmental conditions.
One of the key processes in the effects mentioned above is regulated by aromatase (CYP19), the enzyme that mediates the conversion of androgens to estrogens through a bioconversion process known as aromatization. In humans, this enzyme activity is expressed in various tissues, ovaries, breast, testes, placenta, and fetal adrenals and brain (Pezzi et al., 2003), and it is directly involved in reproduction, development, behavior, and estrogen-dependent carcinogenesis (Simpson et al., 2002). Recently, our laboratory has identified the interaction of several pesticides and suspected in vivo EDCs with aromatase in the H295R human adrenocortical carcinoma cell line (Sanderson et al., 2000). This cell line expresses a significant number of steroidogenic enzymes that are involved in the synthesis of androgens, estrogens, and corticoid steroids (Gazdar et al., 1990).
In this study, using the H295R cell line, we assessed the possible effects and structure–activity relationships of several environmentally relevant BFRs and some of their possible metabolites on the expression or catalytic activity of aromatase.
MATERIALS AND METHODS
Selected brominated flame retardants or their metabolites.
H295R cells were exposed to a range of BFRs, i.e., TBBPA, tetrabromobisphenol A-bis (2,3) dibromopropylether (TBBPA-DBPE), TBP, and a number of PBDEs and their hydroxy or methoxy derivatives (Table 1). With the exception of (TBBPA-DBPE), which was a gift from Broomchemie, Terneuzen, (The Netherlands), chemicals were synthesized at the Wallenberg laboratory of the Stockholm University (Sweden). All BFRs were highly purified (>99% purity), and the presence of brominated dibenzo-p-dioxins or dibenzofurans was eliminated by applying a charcoal column as described elsewhere (rn, 1996; Marsh, 2003). Stock solutions of 2.5 mM were used for further dilution to experimental concentrations that ranged from 0.5 μM to 7.5 μM. This concentration range was initially selected based on earlier experiments in our laboratory with this cell line and a variety of different phytochemicals and pesticides (Sanderson et al., 2001; Heneweer et al., 2004).
Cell culture and treatment.
H295R cells were obtained from the American Type Culture Collection (ATCC #CRL-2128; Manassas, VA) and grown under culture conditions published previously by Sanderson et al. (2000). Briefly, the cells were grown in 1:1 Dulbecco's modified Eagle's medium/Ham's F-12 nutrient mix (DMEM/F12) (GibcoBRL 31300–038). The medium was supplemented with 6.7 μg/l sodium selenite, 10 mg/l insulin, and 5.5 ml/l transfer agent (ITS-G, GibcoBRL 41400–045), 100 U/l penicillin/streptomycin (GibcoBRL 15140–114), and 1% serum Ultroser SF (Sopachem, France). Cells were cultured in 24-well plates (Greiner, The Netherlands) and seeded with 1 ml of cell suspension per well at 37°C and 5% CO2. The culture medium was changed 24 h after plating, during which time the cells attached to the plate and reached near confluence. After this time period the cells were exposed to test chemicals (see Table 1) that were added to the wells at various concentrations (0.5–7.5 μM) using 2.5 mM of stock solutions in DMSO (0.1% v/v).
As positive control for aromatase induction, cells were exposed to 100–300 μM of 8-bromo-cyclic-adenosine-monophosphate (8-Br-cAMP, Sigma B5386). 4-Hydroxyandrostenedione (4-HA, 10 μM) was used as a positive control for aromatase catalytic inhibition, as described elsewhere (Heneweer et al., 2004).
Aromatase assay.
The catalytic activity of aromatase was determined based on the tritiated water-release method of Lephart and Simpson (1991). This method measures the production of 3H2O, which is formed as a result of the aromatization of the substrate [1-3H]-androstenedione. Cells were exposed to 54 nM [1-3H] androstenedione, New England Nuclear Research Products, Boston, MA) and dissolved in serum-free (Ultroser SF-free) culture medium (Sanderson et al., 2001). Following the procedure described elsewhere (Lephart and Simpson, 1991), 200 μl of culture medium was extracted and used for measuring the level of radioactivity after a substrate incubation period of 90 min. Corrections were made for background radioactivity, dilution factor, and specific activity of the substrate.
