Environmental Toxicology and Chemistry, Vol. 34, No. 12, pp. 2833–2840, 2015 # 2015 SETAC Printed in the USA

ENDOCRINE-DISRUPTING EFFECT OF THE ULTRAVIOLET FILTER BENZOPHENONE-3 IN ZEBRAFISH, DANIO RERIO KARIN L. KINNBERG,*y GITTE I. PETERSEN,z METTE ALBREKTSEN,z MITA MINGHLANI,y SUAD MOHAMUD AWAD,y BENTE F. HOLBECH,y JOHN W. GREEN,x POUL BJERREGAARD,y and HENRIK HOLBECHy yDepartment of Biology, University of Southern Denmark, Odense M, Denmark zDHI-Water & Environment, Hørsholm, Denmark xDuPont Applied Statistics, Newark, Delaware, USA

(Submitted 9 March 2015; Returned for Revision 26 April 2015; Accepted 20 June 2015) Abstract: The chemical ultraviolet (UV) filter benzophenone-3 (BP-3) is suspected to be an endocrine disruptor based on results from in vitro and in vivo testing. However, studies including endpoints of endocrine adversity are lacking. The present study investigated the potential endocrine-disrupting effects of BP-3 in zebrafish (Danio rerio) in the Fish Sexual Development Test (Organisation for Economic Co-operation and Development TG 234) and a 12-d adult male zebrafish study. In TG 234, exposure from 0 d to 60 d posthatch caused a monotone dose-dependent skewing of the phenotypic sex ratio toward fewer males and more female zebrafish (no observed effect concentration [NOEC]: 191 mg/L, lowest observed effect concentration [LOEC]: 388 mg/L). Besides, gonad maturation was affected in both female fish (NOEC 191 mg/L, LOEC 388 mg/L) and male fish (NOEC 388 mg/L, LOEC 470 mg/L). Exposure to BP-3 did not affect the vitellogenin concentration in TG 234. After 12 d exposure of adult male zebrafish, a slight yet significant increase in the vitellogenin concentration was observed at 268 mg/L but not at 63 mg/L and 437 mg/L BP-3. Skewing of the sex ratio is a marker of an endocrine-mediated mechanism as well as a marker of adversity, and therefore the conclusion of the present study is that BP-3 is an endocrine-disrupting chemical in accordance with the World Health Organization’s definition. Environ Toxicol Chem 2015;34:2833–2840. # 2015 SETAC Keywords: Endocrine disruptors

UV filters

Vitellogenin

Sex ratio

Aquatic toxicology

exposure study [11] and at 620 mg/L in a 21-d exposure study [12]. In juvenile rainbow trout Oncorhynchus mykiss, vitellogenin was significantly induced after 14 d exposure to 749 mg/L BP-3 [12]. However, BP-3 did not show vitellogenin-inducing effects in adult male zebrafish after 14 d exposure to 312 mg/L [13], in juvenile transgenic zebrafish after 96 h exposure to 2282 mg/L [14], or in juvenile fathead minnow Pimephales promelas after 14 d exposure to 3900 mg/L [15]. In Japanese medaka, the daily average number of eggs produced per female was significantly reduced at 26 mg/L of BP3 after 28 d exposure [11] and in another study at 620 mg/L BP-3 after 7 d, but not after 21 d, when the numbers had returned to control values [12]. The overall percentage of fertilized eggs collected during the 21-d exposure that hatched was significantly lower in the 620 mg/L BP-3 concentration [12]. Whether BP-3 affects sexual development in fish has not yet been documented. A number of definitions for endocrine disruptors have been proposed. One of the most often-cited definitions is that of the World Health Organization, defining an endocrine disruptor as “an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub) populations” [16]. General agreement exists in the various definitions that 3 requirements must be fulfilled for a chemical to be defined as endocrine disrupting: 1) it must work via an endocrine-mediated mechanism; 2) it must have an adverse effect (at population level in wildlife); and 3) there must be a plausible link between the mechanism of action and the adverse effect [17]. Both vitellogenin induction and sex ratio changes are considered endocrine specific, whereas only the sex ratio change is also considered “adverse,” because vitellogenin induction on its own does not necessarily have an impact on population parameters. Based on the information available (also

