ELSEVIER
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Journal of Hazardous Materials
journal homepage: www.elsevier.com/locate/jhazmat
HAZARDOUS MATERIALS
Alteration of sex hormone levels and steroidogenic pathway by several low molecular weight phthalates and their metabolites in male zebrafish (Danio rerio) and/or human adrenal cell (H295R) line
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Juhae Sohnª, Sujin Kimª, Jan Koschorreckb, Younglim Khoc, Kyungho Choia,*
a Department of Environmental Health Sciences, School of Public Heath, Seoul National University, Seoul 08826, Republic of Korea
b Federal Environment Agency (Umweltbundesamt), 06813 Dessau-Rosslau, Germany
” Department of Health, Environment and Safety, Eulji University, Seongnam 13135, Republic of Korea
HIGHLIGHTS
. Low molecular weight phthalates and metabolites were tested for hormone disruption.
. In fish and adrenal cell model, decreased testosterone was observed by phthalates.
. Sex hormone disruption could be explained by altered steroidogenesis.
. Hydrolytic metabolites except MiBP also showed the same disruption mechanisms.
ARTICLE INFO
Article history: Received 19 May 2016
Received in revised form 9 July 2016
Accepted 3 August 2016
Available online 4 August 2016
Keywords: Low molecular weight phthalates Hydrolytic metabolites Fish
Sex hormones
ABSTRACT
Low molecular weight phthalates, such as diethyl phthalate (DEP), benzyl butyl phthalate (BBzP), or diisobutyl phthalate (DiBP), are suspected to disrupt endocrine system. However, their adverse effects on sex steroid hormones and underlying mechanisms are not well-documented. The aim of this study is to investigate the effects of major low molecular weight phthalates (LMWPs), i.e., DEP, BBzP, and DiBP, and their hydrolytic metabolites, on sex steroid hormone system, employing male zebrafish and/or a human adrenocortical carcinoma (H295R) cell. In male zebrafish, 14-day exposure to DEP, BBzP, or DiBP significantly decreased testosterone (T) concentrations. All test compounds significantly up-regulated cyp19a gene expression, and down-regulated star and 36 hsd genes in the male fish. In H295R cell, all test compounds except monoisobutyl phthalate (MiBP) reduced T concentrations and increased E2/T ratio. Gene expression changes in H295R cell, e.g., significant down-regulation of StAR gene and up-regulation of CYP19A gene, supported depressed synthesis of sex hormones in the adrenal cell. Our results show that not only DEP, BBzP, and DiBP, but also their hydrolytic metabolites disrupt sex hormone balances through modulating key steroidogenic genes in the human adrenal cells and in zebrafish.
@ 2016 Elsevier B.V. All rights reserved.
Abbreviations: ANOVA, one-way analysis of variance; 38 HSD2, 3ß- hydroxysteroid dehydrogenase; 17ß hsd, 17ß-hydroxysteroid dehydrogenase; 18s rrna, 18s ribosomal rna; BBzP, benzylbutyl phthalate; CHO cell, Chinese hamster ovary cell; cyp11a, cholesterol side-chain cleavage enzyme gene; cyp17, 17a- monooxygenase; CYP19A, cytochrome P450; DBP, dibutyl phthalate; DEP, diethyl phthalate; DiBP, diisobutyl phthalate; DMEM/F12, Dulbecco’s modified Eagle’s medium and Ham’s F-12 nutrient mixture; DMP, dimethyl phthalate; DMSO, dimethyl sulfoxide; DnBP, di-n-butyl phthalate; E2, 17ß-estradiol; ELISA, enzyme- linked immunosorbent assay; H295R cell, Human adrenalcorticoid cell; LMWPs, Low molecular weight phthalates; MBzP, monobenzyl phthalate; MEP, monoethyl phthalate; MiBP, monoisobutyl phthalate; OH-MiBP, 3OH-mono-methylpropyl phthalate; PVC, polyvinyl chloride; qRT-PCR, quantitative real-time polymerase chain reaction; StAR, steroidogenic acute regulatory protein; T, testosterone; vtg, vitellogenin.
* Corresponding author. E-mail address: kyungho@snu.ac.kr (K. Choi).
http://dx.doi.org/10.1016/j.jhazmat.2016.08.008
0304-3894/ 2016 Elsevier B.V. All rights reserved.
