Endocrine disruption effects of p,p'-DDE on juvenile zebrafish.

Research Article Received: 11 February 2014,

Revised: 13 March 2014,

Accepted: 13 March 2014

Published online in Wiley Online Library

(wileyonlinelibrary.com) DOI 10.1002/jat.3014

Endocrine disruption effects of p,p′-DDE on juvenile zebrafish Marta Sofia Monteiroa*, Maria Pavlakia, Augusto Faustinob, Alexandra Rêmab, Mariana Franchia, Letícia Gedielc, Susana Loureiroa, Inês Dominguesa, Jaime Rendón von Ostena,d and Amadeu Mortágua Velho Maia Soaresa,e ABSTRACT: The persistent organic pollutant p,p′-DDE, the major metabolite of the insecticide DDT, has displayed evidence of endocrine disruption through the inhibition of androgen binding to androgen receptors in different species. Although p,p′-DDE was continuously detected in wild fish with abnormal gonad development such as intersex, little is known about its mode of action during gonad development in fish. To elucidate the potential endocrine effects of this pollutant in zebrafish (Danio rerio), juveniles (30 days post hatch) were exposed to p,p′-DDE during the critical window of sexual differentiation. Fish were exposed to sublethal concentrations ranging from 0.01 to 20 μg l–1 over 14 days and were maintained in control water for an additional 4 months. As core endpoints, the vitellogenin (vtg) concentration was measured at the end of exposure, and sex ratio and the gonadosomatic index were assessed 4 months after the end of exposure. An increase in vtg production in whole body hom*ogenate was observed in fish exposed to 0.2 and 2.0 μg l–1 p,p′-DDE. No signiﬁcant differences were displayed in morphological parameters such as the gonadosomatic index of males and females or sex ratio. However, exposed females presented histopathological changes that include the reduction of the number of mature oocytes, which might impair their successful reproduction. These results demonstrate the ability of p,p′-DDE to cause endocrine disruption in zebrafish exposed during gonad differentiation of juvenile specimens. Furthermore, vtg induction by p,p′-DDE in juvenile zebrafish arises as a predictive marker for adverse effects of this DDT metabolite on the ovarian function of female zebraﬁsh. Copyright © 2014 John Wiley & Sons, Ltd. Keywords: antiandrogen; ecotoxicology; Danio rerio; ﬁsh; histopathology; persistent organic pollutant; pesticide; vitellogenin

Introduction

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* Correspondence to: Marta S. Monteiro, Departamento de Biologia, Universidade de Aveiro, Campus Universitário de Santiago, 3810-193 Aveiro, Portugal. E-mail: [emailprotected] a Department of Biology & CESAM, University of Aveiro, 3810-193 Aveiro, Portugal b Departament of Pathology and Molecular Immunology, ICBAS, University of Porto, Porto, Portugal c Department of Genetics and Morphology, Institute of Biological Sciences, University of Brasília, Brasília, Brazil d

Instituto EPOMEX, Universidad Autónoma de Campeche, 24030 Campeche, Mexico

Programa de Pós-Graduação em Produção Vegetal, Universidade Federal do Tocantins, Campus de Gurupi. Rua Badejós, Zona Rural, Cx. Postal 66, CEP: 77402-970, Gurupi-TO, Brasil

Industrial chemicals and environmental pollutants can disrupt reproductive development in wildlife and humans by mimicking or inhibiting the action of the gonadal steroid hormones even when present at very low concentrations (Kelce and Wilson, 1997). The toxicity of endocrine disruptor compounds (EDCs) is particularly insidious during sex differentiation and development due to the crucial role of gonadal steroid hormones in the regulation of these processes (Hutchinson et al., 2006; Kelce and Wilson, 1997). Within EDCs, antiandrogens are able to bind to the androgen receptor (AR) and prevent the transcription of the associated genes causing abnormal sexual development and demasculinization (Bayley et al., 2002). However, in contrast to estrogenic modes of action, relatively little is known about how environmentally relevant concentrations of (anti)androgenic EDCs affect male reproductive health (Luccio-Camelo and Prins, 2011). The major stable metabolite of the insecticide DDT, p,p′-DDE, is a persistent and highly lipophilic compound, that bioaccumulates in adipose tissue and tends to biomagnify along food chains (Kelce and Wilson, 1997; Oliver and Niimi, 1985). The p,p′-DDE is considered a potent environmental antiandrogen, it has the potential to alter male sex development and reproductive processes in wildlife and human populations (Kelce et al., 1995; Kelce and Wilson, 1997). Although DDT has long been banned in most developed countries, their metabolites are still present in the environment due to their long half-lives (Thomas et al.,

