Potential mechanisms of phthalate ester embryotoxicity in the abalone Haliotis diversicolor supertexta

Environmental Pollution 159 (2011) 1114e1122

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Potential mechanisms of phthalate ester embryotoxicity in the abalone Haliotis diversicolor supertexta Jin Zhou a, Zhong-Hua Cai a, b, *, Ke-Zhi Xing b a b

L-304, Life Sciences Division, Graduate School at Shenzhen, Tsinghua University, Shenzhen University Town, Xili, Shenzhen City 518055, PR China Key Laboratory of Aquatic-Ecology, Tianjin Agricultural University, Lishui Road 112, Tianjin 300384, PR China

Potential mechanisms of PAEs on abalone embryogenesis are osmoregulation disorder, oxidative damage and physiological dysfunction.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 30 June 2010 Received in revised form 7 January 2011 Accepted 8 February 2011

The effects and associated toxicological mechanisms of five phthalate esters (PAEs) on abalone embryonic development were investigated by exposing the embryos to a range of PAEs concentrations (0.05, 0.2, 2 and 10 mg/mL). The results showed that PAEs could significantly reduce embryo hatchability, increase developmental malformations, and suppress the metamorphosis of abalone larvae. The possible toxicological mechanisms of PAEs to abalone embryos included, affecting the NaþeKþ-pump and Ca2þeMg2þpump activities, altering the peroxidase (POD) level and the malondialdehyde (MDA) production, damaging the extraembryonic membranes structure, as well as disrupting endocrine-related genes (gpx, cyp3a, and 17b-hsd 12) expression properties. Taken together, this work showed that PAEs adversely affected the embryonic ontogeny of abalone. The abilities of PAEs affecting the osmoregulation, inducing oxidative stress, damaging embryo envelope structure, and causing physiological homeostasis disorder, are likely to be a part of the common mechanisms responsible for their embryonic toxicity. Ó 2011 Elsevier Ltd. All rights reserved.

Keywords: Phthalate esters Abalone Embryotoxicity Toxic mechanisms

1. Introduction Anthropogenic activities have resulted in a rapid increase in environmental pollution over the last few decades, and the effect of contaminants on the marine ecosystem is growing. Marine populations in some areas are declining at an alarming rate, but the causes of these declines are still not completely clear (Lesley and Clinton, 2005; Reynolds et al., 2005). In addition to over-fishing and habitat loss, other factors that contribute to population decline include introduction of exotic species, diseases, and climate change (Stuart et al., 2004; Raimonda et al., 2009). With progress in ecology research, environmental contamination has come to be known as a factor in biodiversity decline and population quality degeneration that cannot be ignored (Zhou et al., 2010). One of the important candidate contaminants responsible for biodiversity decline in the marine ecosystem is the xenobiotic substances, because of their adverse effect on growth, immunity, reproduction, and development (Segner et al., 2003; Johnson and Yund, 2007). Developmental toxicity induced by xenobiotic pollutants is one of the factors contributing to bio-resource decline, and may have particularly important effects on sensitive developmental stages. In

* Corresponding author. E-mail address: [email protected] (Z.-H. Cai). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.02.016

fish, teleost embryos reached the blastocyst stage at a lower frequency when exposed to bisphenol A (Shao et al., 2005). Estradiol can result in a decline in hatching rate, an increase in the abnormality rate, and even embryonic death in brown trout (Vethaak et al., 2002). Triphenyltin (TPT) can cause reproductive dysfunction and transgenerational toxicity in medaka at environmentally relevant levels (Zhang et al., 2008). In higher vertebrates, tributyltin (TBT) results in hatching success reduction and swimming abnormality in frog embryos (Hopkins et al., 2006). Embryonic development retardation by oil pollution in turtles was observed by Rowe et al. (2009). In addition to vertebrates, invertebrates also suffer from the adverse effect of xenobiotic pollutants. Invertebrates have gradually become a focus in the embryotoxicity assessment field because of their typical ecological representation, high sensitivity, and wide distribution. Increasingly, toxicity assessments based on invertebrate embryos such as bivalves (Bellas et al., 2005; Inoue et al., 2006; Xie et al., 2008), echinoderms (Roepke et al., 2005; Arslan et al., 2007) and crustaceans (Leblanc et al., 2000; Wang et al., 2009) have become available. These studies have demonstrated that xenobiotic contaminants cause multiple developmental toxicities in embryos, such as hatchability reduction, morphological malformation increase, histological damage exhibition, and survival rate decline. However, relatively little information is available on the mechanisms of embryotoxicity, even though these substances