INTRODUCTION

Indications are growing that several ultraviolet (UV) filters exert endocrine-disrupting effects. Benzophenone-3 (BP-3, oxybenzone, 2-hydroxy-4-methoxybenzone) is frequently used as a UV filter in sunscreens and other personal care products, clothing, laundry detergents, and so forth. Benzophenone-3 can reach the aquatic environment either directly through washoff from skin during recreational activities, or indirectly through wastewater. Benzophenone-3 has been detected at levels of up to 125 ng/L in Swiss bathing lakes [1]. In surface water samples collected from beaches at the Mediterranean coast, a maximum BP-3 concentration of 3.3 mg/ L was reported [2]. It also has been detected in treated wastewater at concentrations of up to 700 ng/L and untreated wastewater at concentrations of up to 10.4 mg/L [3]. Moreover, BP-3 has been detected at concentrations of up to 123 ng/g lipid in fish [4]. Benzophenone-3 has shown multiple hormonal activities in vitro. It has been shown to possess estrogenic effects in some in vitro test systems [5–7] but not in others [8,9]. Furthermore, some studies have shown antagonism of both the estrogen receptor and the androgen receptor in vitro [6–8]. For review, see Kim and Choi [10]. Some studies indicate an endocrine activity of BP-3 in fish. However, the effect of BP-3 on induction of vitellogenin synthesis in fish varies between studies. Waterborne BP-3 has been shown to significantly induce vitellogenin synthesis in adult male Japanese medaka Oryzias latipes at 90 mg/L in a 14-d * Address correspondence to [email protected] Published online 25 June 2015 in Wiley Online Library (wileyonlinelibrary.com). DOI: 10.1002/etc.3129 2833

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Environ Toxicol Chem 34, 2015

from epidemiological and mammalian investigations), BP-3 was classified as a suspected endocrine disruptor [18] and placed in group 2a according to the Danish report [19] for criteria for endocrine-disrupting chemicals. In this criteria suggestion, substances can be classified as an endocrine disrupter based on “adverse in vivo effects where an endocrine disrupting mode of action is highly plausible.” However, the evidence regarding the potential endocrine-disrupting effect of BP-3 in fish is ambiguous, and definitive in vivo experiments examining adverse effects are lacking. The European Union report on criteria for endocrine-disrupting chemicals [17] suggests that conclusive results from Organisation for Economic Development and Co-Operation (OECD) test guidelines 416 and 443 for mammals or 234 (Fish Sexual Development Test, TG 234, [20]) may define a chemical as an endocrine disruptor. The purpose of the present study was to examine the endocrine-disrupting effects of BP-3 in an OECD test guideline 234 experiment [20] and a 12-d adult male zebrafish study using the induction of vitellogenin and change in phenotypic sex in juvenile zebrafish as endocrine-specific biomarkers. A change in sex ratio would categorize BP-3 as an endocrine disruptor. The concentrations of BP-3 were chosen for the present study of the endocrine-disrupting potential of BP-3 and were above the concentrations observed in the aquatic environment. MATERIALS AND METHODS

Chemicals

Benzophenone-3 (CAS# 131-57-7, oxybenzone, 2-hydroxy4-methoxybenzone), 17b-estradiol (17b-E; CAS# 50-28-2), and isopropanol (CAS# 67-63-0) were purchased from SigmaAldrich (Sigma-Aldrich H36206, Sigma-Aldrich E8875, and Sigma-Aldrich 33539-2, respectively). Acetone (CAS# 67-641) was purchased from Rathburn Chemicals (Rathburn RG 2001 batch 12A17AB). Experimental animals

Zebrafish (Danio rerio) eggs were obtained from the Department of Cellular and Molecular Medicine, University of Copenhagen, Denmark. The founder zebrafish came from the Zebrafish International Resource Center, Oregon, USA. They were of the AB and Tupfel Long fin (TL) genetic background. The strains were crossed and raised at the zebrafish facility at the University of Copenhagen, Denmark. The AB/TL crossing gives a wildtype zebrafish with a varied genetic background. These are the most widely used zebrafish strains. Adult male zebrafish were obtained from Credo Fish, Denmark. Exposure

TG 234. The experiment was carried out according to OECD TG 234 [20]. The exposure systems were flow-through test systems with 10-L glass aquaria (containing 8 L water). Water exchange was 40 L per 24 h. Administration of water and test compound was controlled by peristaltic pumps (Ole Dich Instrumentmakers). Stock solutions of the test compounds were prepared in isopropanol before administration. Dilution water and stock solution were mixed to the desired concentration before entering the aquaria, and the maximum isopropanol concentration in the exposure water was 100 mL/L, with a 12:12-h light:dark photoperiod. Four replicate aquaria of control, solvent control, and BP-3 exposures were used. Newly fertilized eggs were randomly divided into groups of 30 and placed in 10-L glass aquaria. After hatching, the larvae were fed 2 to 3 times daily with TetraMin

K.L. Kinnberg et al.