1. Introduction
Phthalates are group of several synthetic compounds which are used in polyvinyl chloride (PVC), food containers, commercial prod- ucts, and personal care products [1-3]. Phthalates can be divided into two groups by size; i.e., “low molecular weight phthalates (LMWPs)” with carbon chains of 6 or less, and “high molecular weight phthalates” with more than 6 [4]. LMWPs like diethyl phtha- late (DEP), benzyl butyl phthalate (BBzP), and diisobutyl phthalate (DiBP) (Table 1), have been widely used in toys, childcare prod- ucts, personal-care products, some cosmetics, fragrances, lacquers, and varnishes [5]. In addition, di-n-butyl phthalate (DnBP), and dimethyl phthalate (DMP) have been reported to be used in cos- metics and personal care products [6]. Koo and Lee [7] reported detection of DEP, BBzP, and dibutyl phthalate (DBP) in cosmetic
| Compound | CAS number | Structure | Usage | LogKow | Water solubility (g/L) | Molecular weight (g/mol) |
|---|---|---|---|---|---|---|
| Diethyl phthalate (DEP) | 84-66-2 | cosmetics, fragrances | 2.42 | 1.08 | 222.24 | |
| 1.21 | 194.18 | |||||
| Monoethyl phthalate (MEP) | 2306-33-4 | metabolite of DEP | 1.86 | |||
| Benzylbutyl phthalate (BBzP) | 85-68-7 | PVC, floor tiles | 4.7 | 3.8 × 10-3 | 312.39 | |
| Monobenzyl phthalate (MBzP) | 2528-16-7 | metabolite of BBzP | 3.07 | 0.052 | 256.3 | |
| Diisobutyl phthalate (DiBP) | 84-69-5 | cosmetics, nail polish | 4.11 | 2.0 × 10-5 | 278.35 | |
| 30833-53-5 | metabolite of DiBP | 2.77 | 0.98 | 222.24 | ||
| Monoisobutyl phthalate (MiBP) |
| Matrix | Compound | Country | Location | No. of sampling site | Sampling date | Detection frequency (%) | Median (ng/L) | Lowest level (ng/L) | Maximum level (ng/L) | References |
|---|---|---|---|---|---|---|---|---|---|---|
| Surface water | DEP | China | Songhua River | 10 | Oct., Dec., 2011 | 100 | 2.35 | 1.33 | 6.67 | [42] |
| China | Guangzhou River | 15 | May, 2005 | 100 | 31 | 15 | 320 | [43] | ||
| BBzP | China | Songhua River | 10 | Oct., Dec., 2011 | 18 | 2.49 | n.d. | 4.39 | [44] | |
| DiBP | China | – | 13 | Aug., 2009 | – | 37 | 8.3 | 200 | [45] | |
| China | Guangzhou River | 15 | Apr., 2006 | 100 | 495 | 281 | 1750 | [10] | ||
| 15 | Aug., 2006 | 100 | 1280 | 884 | 2860 | |||||
| 15 | Dec., 2006 | 100 | 321 | 161 | 971 | |||||
| Sediment | DEP | China | Songhua River | 6 | Oct., Dec., 2011 | 100 | 31.51ª | 26.7 | 38.24 | [42] |
| China | Qiantang River | 23 | Apr., 2011 | 95.65 | 7 | n.d. | 218 | [46] | ||
| BBzP | China | Songhua River | 6 | Oct., Dec., 2011 | 33 | 62.36ª | n.d. | 96.32 | [44] | |
| China | Guangzhou River | 15 | May, 2005 | 73 | 34 | n.d. | 280 | [43] | ||
| DiBP | China | Zhujiang River | 11 | – | 100 | 2.29 | 0.561 | 12.4 | [47] | |
| China | Dongjiang River | 21 | – | 100 | 0.811 | 0.108 | 5.28 | |||
| China | Qiantang River | 23 | – | 100 | 0.118 | 0.019 | 0.769 | [46] |
n.d .: not detected; -: not provided.
a Mean.
products including perfume. Frequent use of these phthalates has led to their widespread occurrences in the environment (Table 2); DEP and BBzP have been detected in French rivers at levels up to 1870 ng/L and 12100 ng/L, respectively [8,9]. DiBP was detected up to 2860 ng/L in Guangzhou River [10]. These phthalates have been also reported in sediments of various rivers of China, such as Songhua, Qiantang, Guangzhou, Zhujiang, and Dongjiang rivers (Table 2).
In humans, LMWPs can be easily metabolized in the body. DEP, BBzP, and DiBP are hydrolyzed into their monoester forms of monoethyl phthalate (MEP; metabolism ratio; 73%), monoben-
zyl phthalate (MBzP; 69%), and monoisobutyl phthalate (MiBP; 73%), respectively (Table 1) and excreted in urine (Table 3) [11]. In addition, 3OH-mono-methylpropyl phthalate (OH-MiBP) has often been detected in human urines at the levels around a half of MiBP [12]. These metabolites have been also detected in breast milk. For example, MEP, a major metabolite of DEP, was reported in 100% of breastmilk samples in Korean women (n =62) [13].
