ELSEVIER
Toxicology in Vitro
journal homepage: www.elsevier.com/locate/toxinvit
Toxicology in Vitro
TİV
Steroid hormone related effects of marine persistent organic pollutants in human H295R adrenocortical carcinoma cells
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Myrthe W. van den Dungen a,b,*, Jeroen C.W. Rijk , Ellen Kampman b, Wilma T. Steegenga b, Albertinka J. Murka
a Sub-department of Environmental Technology, Wageningen University, Bornse Weilanden 9, 6708 WG Wageningen, The Netherlands
b Division of Human Nutrition, Wageningen University, Bomenweg 4, 6703 HD Wageningen, The Netherlands
” RIKILT - Institute of Food Safety, Wageningen UR, Akkermaalsbos 2, 6708 WB Wageningen, The Netherlands
ARTICLE INFO
Article history: Received 25 June 2014
Accepted 1 March 2015 Available online 9 March 2015
Keywords:
Persistent organic pollutants (POPs) Gene expression Steroid hormones Adrenal cortex cell line H295R Endocrine disruption
ABSTRACT
Persistent organic pollutants (POPs) such as 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), polychloro- biphenyl (PCB) 126 and 153, perfluorooctanesulfonic acid (PFOS), hexabromocyclododecane (HBCD), 2,2’,4,4’-tetrabromodiphenyl ether (BDE-47), tributyltin (TBT), and methylmercury (MeHg) can be accu- mulated in seafood and then form a main source for human exposure. Some POPs have been associated with changes in steroid hormone levels in both humans and animals. This study describes the in vitro effects of these POPs and mixtures thereof in H295R adrenocortical carcinoma cells. Relative responses for 13 steroid hormones and 7 genes involved in the steroidogenic pathway, and CYP1A1, were analyzed. PFOS induced the most pronounced effects on steroid hormone levels by significantly affecting 9 out of 13 hormone levels measured, with the largest increases found for 17ß-estradiol, corticosterone, and cortisol. Furthermore, TCDD, both PCBs, and TBT significantly altered steroidogenesis. Increased steroid hormone levels were accompanied by related increased gene expression levels. The differently expressed genes were MC2R, CYP11B1, CYP11B2, and CYP19A1 and changes in gene expression levels were more sensitive than changes in hormone levels. The POP mixtures tested showed mostly additive effects, especially for DHEA and 17ß-estradiol levels. This study shows that some seafood POPs are capable of altering steroido- genesis in H295R cells at concentrations that mixtures might reach in human blood, suggesting that adverse health effects cannot be excluded.
@ 2015 Elsevier Ltd. All rights reserved.
1. Introduction
Environmental persistent organic pollutants (POPs), either his- torical or currently in use, accumulate in fatty tissues causing sea- food to be a main source for human exposure (Bilau et al., 2008; Liem et al., 2000; Montano et al., 2013). This is of major concern,
Abbreviations: AhR, aryl hydrocarbon receptor; BDE-47, 2.2’,4,4’-tetrabro- modiphenyl ether; CV, coefficient of variation; DHEA, dehydroepiandrosterone; EIA, enzyme immunoassay; HBCD, hexabromocyclododecane; HPA, hypothalamic- pituitary-adrenal; LOQ, limit of quantification; MC2R, melanocortin 2 receptor; MeHg, methylmercury; OECD, organization for economic co-operation and devel- opment; PCB, polychlorobiphenyl; PFOS, perfluorooctanesulfonic acid; POPs, persistent organic pollutants; SC, solvent control; S/N, signal-to-noise ratio; SPE, solid-phase extraction; TBT, tributyl chlorotin; TCDD, 2,3,7,8-tetrachlorodibenzo- p-dioxin.
* Corresponding author at: Wageningen University, Tuinlaan 5, 6703 HE Wageningen, The Netherlands. Tel .: +31 317 485271; fax: +31 317 484931.
E-mail addresses: myrthe.vandendungen@wur.nl (M.W. van den Dungen), rijk. jeroen@gmail.com (J.C.W. Rijk), ellen.kampman@wur.nl (E. Kampman), wilma. steegenga@wur.nl (W.T. Steegenga), tinka.murk@wur.nl (A.J. Murk).
because many of these marine POPs have been related to negative health effects such as reduced cognitive development, immune toxicity, neurological disorders, cancer, and endocrine disruption (Li et al., 2006; Yu et al., 2010). In addition to the well-known mechanism for dioxin-like compounds via the aryl hydrocarbon receptor (AhR) (Landers and Bunce, 1991; Okey et al., 1994), other mechanisms of toxic action have been suggested to cause adverse health effects, including disruption of thyroid (Brouwer et al., 1998) and steroid hormone systems (WHO, 2012). Such mechan- isms can be inappropriate activation or antagonism of the nuclear steroid receptors, modulating nuclear receptor coactivators, or interference with key enzymes involved in steroid hormone syn- thesis and metabolism (Sanderson, 2006). Several POPs have been shown to affect hormone levels in humans and animals (Steinberg et al., 2008; Turyk et al., 2008; Zimmer et al., 2009). To date, most studies focused on the effects of single POPs on levels of only a few different steroid hormones (Ding et al., 2007; Kraugerud et al., 2010, 2011), or extracted mixtures of only partially identified POPs were tested (Montano et al., 2011; Zimmer et al., 2011).
Investigating single compounds is important to elucidate the mechanisms of action as different compounds may counteract each other’s mechanisms. Humans, on the other hand, are exposed to mixtures so it is important to elucidate combined actions as well.