2008; WHO, 2012). Although p,p′-DDE is often not detectable in water (Tyler et al., 1998), water concentrations of 0.015 μg l–1 (Albanis et al., 1998) and 0.13 μg l–1 p,p′-DDE (Fernández et al., 2000) were measured in surface water samples in Imathia (Greece) and Madrid (Spain), respectively. However, DDT is still in use in some developing countries, where water concentrations may reach 1–10 μg l–1 DDT (Tyler et al., 1998). Indeed, high concentrations of p,p′-DDE have been continuously detected in wild ﬁsh tissues that showed abnormal gonad development such as intersex, namely in chub from the River

M. S. Monteiro et al. Elbe, Czech Republic (Randak et al. 2009), in ﬂatﬁsh from Dogger Bank, North Sea (Stentiford and Feist, 2005), in largemouth bass from the Rio Grande and its United States tributaries (Schmitt et al., 2005), in ﬂounder from the estuary of the River Mersey, England (Leah et al., 1997; Scott et al., 2006), in Yangtze River Chinese sturgeon (Wan et al. 2006; Wei et al., 1997) and in the Mississippi River shovelnose sturgeon (Harshbarger et al. 2000). Several laboratory-based studies have also provided evidence of endocrine disruption induction in male ﬁsh exposed to p,p′DDE (Baatrup and Junge, 2001; Bayley et al., 2002; Mills et al., 2001; Zaroogian et al., 2001). More recently, Zhang and Hu (2008) demonstrated intersex inducement by 100 μg l–1 p,p′DDE in Japanese medaka (Oryzias latipes). In toxicological processes, alterations of subcellular biomarkers occur before the response of higher levels of organization such as conventional histological responses (Hutchinson et al., 2006). For instance, the induction of the protein vitellogenin (vtg) in males have been widely used as a biomarker to evaluate estrogenic EDCs (Hutchinson et al., 2006). Zhang and Hu (2008) have found upregulation of the vtg-1 and vtg-2, choriogenins and estrogen receptor α genes in the liver of Japanese medaka exposed to p,p′DDE. Similarly, Larkin et al. (2002) reported that vtg and choriogenins were upregulated by p,p′-DDE in largemouth bass, whereas in other studies vtg protein levels were not induced by p,p′-DDE exposure in summer ﬂounder (Mills et al., 2001). Therefore, the mode of action (MOA) of p,p′DDE due to its antiandrogenic activity is still controversial and needs to be further studied to be clariﬁed, particularly the link between the disruption of the normal hormonal signals in early life stages and the modiﬁcations in the organization and future functioning of the reproductive system. The development of zebraﬁsh (Danio rerio) from the fertilized egg to full reproductive maturity takes only 3–4 months (Maack and Segner, 2003). This relatively short generation time makes this widely used model animal a particularly useful tool for partial and full life cycle tests to evaluate the effects of EDCs on speciﬁc molecular biomarkers, gonad ontogenetic differentiation and reproduction of ﬁsh contributing to discern their MOA (Maack and Segner, 2003; Segner, 2009; Stegeman et al., 2010). In this study we hypothesized that exposure of zebraﬁsh (Danio rerio) to p,p′-DDE during gonadal differentiation will alter vtg levels and impair gonadal differentiation in adults. To achieve this, zebraﬁsh at 30 days post hatch (dph) were exposed to sublethal concentrations of p,p′-DDE for 14 days, vtg levels were evaluated just after the exposure while the gonadosomatic index (GSI), hepatosomatic index (his), sex ratio and histopathological analysis were analyzed in 150 dph zebraﬁsh after further maintenance in clean medium.

Materials and Methods Chemicals

p,p′-DDE was purchased from Chem Service (West Chester, PA, USA); 17 β-estradiol and dimethyl sulfoxide (DMSO) were supplied from Merck (Darmstadt, Germany); all other chemicals used were of analytical grade quality (Sigma-Aldrich, Germany).