In humans, phthalate exposure has been associated with adverse effects on male reproductive system. Negative associa- tions between urinary phthalate metabolites and anogenital index (AGI), semen quality, or reproductive hormones have been reported
| Compound | Country | Age (year) | Sampling size | Sampling | Detection frequency (%) | Median (µg/L) | Lowest level (µg/L) | 95th percentile (µg/L) | Maximum level (μg/L) | Reference |
|---|---|---|---|---|---|---|---|---|---|---|
| MEP | Korea | 19-27 | 40 (male) | Jul. 2011 | 100 | 9.9 | – | 53.3 | – | [48] |
| 40 (female) | 100 | 22.4 | – | 227 | ||||||
| Norway | 8.8-12.5 | 623 | 2001-2004 | 100 | 56.7 | 8.5 | 360.2 | 6006 | [49] | |
| Germany | 6.8±0.6 | 104 | 2007-2009 | 100 | 33.6 | 2.7 | 391 | 1787 | [50] | |
| 39.2±4.6 | 103 | 100 | 53.8 | 5.1 | 265 | 756 | ||||
| USA | 18-35 | 300 | 1998-2006 | 100 | 143.5 | – | 335.2ª | – | [51] | |
| MBzP | Norway | 8.8-12.5 | 623 | 2001-2004 | 100 | 29.3 | 2.1 | 128.7 | 6710 | [49] |
| Mexico | 813 | 49 (male) | 2010 | 98 | 3.4 | – | 46.8 | 32.5 | [52] | |
| 50 (female) | 100 | 2.9 | – | 29.6 | 48.4 | |||||
| MiBP | Korea | 19-27 | 40 (male) | Jul. 2011 | – | 8.1 | – | 22.2 | – | [48] |
| 40 (female) | – | 9.6 | – | 22.4 | ||||||
| MiBP | Norway | 8.8-12.5 | 623 | 2001-2004 | 100 | 49.2 | 3.1 | 231 | 1480 | [49] |
| Germany | 6.8±0.6 | 104 | 2007-2009 | 100 | 68.7 | 1.8 | 340 | 1284 | [50] | |
| 14-60 | 23 (male) | Apr .- Oct. 2005 | 100 | 47.3 | 23.1 | 107.2 | 119.7 | [53] | ||
| 14-60 | 27(female) | 100 | 36.1 | 15.7 | 109.1 | 163.8 | ||||
| Mexico | 813 | 49 (male) | 2010 | 96 | 2.1 | – | 7.3 | 39.4 | [52] | |
| 50 (female) | 98 | 2.2 | – | 12.1 | 33.9 | |||||
| USA | 24±6.2 | 382 | 1998-2002 | 97.4 | 6.2 | 0.26b | 12ª | 131 | [54] |
-: not provided.
a 75th percentile.
b Value of LOD was determined at 0.26 µg/L.
[14-16]. In addition, associations with decreased semen quality (MBzP and MEP) [17], and with DNA damage and decreased motility of sperm (MEP) have been reported [18-20].
The anti-androgenicity of LMWPs has been reported in vitro and in experimental organisms [21-25]. However, the endocrine disrupting effects of LMWPs in aquatic organisms have not been sufficiently understood. Only one report about estrogenic activity of BBzP in transgenic medaka fish is available [26]. For metabolites of LMWPs, such information is even more limited, and remained as a big knowledge gap (Table 4).
The aim of this study is to investigate the effects of LMWPs and their metabolites on sex steroid hormone balance and the asso- ciated transcriptional changes in steroidogenic pathway using a steroidogenic cell line and a fish. The results of this study will fill information gaps in the potentials and mechanisms of sex hor- mone disruption by LMWPs, and in the future help identify safer alternatives.
2. Materials and methods
2.1. Chemicals
DEP (CAS No. 84-66-2) was purchased from Sigma-Aldrich (St. Louis, MO, USA). MEP (CAS No. 2306-33-4), BBzP (CAS No. 85-68-7), MBzP (CAS No. 2528-16-7), DiBP (CAS No. 84-69-5), and MiBP (CAS No. 30833-53-5) were purchased from AccuStandard (New Haven, CT, USA). Dimethyl sulfoxide (DMSO) was used as solvent in both in vitro and in vivo studies. The final concentration of the solvent in the exposure media was 0.005% (v/v) for zebrafish exposure, and 0.1% (v/v) for H295R cell assays.
2.2. Zebrafish culture and exposure
Wild type adult male zebrafish (~4 months old) were obtained from a commercial vendor (Gangnam Aquaria, Suwon, Korea). Before exposure, fish were acclimated in dechlorinated tap water for 7 days. For DEP or BBzP exposure, 16 male zebrafish were divided into four replicates with four fish per replicate (2 L beaker). For DiBP exposure, 9 male zebrafish were divided into three repli- cates with three fish per replicate. A series of concentrations (0.08,
0.4, 2, or 10 mg/L DEP; 0.02, 0.1, 0.5, or 2.5 mg/L BBzP; and 0.0008, 0.004, 0.02, or 0.1 mg/L of DiBP) was prepared with dechlorinated water. The exposure concentrations were determined at sublethal levels based on preliminary range finding tests. The exposure was carried out for 14 days following OECD test guideline No. 204 [27]. Fish were maintained at 26±1℃ under 14:10h light:dark pho- toperiod. Fish were fed with freshly hatched Artemia nauplii twice a day. The exposure media were renewed every other day, and the water quality parameters including dissolved oxygen, pH, conduc- tivity, and temperature were monitored routinely.
After 14 days of exposure, fish were sacrificed, and liver and testis were dissected out from each fish. Blood samples were col- lected using a glass capillary tube from caudal vein. Due to limited volume of the blood sample, the samples from three fish were pooled for each replicate. Plasma was separated by centrifugation (8000 rpm for 10 min at 4 ℃), and stored at -80℃ until sex steroid hormone analysis. No mortality was observed during the fish expo- sure.