In this study we investigated the effects of single POPs and mix- tures thereof on steroidogenesis. Eight different POPs present in the human seafood chain were chosen based on their relative abundance in polluted fish species. These include the dioxin-like compounds TCDD and PCB 126 (Hoogenboom et al., 2007); the non-dioxin-like PCB 153 (de Boer et al., 2010); the perfluorinated compound heptadecafluorooctanesulfonic acid (PFOS) (Kwadijk et al., 2010); the brominated flame retardants hexabromocyclodo- decane (HBCD) (van Leeuwen and de Boer, 2008) and 2,2’,4, 4’-tetrabromodiphenyl ether (BDE-47) (Voorspoels et al., 2003); the biocide tributyl chlorotin (TBT) (Stab et al., 1996), and the organometallic MeHg (Maes et al., 2008). The composition of mix- tures to which people are exposed depends on the origin and type of polluted seafood that is consumed. Therefore, there is not one mixture that is most relevant to test, but still it is important to test whether compounds in a mixture can induce interactive effects. The in vitro model used was the H295R human adrenocortical car- cinoma cell line, chosen because it expresses all hormones found in the adult adrenal cortex and the gonads and all key enzymes involved in steroidogenesis (Gazdar et al., 1990). The functionality of this bioassay is consistent with most results observed in vivo (Hecker et al., 2006). The endpoints we studied with the H295R model are corticosteroid synthesis and the production of sex ster- oid hormones. The assay has been fully validated to evaluate effects on testosterone and estradiol production (OECD, 2011). However, more elaborate approaches provide a screen for endo- crine disruption and can identify potential targets of the test com- pound (Rijk et al., 2012). In this study we measure steroidogenesis disruption as hormone production of 13 steroid hormones plus expression levels of 7 key steroidogenic genes for 8 individual POPs and mixtures thereof.
2. Materials and methods
2.1. Chemicals
TCDD (2,3,7,8-tetrachlorodibenzo-p-dioxin, CAS 1746-01-6), PCB 153 (2,2’,4,4’,5,5’-hexachlorobiphenyl, CAS 35065-27-1), PFOS (heptadecafluorooctanesulfonic acid potassium salt, CAS 2795-39-3), TBT (tributyl chlorotin, CAS 1461-22-9), and MeHg (methylmercury(II) chloride, CAS 115-09-3) were obtained from Sigma-Aldrich (Zwijndrecht, The Netherlands). PCB 126 (3,3’,4,4’,5-pentachlorobiphenyl, CAS 57465-28-8) was purchased from Promochem (Wesel, Germany). HBCD (hexabromocyclodode- cane, technical mixture) and BDE-47 (2,2’,4,4’-tetrabromodiphenyl ether, CAS 5436-43-1) were kindly provided by Professor Åke Bergman (Stockholm University, Sweden) within the framework of the EU FIRE project (Schriks et al., 2006). Forskolin (CAS 66575-29-9) and prochloraz (CAS 67747-09-5) were obtained from Sigma-Aldrich (St. Louis, MO, USA). Stock solutions for all chemicals were prepared in dimethylsulfoxide (DMSO) (Acros Organics, Belgium). Pregnenolone, 17a-OH-pregnenolone, progesterone, 17a-OH-progesterone, dehydroepiandrosterone (DHEA), androstenedione, testosterone, 11-deoxycorticosterone, corticosterone, 11-deoxycortisol and cortisol were obtained from Steraloids (Newport, RI, USA). The deuterium labeled internal ster- oid standards were from CDN isotopes (Point-Claire, Canada). Derivatisation reagent consisted of 1 mg 4-dimethyl-aminopy- ridine, 5 mg 2-methyl-6-nitrobenzoic anhydride and 3 mg picolinic acid in 1 mL tetrahydrofuran after which 10 uL of triethylamine
was added. All were purchased from Sigma-Aldrich (St. Louis, MO, USA).
2.2. Cell culture and exposure
H295R human adrenocortical carcinoma cells were obtained from the American Type Culture Collection (ATCC, Manassas, VA, USA) and were cultured according to the standardized protocol approved by the OECD (OECD, 2011). Briefly, cells were routinely grown in 75 cm2 culture flasks containing DMEM/F12 (without phenol-red, Sigma, Zwijndrecht, The Netherlands) sup- plemented with 1.2 g/L NaHCO3, 1% Insulin Transferrin Selenium (ITS + premix), and 2.5% NuSerum (BD Biosciences, Bedford, MA) at 37 ℃ and 5% CO2 atmosphere. Cells were subcultured when 80% confluency was reached. For subcultivation, cells were washed twice with PBS, detached using trypsin-EDTA (0.25/0.05%, v/v) (Difco, NJ), and seeded in a 1:3 ratio. After thawing frozen stocks (passage 5), cells were cultured for at least four additional passages prior to testing and cells were not used after passage 13. For experiments, 3 mL of 3 × 105 cells/mL was seeded in 6-well plates (Greiner bio-one, Frickenhausen, Germany). After 24 h, medium was replaced by 2 mL exposure medium containing a single POP or a mixture thereof dissolved in DMSO. The final concentration of DMSO in the medium was 0.25%. Three independent experi- ments were performed and in each experiment three concentra- tions were tested for both the single POPs and POP mixtures. Concentrations were chosen in two steps, a first selection of con- centrations was based on literature study, followed by cytotoxicity testing to ensure experiments were performed at non-toxic con- centrations. Two mixtures of different combinations of these com- pounds were tested, one with only four compounds which all affected steroidogenesis (mixture A) and one with all eight com- pounds (mixture B). Mixtures A1 and B1 consist of the middle con- centration of the concentration range chosen for the individual POP tests based on the absence of cytotoxicity. So mixture A1 consisted of TCDD (30 nM), PCB 126 (3 µM), PCB 153 (3 µM) and PFOS (100 µM) and mixture B1 consisted of these 4 POPs and HBCD (300 nM), BDE-47 (300 nM), TBT (10 nM), and MeHg (30 nM). Mixture A1 or B1 were 3 times diluted to obtain mixture A3 or B3 and 10 times diluted to obtain mixture A10 or B10. Forskolin, a known adenylate cyclase inducer, and prochloraz, a known inhi- bitor of multiple CYPs involved in steroidogenesis such as CYP17, were run as positive and negative controls. After 48 h of exposure, as suggested by the OECD, the medium was collected and stored at -80 ℃ until hormone analysis. The cells were immediately lysed in Trizol (Invitrogen, Breda, The Netherlands), transferred to a fresh vial, and stored at -80 ℃ until RNA extraction.