J. Zhou et al. / Environmental Pollution 159 (2011) 1114e1122

profoundly influence multiple aspects of morphology, physiology, and behavior (Manzo et al., 2006). To better understand the effects of xenobiotic pollutants on embryos, investigations on their varied mechanisms, using suitable target models, are needed. Gastropods are a good candidate for ecotoxicity studies, and their potential as bioindicators of endocrine disruptors has been demonstrated in terrestrial environments (Hall et al., 2009). Among the gastropods, abalone (Haliotis diversicolor supertexta) is an important aquatic univalve mollusk, and a common ecological and economic marine species in many countries. Abalone acts as a slowly moving species that is easily cultured and handled in the laboratory. Moreover, abalone can produce eggs all year round and has a short generation time (Bryan and Qian, 1998). More importantly, the early developmental stages of abalone are highly sensitive to environmental stressors (Kirchman et al., 1981). These traits make this species suitable for in routine and regular monitoring. Phthalate esters (PAEs) are broadly applied in the plastic industry as plasticizer additives (Autian, 1973), and are representative environmental persistent pollutant. PAEs have wide distribution, a long environmental persistence, and a potential tendency for bioaccumulation, so they are regarded as environmental priority pollutants by the Environmental Protection Agency. In the field of aquatic environments, PAEs exist widely in water ecosystems and the reported concentration ranged from 0.02 to 1.8 mg/mL (Suggatt et al., 2003). The developmental dysfunctions caused by PAEs have been reported in various aquatic animals, including impairing spermatogenesis of frog (Lee and Veeramachaneni, 2005), inducing morphological malformations of zebrafish (He et al., 2010), and interfering with the normal ontogenetic process of the brine shrimp (Acey et al., 2002). Virtually, in the culturing industry, abalones are often cultured for long periods with plastic products. For example, polyethylene plastic sheets coated with diatoms are used as settling substrata for living larvae (Bryan and Qian, 1998), and juvenile and adult abalones are often cultured in plastic cages. Therefore, trace PAEs dissolving from plastic sheets or cages inevitably affect the early developmental stages, and the adult abalone. Previously, our laboratory reported the environmental behavior of PAEs in seawater, and found preliminary acute toxic effects on abalone embryos (Liu et al., 2009). However, the mechanisms by which trace plastic-associated contaminants (e.g., PAEs) affect abalone ontogeny are still unknown. Here, we further investigate PAE-induced developmental toxicity and explore the underlying embryotoxic mechanisms.

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2.3. Exposure protocol Fertilized eggs were exposed in porcelain beakers containing 500 mL FSW. Approximately 5  103 embryos were exposed to five different PAEs (DBP, DEP, DMP, DOP, and DEHP) at four concentrations (0.05, 0.2, 2 and 10 mg/mL). The tested doses in this study referenced from our earlier work (Liu, 2010). The same volume of acetone solvent was used as a control and seawater was used as the blank. Three replicates were conducted of all treatments and controls. Fig. 1 shows the experimental flowchart of embryo exposure to PAEs. The appearance of the first polar body marked fertilization success. A multi-cell stage was adopted to determine cleavage block. The blastula stage (about 3 h postfertilization, 3 hpf) was employed as the beginning for detecting abnormal rate, and the monotrochal larva stage (8 hpf, ready to hatch) was employed as the endpoint for detecting abnormalities. At 8 hpf, samples from each treatment group were collected for analyzing physiological parameters, embryo envelope surface structure and reference gene expression. The percentage of hatching rate was calculated at the hatching-out stage (8e9 hpf). After the hatching-out stage, 200 veliger larvae were transferred into 500 mL beakers containing the same test solution at about 25  C, for metamorphosis rate detection. Three thin slides (70  20 mm) coated with diatom films were place in each beaker and installed vertically for larvae settlement. Slides with diatom films were prepared by placing them in an aeration tank containing FSW and diatoms for 24 h. Metamorphosis success was judged according to Vicose et al. (2007), and the metamorphosis rate was recorded after an additional 110 h of incubation. Therefore, the total exposure time for the completion of larval metamorphosis was approximately 120 h. 2.4. Biological endpoints 2.4.1. Developmental parameters 2.4.1.1. Abnormality rate (AR, %). Embryo samples were collected randomly and observed under a dissecting microscope (Olympus, Japan). Each treatment was recorded at least three times per duplication. Any individual with yolk leakage, yolk sac edema, reduction of body pigmentation, mulberry cells bulging into irregular shapes, or other morphological defects, was counted as abnormal. Embryos that developed to a ciliated monotrochal larva stage with no further development were also considered abnormal. The abnormality rate was calculated as: abnormality rate (%) ¼ (number of malformed embryos)/(total counted embryos)  100%. 2.4.1.2. Hatching rate (HR, %). After the monotrochal larva stage (about 9 hpf), embryos hatch and turn into swimming veliger larvae. Embryos that did not hatch were recorded as hatching failures. For each treatment, 300 fertilization eggs (here it refers to all the embryos in development) (100  3 replicates) were chosen for measuring the hatching rate, calculated as hatching rate (%) ¼ (total counted embryos  number of hatching unsuccessful embryos)/(total counted embryos)  100%. 2.4.1.3. Metamorphosis rate (MR, %). Two hundred veliger larvae per replicate were transferred into 500 mL beakers containing the same test solution and slides coated with diatom biofilms, and cultured another 110 h until larvae were fully metamorphosized. The percentages of fully metamorphosized larvae (metamorphosis rate) were recorded after detection with a microscope (Olympus, Japan). The metamorphosis rate was calculated as: metamorphosis rate (%) ¼ (successfully metamorphosized individuals)/(two hundred veliger larvae)  100%.