2.3. H295R cell culture and exposure
H295R cells were cultured in a mixture of Dulbecco’s modi- fied Eagle’s medium and Ham’s F-12 nutrient mixture (DMEM/F12) (Sigma D-2906; Sigma-Aldrich) supplemented with 1% ITS + Premix (BD Biosciences, San Jose, CA, USA), 2.5% Nu-Serum (BD Bio- sciences), and 1.2 g/L Na2CO3 (Sigma Aldrich). Medium was renewed every other day, and cells were subcultured when cell confluence reached 80%. Cells were cultured at 37 ℃ in a 5% CO2 atmosphere [28]. The exposure concentrations were chosen at the levels that resulted in >80% cell viability based on WST-1 cell pro- liferation assay (Roche Applied Science, Mannheim, Germany). For the cell bioassay, H295R cells were seeded into the media at a den- sity of 3.0 x 105 cells/mL in 24-well plates. Next day, the cells were dosed with various levels of LMWPs (0.004, 0.04, or 0.4 mM for DEP and MEP; 0.001, 0.01, or 0.1 mM for BBzP and MBzP; and 0.0001, 0.001, or 0.01 mM for DiBP and MiBP). Each treatment included three replicates. The exposure duration was 48 h. Media were col- lected for sex steroid hormone measurement, and cells for gene expression.
| Compound | Cell/organism | Results | Reference |
|---|---|---|---|
| DEP | H295R cell | T concentration Į | This study |
| Sertoli cell | Androgen receptor activity Į | [41] | |
| H295R cell | E2 concentration 1 T concentration | [24] | |
| Zebrafish (male) | T concentration Į | This study | |
| E2 concentration Į | |||
| Common Carp (male) | VTG 1 | [1] | |
| MEP | H295R cell | T concentration | This study |
| E2 concentration Į | |||
| BBzP | H295R cell | T concentration ¿E2 concentration 1 | This study |
| Yeast | Estrogenic activity 1 | [40] | |
| MCF-7 cell | Estrogenic activity 1 | [25] | |
| MVLN cell | ER transactivation 1 | [24] | |
| Zebrafish (male) | T concentration Į | This study | |
| Transgenic medaka | Estrogenic activity 1 | [26] | |
| Sprague-Dawley rat (female) | Fetal T concentration \ | [23] | |
| Fathead minnow | Number of spawning \ | [55] | |
| Sprague-Dawley rat (male) | Ins13 Į | [56] | |
| T concentration \ | |||
| Albino rat | Sperm count _ | [57] | |
| T concentration Į | |||
| BBzP | Sprague-Dawley rat (male) | T concentration Į AGD Į | [22] |
| Wister rat (female) | LH secretion \ | [58] | |
| Sprague-Dawley rat | T concentration Į | [59] | |
| AGD Į | |||
| MBzP | H295R cell | T concentration Į E2 concentration f | This study |
| DiBP | H295R cell | T concentration Į | This study |
| Yeast | Estrogenic activity 1 | [40] | |
| Zebrafish (male) | T concentration Į star \ cyp19a 1 | This study | |
| Sprague-Dawley rat | T concentration Į | [33] | |
| Sprague-Dawley rat, Wister rat | T concentration Į | [21] | |
| star \ | |||
| cyp11a \ | |||
| Wister rat | Testicular weight \ | [32] | |
| AGD Į | |||
| T concentration Į | [34] | ||
| AGD Į | |||
| Aromatase 1 | |||
| MiBP | H295R cell | T concentration f | This study |
2.4. Measurement of sex hormones and gene expressions
For measurement of sex steroid hormones, enzyme-linked immunosorbent assay (ELISA) was employed with commercial kits (Cayman Chemical; 17ß-estradiol [Cat No. 582251], and testos- terone [Cat No. 582701]). Sex hormone was extracted from the H295R media or the fish plasma, using diethyl ether. Briefly, 500 µL of H295R cell medium or 4 uL of fish plasma was diluted with 400 µL of Ultrapurewater, and extracted with 2 mL of diethyl ether twice by centrifuging at 2000 g for 10 min. Following evaporation of diethyl ether, the residue was dissolved in 120 uL (zebrafish) or 300 µL (H295R cell) of EIA buffer for ELISA assay [29].
For gene expression analysis, H295R cells and fish organ samples were collected and homogenized. Then, the total RNA was extracted using RNeasy mini kit (Qiagen). RNA quality and concentration are determined by a use of a Gen5 2.05 (BioTek, Winooski, VT, USA). RNA was then diluted to 100 ng/pL, and complementary DNA was synthesized using iScriptTM cDNA synthesis kit (BioRad, Hercules, CA, USA) following the manufacture’s protocol. Quantitative real- time PCR (qRT-PCR) was performed with the 20 pL of qRT-PCR reaction mix consisting of 10 pL of LightCycler-DNA Master SYBR Green I mix (Roche Diagnostics Ltd., Lewes, UK), 4.4 pL of nuclease
free water, and 1.8 pL of forward, and reverse primer (10 pmol). After seeding 18 pL of pre-mix, 2 uL cDNA templates are added into each well, and qRT-PCR was performed using Light Cycler 480 (Roche Applied Science, Indianapolis, IN, USA). The relative expression level of target gene was calculated with the threshold cycle (Ct) value using 244 Ct method [30]. Those genes that showed significant alteration in male zebrafish (steroidogenic acute reg- ulatory protein; star, 3ß-hydroxysteroid dehydrogenase; 3Bhsd2, and cytochrome P450; cyp19a) were chosen and were measured in H295R cell assay. As a housekeeping gene, 18 s ribosomal RNA; 18s rrna and beta-actin; B-actin were used for zebrafish and H295R cell, respectively. Primer sequences for selected genes and house- keeping genes are listed in Table S1.