Baby (Tetra GmbH). In addition, newly hatched Artemia sp. nauplii (Sanders) were supplied once or twice daily. The exposure (0 mg/L BP-3, 100 mg/L BP-3, 320 mg/L BP-3, 500 mg/L BP-3 [nominal]) took place from 1 d postfertilization (dpf) until 60 d posthatch (dph). Aquaria were aerated throughout the exposure period. During exposure, pH, oxygen saturation, and temperature were measured once per week. Water temperature was maintained at 27  2 8C. Average oxygen saturation ranged from 88% to 100%. Adult male zebrafish. The exposure systems were flowthrough test systems with 8-L glass aquaria (containing 6 L water). Water exchange was 18 L per 24 h. Administration of water and test compound was controlled by peristaltic pumps (Ole Dich Instrumentmakers). Stock solutions of the test compounds were prepared in acetone before administration. Dilution water and stock solution were mixed to the desired concentration before entering the aquaria, and the maximum solvent concentration in the exposure water was 100 mL/L, with a 12:12-h light:dark photoperiod. Three replicate aquaria of control, solvent control, 17b-E control (20 ng/L), and BP-3 (100 mg/L, 320 mg/L, 500 mg/L [nominal]) exposures were used. Adult male zebrafish were selected from a batch of mixed sex based on secondary sexual characteristics and randomly divided into the aquaria (with 27– 32 fish in each exposure group). The fish were fed every second day with TetraMin flake food (Tetra GmbH). Aquaria were aerated throughout the exposure period. Water temperature was maintained at 27  2 8C, and water pH was 7.6. Sampling

TG 234. At 60 dph, all fish were euthanized in an overdose of buffered MS-222. Length and wet weight of each fish were recorded. Head and tail were cut off behind the operculum and behind the anal fin (as described in OECD [20]), weighed together, and frozen immediately in liquid nitrogen for subsequent quantification of vitellogenin via enzyme-linked immunosorbent assay (ELISA). Remaining trunks were fixed in Bouin’s solution for subsequent histological analyses. By use of this sampling method, vitellogenin and gonadal sex were evaluated on each individual, and a potential change in the vitellogenin level thus could be related to the phenotypic sex of the fish. The actual number of fish analyzed is given in Table 1. Adult male zebrafish. After 12 d exposure, all fish were euthanized in an overdose of buffered MS-222 (100 mg/L). The wet weight of each fish was recorded after each fish was opened and the gonads visually examined to confirm male sex. Female fish exposed by mistake were removed. The head and tail were cut off behind the operculum and behind the anal fin of the male fish as described in OECD TG 234 [20], weighed together, and frozen immediately in liquid nitrogen for subsequent quantification of vitellogenin via ELISA. Remaining trunks were frozen at –80 8C for the possibility of doing further analysis. The actual number of fish analyzed is given in Table 1. Quantification of BP-3 and 17b-E

BP-3. Water samples were passed through 0.45-mm poylvinylidene difluoride filters (Frisenette) before actual exposure concentrations were determined by high-performance liquid chromatography–tandem mass spectrometry (HPLC-MS/ MS; A 1200 Series HPLC and a 6410 Triple Quad LC/MS, both Agilent Technologies). Ten-microliter water samples were injected in the HPLC-MS/MS, and conditions were as follows: column Zorbax Eclipse XDB C18 4.6  50 mm, 1.8 mm Rapid Resolution HT, column temperature 40 8C, (A): 0.1% formic

Endocrine-disrupting effects of benzophenone-3 in fish

Environ Toxicol Chem 34, 2015

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Table 1. Measured concentrations of benzophenone-3, weight, length, survival, and number of zebrafish for analysis, mean  standard error of the mean Treatment group

Measured BP-3 concentration (mg L–1)

Fish weight (mg)

Fish length (mm)

Fish survival (%)

Number of fish for analysis

Juvenile

Control Solvent control 100 320 500

ND (n ¼ 12) ND (n ¼ 12) 191  5 (n ¼ 40) 388  23 (n ¼ 40) 470  30 (n ¼ 40)

108  3 116  3 114  3 123  3b 111  3

19.3  0.2 19.7  0.2 19.7  0.2 20.4  0.2b 19.6  0.1

85  6 85  3 73  6 67  4b 78  1

102 102 88 80 94

Adult

Control Solvent control 17b-estradiol 100 320 500

ND (n ¼ 27) ND (n ¼ 27) ND (n ¼ 27) 63  0.2 (n ¼ 27) 268  23 (n ¼ 27) 437  45 (n ¼ 27)