2. Materials and methods 2.1. Chemical reagents and materials Five PAEs (purity  98%), dibutyl phthalate (DBP), diethyl phthalate (DEP), dimethyl phthalate (DMP), dioctyl phthalate (DOP) and di-(2-ethylhexyl) phthalate (DEHP) were purchased from Shanghai Chemical Reagent Co., Ltd. (Shanghai, China). Individual stock solutions were prepared in acetone at 2  104 mg/mL, and then serially diluted with sand-filtered seawater (FSW) to make test solutions. Molecular biology reagents used in this study were from SigmaeAldrich China (Beijing, China) and Takara Co., Ltd (Dalian, China).

2.2. Animals and fertilization Mature female and male abalones were collected from a local hatchery (Shenzhen, China). After cleaning with FSW, gonad-outstanding individuals were selected. Females and males were kept in separate rectangular tanks containing aerated FSW. Culture conditions were: water temperature 24  2  C, salinity 30  2&, pH 8.0  0.2, and dissolved oxygen no less than 6.0 mg/L. Abalones (20 males; 50 females) were induced to spawn in ultraviolet-radiated seawater (500 mW/h) (Uki and Kikuchi, 1982). The eggs and sperm of a mass spawning were mixed for about 10 min for in vitro fertilization, and then washed three times with FSW to remove excessive sperm. Fertilized eggs (embryos) were used for experiments.

2.4.2. Biochemical assays At the monotrochal larva stage, embryos were sampled from the treated and control groups. Each sample was homogenated and centrifuged at 12,000g for 30 min at 4  C, and the supernatant used to assay for biochemical parameters. All parameters were determined using a commercially available kit from Nanjing Jiancheng Bioengineering Institute (Nanjing, China). The NaþeKþ-ATPase (Naþ pump) and Ca2þeMg2þ-ATPase (Ca2þ pump) activities were determined by quantifying the release of inorganic phosphorus from adenosine triphosphate. Specific activity was expressed as the concentration of adenosine diphosphate liberated per unit time, and standardized to protein content (i.e. units/mg protein). Peroxidase (POD) activity was measured based on the kit instructions (Nanjing, China), for 20 mL sample and 180 mL color-developing buffer (7.3 g C6H8O7$H2O, 11.86 g Na2HPO4$2H2O, 1 L H2O), in 96-well microtiter plates. OD490 (A1) was read at 490 nm, and the 96-well microtiter plate was removed, and 20 mL color-developing reagents (4 mg C6H8N2, 4 ml 30% H2O2, 10 ml color-developing buffer) added. Plates were shaken five times in a microplate spectrophotometer, and color developed for 15 min in dark, before OD490 (A2) was read. Relative POD activity was expressed as A2eA1. Specific activity was in mmol/mg protein. The level of lipid peroxidation was determined by the malondialdehyde (MDA) content in abalone embryos. The thiobarbituric acid reaction method was used to determine MDA, by measuring at 532 nm after reacting with thiobarbituric acid to form a stable pink chromophoric production (Devasagayam, 1986). MDA content was expressed as nanomoles per milligram of protein (nmol/mg protein).

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Fig. 1. The experimental flowchart and timeline for abalone embryos exposure to PAEs. The detailed descriptions are in “Materials and Methods” section.