2.5. Chemical analysis
LMWPs and their metabolites were measured in the exposure media of zebrafish using a high performance liquid chromatog- raphy (Nanospace SI-2, Shiseido, Japan) with triple quadrupole tandem mass spectrometry (LC/MS/MS; API 4000, Applied Biosys- tems, Foster City, CA, USA) following Kho et al. [31]. The separation of phthalate metabolites were accomplished using an analytical
| Exposure chemical | Nominal concentration (mg/L) | Exposure time (h) | Measured concentration (parent; mg/L) | Measured concentration (metabolite; mg/L) |
|---|---|---|---|---|
| DEP | 0 | 0 | n.d. | n.d. |
| 48 | n.d. | n.d. | ||
| 0.08 | 0 | 0.070±0.027 | n.d. | |
| 48 | n.d. | n.d. | ||
| 0.4 | 0 | 0.338±0.071 | n.d. | |
| 48 | n.d. | n.d. | ||
| 2 | 0 | 1.327 ±0.248 | 0.014±0.005 | |
| 48 | n.d. | n.d. | ||
| 10 | 0 | 6.253 ±1.330 | 0.417 ±0.316 | |
| 48 | n.d. | n.d. | ||
| BBzP | 0 | 0 | n.d. | n.d. |
| 48 | n.d. | n.d. | ||
| 0.02 | 0 | 0.010±0.029 | n.d. | |
| 48 | n.d. | n.d. | ||
| 0.1 | 0 | 0.066±0.014 | n.d. | |
| 48 | n.d. | n.d. | ||
| 0.5 | 0 | 0.439±0.072 | 0.001 ±0.003 | |
| 48 | n.d. | n.d. | ||
| 2.5 | 0 | 1.673 ±0.221 | n.d. | |
| 48 | n.d. | n.d. | ||
| DiBP | 0 | 0 | 0.018 ±0.011 | n.d. |
| 48 | 0.012±0.008 | n.d. | ||
| 0.0008 | 0 | 0.016±0.009 | n.d. | |
| 48 | 0.008 ±0.004 | n.d. | ||
| 0.004 | 0 | 0.023 ±0.014 | n.d. | |
| 48 | 0.016±0.005 | n.d. | ||
| 0.02 | 0 | 0.049 ±0.020 | n.d. | |
| 48 | 0.009 ±0.005 | n.d. | ||
| 0.1 | 0 | 0.197±0.128 | n.d. | |
| 48 | 0.024±0.017 | n.d. |
n.d .: not detected.
Limit of quantitation (LOQ) was determined at 0.376, 0.037, 0.079, 0.072, 0.184, 0.177 (µg/L) for DEP, MEP, BBzP, MBzP, DiBP, and MiBP respectively. The results are shown as mean ± SD of three replicates.
column (Imtackt Cadenza CD C18, 75_2.0 mm, 3mm). The 0.1% acetic acid in water (A) and 0.1% acetic acid in acetonitrile (B) were used as mobile phase. The mobile phase was operated in isocratic mode. Injection volume was 5 pL and flow rate was 200 L/min. The target compounds were analyzed using electrospray ionization (ESI) positive mode for parent phthalates, and negative mode for metabolites. LOQ was determined at 0.376 µg/L for DEP, 0.037 µg/L for MEP, 0.079 µg/L for BBzP, 0.072 µg/L for MBzP, 0.184 µg/L for DiBP, and 0.177 µg/L for MiBP.
2.6. Statistical analysis
Normality of data and homogeneity of variances were analyzed by Shapiro-Wilk’s test and Levene’s test, respectively. When nec- essary, log-transformation was conducted. For group comparison with control, one-way analysis of variance (ANOVA) followed by Dunnett’s t-test was employed. For trend analysis, linear regression was performed and the significance of the slope was determined. P-values less than 0.05 were considered statistically significant. All results are shown as mean ± standard deviation (SD). SPSS 20.0 for Windows (SPSS Inc., Chicago, IL, USA) was used.
3. Results
3.1. Chemical analysis
The measured phthalate levels at 0h were slightly lower than the nominal concentrations (Table 5). After 48 h of exposure, how- ever, none of parent compounds were detected except for DiBP. The hydrolytic metabolites of the test LMWPs were not detected in most cases, except for at higher levels of DEP and BBzP at 0h. After 48 h
of exposure, none of the test hydrolytic metabolites were detected in the media.
3.2. Effects of LMWPs on male zebrafish
3.2.1. Plasma sex hormones
The effects of DEP, BBzP, and DiBP on 17B-estradiol (E2), testos- terone (T), and E2/T ratio in male zebrafish are shown in Fig. 1. All three test compounds significantly decreased the concentration of T (Fig. 1). For DEP, E2 levels were also significantly decreased at 10 mg/L. For BBzP exposure, dose-dependent decrease of E2 was observed.