408  13 425  18 411  16 423  14 392  11 421  18

— — — — — —

Experimenta

87  2 83  7 89  5 97  3 93  3 100  0

23 24 21 26 24 26

a

Number of replicates was 4 in the juvenile experiment and 3 in the adult experiment. Significantly different from control group (p < 0.05). ND ¼ not detected. b

acid in water and (B): 0.1% formic acid in acetonitrile isocratic with 70% B electrospray ionization in positive ion mode, drying gas flow 11.0 L/min, nebulizer pressure 35 psig, drying gas temperature 350 8C, and capillary voltage of 4000 V. Fragmentor voltage was 130 V, and collision energies were 20 V (quantifier). The mass-to-charge ratios (m/z) of precursor and quantifier ions were 229.2 and 151.1, respectively. The standards were prepared by dissolving BP-3 in isopropanol followed by dilution in water. The detection limit was approximately 31 mg/L BP-3. 17b-E. Water samples were passed through 0.45-mm poylvinylidene difluoride filters (Frisenette) before actual exposure concentrations were determined by HPLC-MS/MS (A 1200 Series HPLC and a 6410 Triple Quad LC/MS, both Agilent Technologies). Forty-microliter water samples were injected in the HPLC-MS/MS, and conditions were as follows: column Zorbax Eclipse XDB C18 4.6  50 mm, 1.8 mm Rapid Resolution HT, column temperature 25 8C, (A): 0.1% ammonia hydroxide in water, and (B): 0.1% ammonia hydroxide in MeOH with a gradient (0 min: 30% B; 3: 70; 6: 100; 6,5: 100; 6,6: 30; 7,5: 30) electrospray ionization in negative ion mode, drying gas flow 10.0 L/min, nebulizer pressure 50 psig, drying gas temperature 325 8C, and capillary voltage of 4000 V. Fragmentor voltage was 90 V, and collision energies were 30 V (quantifier). The mass-to-charge ratios (m/z) of precursor and quantifier ions were 271.1 and 144.9, respectively. The standards were prepared by dissolving 17b-E in isopropanol followed by dilution in water. The detection limit was approximately 7.8 ng/L 17b-E. Gonad histology

The sex of all sampled fish was determined by histological examination of gonads as described by Kinnberg et al. [21]. In addition, the maturation stages of the gonads were categorized according to the OECD Histopathology Guidance Document [22]. Stage 0 represents gonads with entirely immature phases (females: oogonia to perinucleolar oocytes—no cortical alveoli; males: spermatogonia to spermatids—no spermatozoa). In stage 1, cortical alveolar oocytes are present in the ovaries and spermatozoa are present in the testes. All evaluations of histology were performed blinded to information on treatment received. Vitellogenin analysis

Vitellogenin levels were determined in all sampled fish by species-specific ELISA developed for zebrafish [23].

Vitellogenin levels were determined in the homogenates of the tail and head sections as described by Holbech et al. [23,24] and Morthorst et al. [25]. The detection limit was 0.2 ng/mL vitellogenin. Intra- and interassay coefficients of variation have been determined to be 5.8% and 10.4% for the zebrafish ELISA [23]. Data handling and statistics

For all analysis, a significance level of a ¼ 0.05 was used. Statistics were performed on the replicate level to avoid pseudoreplicate comparison of intra replicate fish. For gonadal staging, the replicate structure was used, but individual subject severity scores were also captured. The actual numbers of fish included in the analysis are presented in Table 1. Control and solvent control were compared using a simple t test or an exact permutation Wilcoxon test and, if no statistically significant difference was found, the controls were combined and used as a basis for determining whether treatment effects occurred. If a statistically significant difference between the controls was found, then the solvent control was used as a basis for determining whether treatment effects occurred [26]. Sex ratio was analyzed as the replicate proportions of male, female, intersex, and undifferentiated fish. Proportions were transformed using an arc-sine square-root transform to normalize the response and stabilize the variance. For vitellogenin, a natural log transform was used. Both transforms are recommended in OECD TG 234. The preferred statistical test for sex ratio and vitellogenin was the step-down Jonckheere-Terpstra test, so long as the data were consistent with a monotone concentration–response [27]. Where the data were not consistent with a monotone concentration–response, Dunnett’s test was used if the data were consistent with normality and variance homogeneity; otherwise Dunn’s test was used. In all cases, these comparisons of treatment groups with controls were performed regardless of the findings of an analysis of variance F-test or Kruskal-Wallis test [27]. Finally, for sex ratio, a generalized linear mixed model with binary errors and a nested variance structure was applied to counts rather than proportions, followed by Dunnett’s test. Gonadal stage was analyzed using the Rao-Scott Cochran-Armitage by Slices test [28], which is recommended by OECD and the US Environmental Protection Agency in the pending medaka extended 1generation reproduction test and larval amphibian growth and development assay (LAGDA) tests. Alternative Mann-Whitney and Cochran-Mantel-Haenszel tests were also performed.