2.4.3. Scanning electron microscope (SEM) examination Embryo samples at monotrochal larva phase from control group and tested groups were fixed with 5% glutaraldehyde in 100 mM phosphate buffer (pH 7.4, 4  C) for 24 h. They were post fixed with 1.5 osmium tetraoxide for 2 h and washed three times with 100 mM phosphate buffer (pH 7.4). After slowly dehydrating with an ethanol series, the embryos were dried at 30e40  C, glued to stubs coated with 20 nm of gold and viewed with scanning electron microscope (Tescan VEGA TS 5136XM, USA) at 15 kV. 2.4.4. RNA extraction and cDNA synthesis At the monotrochal larva stage, approximately 100 mg in 2 mg/mL of the indicated PAE treatment was collected for gene expression analysis. Total RNA was extracted using Trizol Reagent (Invitrogen, USA) according to the manufacturer’s instruction. Total RNA was purified by an mRNA purification column (BBI, Canada), and cDNA synthesized using a cDNA Synthesis Kit following the user manual (TaKaRa, Dalian, China). cDNA was stored at 20  C for real-time quantitative PCR (qRT-PCR) analysis. 2.4.5. Real-time quantitative PCR Three reference genes, glutathione peroxidase (gpx), cytochrome P450 3A (cyp3a), and 17b hydroxysteroid dehydrogenase 12 (17b-hsd 12) were selected to assess the effect of PAEs on the development of abalone embryos. The b-actin gene was used as an internal reference. Specific primers were designed (Table 1) based on the cDNA sequence of the reference gene, using primer premier 5 software. The qRTPCR amplifications were in 20 mL, containing 10 mL 2 SYBR premix ExTaq (Takara, Dalian, China), 5 mL cDNA, 0.2 mL each of 20 mM forward and reverse primers, 0.4 mL

ROX, and 4.4 mL DEPC treated water, in an ABI 7300 real-time PCR machine (Applied Biosystems, USA) with a program of 95  C for 10 s, followed by 40 cycles of 95  C for 5 s, 60  C for 30 s. Amplification products were subjected to melting curve analysis to confirm that a single PCR product had been amplified and detected. Each sample was analyzed three times to confirm reproducibility. b-actin transcript levels were used to normalize the results, and negative control reactions in which template cDNA was omitted were included for every primer set. Relative amounts of target genes and b-actin-RNA in all cDNAs were calculated from a standard curve using a 7300 system SDS software v1.3.0 (Applied Biosystems) to estimate transcript copy numbers for each sample. To maintain consistency, the baseline was determined automatically by the software. Target mRNA levels were expressed in relative levels (target mRNA/b-actin-RNA).

Table 1 Primer sequences for real-time quantitative RT-PCR analyses. Target gene

GenBank accession numbers

Primer sequence (50 e30 )

Product (bp)

cyp3a (Fr) cyp3a (Rr) gpx (Fr) gpx (Rr) 17b-hsd 12 (Fr) 17b-hsd 12 (Rr)

GU984784

TCGAATGGCTAAGAACGACA CAACACCAAACGGAGCAAAA CGCTAACTACTGAGGCAAGA ACCTCCATGATTTGGGTATG GCTTGGGCTGTCAGTAAACG TTGGCAAGATCATCCAGTTT

191

GU984785 GU984783

213 159

J. Zhou et al. / Environmental Pollution 159 (2011) 1114e1122 2.5. Statistical analysis Data on development parameters and physiological features are expressed as mean  standard deviation (SD) and were analyzed by one-way analysis of variance (ANOVA) and Tukey’s multiple comparisons using SPSS 11.0 software. The qRT-PCR data analysis was executed using SDS software v1.3.0 (ABI, USA) according to the 2DDCt method (Livak and Schmittgen, 2001). A student t-test was used to determine whether gene expression differed significantly with PAE concentration. Values are presented as mean  SD. P < 0.05 was considered statistically significant.

3. Results 3.1. Effect of PAEs on developmental parameters 3.1.1. Abnormality rate (AR, %) The main developmental stages of abalone embryos under normal conditions (blank and control groups) are shown in Fig. 2A (1e6). After PAEs exposure, several kinds of morphological

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malformations were observed, such as yolk leakage, yolk sac edema, pigmentation change, morula cells bulging into irregular shapes, and developmental retardation (Fig. 2A, 7e12). AR statistics are in Fig. 2B. The AR showed no obvious elevation in the lowest dose-treated group (0.05 mg/mL) at the end of the experiment. However, the AR of PAE-treated embryos significantly increased at higher doses (2 mg/mL), with the exception of DEHP. For example, for DBP, the AR increased by 2.33-fold at 2 mg/mL, and by 1.92-fold at 10 mg/mL (P < 0.05). Of the different PAEs, the most dramatic effect was observed with exposure to DBP and DOP, which caused significant increases in the AR at 2 mg/mL. In contrast to the effect of DBP and DOP, DMP caused an obvious difference of AR only at 10 mg/mL. DEHP did not significantly affect AR. 3.1.2. Hatching rate (HR, %) HR was greater than 90% in the blank and control groups (Fig. 2C), but trended downward when embryos were exposed to