3.2.2. Steroidogenic gene expressions
Following exposure to DEP, BBzP, or DiBP, in gonad of male zebrafish, expressions of several steroidogenic genes were signifi- cantly affected (Fig. 2). The expression of star gene was significantly down-regulated after exposure to 10 mg/L DEP and 0.1 mg/L DiBP (Fig. 2A). Similarly, a trend of down-regulation in star gene was observed in male gonad by BBzP exposure (B =- 0.17, p<0.05). In addition, exposure to LMWPs led to dose-dependent down- regulation of 30 hsd gene at marginal statistical significance (p<0.10) (Fig. 2C). Gene expression of 17ß-hydroxysteroid dehy- drogenase (17§ hsd; Fig. 2D) was down-regulated at the highest level (0.1 mg/L) of DiBP. Expressions of cholesterol side-chain cleav- age enzyme gene (cyp11a; Fig. 2B) and steroid 17a-monooxygenase (cyp17; Fig. 2E) gene were not affected. In addition, although marginally significant, negative trends were observed in cyp11a gene expressions by DEP exposure (p=0.051), and cyp17 gene expressions by DiBP exposure (p=0.061). In contrast, up-regulation of cyp 19a gene was clearly shown following exposure to DEP, BBzP, and DiBP. In liver, down-regulation of vtg gene was observed after
(A)
DEP
BBzP
DiBP
17ß-estradiol fold change
2.0
B =- 0.041
₿ = - 0.151
B =- 3.551
p = 0.001
p = 0.027
p = 0.052
1.5
1.0
T
*
T
0.5
0.0
(B)
Testosterone fold change
1.4
B =- 0.054
B = - 0.206
B =- 0.306
T
1.2
p = 0.001
p = 0.004
p = 0.005
1.0
T
T
0.8
0.6
T
*
*
0.4
*
*
*
*
0.2
0.0
(C)
B =0.052
B = 0.337
B = 8.155
6
p = 0.04
p = 0.048
p =0.337
E2/T ratio
4
*
2
T
T
T
T
T
T
0
SC 0.08 0.4 2 10
SC
0.02
0.1
0.5
2.5
C 0.0008
00
4 0.0
0.1 mg/L
exposure to 10 mg/L of DEP, for the other LMWPs, no significant change was observed (Fig. 3).
3.3. Effects of LMWPs and their metabolites on H295R cells
3.3.1. Sex hormones
In H295R cells, all test chemicals except MiBP showed increase of E2/T ratio (Fig. 4). While the changes in E2 level varied, e.g., E2 level increased by BBzP and its metabolite, but decreased by MEP, these test chemicals except for MiBP led to significant dose- dependent decrease of T (Fig. 4). T levels measured following the exposure to MiBP showed the opposite pattern; T levels increased by MiBP exposure, and E2/T ratio decreased by MiBP exposure.
3.3.2. Steroidogenic gene expressions
Significant changes in expression were observed for several genes in H295R cells following exposure to six compounds (Fig. 5). In general, StAR gene was down-regulated, and CYP19A gene expressions were up-regulated in all test LMWPs and their metabo-
lites, except for MiBP. The 3BHSD2 gene was down-regulated by DEP and its metabolite, but for other test compounds such trend was not evident. Again, the regulatory change by MiBP was differ- ent from the other chemicals: While StAR and 3ßHSD2 genes were not affected significantly, the CYP19A gene showed clear down- regulating trend (Fig. 5C).
4. Discussion
4.1. Reduced androgenicity by alteration of steroidogenic pathway in male zebrafish
DEP, BBzP, and DiBP exposure led to unanimous decrease of both E2 and T levels in male zebrafish (Fig. 1). Because the extent of T decrease was much greater than that of E2 decrease, E2/T ratio, a measure of relative estrogenicity, generally increased in the male fish (Fig. 1C). Decrease of T levels by exposure to LMWPs including their mixture has been reported elsewhere in many experimen- tal models, but most studies are limited to rodents [21,23,32,33].
(A)
DEP
BBzP
DiBP
(B)
DEP
BBzP
DiBP
2.0
₿ = - 0.043
₿ = - 0.170
₿ = - 8.649
2.5
B = - 0.046
B = - 0.124
B = - 4.679
star fold change
p=
0.004
p= 0.009
p= 0.002
cyplla fold change
2.0
p= 0.051
p= 0.105
p= 0.136
1.5
1.5
1.0
*
T
1.0
0.5
#
0.5
0.0
0.0
SC
0.08
0.4
2
10
SC 0.02 0.1 0.5 2.5
SC 0.0008 0.004 0.02 0.1 mg/L
SC 0.08 0.4 2 10
SC 0.02 0.1 0.
2.5
SC 0.0008 0.004 0.02 0.1mg/L
(C)
DEP
BBzP
DiBP
(D)
DEP
BBzP
DiBP
2.0
3 B hsd fold change
B = - 0.036
B = - 0.225
B = - 5.047
2.5
17ß hsd fold change
B = 0.011
B = - 0.060
B = - 4.018
p=
0.075
p < 0.001
p= 0.065
2.0
p=0.620
p= 0.472
p =
0.005
1.5
1.5
1.0
T.