Protein determination.
The cells were stored at 4°C for measurement of protein content within 2–4 days, according to established methods (Lowry et al., 1951). Protein levels were extrapolated from a standard curve that was generated using bovine serum albumin (Sigma A7030).
Cytotoxicity measurements.
Cell viability, as an indicator of cytotoxicity, was determined by measuring the capacity of H295R cells to reduce MTT (3-[4,5-dimethylthiazol-2-yl]-2,5-diphenyltetrazolium bromide) to formazan (Denizot and Lang, 1986). MTT is reduced to blue-colored formazan by the mitochondrial enzyme succinate dehydrogenase, which is considered a sensitive measure of the mitochondrial function. Briefly, the cells in each well on the 24-well plate were incubated for 30 min, at 37°C with 0.5 ml of MTT (1 mg/ml) dissolved in culture medium without serum. Then, the MTT solution was removed, after which the cells were washed twice with PBS. The formazan formed in the cells was extracted by adding 1 ml of isopropanol and incubation for 10 min at room temperature. The alcohol fraction was measured spectrophotometrically at 595 nm (FLUOstar Galaxy V4.30–0/ Stacker Control V1.02–0, BMG Labtechnologies).
Lactate dehydrogenase (LDH) leakage was also used as an indicator of cytotoxicity for measuring the membrane integrity of the cells. The amount of LDH was measured according to the method of (Bergmeyer et al., 1965). The percentage of LDH leakage was determined by measuring the amount present in the well medium, relative to the total sum of the LDH in the medium and the cellular lysate.
Data analysis.
All experiments were done in triplicate, and within an individual experiment each concentration was tested in quadruplicate. Experiments were found to be reproducible. Data shown are those of one representative experiment. All results are presented as means with their standard deviations. Statistically significant differences from control groups were calculated using a two-tailed t-test analysis.
RESULTS
Induction of Aromatase Activity
Thirty BFRs and derivatives were studied individually to determine the effects on aromatase (CYP19) in the H295R cell line. Nineteen different BDEs were tested, and none of them showed a significant inductive effect on aromatase activity (Table 1). Only BDE28 and BDE38 (Table 1) showed minor induction of aromatase, which was observed only at the highest concentration of 7.5 μM, and exceeded no more than two times the control activity.
However, significant dose-dependent induction of aromatase was observed for TBP in the concentration range 0.5–7.5 μM. A maximum of 3.8-fold induction of aromatase (Fig. 1) was measured at 7.5 μM. We also studied the structure–activity relationship for bromophenol-mediated aromatase induction with respect to the presence of bromine atoms adjacent to the OH-group, and the replacement of the OH-group with a CH3O-group.
Structurally related 2,4,6-tribromoanisole (TBA) and 4-bromophenol (4-BP) were used for this purpose (Table 1). At 2.5, 5, and 7.5 μM, TBA and 4-BP did not show any induction of aromatase (data not shown). This indicates that the hydroxy group, combined with adjacent bromine(s), is an essential structural requirement for the induction of aromatase in the H295R cells. No change in protein content or cytotoxicity was observed in experiments with TBP, 2,4,6-TBA, or 4-BP treatments. As a positive control for these induction experiments, 8-Br-cAMP was used at 100 μM, which caused a fourfold induction of aromatase activity as reported elsewhere (Heneweer et al., 2004).
Inhibition of Aromatase Activity
Of the 19 (Table 1) different BDEs tested at concentrations ranging from 2.5 to 7.5 μM, none of the congeners showed a significant inhibitory effect on aromatase activity except BDE206 and BDE209, but only at 7.5 μM (Table 1). For BDE206, the aromatase activity was 61% of the control activity at a concentration of 7.5 μM. For some of these congeners, e.g., BDE99, significant cytotoxicity was found at the highest concentrations (>7.5 μM). Initial range-finding experiments were done with five different OH-BDEs. Some of these compounds, e.g., 6-OH-BDE99 and 6-OH-BDE47, caused a significant reduction of aromatase activity, 46.4% and 1.8%, respectively, at 7.5 μM (Table 1).