Fig. 2. Effect of PAEs on developmental parameters of abalone embryos. Data are represented as mean  SD (n ¼ 3). Asterisk denotes statistically significant differences between controls and treatment groups by one-way ANOVA. (A), developmental process of abalone embryo (mean size was 150  10 mm). 1e6 denotes the normal embryo at various stages. During PAEs environment (taking 2 mg/mL DBP as an example), several gross morphological malformations were observed. 7e12 point out the abnormal embryos, such as egg-yolk leakage (7), yolk sac edema (8), pigmentation change (9), morula cells bulging into irregular shapes (10), turgid vesicles in perivitelline space (11) and developmental retardation (12). (B)e(D) represent abnormal rates (%), hatching success rate (%) and metamorphosis rate (%), respectively.

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B

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3.2. Effect of PAEs on physiological traits of abalone embryos

60

Abnormal rate (%)

50

Seawater

Acetone

DBP

DEHP

DMP

DOP

DEP

40 30 20 10 0

Hatching success rate (%)

C

Control

110 100

0.05

Seawater DEHP

0.2 Concentration (µ g/mL)

Acetone DMP

2

DBP DOP

10

DEP

90 80 70 60 50 Control

D

21

Metamorphosis rate (%)

18

0.05

0.2 2 Concentration (µ g/mL)

S eawater

Acetone

DBP

DEHP

DM P

DO P

10

DEP

15

The activities of the Ca2þeMg2þ-ATPase (Ca2þ pump) and the NaþeKþ-ATPase (Naþ pump) were examined, because they are involved in membrane integrity and osmotic balance. The results showed that the activities of the ATPases did not change significantly between the control and the low-dose PAEs-treated groups, with the exception of DOP (Fig. 3A and B). The levels of the Ca2þ pump and the Naþ pump significantly increased when embryos were treated with higher doses of PAEs. Enzyme activities exhibited dose dependence. The enzyme activity levels of the embryos showed a large elevation (P < 0.05) after exposure to PAEs at the highest concentration, with an elevation range of 29.3e110.2%. Under PAEs exposure, the POD contents were essentially the same as the controls at the low-dose PAE levels of 0.05 and 0.2 mg/mL. However, POD levels increased (1.43e2.01-fold) with dose elevation (P < 0.05) (Fig. 3C). These results showed that the POD induced by PAEs was characterized by low-dose maintenance and high-dose stimulation, and that oxidative stress in embryos occurred after PAE exposure. From the MDA results shown in Fig. 3D, we can see that various PAEs have different effects on MDA value. The basic trend was similar to the POD results, although significant changes were not found at the two lower concentrations of 0.05 and 0.2 mg/mL. When the embryos were incubated with 2 or 10 mg/mL, relative to controls, the intracellular MDA increased 1.58-fold with DMP (2 mg/mL), 2.31-fold with DBP (10 mg/mL), 1.92-fold with DMP (10 mg/mL), and 1.86-fold with DOP (10 mg/mL). The results further indicated a disturbance in antioxidative balance after PAE exposure. 3.3. Effects of PAEs on embryo surface structure

12 9 6 3 0

C ontrol

0.05

0.2

2

10

C oncentration (µ g/mL) Fig. 2. (continued).

PAEs, although no obvious changes were observed at the lower concentrations of 0.05 and 0.2 mg/mL. Interestingly, embryos treated with PAEs from 2 to 10 mg/mL showed an HR reduction from 23.5 to 36.7% (P < 0.05). The lowest HR value, having been reduced to 60% of the control level, appeared in the 10 mg/mL DBP group. DBP and DEP displayed higher toxicity on HR, while DOP displayed second toxicity, followed in effect by DMP and DEHP. 3.1.3. Metamorphosis rate (MR, %) Metamorphosis is the endpoint for larvae development. The mean MR of abalone larvae is shown in Fig. 2D. Although the metamorphosis index did not significantly change in the lowest dose-treatment group (0.05 mg/mL) compared to the control groups, the MR showed an obvious decrease with dose increase, reflecting a relationship with dosage. At 2 mg/mL, the MR in the DBP group was reduced by 3.1-fold and the DEP group was reduced by 2.9-fold compared to the controls. The most striking differences in MR appeared at 10 mg/mL group, which caused a dramatic enhancement of toxicity, reducing the MR to 20% of the controls. Also, in the highest dose-treatment group, the PAEs induced retardation in swimming behavior, and a decrease in metamorphosis, eventually inducing larval death (data not shown).