2.3. Cell viability assays
For cytotoxicity testing, cells were seeded in a 96-well plate using 100 µL of 3 × 105 cells/mL. After 24 h, medium was replaced by 200 uL medium containing the individual POPs or mixtures thereof in triplicate. The highest concentrations tested were 100 nM for TCDD, 10 µM for both PCBs, 600 uM for PFOS, 1 µM for HBCD, 1 µM for BDE-47, 100 nM for TBT, 100 nM for MeHg, and the highest test concentration of both mixtures. Cytotoxic effects from three independent experiments were compared to the sol- vent control (SC) and Triton X-100 (0.1%) was included as a posi- tive control. After 48 h, cytotoxic effects of the test compounds were evaluated with two cytotoxicity assays that measure differ- ent endpoints. The assays applied were the WST-1 cytotoxicity assay (Roche Diagnostics, Mannheim, Germany) based on the activity of mitochondrial dehydrogenase enzymes and the ATPlite assay (PerkinElmer, Groningen, The Netherlands) based on the principle that ATP is present in all metabolically active cells. For
the WST-1 assay, 20 uL of WST-1 reagent (WST-1 kit content) was added, and after 1 h absorbance was measured at 450 nm using a Synergy HT multi-detection microplate reader (BioTek Instruments Inc.). For the ATPlite assay, 100 uL medium was removed from every well, and 50 uL mammalian cell lysis buffer (ATPlite kit content) was added. After 5-min incubation on an orbi- tal plate shaker, 50 µL of substrate solution (ATPlite kit content) was added, and the plate was shaken again for 5 min. After 10- min incubation in the dark, luminescence was measured at 590 nm using the same microplate reader. To ensure that further experiments would not be performed at cytotoxic concentrations, the highest concentrations included did not deviate more than 20% from the SC in either the WST-1 or the ATPlite assay.
2.4. LC-MS/MS hormone analysis
Hormone levels of pregnenolone, 17a-OH-pregnenolone, pro- gesterone, 17a-OH-progesterone, DHEA, androstenedione, testos- terone, 11-deoxycorticosterone, corticosterone, 11-deoxycortisol, and cortisol were measured in the cell culture medium following a method previously described (Wang et al., 2014). Calibration standards were prepared by spiking 900 uL of supplemented DMEM/F12 medium with a mix of steroid standards, resulting in concentrations of 0; 10; 25; 50; 100; 250; 500; 1000; 2500; 5000; 10,000; 25,000; 50,000 and 100,000 pg/mL. Standards as well as samples (also 900 µL) were spiked with 22.5 µL 13C labeled internal standard mix (10 ng/ml) and filled up to 1 mL with MilliQ water. Calibration standards and samples were subjected to solid-phase extraction (SPE) using OASIS HLB cartridges (Waters, 60 mg) in 96-wells format previously conditioned with methanol and Milli- Q water. After washing with Milli-Q water, methanol/water/acetic acid (55:43:2, v/v/v %), methanol/water/25% ammonia (30:62:8, v/v/v %), and acetonitrile/water (35:65, v/v %) the free steroids were eluted with acetone. The eluate was evaporated to dryness at 45 ℃ under nitrogen, and picolinoyl derivatisation was achieved by incu- bating the dried sample extracts with 35 uL of derivatisation reagent (see chemicals section) for 45 min at room temperature. The reaction was terminated by adding a 5% ammonia solution and samples were directly analyzed on the LC-MS/MS as described by (Blokland et al., 2012)), with a change in the gradient of the LC mobile phase. The gradient was: 0-0.2 min 5% B (90/10 v/v-% ace- tonitrile/0.1% formic acid), 0.2-0.5 min, linear increase to 20% B, 0.5-3.5 min, linear increase to 80% B, 3.5-3.6 min, increase to 95% B with a hold of 0.45 min, after which the gradient returned in 0.05 min to 5% B with a final hold of 0.9 min. For androstenedione and 11-deoxycortisol analysis samples were diluted with 200 µL tetrahydrofuran/ammonia (35:50, v/v %) before measurement. Steroid hormone concentrations were calculated using a linear calibration curve and corrected for loss during the sample extrac- tion based on the isotope labeled internal standard. Accuracy was calculated by dividing the mean of the determined concentration by the actual concentration times 100% and should be between 70% and 130%. The precision of the method was calculated as the coefficient of variation (CV) and should not exceed 30%. The range was set at the concentrations of the calibration curve where the method could operate with acceptable accuracy, precision, within the linear range of the curve (r2 > 0.99), and with a signal-to-noise (S/N) ratio greater than 10. The limit of quantification (LOQ) was set at the lowest value of the range and samples with an S/N ratio lower than 10 were assigned a value equivalent to the LOQ in order to calculate fold changes.