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Test Fish and Culture Conditions The zebraﬁsh used in this study were obtained from the facility established at the Biology Department, Aveiro University (Portugal). Fish were maintained in ﬂow-through aquaria in carbon-ﬁltered water complemented with salt “Instant Ocean Synthetic Sea Salt” (conductivity 550 ± 50 μS, pH 7.5 ± 0.3 and dissolved oxygen 95% saturation), and were fed twice daily with brine shrimp (Artemia nauplii) and artiﬁcial diet. Fish were kept at 26 ± 1.0 ºC under a photoperiod of 16 h light/8 h dark. Similar temperature and photoperiod conditions were maintained during all the assays performed.

Preparation of Test Solutions The water from zebraﬁsh culture was used as the control and as dilution water in the preparation of test solutions in all assays performed. Stock solutions were prepared by dissolving p,p′DDE in DMSO and test solutions were obtained by consecutive dilutions of the stock solution. All solutions were prepared using glass recipients. The concentrations of p,p′-DDE mentioned in the text refer to the nominal concentrations tested.

Acute Test A ﬁsh acute toxicity test of 96 h was performed in semistatic conditions following the OECD Guideline 203 (OECD, 1992). Three groups of 12–15 juvenile zebraﬁsh (30 dph) were allocated to each treatment, 0.01, 0.022, 0.045, 0.1, 0.22, 0.45 and 1.0 mg l–1 p,p′-DDE, the negative control and the control for solvent (50 μl l–1 DMSO). Each group was exposed in 1 liter glass tanks and test solutions were completely renewed after 48 h. Mortality was recorded daily and the LC50 and LC10 were determined.

Sublethal Test Juvenile zebraﬁsh (30 dph) were exposed for 14 days to sublethal concentrations of p,p′-DDE (0.01, 0.02, 0.2, 2.0 and 20 μg l–1) following the OECD guideline 204 (OECD, 1984). Four replicates of 10 ﬁsh each were allocated to each concentration of p,p′DDE, the negative control, solvent control (50 μl l–1 DMSO) and a positive control (100 ng l–1 β-estradiol), using 1 liter glass tanks. Test solutions were completely renewed every 2 days. After the exposure, ﬁsh were maintained until 150 dph in clean water. Survival was checked daily and ﬁsh were fed twice a day.

Sampling of Fish After the end of the 14 day exposure (44 dph) three ﬁsh per each of the four replicates per treatment (n = 4; total of 12 ﬁsh) were anesthetized with MS-222 and killed by decapitation, frozen in liquid nitrogen and kept at – 80 ºC until vtg analysis. Survival and weight were recorded. All the remaining ﬁsh were maintained in clean water until 150 dph. At that time, all ﬁsh were killed as described above. Length and weight of ﬁsh were recorded and, after dissection, ﬁsh were inspected for sex ratio determination using a stereomicroscope. Liver and gonads were removed and weighed for GSI and HSI determination. Gonads were preserved for histological analysis.

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Endocrine disruption effects of p,p′-DDE on juvenile zebraﬁsh Gonadosomatic and Hepatosomatic Indices GSI and HSI of female and male zebraﬁsh were calculated as follows, using the weight of gonads and liver, respectively: (organ weights (g)/body weights (g)) × 100. Vitellogenin Analysis Three ﬁsh per replicate (n = 3) were weighed and hom*ogenized in buffer solution 50 mM Tris-HCl, 0.02% aprotinin, 0.1 M PMSF. The hom*ogenate was centrifuged (14 000 g, 4 ºC, 1 h). The supernatant recovered was diluted and vtg concentrations were measured using a precoated vtg ELISA kit following the manufacturer’s instructions (Biosense Laboratories, Bergen, Norway).

were found in weight of exposed juvenile ﬁsh (44 dph) just after the exposure to p,p′-DDE (Fig. 1A), nor later, in the weight or length of mature ﬁsh (150 dph) maintained in clean medium after the exposure (P > 0.05) (Fig. 1B,C). Vitellogenin. The vtg levels measured in juvenile ﬁsh exposed to the different treatments are presented in Fig. 2. Vtg was signiﬁcantly induced in juvenile zebraﬁsh exposed to the positive control treatment (100 ng l–1 β-estradiol, P < 0.05) and to 0.2 and