The most potent OH-BDE, 6-OH-BDE47, was then further tested in more detail at various concentrations ranging from 0.5 μM to 7.5 μM, which confirmed a strong concentration-dependent inhibitive effect on aromatase (Fig. 2A). However, cell viability measurements using MTT and LDH clearly indicated that the decreased aromatase activities occurred at concentrations of 6-OH-BDE47 that were also cytotoxic (Fig. 2B). In contrast, no cytotoxic effects were observed for 6-OH-BDE99, not even at the highest concentrations (Table 1). These results suggest that position of the OH-group and possible adjacent bromine atoms can have a differential influence on the inhibition of aromatase as well as cytotoxicity in this in vitro system. The CH3O-analog of 6-OH-BDE47, 6-CH3O-BDE47, showed no cytotoxicity, which demonstrates that cytotoxicity of 6-OH-BDE47 is a function of the OH-group. Nevertheless, an aromatase inhibition of approximately 50% still remained with 6-CH3O-BDE47 at concentrations of 5 μM and 7.5 μM (Fig. 3). This experiment indicates that the presence of a 6-CH3O-group reduces, but does not eliminate completely, the inhibitive potency of 6-OH-BDE47 on aromatase activity.
Two brominated bisphenol A compounds were also tested for inhibition of aromatase. TBBPA did not have a significant effect on aromatase activity and was not cytotoxic at concentrations of 2.5 μM and 7.5 μM. However, TBBPA-DBPE decreased the catalytic activity of aromatase with 50% at 7.5 μM (Fig. 4), and this was not caused by cytotoxicity, which was absent at the concentrations tested (data not shown).
DISCUSSION
Brominated flame retardants are widespread environmental contaminants that can commonly be found in the abiotic as well as biotic environment. A number of these compounds, such as certain PBDE congeners and HBCD isomers (e.g., -HBCD), bioaccumulate in the (human) food chain and are found in human milk and blood at concentrations that approach those of PCBs (Bocio et al., 2003).
Detailed information about possible mechanisms of toxicological or biological action is mostly lacking for these BFRs. However, both in vitro and in vivo experiments with, for example, PBDEs and TBPPA, have shown that these compounds or their hydroxylated metabolites or analogs can have an effect on thyroid hormones, the estrogen receptor, and neurobehavioral development (Meerts et al., 2000; Legler and Brouwer, 2003; Viberg et al., 2003). Based on these earlier results, some BFRs are considered to be potential EDCs (Darnerud et al., 2001), for which more information is necessary with respect to mechanism of action and concentration–effect relationships. In view of the accumulation of some PBDEs in humans and wildlife, there is concern about possible effects on reproduction and development by these compounds. One of the obvious end points to assess possible effects on the steroidogenesis by BFRs is aromatase, the key enzyme for the production of estrogens that is involved in many developmental and reproductive processes.
In our study we observed interactions of some BFRs and related compounds with aromatase activity in the H295R human adrenocortical carcinoma cell line, which also possesses a wide range of steroidogenic enzymes (CYP17, 3-HSD, among others). Using the same H295R cell line, our laboratory previously demonstrated that a number of pesticides, herbicides and fungicides are capable of both inhibiting or inducing aromatase activity (Sanderson et al., 2001, 2002; Heneweer et al., 2004).
Only a few of the 30 different BFRs (mainly PBDEs or their structural analogs) had an inhibiting or inducing effect on aromatase activity. However, the presence and position of an OH-group, or substitution with a CH3O-group, had a modulating effect on the potency of the inhibition. Two lower brominated PBDEs (28 and 38) and TBP could induce aromatase in vitro. Whereas both PBDE congeners induced aromatase slightly, TBP was found to be a more potent inducer, closely resembling 8-Br-cAMP in H295R cell line (Sanderson et al., 2002; Heneweer et al., 2004). It was also found that the OH-group in combination with adjacent bromine atoms plays an essential role in this induction mechanism. This was indicated by further experiments with 2,4,6-tribromoanisole (TBA) or 4-bromophenol (4-BP), neither of which caused induction of aromatase.