The effects of PAEs on embryos envelope structure were showed in Fig. 4. In this study, we took DBP as an example. In the control groups, the surface of abalone embryos was relatively smooth under scanning electron microscopy (SEM) (Fig. 4A). However, in DBP environment (0.05 mg/mL), the microvilli were observed on the external layer of abalone embryos (Fig. 4B). An occasional filopodium were found on the surface of these DBP-treated embryos (Fig. 4C) (0.2 mg/mL). The most obvious change was the microporous structure, which appeared on the embryonic membrane in the higher concentration of DBP (2 and 10 mg/mL), and the scale and number displayed a dose-dependent pattern (Fig. 4D and E). 3.4. qRT-PCR The mRNA expression patterns of three physiological-related genes (cyp3a, gpx and 17b-hsd 12) were examined in this study. We took 2 mg/mL PAEs as an example. As shown in Fig. 5, expression of cyp3a and gpx was up-regulated, while expression of 17b-hsd 12 was down-regulated. Among the three genes, the most sensitive one was cyp3a, whose expression increased from 2.1- to 6.3-fold after treatment, although the differences were not statistically significant for DEHP. A similar tendency was found in gpx expression, but the increase was not as large as for cyp3a. The 17b-hsd 12 was down-regulated in all treatments, and DBP and DMP exposure resulted in the lowest expression of it (52.6e63.7% level of the controls) (P < 0.05). 4. Discussion Recently, the biological effect of persistent pollutant has been recognized as more serious for early developmental events in the embryo or larva, than in the adult, because embryos and larvae are more sensitive and fragile (Gopalakrishnan et al., 2007; Libralato

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Fig. 3. Effect of PAEs on physiological status of abalone embryos. (A) Ca2þeMg2þ-ATPase activity, (B) NaþeKþ-ATPase activity, (C) POD level and (D) MDA production. Each bar represents the mean  SD (n ¼ 3). Significant differences (P < 0.05) of biochemical parameters between the trial and the control groups are indicated with asterisks.

et al., 2007). In this work, abalone embryos exposed to PAEs showed an increase in the frequency of malformations, a decrease in hatching success rate, and disturbances in metamorphosis behavior. Developmental abnormality and hatching success rate are common indicators of the effects of contaminants on embryogenesis. Several embryotoxicity studies have demonstrated that xenoestrogens result in morphological malformation in aquatic organisms, including sturgeon treated with TPT (Hu et al., 2009), sheephead minnow treated with 17b-estradiol II (Raimonda et al., 2009), sea urchins exposed to octylphenol (Arslan et al., 2007), and clams exposed to TBTO (Inoue et al., 2006). Similarly, in field conditions, Klumpp et al. (2002) documented a higher rate of embryonic malformation in fish in polluted coastal waters in Xiamen, China. Previous studies demonstrated that DAP potentially induces developmental dysfunction in Hyalella azteca, and DEHP impair hatchability in Daphnia magna (Adams et al., 1995). A similar result was obtained in this study, in which PAEs interfered with abalone embryonic development, causing malformations that included yolk

leakage, yolk sac edema, reduction of body pigmentation, and developmental retardation (Fig. 2A, 7e12). Attachment and metamorphosis behaviors occur widely in the embryonic developmental process of gastropods such as snails and abalones (Jackson et al., 2002; Gapasin and Polohan, 2004). Metamorphosis is the endpoint of larvae development and a crucial stage for evaluating the adverse effects of pollutants on gastropods, because it reflects both physiological and developmental status, including differentiation, settlement, locomotor ability, and feeding behavior (Li et al., 2006). Previously, in laboratory field, malformed monotrochal larvae were found to hatch and swim in seawater, and even settle on slides, but could not finish metamorphosis (Liu et al., 2009). In this study, we further demonstrated that metamorphosis significantly decreased after PAE exposure, in a dose-dependent pattern. In wild field, Hodin (2006) pointed out that the life-cycle transition from settlement to metamorphosis might be especially sensitive to contaminant exposure during developmental periods; and, he deemed that there was a risk of losing attachment capability

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B

5

Seawater

Acetone

DBP

4

DEHP

DMP

DOP

Sodium/Potassium ATPase activit (U/mg pro)

Calcium/Magnesium-ATPas activity (U/mg Pro)

A DEP

3 2 1 0

Control

0.05 0.2 2 Concentration (µg/mL)

10

90 80 70 60 50 40 30 20 10 0

Seawater DEP DOP

Control

Acetone DEHP

DBP DMP

0.05 0.2 2 Concentration (µg/mL)

10

D

Seawater DEP DOP

Control

Acetone DEHP

0.05

MDA production (nmol/mg pro)

Total POD activity (umol/mg Pro)

C

4.5 4 3.5 3 2.5 2 1.5 1 0.5 0

DBP DMP

0.2 2 Concentration (µg/mL)