Gonad Histological Analysis Fish gonads were ﬁxed in Davidson’s solution, transferred to 70% ethanol after 24 h and processed according to standard histological methods (Kiernan, 1990; Wolf et al., 2004). The ﬁxed tissue was embedded in parafﬁn wax; sections were cut at 3–4 μm transversally along the gonads, stained with hematoxylin and eosin, and observed on an optical microscope (Zeiss, Germany) and digitized with a AxioCam digital camera (Oberchoken, Zeiss). Gonadal sections were examined for the presence of pathological changes following the OECD guidelines (OECD, 2009). Zebraﬁsh oocytes were classiﬁed into ﬁve stages according to their maturation scale (OECD, 2009). Oocytes at stage V (mature oocytes) were counted (number per gonadal section) and analyzed for the accomplishment of the comparative analysis of the diameter of the oocytes (Arocha, 2002). Concerning male gonads, the average perimeter containing the spermatozoa was analyzed in the different treatments. All morphometric analyses were performed with the software Image-pro Plus 5.1 (Media Cybernetics, Silver Spring, MD, USA). Statistical Methods LC50 and LC10 for 96 h (and 95% CI) were determined by probit analysis (Minitab 14.0). The remaining statistical analyses were performed in Sigma Stat 3.5. The chi-squared test was used to test differences in sex ratios between control and each treatment group. The remaining data were analyzed using one-way ANOVA followed by Dunnett’s test to evaluate signiﬁcant differences (P < 0.05) between treatments and control.

Results Acute Test The LC10 and LC50 for 96 h exposure of juvenile zebraﬁsh (30 dph) to p,p′-DDE were determined as 20 μg l–1 (3–31, 95% CI) and 59 μg l–1 (49–72, 95% CI), respectively. Sublethal Test

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Figure 1. Zebraﬁsh weight (44 dph) after the exposure to p,p′-DDE during 14 days (n(number of replicates) = 4; total of 12 ﬁsh: 3 ﬁsh × 4 replicates in each treatment) (A) and weight (B) and length (C) of zebraﬁsh (150 dph) maintained in clean medium after the 14-day exposure (n = 4; total of 9–13 ﬁsh per treatment). C - negative control; CS - control of solvent; C+ - positive control.

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Survival and growth. Survival of zebraﬁsh juveniles was 92.5% in the control and control for solvent, validating the test that requires less than 10% of mortality in these groups. In the remaining treatments, survival varied between 82.5 and 92.5%, and did not present signiﬁcant differences relative to the survival registered in the control groups. No signiﬁcant differences

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Figure 2. Vtg levels in juvenile ﬁsh (44 days post hatch) exposed to p,p′-DDE during 14 days. Number of replicates = 4 (3 ﬁsh × 4 replicates in each treatment = 12 ﬁsh per treatment). *Signiﬁcant differences from control (P < 0.05). vtg, vitellogenin. C - negative control; CS - control of solvent; C+ - positive control.

2.0 μg l–1 p,p′-DDE (P < 0.05). Therefore, a no observed effect concentration value of 0.02 μg l–1 and a lowest observed effect concentration value of 0.2 μg l–1 were derived from the statistical analysis. Gonadosomatic index, hepatosomatic index and sex ratio. GSI and HSI were determined in 150 dph ﬁsh previously exposed to p,p′-DDE. The GSI and HSI of the different treatment groups are displayed in Table 1. No signiﬁcant differences were observed in HSI or in GSI of exposed ﬁsh in comparison to the negative control (P > 0.05). Sex ratio proportion in 150 dph ﬁsh were not signiﬁcantly affected by the p,p′-DDE exposure (P > 0.05) as depicted in Fig. 3. Histopathology. The macroscopic examination during dissection of 150 dph ﬁsh revealed normal female and male gonads from both control and exposed ﬁsh; no intersex gonads were identiﬁed macroscopically. The histological examination of male gonads revealed the presence of all development phases of spermatogenic cells (Fig. 4a), with no histopathological changes present in any treatment. In particular, determination of the perimeter containing the spermatozoa resulted in no signiﬁcant differences (P > 0.05) between control and exposed ﬁsh testes (Table 2). As seen in Fig. 4(b) female gonads of control ﬁsh presented oocytes in all phases of oogenesis, with the predominance of later stages, namely vitellogenic (stage IV) and mature oocytes (stage V). However, female gonads of p,p′-DDE exposed ﬁsh were found to exhibit gonadal regression when compared to those of control ﬁsh (Fig. 4b–d). Histologically, this regression was characterized by the presence of atretic oocytes and by the predominance of