Another of the brominated triphenols, TBP, is widely found in the environment. It is both a naturally occurring compound in the marine environment (Gribble, 1996) and a synthetic flame retardant. Other studies with TBP showed that it can bind to the estrogen receptor and transthyretin, a thyroid hormone transport protein (Olsen et al., 2002). Our experiments now show that TBP can, at least in vitro, interfere in steroidogenesis by inducing aromatase. Thus, this compound clearly shows a number of in vitro effects that would classify it as a potential in vivo endocrine disruptor that needs to be investigated more. Furthermore, the exact mechanism of induction of aromatase in these H295R cells has to be elucidated to determine which different promoter genes can be responsible for the endocrine disruption (Heneweer et al., 2004). The human aromatase gene is under the control of different tissue-specific promoters with an apparently different physiological and biological relevance (Simpson et al., 2001). In H295R cells several promoters are involved in the induction of aromatase activity through different activation pathways, depending on the inducer added to the cells (e.g., cAMP or prostaglandins). Previous studies in our laboratory showed that amplification responses of aromatase pII and I.3 promoters could be measured in H295R cells after exposure to different inducers (Heneweer et al., 2004). Based on those results, it could be suggested that induction of aromatase by TBP might take place via one of these promoter regions (p.II or pI.3), but further experiments are needed to elucidate the possible induction pathway.
a physiological point of view, induction of aromatase activity will increase estrogen production. Estrogens are involved in numerous processes that play a crucial role in reproduction, development, and hormone-dependent carcinogenesis, among other processes (Clemons and Goss, 2001). Our results provided a basis for more research on BFRs and their potential modulating role in in vivo estrogen synthesis. For most BFRs studied in our experiments, an antiestrogenic effect caused by inhibition appears to be more likely, with the exception of TBP, which was found to be a clear inducer. However, further in vivo experiments should confirm if such an effect is relevant.
Apparent hydroxymetabolites of BDE47 and BDE99 (OH-BDE47/OH-BDE99) and a derivative of TBBPA, TBBPA-DBPE, showed potent inhibitory effects, which caused as much as a 95% reduction of the aromatase activity at a concentration of 7.5 μM for 6-OH-BDE47. However, simultaneous cytotoxicity tests in the same concentration range demonstrated a significant reduction of MTT and an increase in LDH leakage by 6-OH-BDE47, indicating that the inhibition of aromatase by 6-OH-BDE47 may be explained, at least in part, by the cytotoxicity of this compound.
To differentiate possible cytotoxic effects from aromatase inhibition by the hydroxy group in a PBDE molecule, we also studied the methoxyderivative of 6-OH-BDE47. Our experiments with 6-CH3O-BDE47 did not show cytotoxicity, but a significant inhibition of aromatase activity still remained. The two OH-BDE metabolites with an OH-group in the 6 position, 6-OH-BDE47 and 6-OH-BDE99, which only differ by one additional adjacent bromine atom, showed differences in inhibition and cytotoxicity. This structural difference is apparently sufficient to convert the highly cytotoxic 6-OH-BDE47 into the noncytotoxic compound 6-OH-BDE99, while conserving its aromatase-inhibiting properties.
Combined, our results indicate that, at least in vitro, some environmentally relevant OH-BDEs can influence steroidogenesis through aromatase inhibition and concurrent cytotoxicity. In our studies, however, this influence depended strongly on the position of the OH-group and bromine atoms. Several hydroxylated and methoxylated BDE derivatives have been found in Atlantic salmon, herring, and ringed and gray seals from the Baltic Sea, and their concentrations were of the same order of magnitude as PBDEs present in the sample (Haglund et al., 1997; Marsh et al., 2004; Valters, 2005). Furthermore, 6-CH3O-BDE47 has been found in human blood (Hovander et al., 2002; Marsh et al., 2004; Valters, 2005). However, the occurrence of OH-BDEs in the environment is complicated by the fact that many now appear to be natural products, at least in marine environments. Particularly OH-BDEs with the OH-group in an ortho-position are known endogenous compounds in organisms from the marine environment (Marsh et al., 2004). In addition, both OH-BDEs and TBBPA have been found to have the necessary structural properties to interact with TTR and thyroid receptor, which could lead to a cascade of effects on the thyroid hormones (Legler and Brouwer, 2003).