1800 1500

Seawater DEP DOP

Acetone DEHP

DBP DMP

1200 900 600 300 0

10

Control

0.05 0.2 2 Concentration (µg/mL)

10

Fig. 4. Effects of PAEs (taking DBP as an example) on envelope surfaces of abalone embryos. (A)e(E) indicated the SEM images of 0 (control), 0.05, 0.2, 2 and 10 mg/mL DBP, respectively (AeD: 3,000; E: 3,500). The solid arrows denote the microvillus or filopodium, the dotted arrows denote the microporous in the surface of the embryo envelope.

et al., 2005). In this study, the main types of embryo malformation were yolk sac edema, irregular shapes and pigmentation changes. The morphological abnormality of yolk sac indicated that PAEs could destroy the embryonic membrane, and disturb the osmotic balance. Indeed, in this work, the surface microstructure of embryo envelope was changed (Fig. 4), which suggested that PAEs affected the membranous integrity or permeability. We can speculate that the membrane damage may supply a gateway for PAEs to

of juveniles when they were exposed to xenobiotic pollutants. Similar to the action of organotin on echinoderms (Hadfield and Paul, 2001), PAEs may act as anti-metamorphic agents on abalone larvae. Metamorphosis behavior is an integrative evaluation indicator, and the embryo development toxicity of PAEs was ranked as DBP > DEP > DMP > DOP > DEHP by metamorphosis rate (%). Osmoregulation and membrane permeability are important issues to understand the mechanisms of embryotoxicity (Wood

Relative expression of mRNA of various genes

13

Acetone

Seawater

DBP

DEP

DEHP

DMP

DOP

12 Target gene

11

β-actin

10 9 8 7 6 5 4 3 2 1 0

cyp3a

gpx

17b-hsd 12

RT-PCR Fig. 5. Effect of PAEs on reference genes expression in abalone embryos. Each bar represents the mean  SD (n ¼ 3). All real-time reacting were performed three times. Significant differences (P < 0.05) of target genes expression between the experimental and control groups are indicated with asterisks. Amplification of b-actin served as a control for a constitutively expressed gene. The insect pictures are corresponding agarose gel maps, up are target genes, and down are b-actin.

J. Zhou et al. / Environmental Pollution 159 (2011) 1114e1122

penetrate into fertilized eggs and form PAE-contaminated embryos, and ultimately bring about developmental malfunction. In addition, the activities of the Ca2þ and Naþ pumps were significantly enhanced under high-dose PAE treatment, and appeared to positively correlate with the morphological results. These physiological responses indicated that PAEs induced embryonic membrane ionic channel changes in abalone. Previously, Girard et al. (1996) demonstrated that Ca2þ homeostasis was a critical parameter for embryotoxicity in marine invertebrates. Roepke et al. (2005) also reported that the potential alteration of calcium signaling by TBT might be an important cause of developmental toxicity in sea urchin embryos and larvae. Osman et al. (2007) further pointed out that heavy metal (lead) result in metabolic and osmotic disturbances and caused subsequently yolk sac edema. In the present study, the activity changes of the Ca2þ and Naþ pumps induced by PAEs indicated alterations in ion regulation and osmoregulation. From the above results, it can be concluded that PAEs affected embryo envelop structure and disrupted its ion homeostasis, and thereby influenced normal morphogenesis or hatching behavior. Other minor deformities were also observed in this study, such as the reduction pigmentation in embryos. Previously, reduction in pigmentation was observed in Clarias gariepinus and Danio rerio when they were exposed to lead and clozapine, respectively (Nguyen and Janssen, 2002; Akande et al., 2010). Pigmentation of the tissue is controlled by melanocyte stimulating hormone (aMSH) and melanin-concentrating hormone (MCH), which are known to be stress regulated (Osman et al., 2007). In this work, the weak pigmentation was exhibited in PAEs-treated abalone embryos. We speculated that reduction in pigmentation of abalone embryos may either result from PAEs stress at the cellular, organ or individual level or alter the function of aMSH and MCH directly. Whereas, identification of aMSH and MCH in abalone and clarification of their relations with pigmentation need to be further explored. Oxidant stress would be another crucial parameter to elucidate the possible mode of action when organisms are in stress. In early embryonic stages, anti-oxidant stress is the main protecting mechanism of organism to eliminate various hazards by triggering redox reactions to generate free radicals (e.g., ROS) and induce physiological alterations (Lister et al., 2010). Escoffier et al. (2007) reported that okadaic acid-induced oxidative damage disrupted normal physiological status, and resulted in developmental malfunction of medaka. Hopkins et al. (2006) found that oxidative stress caused teratogenesis and swim-up failure in frog larvae, and was more damaging than in adults. Kalaimani et al. (2009) further demonstrated that anti-oxidant status was extremely important to the embryonic development process, as an indicator of successful ontogeny. In this work, both POD and MDA significantly increased under high doses environment, suggesting that PAEs triggered oxidative stress. As oxidative stress is the first response to environment stressors (Livingstone, 2001), we think that the embryo initially initiates anti-oxidant and detoxification responses to counteract a PAE attack, but when the anti-oxidant and detoxification systems break down, PAEs may cause other toxicologically relevant responses in abalone embryos, such as respiratory impairment, ion balance disruption, and physiological regulation disorder. In combination with the results in this study, it can be concluded that the oxidative damage maybe one of the primary factors to lead to teratogenesis or loss of developmental potential. In addition to physiological mechanisms, gene expression analysis by qRT-PCR probably provided us some new insights for understanding the mechanism of embryotoxicity at molecular level. Two enzymes encoded by gpx and cyp3a are key anti-oxidant and detoxification protein (Buttke and Sandstrom, 1994; Plant, 2007). We found that gpx was induced by PAEs, in agreement