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Environ Toxicol Chem 34, 2015 RESULTS

K.L. Kinnberg et al. *

100%

* Intersex

90%

BP-3 and 17b-E water concentrations

The measured concentrations of BP-3 are presented in Table 1. TG 234. No statistically significant differences were seen between the BP-3 concentrations in the 4 replicates of any of the exposure concentrations. The 3 exposure concentrations were all significantly different from each other (p < 0.001). Mean measured concentrations of BP-3 ranged between 94% and 191% of the nominal concentrations. Adult male zebrafish. Mean measured concentrations of BP3 ranged between 63% and 87% of the nominal concentrations. The measured concentration (n ¼ 9) of 17b-E used as a positive estrogenic control in the adult experiment was 17  2 ng/L 17bE, which corresponded to 85% of the nominal concentration.

Undifferentiated

80%

Male

70%

Female

60% 50% 40% 30% 20% 10%

102

102

Control

Solvent control

0%

88

80

94

191 388 470 Benzophenone-3 (µg/L)

Figure 1. Percentages of phenotypic females, males, undifferentiated, and intersex zebrafish after exposure to benzophenone-3 in the water from 0 d posthatch (dph) to 60 dph. n is given in histogram inserts. *p < 0.01. Mean measured concentrations of benzophenone-3 are shown.

Fish survival and growth

TG 234. The total survival from fertilized eggs until 60 dph for all of the groups was 77.7% (466/600). Survival in the control and solvent control did not differ, and survival in the pooled control group was 85  3% (mean  standard error of the mean [SEM] for the 8 replicates). Survival was lower (p ¼ 0.04) in the group exposed to 388 mg/L BP-3 than in the pooled control group, but survival was not related to exposure concentrations, because survival in the 2 groups exposed to 191 mg/L and 470 mg/L BP-3 did not show statistically significant differences from survival in the pooled control group (Table 1). Length and weight of all of the surviving fish were 19.7  0.09 mm and 114  1 mg, respectively (mean  SEM for 466 fish). Length and weight did not differ between the control and the solvent control groups (Table 1). Both length (p ¼ 0.001) and weight (p ¼ 0.03) were higher in the group exposed to 388 mg/L BP-3 than in the pooled control group, but neither length nor weight was related to exposure concentrations, because the groups exposed to 199 mg/L and 470 mg/L BP-3 showed no statistically significant differences from length and weight in the pooled control group (Table 1); length in the group exposed to 388 mg/L BP-3 was higher (p ¼ 0.04) than length in the group exposed to 470 mg/L BP-3. Adult male zebrafish. The survival rate and weight at sampling time is presented in Table 1. No exposure-related mortality occurred, and the maximum mortality of 17% was observed in the solvent control group. No significant differences in the mean weight of the adult male zebrafish were observed after 12 d exposure to BP-3. Sex ratio, gonadal staging, and vitellogenin concentration

TG 234. The control and solvent control groups both contained 53.8% males, and no statistically significant differences were seen between the percentages of females (38.8%), undifferentiated fish (6.4%), and intersex fish (1%) in the 2 control groups. These were consequently pooled in the statistical analysis. The group exposed to 191  5 mg/L BP-3 contained 48.1% females and 46.9% males, but no statistically significant difference from the control group was found (Figure 1). The groups exposed to 388  23 mg/L BP-3 and 470  30 mg/L BP-3 both had a significantly increased number of females (p ¼ 0.006 and p ¼ 0.009) and a statistically significantly reduced number of males (p ¼ 0.004 and