2.5. Hormone measurement EIAs
17ß-Estradiol and aldosterone levels in medium were analyzed using enzyme immunoassays (EIAs) according to the
manufacturer’s recommendations. Cross-reactivity of the 17ß- estradiol kit (Oxford Biochemical Research) reported by the manufacturer was 1.0% for testosterone and lower for other hor- mones. Cross-reactivity for the aldosterone measurement (Cayman Chemical) was reported to be 1.1% for corticosterone and lower for other hormones. Prior to hormone analysis, the medium samples were thawed and kept on ice. The kits were mod- ified by replacing the standard curve with standards prepared in cell culture medium. Data were linearized using a logit transforma- tion followed by a linear regression fit. Accuracy and precision (see LC-MS/MS hormone analysis) were calculated for both assays.
2.6. RNA isolation and cDNA synthesis
Total RNA was isolated according to the manufacturer’s (Invitrogen) instructions using a chloroform/isopropyl alcohol extraction. RNA was dissolved in Milli-Q water and the RNA con- centration was measured using a NanoDrop ND-1000 UV-vis spec- trophotometer (Isogen, Maarsen, The Netherlands). Samples with a 260/280 value around 2.0 and with a 260/230 ratio > 2.0 were regarded to be of good quality resulting in good PCR results. First-strand cDNA was synthesized from 1 µg of total RNA using RevertAid First Strand cDNA Synthesis Kit (Thermo Scientific, Germany) following the supplier’s protocol.
2.7. Real-time quantitative PCR
Primer sequences for CYP1A1 and 7 key enzymes involved in steroidogenesis were retrieved from literature or from the Harvard PrimerBank (http://pga.mgh.harvard.edu/primerbank) (Table S1). Primers were tested for specificity by BLAST analysis using NCBI PrimerBLAST (http://www.ncbi.nlm.nih.gov/tools/ primer-blast). Real-time quantitative PCR (qPCR) was performed using SensiMix SYBR (GC Biotech, Alphen A/D Rijn, The Netherlands) and a CFX384 Real-Time PCR detection system (Bio- Rad laboratories BV, Veenendaal, The Netherlands). 2 uL cDNA was used in a total volume of 8 uL, with a primer concentration of 600 nM. The following thermal cycling conditions were used: 10 min at 95 ℃, followed by 40 cycles of 95 ℃ for 10 s and 60 ℃ for 15 s. Melting curve analysis was performed to assure no non- specific signals were detected. Relative expression was calculated using a threefold dilution series of pooled cDNA to correct for the efficiency of the PCR reaction. The reactions were performed in duplicate and all samples were normalized to HPRT (hypoxanthine phosphoribosyltransferase) expression. Relative changes in gene expression levels were expressed as fold change compared to the SC.
2.8. Statistical analysis
Statistical significance of reduced cell viability was calculated using an unpaired Student’s t-test. Differences between treatments and SCs for hormone levels and gene expression analysis were cal- culated as fold change compared to the SC and evaluated using a one way ANOVA followed by a two-sided Dunnett’s test. Differences were considered significant at a p-value <0.05 and highly significant at a p-value <0.01. Statistical analyses were per- formed with IBM SPSS Statistics version 20.
3. Results
3.1. Cell viability
Upon exposure of H295R cells no decrease in cell viability was observed with either the WST-1 or the ATPlite assay for any of
the concentrations tested for TCDD, PCB 126, PCB 153, HBCD, BDE- 47, TBT, MeHg, and both mixtures. PFOS showed decreased viabi- lity for concentrations ≥300 µM (data not shown). The highest non-cytotoxic concentration of PFOS in our experimental setup (200 µM) was used for further experiments.
3.2. Quality controls
The basal hormone productions in H295R cells exposed for 48 h to the SC can be found in Table 1. Production was highest for 11-deoxycortisol (200 ng/ml), followed by androstenedione (43 ng/ml), and lowest for 17ß-estradiol (143 pg/mL) and proges- terone (600 pg/mL). The LOQ, reported as the lower range, was low enough for all hormones to significantly determine down- regulation compared to the SC (Table 2A and Fig. 1). The hormone measurements for 17ß-estradiol and testosterone were in compli- ance with the requirements of the OECD guidelines regarding LOQ, fold change, precision and accuracy (OECD, 2011). Forskolin sig- nificantly induced the levels of all hormones, except progesterone, as well as the gene expression levels of all measured genes involved in steroidogenesis (Tables 2A and 2B). Prochloraz sig- nificantly reduced all hormone levels, except for pregnenolone and 11-deoxycorticosterone, while none of the gene expression levels were decreased (Tables 2A and 2B).
3.3. Effects of POPs on steroid synthesis
Three POPS (HBCD, BDE-47 and MeHg) did not significantly influence any of the hormone levels (Fig. 1 and Table S2). Interestingly, PFOS had the most pronounced effects of all POPs tested and altered almost all hormones of the steroid biosynthesis pathway. The hormones at the end of the pathway, aldosterone, cortisol, testosterone and 17ß-estradiol, as well as corticosterone and androstenedione, were up-regulated after exposure to PFOS. Hormones in the beginning of the pathway, 17a-OH-pregnenolone, pregnenolone and 11-deoxycorticosterone were significantly down-regulated. TCDD and PCB 126 significantly up-regulated pro- gesterone (1.4-fold) and 17ß-estradiol (1.8-fold), respectively. PCB 153 and TBT did not up-regulate any hormone levels, but decreased several hormone levels. PCB 153 had the most pro- nounced effect on 17a-OH-progesterone (1.8-fold down- regulation) and testosterone (1.5-fold down-regulation). TBT down-regulated pregnenolone (-1.7-fold), progesterone and
17a-OH-progesterone (both -1.5-fold). Mixtures A and B had com- parable effects and each raised or lowered levels of several of the hormones produced (Fig. S1). 17a-OH-pregnenolone and DHEA were significantly down-regulated after exposure to mixtures of POPs, and progesterone, 17ß-estradiol and corticosterone levels were significantly increased. In addition, mixture B decreased pregnenolone levels and raised aldosterone levels.