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with the results of antioxidative enzyme activity assays (Fig. 3C and D). At the same time, the clear response of cyp3a to PAEs was also observed, which indicated that PAEs exposure triggered the detoxification defense. These results further demonstrated that PAEs disturbed the antioxidative balance of embryo individual. The abnormal expression of gpx and cyp3a is possible molecular signal for teratogenicity of abalone embryos. Like the gpx and cyp3a, the 17b-hsd 12 (17b hydroxysteroid dehydrogenase 12) has also obvious response to PAEs exposure. The 17b-hsd 12 participates in steroid biosynthesis, which is important for development, growth and reproduction (Mindnich et al., 2004). Previous research demonstrated that 17b-hsd 12 was a key regulator in early embryo development, because it enabled the embryos to maintain steroid homeostasis, which provides a convenient pathway for cellecell communication, and therefore functions to regulate cell differentiation, organogenesis, and embryonic ontogenesis (Khan et al., 1997). In our study, the down-regulation of 17b-hsd 12 was exhibited, indicating that PAEs could affect steroid biosynthesis or metabolic pathways during abalone embryogenesis. Considering the important role of steroid in mediating environmental effects on developmental plasticity, we conclude that endocrine regulation disruption provides a potential explanation for the alteration of developmental parameters in abalone embryos after PAE exposure. 5. Conclusions This study demonstrated that PAEs adversely affected abalone embryonic development, including increasing the developmental malformation rate, decreasing the hatching success rate, and inhibiting the normal metamorphosis behavior of larvae. The potential toxicological mechanisms of PAEs on abalone embryos may include membrane dysfunction (extraembryonic membrane lesion, ion homeostasis disruption and osmoregulatory alteration), oxidative damage, and physiological equilibrium disorder. The five PAE compounds tested share similar modes of action on abalone embryos, and the general toxicity order was calculated as DBP > DEP > DMP z DOP > DEHP. Additionally, the results of this study imply that the abalone deserves more attention in the future, and can be regarded as a suitable bioindicator for embryonic bioassay exposure to plastic-derived pollutants. Acknowledgments We would like to express our heartfelt thanks to Mr. Zheng-Ping Cai for his help in collecting the embryo samples and assaying developmental parameters. This study was supported by PhD Startup Fund of Natural Science Foundation of Guangdong Province, China (10451805702004177), and International Cooperative Project of SZSITIC Commission of Shenzhen (ZYA200903260036A). References Acey, R.A., Bailey, S., Healy, P., Jo, C., Unger, T.F., Hudson, R.A., 2002. A butyrylcholinesterase in the early development of the brine shrimp (Artemia salina) larvae: a target for phthalate ester embryotoxicity? Biochemical and Biophysical Research Communications 299, 659e662. Adams, W.J., Biddinger, G.R., Robillard, K.A., 1995. A summary of the acute toxicity of 14 phthalate esters to representative aquatic organisms. Environmental Toxicology & Chemistry 14, 1569e1574. Akande, M.G., Orn, S., Norrgren, L., 2010. Evaluation of the toxic effects of clozapine in zebra fish (Danio rerio) embryos with the fish embryo toxicity test. International Journal of Pharmaceutical and Biomedical Research 1, 90e94. Arslan, O.C., Parlak, H., Katalay, O.S., 2007. The effects of nonylphenol and octylphenol on embryonic development of sea urchin (Paracentrotus lividus). Archives of Environmental Contamination and Toxicology 53, 214e219. Autian, J., 1973. Toxicity and health threats of phthalate esters: review of the literature. Environmental Health Perspectives 4, 3e26.

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