p ¼ 0.00001; Figure 1). The percentage of undifferentiated fish was not significantly different between the groups (Figure 1). Seven intersex fish were detected among the 466 fish in the experiment; no clear relation was seen between exposure and occurrence of intersex fish. The percentages of males and females were negatively and positively correlated, respectively, with the concentration of BP-3 in the entire range of exposure concentrations (Figure 2). The actual number of females and males in each exposure group is shown in Figure 3. Benzophenone-3 exposure affected the gonad maturation in both female fish (no observed effect concentration [NOEC] 191 mg/L; lowest observed effect concentration [LOEC] 388 mg/L) and male fish (NOEC 388 mg/L, LOEC 470 mg/L). A significantly increased incidence of fish with ovaries at the least advanced maturation stage (gonadal stage 0) was found in the 2 highest test concentrations (p ¼ 0.02; Figure 3A). For males, the incidence of fish with testes at the least advanced maturation stage (gonadal stage 0) was significantly increased in the highest test concentration (p ¼ 0.0008; Figure 3B). Vitellogenin concentrations were not affected by the exposure to BP-3 in males, females, undifferentiated fish, or intersex fish (p > 0.25, Figure 4). Adult male zebrafish. Vitellogenin concentrations were significant higher (p ¼ 0.02) in the group exposed to 268 mg/L BP-3 and 17 ng/L 17b-E, but not at 63 mg/L and 437 mg/L BP-3 (Figure 5).

70% Females % = 0.048[BP-3] + 38.9 R² = 0.9926 p = 0.0003

60%

50%

40% Males % = -0.056[BP-3] + 54.8 R² = 0.9645 p = 0.0029

30%

20% 0

100

200 300 Benzophenone-3 (µg/L)

400

500

Figure 2. Percentages of phenotypic female and male zebrafish at 60 d posthatch (dph) as functions of the concentration of benzophenone-3 in the juvenile experiment.

Endocrine-disrupting effects of benzophenone-3 in fish

A

Ovarian maturation stage *

100%

Environ Toxicol Chem 34, 2015

* Stage 1

90%

Stage 0

80% 70% 60% 50% 40% 30% 20% 10% 0%

39

40

Control

Solvent control

B

42

47

57

191 388 470 Benzophenone-3 (µg/L)

Testicular maturation stage *

100%

Stage 1

90%

Stage 0

80% 70% 60% 50% 40% 30% 20% 10% 0%

55

54

Control

Solvent control

41

28

24

191 388 470 Benzophenone-3 (µg/L)

Figure 3. Percentages of gonads in the 2 observed maturity stages in female (A) and male (B) zebrafish at 60 dph in the juvenile experiment. Stage 0 represents gonads with entirely immature phases (females: oogonia to perinucleolar oocytes—no cortical alveoli; males: spermatogonia to spermatids—no spermatozoa). In stage 1, cortical alveolar oocytes are present in the ovaries and spermatozoa are present in the testes. n is given in histogram inserts. *p < 0.05. Mean measured concentrations of benzophenone-3 are shown.

DISCUSSION

The overall survival (77.7%) in the TG 234 study fully meets the recommendation in the test guideline (80% hatch and 70% posthatch survival—total 56% [20]), and the average weight of the fish (114 mg) also meets the recommended minimum value of 75 mg. The TG 234 study showed that benzophenone-3 exposure from 0 dph to 60 dph caused an adverse endocrine-disrupting effect in developing zebrafish by skewing phenotypic sex toward females. This effect was manifested as a monotone dosedependent skewing of the phenotypic sex ratio toward less male and more female zebrafish, with a NOEC of 191 mg/L and a LOEC 388 mg/L BP/3 (measured concentrations). Sexual differentiation in zebrafish is very sensitive to exposure to exogenous hormones, and altered sex ratios are inducible with estrogenic substances [24,29,30,31], androgens [25], and aromatase inhibitors [21,32]. The change in sex ratio toward more phenotypically female and less phenotypically male zebrafish observed in the present study indicates an overall estrogenic or an antiandrogenic activity of BP-3. The very weak nonmonotone increase in vitellogenin concentration in the adult male zebrafish exposed for 12 d to 268 mg/L but not to 437 mg/L could suggest a very weak estrogenic action of BP-3 affecting only the most sensitive fish