3.4. Effects of POPs on steroid related gene expression levels
Fig. 2 gives an overview of the changes in gene expression levels found. CYP1A1, an indicator for exposure to AhR agonists such as dioxin-like compounds, was up-regulated after exposure to TCDD, PCB 126 and both mixture A and B. Expression levels of the steroid related genes MC2R (3-9-fold change), CYP11B1 (3-27-fold change), and CYP11B2 (6-81-fold change) were up-regulated by PCB 126 (10 µM), PFOS (200 µM) and both mix- tures (highest concentration). Furthermore, PCB 126 and mixture A and B up-regulated CYP19A1 resulting in a 2.9-, 3.9-, and 2.3-fold induction, respectively. No significant changes in expression levels were found for StAR, CYP11A1, and HSD3B2 after exposure to any of the tested POPs (Fig. S2).
4. Discussion
The effects of persistent organic pollutants (POPs) and two mix- tures thereof on steroidogenesis were studied using the enhanced H295R screening assay. For this purpose, 13 steroid hormone levels and 7 steroid-related genes were measured.
In absolute quantities the H295R cells produced higher levels of adrenal steroids than of sex steroids, which is in accordance with the basal hormone levels reported by Xing et al. (2011). Forskolin and prochloraz also gave comparable changes in hormone levels and gene expression levels as reported in previous studies (Gracia et al., 2006; Hecker et al., 2006), with the exception that prochloraz did not decrease gene expression levels after 48 h expo- sure. Earlier research reported small decreases in expression levels of steroidogenic related genes, but not for all genes a dose-respon- se relationship was observed (Ohlsson et al., 2009). In our study prochloraz, which is a known AhR agonist, did induce CYP1A1 as was reported before (Long et al., 2003). The main endocrine modulating mode of action of prochloraz, however, is not by
| Hormone | Range standard curve (pg/mL)a | Maximum RSDb(%) | Mean accuracy (%) | Hormone levels (pg/mL)" | Relative hormone levels (%)d |
|---|---|---|---|---|---|
| Pregnenolone | 2500-25,000 | 12 | 100-125 | 5229 ± 274 | 1.7 ±0.2 |
| 17a-OH-pregnenolone | 2500-25,000 | 13 | 89-105 | 9640 ± 472 | 3.1 ± 0.3 |
| Progesterone | 50-10,000 | 29 | 81-111 | 600 ± 52 | 0.2 ± 0.0 |
| 17a-OH-progesterone | 25-25,000 | 26 | 87-130 | 4360 ± 221 | 1.4 ±0.1 |
| DHEA | 1000-25,000 | 31 | 94-127 | 3538 ± 154 | 1.2 ±0.1 |
| Androstenedionee | 500-100,000 | 8 | 78-105 | 42,897 ± 4073 | 14 ±2.3 |
| Testosterone | 100-10,000 | 9 | 94-118 | 1689 ± 305 | 0.6 ± 0.2 |
| 17ß-Estradiolf | 20-2000 | 16 | 94-112 | 143 ± 8 | 0.04 ± 0.0 |
| 11-Deoxycorticosterone | 50-10,000 | 11 | 93-125 | 9383 ± 163 | 3.1 ± 0.1 |
| Corticosterone | 500-25,000 | 24 | 93-117 | 3616 ± 515 | 1.2 ± 0.3 |
| Aldosteronef | 7.8-1000 | 17 | 97-108 | 606 ± 104 | 0.2 ± 0.1 |
| 11-Deoxycortisole | 2500-100,000 | 11 | 89-108 | 200,451 ± 13,120 | 66 ± 7.4 |
| Cortisol | 250-25,000 | 30 | 92-107 | 24,402 ± 5704 | 8.0 ± 3.2 |
a Range was determined as those concentrations of the calibration curve where the method could operate with acceptable precision, accuracy and within the linear range of the curve (r2 > 0.99), and with a signal-to-noise (S/N) ratio greater than 10.
b Relative standard deviation (RSD) is calculated as a measurement of precision.
” Basal hormone levels measured in culture medium after 48 h exposure to the solvent control (0.25% DMSO) (mean ± SE).
d Relative hormone levels (%) compared to the sum of all measured basal hormones levels.
e Androstenedione and 11-deoxycortisol levels were measured after dilution of the samples.
f 17B-Estradiol and aldosterone levels were measured using an EIA kit.