The question arises: can the observed in vitro inhibition potency of OH-PBDEs in our study possibly be relevant for the in vivo situation To have a first indication of the relevance of this effect for the human situation, a comparison with pharmacological aromatase inhibitors can be made. Earlier, in vitro experiments in our laboratory with the H295R human adrenocortical carcinoma cell line and the aromatase inhibitor Fadrozole showed an IC50 value of 3 nM (Heneweer et al., 2004). Thus, the inhibition potency of Fadrozole is many order of magnitudes higher than those observed for some OH- or CH3O-derivatives of PBDEs and TBBPA-BDPE in the present study.
No information about the occurrence of BDE-47 derivatives in human blood is yet available. From literature it is known that certain PBDEs, like BDE47, can be measured in human blood and serum samples at levels ranging between 4 and 16 ng/g lip.w. (Sjodin et al., 1999; Thomsen et al., 2001).
As a worst-case scenario, the hydroxy and methoxylated BDEs would be present in the human body at concentrations similar as those of BDE47. In this case, a maximum concentration of metabolites that might be found in whole blood would be about 0.35 nM (parameters used for calculation: plasma weight = 1.02 kg/l, plasma fat = 2%, total blood volume = 5.1 l, of which 2.8 l is plasma). This estimated blood concentration is still several orders of magnitude lower than the medium concentrations of hydroxylated and methoxylated BDEs (2.5–7.5 μM) that had a significant inhibitory effect in our in vitro experiments. Thus, assuming that median concentrations in our in vitro experiments can be compared with our maximal estimated blood concentrations in humans, our calculations could indicate that in vivo aromatase interaction by these PBDE metabolites would not easily be achieved. However, a more precise assessment can only be made when more in vivo information is available—e.g., about human tissue and plasma concentrations of these PBDE metabolites and TBBPA-BDPE.
In conclusion, we have identified some BFRs that are able either to inhibit or to induce in vitro aromatase activity in the H295R human adrenocortical carcinoma cell line. The importance of a phenolic structure, as well as the position of the hydroxyl group and the bromine atoms was established, but based on these results a complete structure–activity relationship cannot yet be determined. At present insufficient information is available about these BFRs and their metabolites, specifically in humans, to actually determine the possible in vivo relevance of this effect.
ACKNOWLEDGMENTS
This work described in this paper was fully supported by FIRE European project under contract number QLRT-2001–00596.
REFERENCES
Alaee, M., Arias, P., Sjodin, A., and Bergman, A. (2003). An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 29, 683–689.
Asplund, L. A. M., Sjodin, A., Bergman, A., Borjeson, H. (1999). Organohalogen substances in muscle, egg and blood from healthy Baltic salmon (Salmo salar) and Baltic salmon that produced offspring with the M74 syndrome. Ambio 28, 67–76.
Bergmeyer, H. U., Bent, U., and Hess, B. (1965). Lactic dehydrogenase. In Methods in Enzymatic Analysis (H. U. Bergmeyer, Ed.), Verlag Chemie GmbH, Wienheim, pp. 736–743.
Betts, K. S. (2002). Rapidly rising PBDE levels in North America. Environ. Sci. Technol. 36, 50A–52A.
Bocio, A., Llobet, J. M., Domingo, J. L., Corbella, J., Teixido, A., and Casas, C. (2003). Polybrominated diphenyl ethers (PBDEs) in foodstuffs: Human exposure through the diet. J. Agric. Food Chem. 51, 3191–3195.