T
1.0
T
*
0.5
0.5
0.0
0.0
SC
0.08
0.4
2
10
SC 0.02 0.1 0.5
2.5
SC 0.0008 0.004 0.02 0.1 mg/L
SC
0.08
0.4
2
10
SC
0.02
0.1
0.5
2.5
SC 0.0008 0.004 0.02 0.1 mg/L
(E)
DEP
BBzP
DiBP
(F)
DEP
BBzP
DiBP
3.0
B= 0.033
₿ = - 0.047
B = - 6.400
5
cyp17 fold change
2.5
p = 0.286
p = 0.610
p = 0.061
cyp19a fold change
B= 0.150
B = 0.391
₿ = 9.459
4
P
E
0.001
p= 0.023
p= 0.001
2.0
*
3
*
*
1.5
*
容
1.0
2
0.5
1
T
T
T
0.0
0
SC 0.08 0.4 2
10
SC 0.02 0.1 0.5
2.5
SC 0.0008 0.004 0.02 0.1 mg/L
SC
0.08
0.4
2
10
SC 0.02 0.1 0.5
2.5
SC 0.0008 0.004 0.02 0.1 mg/L
DEP
BBzP
DiBP
2.5
B = - 0.079
B = - 0.078
B = 0.668
vtg fold change
2.0
p < 0.001
p = 0.283
p = 0.688
1.5
1.0
0.5
*
0.0
SC
: 0.08 0.4 2
10
SC
0.02
0.1
0.5
2.5
SC 0.0008 0.004 0.02 0.1 mg/L
Decrease of T level can be explained by down-regulation of star gene (Fig. 2A) and up-regulation of cyp19 gene (Fig. 2F). The star gene is an important player in the first stage of steroidogenic pathway, and its protein is responsible for the transport of cholesterol into the mitochondrial membrane [34]. Therefore, down-regulation of star gene is associated with reduced cholesterol uptake [35], leading to decreased steroid hormone synthesis (Fig. 1). Down-regulation of star gene by exposure to phthalates has been reported in male Sprague Dawley rat [33]. The cyp19 or aromatase is an enzyme responsible for conversion of T to E2 [28]. Increased expression of cyp19 gene can lead to increased conversion of T to E2, aggravat-
ing deficiency of T level in the male fish. In Wistar rat, following exposure to DiBP, decrease of T levels was observed along with up- regulation of ovarian aromatase gene [32,36]. Down-regulation of 3 hsd gene (Fig. 2C) is also associated with decreased T levels, as 3ß hsd is a steroid-metabolizing enzyme essential for conversion of dehydroepiandrosterone to androstendione. As androstendione is converted to T by enzyme 178 hsd [37], the down-regulation of both 3B hsd and 170 hsd genes can lead to reduced T level.
While steroidogenesis is generally suppressed, significant up- regulation of cyp19a gene could render E2 level relatively less influenced. That is, even though cyp19a gene was up-regulated,
DEP
MEP
BBzP
MBzP
DiBP
MiBP
17ß-estradiol (pg/ml media)
(A)
250
ß = - 0.505
₿ = - 1.211
B = - 3.322
₿ = - 5.738
B = 2.831
₿ = 5.421
200
p = 0.091
p = 0.002
p = 0.002
p < 0.001
p = 0.830
p =0.579
150
*
100
50
0
(B)
Testosterone (pg/ml media)
4000
₿ = - 1.282
₿ =- 1.582
B = - 6.095
₿ = - 5.646
₿ = - 63.662
₿ = 24.961
3000
p = 0.017
p = 0.001
p < 0.001
p = 0.001
p = 0.009
p = 0.187
2000
* *
*
*
1000
*
*
*
*
*
0
(C)
12
B = 10.443
฿ = 5.449
₿ =43.177
B = 44.378
₿ = 195.598
₿ = - 14.232
10
p < 0.001
p = 0.001
p <0.001
p < 0.001
p = 0.003
p = 0.321
E2/T ratio
8
*
*
6
*
4
*
2
0
SC 0.0040.04 0.4 SC 0.004 0.04 0.4 SC 0.0010.01 0.1 SC 0.001 0.01 0.1 SC 0.0001 0.001 0.01 SC 0.0001 0.001 0.01mM
(A)
DEP
MEP
BBzP
MBzP
DiBP
MiBP
1.8
B = - 0.814
₿ = - 0.421
₿ = - 0.417
₿ = - 5.577
₿ = - 9.236
StAR fold change
1.6
B = - 13.84 p = 0.354
1.4
p = 0.007
p = 0.369
p = 0.605
p = 0.002
p = 0.472
1.2
1.0
0.8
*
*
0.6
*
*
$
*
*
*
0.4
0.2
0.0
(B)
2.5
3BHSD2 fold change
B = - 0.886
₿ = - 0.884
₿ = - 0.809
₿ = - 1.329
₿ = 33.959
B = 19.721
2.0
p = 0.094
p = 0.054
p = 0.315
p = 0.341
p = 0.067
p = 0.133
1.5
1.0
*
*
0.5
*
*
*
*
0.0
(C)
4
CYP19A fold change
₿ = 2.474
₿ = 1.653
₿ = 3.274
₿ = 5.845
B = 72.338
₿ = - 26.981
3
p = 0.005
p=0.001
p =0.402
p=0.001
p < 0.001
p=0.169
*
*
2
*
*
*
$
*
*
se
1
*
*
0
SC 0.004 0.04 0.4
SC 0.004 0.04 0.4
SC 0.001 0.01 0.1
SC 0.001 0.01 0.1
SC 0.001 0.01 0.1
SC 0.0001 0.001 0.01 mM
overall down-regulation of star gene, a key steroidogenic gene, along with general down-regulation of 3 hsd or 178 hsd caused significant decrease of T (Fig. 1B), a precursor of E2, resulting in subsequence decrease of E2 levels. Moreover, facilitated conver- sion of T to E2 by cyp19a could aggravate deficiency of T level in the male fish. Our data indicate that most test LMWPs can reduce T levels in the male zebrafish through down-regulation of star, and upregulation of cyp19 gene.