TCDD
PCB 126
PCB 153
PFOS
HBCD
BDE-47
TBT
MeHg
Mix A
Mix B
(10 nM)
(30 nM)
(100 nM)
(1 μΜ)
(3 ΜΜ)
(10 μΜ)
(1 μΜ)
(3 ΜΜ)
(10 μΜ)
(30 μM)
(100 μΜ)
(200 μΜ)
(100 nM)
(300 nM)
(1000 nM)
(100 nM)
(300 nM)
(1000 nM)
(3 nM)
(10 nM)
(30 nM)
(10 nM)
(30 nM)
(100 nM)
(A10)
(A3)
(A1)
(B10) (B3)
(B1)
Pregnenolone
17a-OH-Pregnenolone
Progesterone
17a-OH-Progesterone
DHEA
Androstenedione
Testosterone
17ß-Estradiol
11-Deoxycorticosterone
Corticosterone
Aldosterone
11-Deoxycortisol
Cortisol
| Hormone | Forskolin | Prochloraz | ||
|---|---|---|---|---|
| Fold changeª | SE | Fold changea,b | SE | |
| Pregnenolone | 2.1 | 0.2 | 1.2 | 0.1 |
| 17a-OH-pregnenolone | 6.4 | 1.0 | <- 5.1 | 1.2 |
| Progesterone | 1.3 | 0.1 | 10.0 | 0.6 |
| 17a-OH-progesterone | 3.8 | 0.8 | -2.0 | 0.4 |
| DHEA | 28.8 | 9.6 | <- 5.8 | 2.3 |
| Androstenedione | 6.5 | 1.1 | -31.9 | 5.5 |
| Testosterone | 3.6 | 0.8 | <- 17.9 | 6.1 |
| 17ß-Estradiol | 11.0 | 1.6 | -2.6 | 0.4 |
| 11-Deoxycorticosterone | 2.6 | 0.1 | -1.1 | 0.1 |
| Corticosterone | 6.6 | 1.9 | <- 7.1 | 1.9 |
| Aldosterone | 3.9 | 0.7 | -3.5* | 0.3 |
| 11-Deoxycortisol | 3.1 | 0.6 | -16.0 | 4.6 |
| Cortisol | 4.8 | 0.7 | <- 57.1 | 17.7 |
a Fold changes are calculated compared to the DMSO solvent control.
b Fold induction indicating smaller than (<) are calculated using the limit of quantification.
* p value <0.05.
p value <0.01.
| Gene | Forskolin | Prochloraz | ||
|---|---|---|---|---|
| Fold changeª | SE | Fold changeª | SE | |
| CYP1A1 | 1.5 | 0.3 | 6.7 | 1.5 |
| MC2R | 29.5 | 11.7 | 2.2 | 1.1 |
| StAR | 6.4 | 0.6 | -1.1 | 0.1 |
| CYP11A1 | 3.1 | 0.1 | 1.0 | 0.1 |
| HSD3B2 | 28.4 | 5.0 | 1.2 | 0.2 |
| CYP11B1 | 131.2 | 40.3 | 1.1 | 0.3 |
| CYP11B2 | 29.1 | 7.3 | 2.0 | 0.4 |
| CYP19A1 | 18.4 | 3.9 | 1.8 | 0.2 |
a Fold changes are calculated compared to the DMSO solvent control.
* p value <0.05.
p value <0.01.
inducing CYP1A1 but by decreasing CYP17 and CYP21 activity (Ohlsson et al., 2009). All three test concentrations of both diox- in-like compounds (TCDD and PCB126) induced CYP1A1
expression approximately ten-fold, suggesting that the lowest con- centrations already induced maximal AhR activation. Since TCDD and PCB 126 induced very different steroidogenic profiles, this sug- gests that the hormone alteration mechanisms in H295R cells are not, at least not solely, AhR related. This observation is in agree- ment with previous research (Lin et al., 2006).
The compounds HBCD, BDE-47, and MeHg did not significantly alter any hormone level, which is in accordance with the finding that they hardly changed expression of genes involved in steroido- genesis. Previous research showed that metabolites of BDE-47 can affect steroidogenesis in H295R cells (He et al., 2008; Song et al., 2008). It is known that H295R cells possess little CYP1B activity (Sanderson et al., 2001), which is apparently not enough to pro- duce metabolites of BDE-47 for measurable effects on steroido- genesis. The limited production of OH-metabolites in vitro has been shown before (Murk et al., 2013). In our experiments MeHg (up to 100 nM) had no effect, while (Knazicka et al., 2013) reported lowered testosterone production in H295R cells exposed to ≥1 uM mercury chloride. It cannot be excluded that the MeHg concentra- tions in our study (up to 100 nM) were below effect levels.
TCDD had little effect in the H295R cells, while TCDD exposure in humans is negatively correlated with testosterone levels (Egeland et al., 1994; Gupta et al., 2006). In vivo TCDD has been shown to decrease hormone production in mouse antral follicles (Karman et al., 2012) and the number and size of Leydig cells in rats, which is likely to affect testosterone production (Johnson et al., 1994). TCDD might affect steroidogenesis by interfering with the hypothalamic-pituitary-adrenal (HPA) axis, but these effects cannot be investigated by the current H295R cell model.
PCB 126 exposure increased CYP19A1 expression which corre- sponds with previous research (Li, 2007). This also is in agreement with the increased 17ß-estradiol levels we found, comparable to another study, exposing H295R cells for 48 h to 4 uM PCB126 (Kraugerud et al., 2010). In addition, Kraugerud and colleagues reported increased aldosterone and small increases in cortisol and progesterone levels. In a different study, exposing H295R cells for 10 days, the cortisol and aldosterone levels increased time-de- pendently (Li and Wang, 2005). Therefore, the 48 h exposure in the current study could have been too short. In our study PCB 153 not only decreased testosterone levels, which is in accordance to ear- lier research in H295R cells (Kraugerud et al., 2010), but also androstenedione, 17a-OH-progesterone, and 11-deoxycortisol levels.
Gene expression (fold change)
Gene expression (fold change)
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TCDD PBC126 PCB153 PFOS HBCD BDE-47 TBT MeHg
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Gene expression (fold change)
M.W. van den Dungen et al./Toxicology in Vitro 29 (2015) 769-778
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Fig. 2. Gene expression changes in H295R cells exposed for 48 h to eight POPs and mixtures A and B. Gene expression levels were normalized for HPRT expression. Values
represent fold change (mean + SE) compared to the solvent control (0.25% DMSO) from three independent experiments. * p-value <0.05 and ** p-value <0.01.