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in the actual exposure concentration area and leading to the relatively higher variation between fish seen at the higher exposure levels. A test design with more fish and replicates would be needed to confirm this. The effect could also be coincidental, but because of a statistical significance (p ¼ 0.02) below the defined level of 0.05 the result is discussed as a very weak inducing effect of BP-3 on vitellogenin in adult male zebrafish. Inducing effect of BP-3 on vitellogenin levels has been shown in some previous studies with Japanese medaka, fathead minnow, and juvenile rainbow trout [11,12], whereas other studies have found no effect of BP-3 on the vitellogenin levels in zebrafish and fathead minnow [13,15]. The present study’s result on weak vitellogenin induction in adult male zebrafish is not in contrast with the previous study showing no effect on the vitellogenin levels after 14 d exposure of adult male zebrafish to 312 mg/L BP-3 [13], because the increase in the vitellogenin concentration is so low that the ELISA methodology used in the study by Bl€uthgen et al. [13] would not detect it. A significant increase in plasma vitellogenin concentrations was observed in male Japanese medaka after 14 d exposure to 90 mg/ L BP-3 [11] and 21 d exposure to 620 mg/L BP-3 [12]. Vitellogenin induction was also observed in juvenile rainbow trout after exposure to 749 mg/L BP-3 [12]. However, BP-3 did not induce vitellogenin in juvenile fathead minnow after 14 d exposure to 3900 mg/L [15]. The lack of induction of vitellogenin synthesis in the present study’s TG 234 experiment (p > 0.25) could suggest that vitellogenin induction after exposure to BP-3 is slightly more sensitive in the adult stage than in the juvenile stage. This higher sensitivity of vitellogenin induction after exposure to estrogenic substances in adults compared with juveniles has previously been observed in fish [29,33–35] and has been ascribed to immaturities in the juveniles leading to differences in expression of the estrogen receptor [34] or differences in the biotransformation reaction rates of the substances [35]. In the present study, juveniles may differ from adults in their metabolism of BP-3. The study by Bl€uthgen et al. [13] showed that BP-3 is metabolized to benzophenone-1 (BP-1) in adult zebrafish, but not in eleuthero-embryos (up to 5 dpf), probably because BP-3 metabolizing enzymes are not yet fully active at this early life-stage. In the study with Japanese medaka, where no induction of vitellogenin by BP-3 was observed [15], a significant induction of vitellogenin was found after exposure to BP-1. This indicates that the metabolite BP-1 may have higher estrogenic potency in fish than BP-3 itself. The differences in vitellogenin induction between adults and juveniles in the present study thus may be attributable to a higher level of the more potent estrogenic metabolite BP-1 in adults. The marked shift in sex ratio in spite of no effects on vitellogenin concentrations in the present study’s juvenile experiment and only very weak effects in the adult experiment suggests that BP-3 may have shown antiandrogenic activity in addition to the weak estrogenic effects. This is also in agreement with the conclusions of a study dealing with the effects of BP-3 in adult male zebrafish and in eleuthero-embryos on gene expression focusing on the sex hormone system [13]. An overall down-regulation of the hsd3b, hsd17b3, hsd11b2, and cyp11b2 transcripts was observed in the testes, which indicates an antiandrogenic activity of BP-3. Besides the shift in sex ratio, the present study also demonstrated that BP-3 interferes with the maturation stages of the gonads. Histological evaluation showed that fewer female and male zebrafish had reached an advanced stage of gametogenesis by the end of exposure to the highest

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Environ Toxicol Chem 34, 2015

K.L. Kinnberg et al.

ng vitellogenin/mL homogenate

Test Guideline 234 female vitellogenin 1e+5

A

1e+4

1e+3

4 ol r t n Co

1e+2

c nt lve So

4 ol r t on

4

4

4

g/L 1µ 19

g/L 8µ 38

g/L 0µ 47

ng vitellogenin/mL homogenate

Test Guideline 234 male vitellogenin 1000

B

100

10

4

4

4

4

4

ol ntr Co

l tro on

g/L 1µ 19

g/L 8µ 38

g/L 0µ 47

c nt lve So

ng vitellogenin/mL homogenate

Test Guideline 234 undifferentiated and intersex vitellogenin 1000

C

100

10

4

4

4

3

5

ol ntr Co

l tro on

g/L 1µ 19

g/L 8µ 38

g/L 0µ 47

c nt lve So

Benzophenone-3 exposure concentrations Figure 4. Vitellogenin concentrations in zebrafish exposed to benzophenone-3 in the water from 0 d posthatch (dph) to 60 dph. (A) Phenotypic female fish. (B) Phenotypic male fish. (C) Phenotypic intersex and undifferentiated fish. Fiftieth percentile and 90th percentile box plot with median and outliers. *p < 0.05. The number of replicates in each group is shown above the x-axis. Mean measured concentrations of benzophenone-3 are shown.