pre-vitellogenic stages, namely perinucleolar oocytes (stage II) and cortical alveolus oocytes (stage III) as observed in Fig. 4(c,d). Indeed, the number of mature oocytes (stage V) per gonadal section registered a signiﬁcant reduction in ﬁsh exposed to the concentrations of 0.01, 0.2 and 20 μg l–1 p,p′-DDE (p < 0.05; Table 2).

Discussion The main goal of this study was to determine the effects of p,p′-DDE exposure during gonadal differentiation of zebraﬁsh, focusing on the impairment of gonadal development and differentiation in adults.

Figure 3. Sex ratio in zebraﬁsh (150 days post hatch) after the exposure to p,p′-DDE during 14 days and further maintenance in clean medium (number of replicates = 4; total of 9–13 ﬁsh per treatment). C - negative control; CS - control of solvent; C+ - positive control.

Table 1. Gonadosomatic (GSI) and hepatosomatic index (HSI) of ﬁsh (150 days post hatch) in control (C), solvent control (CS, 50 μg l–1), positive control (PC, 100 ng l–1 β-estradiol) and p,p′-DDE (0.01, 0.02, 0.2, 2.0, 20 μg l–1) exposure groups (mean ± SE)

GSI HSI

p,p′-DDE (μg l–1)

0.01

0.02

0.2

2.0

Female Male Female Male

7.6 ± 2.00 0.9 ± 0.18 1.2 ± 0.35 0.8 ± 0.35

7.7 ± 1.77 1.1 ± 0.41 0.5 ± 0.06 0.3 ± 0.13

7.3 ± 0.77 0.7 ± 0.09 1.4 ± 0.53 0.7 ± 0.19

7.1 ± 0.53 1.5 ± 0.72 1.1 ± 0.08 0.5 ± 0.07

7.9 ± 1.03 2.3 ± 0.76 0.7 ± 0.21 0.7 ± 0.09

5.4 ± 1.06 1.4 ± 0.43 1.1 ± 0.25 0.4 ± 0.10

7.2 ± 1.81 0.9 ± 0.20 1.1 ± 0.19 0.6 ± 0.20

6.1 ± 0.70 1.4 ± 0.56 0.8 ± 0.14 0.5 ± 0.07

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Endocrine disruption effects of p,p′-DDE on juvenile zebraﬁsh

Figure 4. Representative gonad sections of 150 days post hatch zebraﬁsh stained with hematoxylin and eosin. (a) Male section of control group showing all stages of germ cell development: oc, oocytes; psc, primary spermatocyte; sd, spermatid; sg, spermatogonia; ssc, secondary spermatocyte; sz, –1 spermatozoa. (b) Female section of control group showing four stages of oocyte development. (c,d) Female section of 20 μg l p,p′-DDE-treated group: a, atretic oocytes; CA, Cortical alveolus oocytes; M, Mature oocytes; PN, Perinucleolar oocytes; V, vitellogenic oocytes.