Down-regulation of vtg gene expression in liver (Fig. 3) is explained by decrease of E2 following DEP exposure (Fig. 1A). Expression of vtg gene is well-known response following exposure to either E2 or E2 mimicking compounds [25]. The vtg induction by exposure to LMWPs has not been frequently investigated in fish; one available report showed increased vtg in male common carp by DEP [1]. But this study reported only vtg protein, and did not measure sex hormones or expression of steroidogenic genes.
The hormonal disruption caused by exposure to LMWPs may lead to reproduction damages. Following exposure to mixture of contaminants including DEP, failure of pregnancy and delivery was observed in mice [38]. Also in Swiss CD-1 mice, delayed sexual mat- uration, sperm abnormality, or decreasing number of pups were observed following two-generation exposure to DEP [39]. Repro- duction related consequences of sex hormone disruption observed in the present study following exposure to LMWPs warrant further investigation.
4.2. Reduced T in H295R cells by alteration of steroidogenic pathway
Disruption of sex hormone levels in H295R cells (Fig. 4) cor- responds well with those observed in male zebrafish (Fig. 1). Significant decrease of T levels and increase of E2/T ratio by all test phthalates observed in the cells support antiandrogenic effects of LMWPs. Interestingly, hydrolytic metabolites of the test LMWPs showed generally similar patterns of change in the sex hormones except for MiBP (Fig. 4). Transcriptional changes of three steroido- genic genes were also very similar to those observed in the fish. Such pattern of changes was the same with the hyodrolytic metabo- lites of the test LMWPs (Fig. 5). Down-regulation of StAR and 3HSD2 genes can be linked to reduced cholesterol uptake and decreased androstendione synthesis, respectively. Despite reduced steroidogenesis, up-regulation of CYP19A gene by DEP, BBzP, and their metabolites (Fig. 5C) maintained the levels of E2 concentration in the adrenal cell (Fig. 4A).
Our observation of sex hormone changes in H295R cell is often different from those reported in Mankidy et al. [24]. Mankidy et al. [24] reported that DEP shifted sex hormone balance toward more E2, e.g., increased E2 but decreased T level in H295R cells. Since Mankidy et al. used the exposure doses that were quite different from those of ours, and did not measure transcriptional changes of key steroidogenic genes, direct comparison is not possible. Most of the current in vitro studies that investigated endocrine disrupt- ing effects of LMWPs are limited to their effects on sex hormone receptor transactivation. Harris et al. [40] conducted recombined yeast screening assay and reported that BBzP and DBP are estro- genic compounds in terms of estrogen receptor transactivation. In addition, in Chinese hamster ovary (CHO) cell, exposure to DEP inhibited the T induced androgen receptor transactivation [41]. Our data obtained in the present study clearly show that several LMWPs and their hydrolytic metabolites can disrupt sex hormone balances through altering steroidogenic pathway in the human adrenal cells.
One of the interesting findings in the present study is the dif- ferent endocrine disrupting effects of MiBP, which is generally in opposite direction of those observed from the other test phthalates including its own parent LMWP, DiBP (Figs. 4 and 5). In both the fish and the cells, DiBP exposure led to decrease of T and up-regulation
of CYP19A (Figs. 1 B, 2 F, 4 B, 5 C), but MiBP exposure resulted in increase of T and down-regulation of CYP19A in the cells (Figs. 4 B and 5 C). It is clear that MiBP causes sex endocrine disruption that is different from the other test phthalates and their metabolites. Unlike DEP and BBzP which are metabolized primarily to hydrolytic monoesters such as MEP and MBzP, MiBP can be further metabo- lized into 3OH-mono-methylpropyl phthalate (OH-MiBP) [12]. It is possible that this secondary metabolite may play more potent role in endocrine disruption than MiBP, and cause antiandrogenicity in the fish. Further studies are warranted as endocrine disruption effects of OH-MiBP and other possible metabolites of DiBP have not been investigated yet.
4.3. Implications of the study
Our results show that DEP, BBzP, and DiBP reduce sex hormone synthesis through alteration of steroidogenic pathway, and lead to antiandrogenicity in male fish. Primary metabolites of these LMWPs also play important role in alteration of steroidogenesis. Endocrine disruption by LMWPs has been well recognized, but mostly based on rodent models (Table 4). To our knowledge, this is the first study that demonstrated endocrine disruption of major hydrolytic metabolites of LMWPs and alteration of steroidogenic gene transcription. Consequences of long-term exposure to these LMWPs warrant further investigation.
Acknowledgment
This work was supported by a grant from National Research Foundation (NRF) of Korea (NRF-2014R1A2A1A11052838).
Appendix A. Supplementary data
Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jhazmat.2016.08. 008.
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