concentrations of BP-3 in the present study’s TG 234 experiment (LOEC for females: 388 mg/L, LOEC for males: 470 mg/L). Inhibition of gametogenesis has previously been observed in both female and male fish after exposure to estrogenic [36] or antiandrogenic substances [37] during development. Furthermore, spermatocyte and oocyte development was significantly inhibited in reproductively mature fathead minnows exposed to 1.2 mg/L benzophenone-2 for 15 d [38]. No histological changes were observed in testes of adult male zebrafish exposed to 312 mg/L BP-3 for 14 d [13], which is in agreement with the present study finding a NOEC of 388 mg/ L for effects on the testes. The mode of action of BP-3 in fish could be further evaluated using the androgenized female stickleback screen [39] in which

inhibition of androgen-induced spiggin production serves as a biomarker of antiandrogenic activity. However, regardless of the precise mode of action of BP-3 in the present study, previous studies show that BP-3 can exert multiple endocrine activities. The present study is the first to show a significant effect on the sex ratio, which is considered to be an adverse populationrelevant effect. Consequently, both the criteria of adverse in vivo effects and a plausible endocrine mode of action are met. Therefore, the conclusion of the present study is that BP-3 is an endocrine disruptor according to the Danish criteria report [19] and the World Health Organization definition [16] and in accordance with the OECD guidance document on Standardized test guidelines for evaluating chemicals for endocrine disruption [40].

Endocrine-disrupting effects of benzophenone-3 in fish

Environ Toxicol Chem 34, 2015

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Figure 5. Vitellogenin concentrations in adult male zebrafish exposed to benzophenone-3 in the water for 12 d. Fiftieth percentile and 90th percentile box plot with median and outliers. *p < 0.05. The number of replicates in each group is shown above the x-axis. Mean measured concentrations of benzophenone-3 are shown.

Acknowledgment—The present study was supported by grants from the Danish Environmental Protection Agency via the Danish Centre on Endocrine Disrupters (#MST-621-0116) and the Danish Natural Science Research Council (#09-062337). Data availability—Data are available on request ([email protected]). REFERENCES 1. Poiger T, Buser HR, Balmer ME, Bergqvist PA, Muller MD. 2004. Occurrence of UV filter compounds from sunscreens in surface waters: Regional mass balance in two Swiss lakes. Chemosphere 55:951–963. 2. Tarazona I, Chisvert A, Leon Z, Salvador A. 2010. Determination of hydroxylated benzophenone UV filters in sea water samples by dispersive liquid-liquid microextraction followed by gas chromatography-mass spectrometry. J Chromatogr A 1217:4771–4778. 3. Loraine GA, Pettigrove ME. 2006. Seasonal variations in concentrations of pharmaceuticals and personal care products in drinking water and reclaimed wastewater in Southern California. Environ Sci Technol 40:687–695. 4. Balmer ME, Buser HR, Muller MD, Poiger T. 2005. Occurrence of some organic UV filters in wastewater, in surface waters, and in fish from Swiss lakes. Environ Sci Technol 39:953–962. 5. Schlumpf M, Cotton B, Conscience M, Haller V, Steinmann B, Lichtensteiger W. 2001. In vitro and in vivo estrogenicity of UV screens. Environ Health Perspect 109:239–244. 6. Schreurs R, Sonneveld E, Jansen JHJ, Seinen W, van der Burg B. 2005. Interaction of polycyclic musks and UV filters with the estrogen receptor (ER), androgen receptor (AR), and progesterone receptor (PR) in reporter gene bioassays. Toxicol Sci 83:264–272. 7. Suzuki T, Kitamura S, Khota R, Sugihara K, Fujimoto N, Ohta S. 2005. Estrogenic and antiandrogenic activities of 17 benzophenone derivatives used as UV stabilizers and sunscreens. Toxicol Appl Pharmacol 203:9–17. 8. Molina-Molina JM, Escande A, Pillon A, Gomez E, Pakdel F, Cavailles V, Olea N, Ait-Aissa S, Balaguer P. 2008. Profiling of benzophenone derivatives using fish and human estrogen receptor-specific in vitro bioassays. Toxicol Appl Pharmacol 232:384–395. 9. Nakagawa Y, Suzuki T. 2002. Metabolism of 2-hydroxy-4-methoxybenzophenone in isolated rat hepatocytes and xenoestrogenic effects of its metabolites on MCF-7 human breast cancer cells. Chem Biol Interact 139:115–128. 10. Kim S, Choi K. 2014. Occurrences, toxicities, and ecological risks of benzophenone-3, a common component of organic sunscreen products: A mini-review. Environ Int 70:143–157.

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Endocrine-disrupting effect of the ultraviolet filter benzophenone-3 in zebrafish, Danio rerio.

The chemical ultraviolet (UV) filter benzophenone-3 (BP-3) is suspected to be an endocrine disruptor based on results from in vitro and in vivo testin...
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