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biomarker for detecting the presence of endocrine disruptors in the aquatic compartment (Hutchinson et al., 2006; Segner et al., 2003). The results from this study clearly demonstrated an induction of vtg by p,p′-DDE in a dose–response manner in juveniles. Similarly to our results, other authors have reported vtg upregulation by p,p′-DDE, together with other estrogenic-responsive genes, in male largemouth bass injected with 100 mg kg–1 p,p′-DDE (Larkin et al., 2002) and male juvenile (20 dph) Japanese medaka exposed for 2 months to up to 100 μg l–1 p,p′-DDE (Zhang and Hu, 2008). The vtg results here presented are consistent with the known estrogenic potential of p,p′-DDE. One MOA of p,p′-DDE is through the induction of aromatase, the enzyme essential for the conversion of testosterone to 17β-estradiol. This has been observed in adult male rats (You et al., 2001) and in ﬁsh (Garcia-Reyero et al., 2006). Accordingly, O’Connor et al. (1999) have registered an increase of 17β-estradiol circulating levels due to p,p′-DDE (O’Connor et al., 1999). The p,p′-DDE can also act as a weak estrogen, increasing the expression of vtg and estrogen receptors in the liver and can alter the expression of genes involved in the synthesis of endogenous hormones as well as their metabolism (Garcia-Reyero et al., 2006). On the other hand, Mills et al. (2001) reported no induction of vtg genes in juvenile male summer ﬂounder injected with up to 120 mg kg–1 p,p′-DDE. The alterations in vtg levels due to the estrogenic effect of p,p′-DDE are still not very clear. Therefore, further studies are needed to clarify these underlying mechanisms. Regarding the endocrine effects above the molecular level, it therefore required the evaluation of parameters directly involved in reproduction. We have therefore assessed p,p′-DDE by measuring sex ratio, GSI and histological analysis of the

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First, to select the concentration range of p,p′-DDE to perform the sublethal test, juvenile zebraﬁsh were exposed to lethal concentrations of this compound for 96 h. The LC50 for 96 h exposure to p,p′-DDE obtained for zebraﬁsh (59 μg l–1) is within the range obtained for other ﬁsh species, as reported in the database generated by the EU-funded FOOTPRINT project for Oncorhynchus mykiss with a LC50 value for 96 h of 32 μg l–1 (PPDB, 2009). Concentrations below 20 μg l–1, the LC10 determined in this study, were then used in the sublethal test performed with zebraﬁsh. The transient exposure to p,p′-DDE during the critical window of gonad differentiation in zebraﬁsh led to vtg induction measurements just after exposure. This alteration observed at 44 dph zebraﬁsh was then followed by histopathological alterations observed later in female gonads. Using similar experimental designs, Segner et al. (2003) have reported that the induced effects of selected EDCs during sexual differentiation (e.g. 42–75 dpf) were identical to those observed in the full life cycle exposures, e.g. elevated vtg levels and altered gonadal differentiation. These were considered promising results concerning the substitution of full life cycle tests in zebraﬁsh by partial life cycle tests during the period of ﬁnal gonad differentiation in EDC testing (Segner et al., 2003). The impact of p,p′-DDE treatment was ﬁrst investigated on circulating vtg levels in 44 dph juvenile zebraﬁsh. Vtg is the egg yolk precursor protein produced in the liver in response to estrogenic activity and is transported to the ovaries, where it is incorporated into maturing oocytes (Hutchinson et al., 2006). Very low levels of vtg are present in sexually immature and male animals. However, vtg can be induced in juvenile or male ﬁsh by estrogen treatment, which makes it a suitable and speciﬁc

0.4 ± 0.30a 4.4 ± 1.21 3.4 ± 0.87 7.0 ± 1.00 8.0 ± 2.68

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Values in bold and a refer to signiﬁcant differences from control (P < 0.05). b n = 1. c Data not available.

523.5 ± 3.13 623.6 ± 70.65

8.3 ± 2.39

1.8 ± 1.03a

2.0 ± 1.41a

766.9 ± 1.10 501.3 ± 17.14 489.7 ± 90.51 420.5 ± 66.63 522.0 ± 46.93

149.9 c 252.0 213.6 ± 28.46 176.1 ± 35.24

Perimeter containing spermatozoa (μM) Diameter of oocytes stage V (μM) Oocytes stage V per gonad section

164.1 ± 93.38

490.9 ± 39.12

195.3 ± 15.70 158.2 ± 17.19

20 2.0

0.2 0.02

0.01 PC SC C p,p′-DDE (μg l–1)

Table 2. Histological quantiﬁcation of different parameters in male and female zebraﬁsh gonads (150 days post hatch) in control (C), solvent control (SC), positive control (PC, 100 ng l–1 β-estradiol) and p,p′-DDE (0.01, 0.02, 0.2, 2.0, 20 μg l–1) exposure groups (mean ± SE)