Clemons, M., and Goss, P. (2001). Estrogen and the risk of breast cancer. N. Engl. J. Med. 344, 276–285.
Darnerud, P. O. (2003). Toxic effects of brominated flame retardants in man and in wildlife. Environ. Int. 29, 841–853.
Darnerud, P. O., Eriksen, G. S., Johannesson, T., Larsen, P. B., and Viluksela, M. (2001). Polybrominated diphenyl ethers: Occurrence, dietary exposure, and toxicology. Environ. Health Perspect. 109(Suppl. 1), 49–68.
de Wit, C. A. (2002). An overview of brominated flame retardants in the environment. Chemosphere 46, 583–624.
DeCarlo, V. J. (1979). Studies on brominated chemicals in the environment. Ann. N. Y. Acad. Sci. 320, 678–681.
Denizot, F., and Lang, R. (1986). Rapid colorimetric assay for cell growth and survival. Modifications to the tetrazolium dye procedure giving improved sensitivity and reliability. J. Immunol. Methods 89, 271–277.
Eriksson, P., Jakobsson, E., and Fredriksson, A. (2001). Brominated flame retardants: A novel class of developmental neurotoxicants in our environment Environ. Health Perspect. 109, 903–908.
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., and La Rocca, R. V. (1990). Establishment and characterization of a human adrenocortical carcinoma cell line that expresses multiple pathways of steroid biosynthesis. Cancer Res. 50, 5488–5496.
Gribble, G. W. (1996). Naturally occurring organohalogen compounds A comprehensive survey. Fortschr. Chem. Org. Naturst. 68, 1–423.
Haglund, P. S., Zook, D. R., Buser, H. R., and Hu, J. W. (1997). Identification and quantification of polybrominated diphenyl ethers and methoxy-polybrominated diphenyl ethers in Baltic biota. Environ. Sci. Technol. 31, 3281–3287.
Hakk, H., and Letcher, R. J. (2003). Metabolism in the toxicokinetics and fate of brominated flame retardants—A review. Environ. Int. 29, 801–828.
Heneweer, M., van den Berg, M., and 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.
Hovander, L., Malmberg, T., Athanasiadou, M., Athanassiadis, I., Rahm, S., Bergman, A., and Wehler, E. K. (2002). Identification of hydroxylated PCB metabolites and other phenolic halogenated pollutants in human blood plasma. Arch. Environ. Contam. Toxicol. 42, 105–117.
Legler, J., and Brouwer, A. (2003). Are brominated flame retardants endocrine disruptors Environ. Int. 29, 879–885.
Lephart, E. D., and Simpson, E. R. (1991). Assay of aromatase activity. Methods Enzymol. 206, 477–483.
Lowry, O. H., Rosebrough, N. J., Farr, A. L., and Randall, R. J. (1951). Protein measurement with the Folin–phenol reagent. J. Biol. Chem. 193, 265–275.
Malmberg, T. (2004). Identification and characterisation of hydroxylated PCB and PBDE metabolites in blood. Congener specific synthesis and analysis. Department of Environmental Chemistry, Stockholm University.
Marsh, G. (2003). Polybrominated diphenyl ethers and their hydroxy and methoxy derivatives. Department of Environmental Chemistry. Stockholm University.
Marsh, G., Athanasiadou, M., Bergman, A., and Asplund, L. (2004). Identification of hydroxylated and methoxylated polybrominated diphenyl ethers in Baltic Sea salmon (Salmo salar) blood. Environ. Sci. Technol. 38, 10–18.
Meerts, I. A., van Zanden, J. J., Luijks, E. A., van Leeuwen-Bol, I., Marsh, G., Jakobsson, E., Bergman, A., and Brouwer, A. (2000). Potent competitive interactions of some brominated flame retardants and related compounds with human transthyretin in vitro. Toxicol. Sci. 56, 95–104.
Meironyte, D., Noren, K., and Bergman, A. (1999). Analysis of polybrominated diphenyl ethers in Swedish human milk. A time-related trend study, 1972–1997. J. Toxicol. Environ. Health A 58, 329–341.