PFOS induced the most pronounced effects of the tested POPs on steroid hormone levels and gene expression. Earlier studies
reported similar effects of PFOS on 17ß-estradiol, at lower concen- trations (up to 300 nM) (Du et al., 2013), and testosterone, at high- er concentrations (up to 600 µM) (Kraugerud et al., 2011). At lower concentrations than we tested (300 nM), PFOS was shown to decrease testosterone levels in H295R cells (Du et al., 2013). Kraugerud et al. reported no change in cortisol levels after expo- sure to PFOS (Kraugerud et al., 2011), which we did find, but this could also be due to their less sensitive detection method using a radioimmunoassay with cross-reactivity, which is not a problem in LC-MS/MS measurements as we applied (Taylor et al., 2002). Levels of PFOS in human serum of occupationally exposed workers were as high as 114 µg/mL (=212 µM), but for the general popula- tion median values are not higher than 53 ng/ml (around 100 nM) (Fromme et al., 2009). Indeed, PFOS is negatively associated with serum testosterone levels in humans, but no effects on other repro- ductive hormones have been shown (Joensen et al., 2013). Already our lowest test concentration (30 uM) affected steroid hormone production, and this is the first study to show that PFOS has the
ability to alter most of the steroid hormone levels and gene expres-
sion levels in H295R cells.
The effects of TBT on hormone production and gene expression have, to the best of our knowledge, not been shown before in H295R cells. TBT lowered pregnenolone, 17a-OH-progesterone, and progesterone levels significantly, but in bovine adrenal cells increased 17a-OH-progesterone levels have been seen (Yamazaki et al., 2005). The fact that CYP19A1 gene expression levels were not significantly altered in our study was to be expected, as our exposure concentrations were lower (up to 30 nM) than reported for inhibition of aromatase activity by TBT (Kotake, 2012). The TBT levels reported for human blood, as high as 85 ng/ml (261 nM) with a geometric mean of 6.6 ng/ml (20 nM) (Kannan et al., 1999), are higher than the lowest observed effect concentra- tion of 10 nM in the H295R cells. As in vivo studies also indicate that TBT targets multiple steroidogenic tissues, including testos- terone levels and steroidogenic enzymes in the testes (Kim et al., 2008), endocrine disruption in humans by TBT cannot be excluded. DHEA levels were significantly decreased after exposure to either of both mixtures, but were not significantly changed after
exposure to any of the single POPs. As no cytotoxicity was observed, this suggests an additive effect if this effect can be repli- cated. Also for 17ß-estradiol the mixture effect of the compounds was additive. Mixture B consisted of mixture A plus HBCD, BDE-47, and MeHg that individually did not alter hormone levels, and TBT that individually decreased pregnenolone, 17a-OH-pro- gesterone, and progesterone. The latter were also decreased by exposure to mixture B, suggesting that TBT is responsible for the difference in response between the mixtures A and B. Aldosterone levels, which were only induced with the highest
PFOS concentration tested (200 µM), were significantly increased after exposure to mixture B (PFOS concentration of 100 µM) and not mixture A. Although synergistic effects in mixture B cannot be excluded, the effects might also be explained by addition of effects of the individual compounds just below the level of quantification.
An overview of how POPs modulated the steroidogenic pathway is presented in Figs. 3 and S3. Increased expression of steroido- genesis related genes did not always correspond to increased hor- mone levels, but only when the gene expression changes were
A
-2.1
-2.6
-1.5
StAR
Cholesterol
CYP11A1
Pregnenolone
(CYP17)
17a-OH-Pregnenolone
(CYP17)
DHEA
HSD3B2
HSD3B2
HSD3B2
PFOS (200 μM) exposure
1.4
1.4
1.4
CYP19A1
Progesterone
(CYP17)
17a-OH-Progesterone
(CYP17)
Androstenedione
1.5
(Estrone)
-1.7
(CYP21)
(CYP21)
1.5
(HSD17B)
(HSD17B)
CYP19A1
11-Deoxycorticosterone
11-Deoxycortisol
Testosterone
1.5
17ß-Estradiol
CYP11B1 /B2
CYP11B1
1.6
2.8
27.2
Corticosterone
Cortisol
2.7
CYP11B2
2.1
80.9
Aldosterone
1.5
B
-1.4
-1.8
-1.7
StAR
Cholesterol
CYP11A1
Pregnenolone
(CYP17)
17a-OH-Pregnenolone
(CYP17)
DHEA
PCB 153 (10 μM) exposure
HSD3B2
-1.8
HSD3B2
-1.4
HSD3B2
Progesterone
(CYP17)
17a-OH-Progesterone
(CYP17)
Androstenedione
OCYP19A1
(Estrone)
(CYP21)
-1.5
(CYP21)
-1.5
(HSD17B)
(HSD17B)
11-Deoxycorticosterone
11-Deoxycortisol
Testosterone
CYP19A1
17ß-Estradiol
CYP11B1 / B2
CYP11B1
Corticosterone
Cortisol
CYP11B2
Aldosterone
C
-1.9
-2.4
-2.9
StAR
Cholesterol
CYP11A1
Pregnenolone
(CYP17)
17a-OH-Pregnenolone
(CYP17)
DHEA
HSD3B2
HSD3B2
HSD3B2
Mix B1 exposure
1.5
(CYP17)
1.5
(CYP17)
1.5
CYP19A1
Progesterone
17a-OH-Progesterone
Androstenedione
2.3
(Estrone)
(CYP21)
(CYP21)
(HSD17B)
(HSD17B)
CYP19A1
11-Deoxycorticosterone
11-Deoxycortisol
Testosterone
2.3
17ß-Estradiol
CYP11B1 /B2
CYP11B1
3.1
18.7
Corticosterone
Cortisol
2.3
CYP11B2
1.4
55.4
Aldosterone
1.8
No difference in hormone levels / gene expression levels.