M. S. Monteiro et al. gonads in 150 dph zebraﬁsh. The induction of vtg observed at 44 dph has possibly triggered the functional alterations observed later in mature ﬁsh, namely the regression of gonadal development in females. However, no major alterations were observed in the other parameters analyzed at 150 dph. In ﬁsh populations exposed to environmental estrogens or antiandrogens, the sex ratio is typically biased in favor of females and can be also affected in juvenile ﬁsh exposed during the period of sex determination and differentiation (Milnes et al., 2006). Female-biased sex ratios were indeed observed in juvenile guppies (Poecilia reticulata) exposed during development to p,p′-DDE (Bayley et al., 2002); however, no signiﬁcant differences were observed in the sex ratio in the present study. Gross indexes such as GSI and HSI are commonly used as indicative of toxicant effects (van der Oost et al., 2003). More speciﬁcally, the GSI is commonly used as an endpoint in reproduction studies with ﬁsh. The GSI values observed for control ﬁsh in this study are comparable with those reported for zebraﬁsh in other studies (Arcand-Hoy and Benson, 1998; Örn et al., 1998; Van den Belt et al., 2002). In general, male GSI is reduced by estrogenic and/or antiandrogenic chemicals (Milnes et al. 2006). Indeed, lower GSI has been induced in adult male guppies (P. reticulata) exposed for 30 days to 1 μg mg–1 p,p ′-DDE administered in food (Baatrup and Junge, 2001). However, in the present work no signiﬁcant differences were displayed in GSI of exposed zebraﬁsh. Other studies have also failed to get a discernible effect from this chemical on the GSI of juvenile guppies (P. reticulata) fed sublethal doses of p,p′-DDE, 0.1 and 0.01 μg mg–1 (Bayley et al., 2002), nor in juvenile male summer ﬂounder (Paralichthys dentatus) injected up to 120 mg kg–1 p,p′DDE (Bayley et al., 2002; Mills et al., 2001) even when having signiﬁcant reductions in other endpoints, such as display coloration, gonopodium development, reduced sperm count and courtship behavior (Bayley et al. 2002). It should be highlighted that a considerable variation in GSI was observed in all these studies, including in the control zebraﬁsh (ArcandHoy and Benson, 1998; Örn et al., 1998; Van den Belt et al., 2002). In general, there is a causal relationship between exposure to chemical pollutants and liver enlargement (van der Oost et al., 2003). Responses to p,p′-DDE included increased relative liver weight (O’Connor et al., 1999) and pronounced hepatocellular hypertrophy (You et al., 1999) in rats. However, in the present study no effect was detected on HSI. The overproduction of the protein vtg due to estrogenic exposure has been related to the induction of both hypertrophy and hyperplasia of hepatocytes, which might result in hepatomegaly (e.g. Herman and Kincaid, 1988); a similar response of HSI could be expected to occur in the present work. Although, the transient induction of the protein vtg registered in the 44 dph exposed zebraﬁsh cannot be directly related to liver alterations in 150 dph zebraﬁsh as vtg is known to be secreted by the liver shortly after synthesis and transported through the plasma to the gonads. Considering gonadal histopathology Zhang and Hu (2008) demonstrated that p,p′-DDE can also induce intersex in medaka and postulated that this result should be one of the clues in explaining the intersex observed in wild ﬁsh. Indeed, these authors demonstrated the ability of p,p′-DDE to induce intersex in laboratory experiments with juvenile male ﬁsh, but only at the concentration of 100 μg l–1 p,p′-DDE, one order of magnitude above the usual water concentration of DDT detected in the environment, 1–10 μg l–1 DDT (Tyler et al., 1998). A more

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Endocrine disruption effects of p,p′-DDE on juvenile zebraﬁsh