Norstrom, R. J., Simon, M., Moisey, J., Wakeford, B., and Weseloh, D. V. (2002). Geographical distribution (2000) and temporal trends (1981–2000) of brominated diphenyl ethers in Great Lakes hewing gull eggs. Environ. Sci. Technol. 36, 4783–4789.
Olsen, C. M., Meussen-Elholm, E. T., Holme, J. A., and Hongslo, J. K. (2002). Brominated phenols: Characterization of estrogen-like activity in the human breast cancer cell-line MCF-7. Toxicol. Lett. 129, 55–63.
rn, U., Erikson, L., Jakobsson, E., and Bergman, . (1996). Synthesis and characterisation of polybrominated diphenyl ethers—Unlabeled and radiolabelled tetra-, penta- and hexabrominated diphenyl ethers. Acta Chem. Scand. 50, 802–807.
Pezzi, V., Mathis, J. M., Rainey, W. E., and Carr, B. R. (2003). Profiling transcript levels for steroidogenic enzymes in fetal tissues. J. Steroid Biochem. Mol. Biol. 87, 181–189.
Sanderson, J. T., Boerma, J., Lansbergen, G. W., and 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.
Sanderson, J. T., Letcher, R. J., Heneweer, M., Giesy, J. P., and van den Berg, M. (2001). Effects of chloro-s-triazine herbicides and metabolites on aromatase activity in various human cell lines and on vitellogenin production in male carp hepatocytes. Environ. Health Perspect. 109, 1027–1031.
Sanderson, J. T., Seinen, W., Giesy, J. P., and van den Berg, M. (2000). 2-Chloro-s-triazine herbicides induce aromatase (CYP19) activity in H295R human adrenocortical carcinoma cells: A novel mechanism for estrogenicity Toxicol. Sci. 54, 121–127.
Sellstrom, U., Bignert, A., Kierkegaard, A., Haggberg, L., de Wit, C. A., Olsson, M., and Jansson, B. (2003). Temporal trend studies on tetra- and pentabrominated diphenyl ethers and hexabromocyclododecane in guillemot egg from the Baltic Sea. Environ. Sci. Technol. 37, 5496–5501.
Simpson, E. R., Clyne, C., Rubin, G., Boon, W. C., Robertson, K., Britt, K., Speed, C., and Jones, M. (2002). Aromatase—A brief overview. Annu. Rev. Physiol. 64, 93–127.
Simpson, E. R., Clyne, C., Speed, C., Rubin, G., and Bulun, S. (2001). Tissue-specific estrogen biosynthesis and metabolism. Ann. N. Y. Acad. Sci. 949, 58–67.
Sjodin, A., Hagmar, L., Klasson-Wehler, E., Kronholm-Diab, K., Jakobsson, E., and Bergman, A. (1999). Flame retardant exposure: Polybrominated diphenyl ethers in blood from Swedish workers. Environ. Health Perspect. 107, 643–648.
Sjodin, A., Patterson, D. G., Jr., and Bergman, A. (2003). A review on human exposure to brominated flame retardants—Particularly polybrominated diphenyl ethers. Environ. Int. 29, 829–839.
Thomsen, C., Leknes, H., Lundanes, E., and Becher, G. (2001). Brominated flame retardants in laboratory air. J. Chromatogr. A 923, 299–304.
Valters, K. M., Alaee, H., Li, G., Marsh, ., Bergman, I. D'Sa, and Letcher, R. J. (2005). Hydroxylated polybrominated and polychlorinated diphenyl ethers in the plasma of Detroit River fish. Environ. Sci. Technol. 39, 5612–5619.
Viberg, H., Fredriksson, A., and Eriksson, P. (2003). Neonatal exposure to polybrominated diphenyl ether (PBDE 153) disrupts spontaneous behaviour, impairs learning and memory, and decreases hippocampal cholinergic receptors in adult mice. Toxicol. Appl. Pharmacol. 192, 95–106.(Rocío F. Canton, J. Thomas Sanderson, Ro)