Statistical significant increased or decreased hormone levels / gene expression levels
Trend towards increased or decreased hormone levels / gene expression levels.
TCDD
PCB 126
PCB 153
PFOS
HBCD
BDE-47
TBT
MeHg
Mix A
Mix B
(10 nM)
(30 nM)
(100 nM)
(1 μΜ)
(3 ΜΜ)
(10 μM)
(1 μΜ)
(3 ΜΜ)
(10 μM)
(30 μΜ)
(100 μM)
(200 μM)
(100 nM)
(300 nM)
(1000 nM)
(100 nM)
(300 nM)
(1000 nM)
(3 nM)
(10 nM)
(30 nM)
(10 nM)
(30 nM)
(100 nM)
(A10)
(A3)
(A1)
(B10)
(B3)
(B1)
Cholesterol side-chain cleavage enzyme (CYP11A1)
3ß-Hydroxysteroid dehydrogenase (HSD3B2)
17a-Hydroxylase (CYP17A1)
17,20-Lyase (CYP17A1)
21-Hydroxylase (CYP21)
11ß-Hydroxylase (CYP11B1/2)
18-Hydroxylase (CYP11B2)
17ß-Hydroxysteroid dehydrogenase (HSD17B)
Aromatase (CYP19A1)
relatively high (>10-fold for CYP11B1 and CYP11B2), indicating that gene expression changes might be more sensitive to POP exposure than changes in hormone levels. Many steroidogenic related genes are still up-regulated after 48 h exposure, except HSD3B2 which was shown to return to basal levels after 48 h (Hilscherova et al., 2004). Increased hormone levels were always accompanied by increased gene expression levels, suggesting that increased hormone levels are, at least partly, caused by increased gene expression levels which is biologically plausible. On the other hand, decreased hormone levels did not coincide with decreased gene expression levels. This is not surprising, as a decrease in hor- mone levels is not necessarily due to decreased hormone produc- tion, but could also be the consequence of increased conversion into subsequent steroid hormones further down the pathway. This was also shown by the enzyme conversion flows in Fig. 3B. PFOS clearly up-regulates the enzymes toward the end of the path- way (e.g. 11ß-hydroxylase and 18-hydroxylase) without lowering the total hormone levels as shown by the stable cholesterol side- chain enzyme. This suggests that PFOS lowers hormone levels by increasing the conversion toward subsequent hormones.
Most in vitro studies, including the H295R assay, lack certain in vivo features such as the formation of POP metabolites and feed- back mechanisms at the level of the HPA axis. For example, in our study melanocortin 2 receptor (MC2R) gene expression levels were induced by PCB 126 and PFOS. This receptor is activated by adreno- corticotropic hormone (ACTH) (Dores, 2009) which is not produced by H295R cells. Therefore up-regulation of MC2R in this study is not expected to result in a functional effect. Up-regulation of MC2R in vivo, however, could result in over-stimulation of the adrenocortical cells in the presence of ACTH, leading to an unbal- anced steroid hormone synthesis. These, and other indirect effects among different tissues, make development of an in vitro battery of tests to study endocrine disruption with relevance for the in vivo situation very challenging, but not impossible as has been rea- soned for studying thyroid hormone disruption (Murk et al., 2013). In addition to multiple effect mechanisms, also multiple xenobiotic compounds, including their bioactive OH-metabolites, are present together in real life and, as shown in this study, may lead to additive effects. For example, average PCB serum levels in humans are around 210 ng/ml (approximately 0.6 uM) (Humblet et al., 2011), which is only about 10 times lower than the lowest observed effect concentration of 10 uM in our experiments for the individual compounds PCB 126 and PCB 153. Therefore, it can- not be excluded that human steroidogenesis can be affected by
PCBs and related (OH-) POPs, present as mixtures at effect levels (Montano et al., 2013).
In conclusion, this study shows that ubiquitous seafood POPs, such as TCDD, PCB 126, PCB 153, PFOS, TBT and mixtures thereof, are capable of altering steroidogenesis in H295R cells at non cyto- toxic and relevant concentrations. The overall effects of the mix- tures in this study suggest additivity, not synergy. As POPs have been reported to affect steroidogenesis via several mechanisms in different tissues, adverse effects in humans cannot be excluded.
Funding sources
This research was supported by the graduate school VLAG (advanced studies in food technology, agrobiotechnology, nutrition and health sciences) and in part by the Netherlands Genomics Initiative, the Netherlands Organisation for Scientific Research, and the Netherlands Toxicogenomics Centre (Grant No. 05060510).
Author contributions
The manuscript was written through contributions of all authors. All authors have given approval to the final version of the manuscript.
Conflict of Interest
The authors declare that there are no conflicts of interest.
Transparency Document
The Transparency document associated with this article can be found in the online version.
Appendix A. Supplementary material
Primer sequences and their product length are provided in Table S1. Hormone levels measured after exposure to the POP mix- tures are presented in Fig. S1. Gene expression data of genes not affected by any of the POPS are displayed in Fig. S2. Table S2 con- tains information of all tested POPs and their corresponding hor- mone levels, which is the basis of both heat maps (Figs. 1 and 3B). Supplementary data associated with this article can be found,
in the online version, at http://dx.doi.org/10.1016/j.tiv.2015.03. 002.
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