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In conclusion, the results of exposure to p,p′-DDE during the critical window of gonadal differentiation in zebraﬁsh demonstrated that vtg induction in juveniles may be predictive for adverse effects of this DDT metabolite on ovarian differentiation and maturation in female zebraﬁsh. Furthermore, these results suggest that exposure during gonadal differentiation to p,p′-DDE might impair reproductive function in zebraﬁsh. Acknowledgments This study was supported by FEDER through COMPETE e Programa Operacional Factores de Competitividade and by National funding through Fundação para a Ciência e Tecnologia (FCT), within the research projects FUTRICA – Chemical Flow in an Aquatic Trophic Chain (FCOMP-01-0124-FEDER-00008600; Ref. FCT PTDC/AAC-AMB/104666/2008) and DOMINO EFFECT – Degradation of lotic ecosystems associated with plantation forestry: an evaluation of plantation forest food-web communities (FCOMP01-0124-FEDER-008727; Ref. FCT PTDC/AGR-AAM/104379/2008). The FCT supported the postdoctoral fellowships of MS Monteiro (FCT/SFRH/BPD/45911/2008) and I Domingues (SFRH/BPD/ 90521/2012). AMVM Soares is “Bolsista CAPES/BRASIL,” Project NºA058/2013.

Conﬂict of Interest All of the authors declare that they have no actual or potential competing ﬁnancial interests.

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environmental realistic concentration of p,p′-DDE, 20 μg l–1, did not induce intersex gonads neither in the present study with zebraﬁsh neither in juvenile male Japanese medaka (Zhang and Hu, 2008). Therefore we consider that the intersex caused by concentrations of p,p′-DDE above those that could be found in aquatic compartments do not completely explain the intersex observed in wild ﬁsh. Considering other gonadal histopathological alterations, p,p′DDE altered gonadal development of 150 dph female zebraﬁsh, but failed to cause histopathological changes in males. Similarly, no alterations in male gonads were observed in summer ﬂounder injected with up to 60 mg kg–1 p,p′-DDE (Zaroogian et al., 2001), while thickening of the tubule wall and reduction of sperm were observed in p,p′-DDE-treated ﬁsh in a dose-dependent manner (Zhang and Hu, 2008). The 14-day exposure duration in the case of zebraﬁsh might be too short to cause alterations in male gonad development, that in fact contrast with the prolonged exposure of 2 months performed by Zhang and Hu (2008). The gonadal regression observed in p,p′-DDE exposed females was characterized histologically by the absence of mature vitellogenic and the predominance of pre-vitellogenic oocytes and atretic follicles. This response is consistent with estrogenic MOA registered for instance in zebraﬁsh exposed to 17α-ethinylestradiol, where vtg elevation in plasma was related to the similar histological alterations registered in this study, namely vitellogenic oocyte stages became atretic and only pre-vitellogenic stages remained (Van den Belt et al., 2002). However, considering the lower weights (not signiﬁcant) of almost all female treated ﬁsh groups compared to controls, a general toxic effect of DDE through impairment in ﬁsh growth might be impacting ovarian development and contributing to the observed histological alterations. Considering the overall results obtained in the present study, p,p′-DDE appeared to act as an estrogen, whereas its antiandrogenic potential was not demonstrated. The global differences and potency of responses observed in different ﬁsh species exposed to p,p′-DDE might be due not only to the exposure route (water/food-borne exposure/injection), organism life stage, dose and duration of exposure to p,p′-DDE, but also to the inter- and intraspecies afﬁnity of p,p′-DDE to hormone receptors, which might play an important role in this. One MOA of p,p′-DDE is mediated at the level of the AR acting similarly to the therapeutic antiandrogen ﬂutamide (Kelce et al., 1995; Kelce and Wilson, 1997). Several studies in male rats have demonstrated that p,p′-DDE prevents competitive binding to the AR, preventing the transcription of androgen-dependent genes, consequently causing abnormal sexual development and demasculinization (Kelce et al., 1995; Wolf et al., 1999). However, in ﬁsh, it is still unclear how p,p′-DDE interacts with ARs and cause reproductive abnormalities. The p,p′-DDE afﬁnity to AR is known to differ among ﬁsh species and even among target tissues within the same ﬁsh species. For instance, Wells and Van der Kraak (2000) have demonstrated p,p′-DDE afﬁnity to ARs in goldﬁsh (Carassius auratus) testes but not in its brain tissue nor in rainbow trout (Oncorhynchus mykiss) tissues. Moreover, the direct links between in vitro receptor binding and in vivo activity remain unclear. These differences in the hormone system together with the high diversity of reproduction strategies observed in teleosts may be the cause of the observed differences in the effects and antiandrogen potency of p,p′-DDE among different ﬁsh species.

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Endocrine disruption effects of p,p'-DDE on juvenile zebrafish. - PDF Download Free (2024)

References